EPA/600/A-95/112
Intrinsic Bioattenuation for Subsurface Restoration
Hanadi S. Rifai, Robert C. Borden, John T. Wilson, and Calvin H. Ward
ABSTRACT
Intrinsic bioattenuation has recently evolved as a viable remediation
alternative at a number of sites where the risk of exposure to contaminants is
within acceptable standards. Important mechanisms controlling the instrinsic
bioattenuation include advection, dispersion, sorption, dissolution from a residual
source, and abiotic and biological transformations. Since intrinsic bioattenuation is
a plume management strategy, it requires characterizing and monitoring these
processes. Intrinsic bioattenuation involves an assessment of risks to public health
and the environment, and consequently requires prediction of the fate and transport
of contaminants at the candidate sites. This paper reviews the processes controlling
intrinsic bioremediation and summarizes case histories where intrinsic
bioattenuation has been observed at sites contaminated with petroleum
hydrocarbons and chlorinated solvents. The key steps in evaluating natural
attenuation as a remedial alternative are summarized.
introduction
In the absence of human intervention, many contaminant plumes will
develop until they reach a quasi-steady-state condition. At steady-state, the
contaminant plume is no longer growing in extent and may shrink somewhat over
time. Major processes controlling the size of the steady-state plume include (1)
release of dissolved contaminants from the source area, (2) downgradient transport
of the contaminants and mixing with uncontaminated groundwater, (3)
volatilization, and (4) abiotic and biologically mediated transformations of the
contaminants of concern.
Intrinsic bioattenuation is a plume management strategy where the natural
assimilation processes are monitored and used to limit adverse impacts of
groundwater contamination. This strategy also requires an assessment of risks to
public health and other environmental receptors. A successful implementation of
intrinsic bioattenuation at a field site requires adequate site hydrogeological,
chemical, and biological characterization; detailed data analysis to determine
whether contaminants are being attenuated and/or removed from the aquifer;
modeling of the fate and transport of the dissolved groundwater plume; and, finally,
long-term monitoring to confirm and ensure protection of human health and the
environment.
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PROCESSES CONTROLLING THE STEADY-STATE CONTAMINANT
DISTRIBUTION
Physical
The primary physical processes affecting the distribution of contaminants in
groundwater include advection, dispersion, sorption, volatilization, and dissolution
from residual contaminants located in the source area. Advection is the process by
which contaminants are transported with the flow of groundwater. Dispersion
accounts for mechanical and molecular mixing processes. Both advection and
dispersion reduce contaminant concentrations but do not cause a net loss of mass of
contaminants in the aquifer. Higher advection and dispersion cause more
spreading and more dilution of a dissolved contaminant plume.
Sorption describes the partitioning of contaminants between the aqueous
phase and the solid aquifer matrix. Sorptive processes also tend to reduce the
dissolved contaminant concentrations and limit the migration of the aqueous-phase
plume, but they do not result in a loss of contaminant mass from the aquifer.
Under steady-state conditions, sorption will not affect the final contaminant
distribution; however, sorption will delay the development of a steady-state plume.
Volatilization refers to the partitioning of a contaminant between the
aqueous phase in the saturated zone and the vapor phase in the unsaturated zone.
While volatilization actually removes mass from the aquifer, it is not thought to be
a significant attenuation mechanism except in situations where the groundwater
table is less than 15 ft deep and the unsaturated zone consists of relatively
transmissive soils. Chiang et al. (1989) estimated that volatilization resulted in a
mass loss of benzene of less than 5% at a gas plant facility in Michigan.
Dissolution from residual contaminants located in the source area is by far
the most significant physical process that controls the extent of a contaminant
plume. In the case of petroleum hydrocarbons, contaminants may dissolve from a
lens of mobile hydrocarbons floating on the water table or from residual
hydrocarbons trapped in the soil matrix above and/or below the water table.
Seasonal fluctuations in the water table can cause additional "smearing" and
dissolution from residual source areas. Until sources are depleted, a contaminant
plume will expand until it reaches a quasi-steady-state.
Abiotic and Biologically Mediated Transformations
Aerobic and anaerobic biodegradation processes are believed to account for
both contaminant concentration reduction and loss of pollutant mass from the
aquifer. Abiotic transformations, such as hydrolysis and dehydrohalogenation, also
attenuate concentrations and contaminant mass in an aquifer; but are only
significant for specific chemicals such as chlorinated solvents.
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Aerobic biodegradation relies on dissolved oxygen as the electron acceptor
used by the microorganisms. Petroleum hydrocarbons are generally very amenable
to aerobic biodegradation in aquifers with dissolved oxygen concentrations
exceeding 1 to 2 mg/L. Many shallow water table aquifers will contain background
dissolved oxygen concentrations between 1 to 12 mg/L depending on the
temperature of the groundwater. While aerobic biodegradation takes place at
relatively higher rates than anaerobic processes, it is often limited by the available
supply of oxygen to a contaminant plume. Once the background dissolved oxygen is
consumed in the center of the plume, aerobic biodegradation is limited to the edges
of the contaminant plume where the dissolved contaminants come into contact
with oxygen-rich groundwaters.
Anaerobic processes refer to a variety of biodegradation mechanisms that use
NO3 SO4 Fe+3, and CO2 as terminal electron acceptors. Anaerobic
biodegradation dominates the interior of a contaminant plume. Both petroleum
hydrocarbons and chlorinated solvents are believed to biodegrade to varying degrees
under anaerobic conditions. However, the rates of biodegradation are often slower
than under aerobic conditions.
Hydrocarbon Biodegradation. Most organic compounds found in crude,
refined oil and fuels are known to degrade under aerobic conditions. The aerobic
biodegradation of benzene, toluene, ethylbenzene, xylene, napthalene, methyl-
napthalenes, dibenzofuran, and fluorene has been confirmed by a large number of
laboratory and field studies. Aerobic processes are relatively fast and limited by the
rate at which oxygen is supplied to a contaminant plume. Because of this
phenomenon, Rifai et al. (1988) modeled aerobic biodegradation as an instantaneous
reaction between oxygen and the hydrocarbons.
Current research efforts have also shown that monoaromatic compounds
degrade under anaerobic conditions. This biodegradation occurs with NO3 " (Evans
et al. 1991), Fe^+ (Lovley and Lonergan 1990), SO4 (Edwards et al. 1991), and
carbon dioxide (Grbic-Galic and Vogel 1987; Wilson et al. 1986) as electron acceptors.
Benzene has been found to be recalcitrant to anaerobic biodegradation in
laboratory studies using nitrate and sulfate as electron acceptors (Kuhn et al. 1989;
Edwards et al. 1991). However, some laboratory and field studies demonstrated the
degradation of all monoaromatic hydrocarbons under denitrifying, sulfate reducing
and methanogenic conditions (Major et al. 1988; Cozzarelli et al. 1990; Vogel and
Grbic-Galic 1986; Barker and Wilson 1992; Wilson et al. 1994a). The aromatic
compounds may be oxidized first to phenols or organic acids, then transformed to
volatile fatty acids before complete mineralization. Anaerobic biodegradation of
aromatic hydrocarbons is therefore associated with the production of fatty acids,
methane, carbon dioxide, solubilization of iron, and reduction of nitrate and sulfate.
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Adaptation of an aromatic plume to nitrate as a terminal electron acceptor
seems to occur readily once oxygen is depleted. Natural biodegradation through
nitrate respiration should be similar to oxidative biodegradation. Depending on its
concentration, sulfate can also be an important electron acceptor. Acton and Barker
(1992) demonstrated the sulfate reduction of toluene and m-xylene in a forced
gradient injection experiment at an active landfill in Ontario, Canada. Benzene, q-
xylene, ethylbenzene, and 1, 2,4- trimethylbenzene were not degraded at the site.
CAH Transformation. Chlorinated solvents, consisting primarily of
chlorinated aliphatic hydrocarbons (CAHs), can be transformed by chemical and
biological processes to form a variety of other CAHs (McCarty, 1994). Reduction of
tetrachloroethene (PCE) and trichloroethene (TCE) to ethene has occurred at many
sites, although transformations are often not complete. Freedman and Gossett
(1989) provided evidence for the conversion of PCE and TCE to ethene and de Bruin
et al. (1992) reported complete reduction to ethane. McCarty (1994) lists the possible
transformations for a number of the predominant chlorinated solvents in
groundwater (Table 1). McCarty (1994) indicates that methanogenesis is the most
favorable mechanism for complete reduction of PCE and TCE to ethene.
A number of researchers have confirmed the biological transformation of
CAHs at field sites. Major et al. (1991) reported evidence for bioattenuation of PCE to
ethene and ethane at a chemical transfer facility in North Toronto. Fiorenza et al.
(1994) presented data on the chemical and biological transformation of
trichlorethane (TCA) to 1,1-DCA and that of PCE and TCE to cis-DCE, vinyl chloride
(VC), and ethene at a manufacturing plant in Ontario. Beck (1994) reported the
degradation of 1,1,1-trichloroethane (TCA), PCE, and TCE to ethene and methane at
the Dover AFB in Delaware. McCarty and Wilson (1992), Haston et al. (1994),
Kitariidis et al. (1993), McCarty et al. (1991), and Wilson et al. (1994b) confirmed the
intrinsic biodegradation of chlorinated solvents at the St. Joseph, Michigan,
Superfund site.
In addition to biological transformations, chemical transformations of some
CAHs can occur in groundwater through elimination or hydrolysis. TCA is one of
the main chlorinated solvents that can be transformed chemically in groundwater
under all conditions likely to be found and within a reasonable time frame (McCarty
1994). The rate of chemical transformations is usually expressed using a first-order
reaction. TCA chemical transformation, for example, leads to the formation of 1,1-
DCE and acetic acid with a reported average half-life of less than 1 year at a
temperature of 20°C.
CASE STUDY — INTRINSIC BIOREMEDIATTON OF A UST RELEASE
The underground storage tank release in Rocky Point, North Carolina,
provides a representative example of a dissolved benzene, toluene, ethyl benzene,
and xylene (BTEX) plume undergoing intrinsic biodegradation using oxygen, nitrate,
iron, and sulfate as terminal electron acceptors (Borden et al. 1995). The water table
aquifer consists of mostly fine-grained, dark gray or greenish gray, micaceous,
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glauconitic, slightly silty, and compact quartz sand. The sand appears to be very
homogeneous throughout the site with only a few exceptions. This sand is overlain
by lower permeability clays and clayey sands that form a surface confining layer
throughout the site. The average groundwater velocity is approximately 30 m/yr.
The organic carbon content of the sand is relatively low (0.1%), and consequently
sorption is not a major attenuation mechanism for the moderately soluble BTEX
fraction.
Spatial Distribution of BTEX and Indicator Parameters
Background groundwater contains moderate levels of dissolved oxygen (2 to 3
mg/L), nitrate (1 to 6 mg/L as N), and sulfate (20 to 30 mg/L). Background dissolved
iron is low (< 0.5 mg/L) and the groundwater is acidic (pH < 5) with low buffering
capacity (alkalinity ~ 6 mg/L as CaC03) and low levels of dissolved CO2 (15 to 30
mg/L as C). At the upgradient edge of the BTEX plume, residual hydrocarbon is
trapped below the water table in the sand aquifer. As uncontaminated groundwater
enters this region, soluble hydrocarbons partition out of the nonaqueous-phase
(NAPL) and into the aqueous phase. Figure 1 shows the observed variation in BTEX
components, electron acceptors, and indicator parameters in a profile along the
dissolved hydrocarbon plume centerline. Several distinct zones can be identified
where different oxidation-reduction processes dominate.
At the upgradient edge of the BTEX plume, a portion of the soluble
hydrocarbons released from the residual NAPL are immediately degraded using
oxygen and nitrate carried into this zone by the flowing groundwater. Dissolved
iron increases to 29 mg/L due to reduction of insoluble iron oxides associated with
the sediment. In this region, the dominant electron acceptor is nitrate followed by
iron and oxygen. Dissolved CO2 increases from 16 to 60 mg/L as carbon (C) due to
oxidation of organic matter. The pH also rises from - 4.7 to 5.8 due to consumption
of H+ during iron reduction.
During transport downgradient from the source, toluene and ^-xylene decline
rapidly followed by m, ^-xylene and benzene. Ethylbenzene does not decline notably
with distance. This pattern is apparently due to preferential biodegradation of the q-
xylene isomer (and toluene) by subsurface microorganisms. Sulfate decreases from
34 to 0.7 mg/L, total sulfur decreases from 37 to 5 mg/L as SO4, and dissolved iron
increases from 29 to 65 mg/L. The large decline in sulfate and increase in dissolved
iron indicate that both sulfate and iron reduction are occurring. The smaller decline
in total sulfur suggests that sulfate is being reduced but is not being removed as
ferrous sulfide (FeS) precipitates. Ion chromatographic analysis of the groundwater
indicates that thiosulfate is a major component of the nonsulfate sulfur. Oxygen
and nitrate are not significant electron acceptors in this portion of the plume since
they were already consumed during transport through the source area.
The dissolved hydrocarbon plume becomes slightly narrower with distance
downgradient. The limited spreading of the plume is apparently due to the
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combined effects of anaerobic biodegradation in the plume center and aerobic
biodegradation at the sides of the plume. The background dissolved oxygen
concentration varies from 2 to 3 mg/L, whereas in the center of the plume dissolved
oxygen is below the field detection limit (0.3 mg/L). As the plume spreads in width
due to dispersion, oxygen in the uncontaminated groundwater mixes with BTEX,
enhancing biodegradation at the plume sides. The zone of highest BTEX
concentration also moves vertically downward with increasing distance
downgradient. The vertical drop along the length of the plume is due to surficial
groundwater recharge that both adds a layer of dean, uncontaminated water on top
of the plume and enhances biodegradation by introducing oxygen and other electron
acceptors with the recharge water.
Rates of Intrinsic Bioremediation
Field implementation of intrinsic bioremediation requires an accurate
estimation of in situ biodegradation rates. Three different approaches were applied
at the Rocky Point site to estimate intrinsic bioremediation rates: (1) comparing of
peak contaminant concentrations in monitoring wells versus travel time from the
source (Borden et al. 1994); (2) monitoring laboratory microcosms under ambient
(anaerobic) aquifer conditions (Hunt et al. 1995); and (3) monitoring of in-situ test
chambers to determine compound loss over time (Hunt et al. 1995). The in situ test
chambers were installed approximately midway down the plume in the iron-
reducing zone at the same location used to collect the sediment for the laboratory
microcosms. Results from each of these approaches are compared in Table 2.
In the laboratory microcosms, a distinct order of biodegradation was observed.
Toluene and ^-xylene appeared to be the most biodegradable followed by m-, j>-
xylene and benzene, with ethylbenzene being the least biodegradable. This same
order of disappearance is also seen in the field data. Unfortunately, the rates of
biodegradation estimated from the field data are vastly different from the laboratory
data. In the laboratory microcosms, the benzene, toluene, and xylene isomers were
degraded from between 2,000 and 3,000 |ig/L each to below detection limit in 400
days. Ethylbenzene was only slightly degraded over this period. If similar
biodegradation rates were observed in the field, the benzene, toluene, and xylene
plumes would completely biodegrade over 100 to 200 ft. Yet significant
concentrations of these dissolved hydrocarbons persist over 1,300 ft downgradient
from the source. The cause of this discrepancy is not well understood.
Figures 2a and 2b show the vertical distribution of benzene, ethylbenzene, m.-,
p-xylene, and two trimethylbenzene isomers (mesitylene and pseudocumene) in a
multilevel sampler located 800 ft downgradient from the source near the location of
the in situ test columns. The concentrations of toluene and Q-xylene were too low
to be shown on these figures. In both figures, the vertical distribution of the more
recalcitrant compounds (benzene, ethylbenzene, and mesitylene) is relatively
consistent. In contrast, there are large changes in the concentration of the more
biodegradable compounds (m-, ^-xylene and pseudocumene). This suggests that the
rate of biodegradation may be significantly different between adjoining layers.
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The observed differences in field and laboratory biodegradation rates could be
due to changes in the activity of different layers against the pollutants. In the field,
if one layer is not active against the pollutants, the vertically averaged concentration
measured with long-screened wells would remain high and the apparent
biodegradation rate would be low. In contrast, within in situ test columns and
laboratory microcosms, groundwater is forced into contact with sediment of
differing activities. This would result in a higher apparent biodegradation rate.
CASE STUDY — INTRINSIC BIOREMEDIATION OF TCE IN GROUND WATER
The groundwater at the St. Joseph, Michigan, site is contaminated with
chlorinated aliphatic hydrocarbons at concentrations ranging from 10 to 100 mg/L.
The contaminants are divided into eastern and western plumes as the suspected
sources were situated over a groundwater divide. Both plumes contain TCE, £i£-
and trans-1, 2-dichloroethene (c-DCE and t-DCE), 1,1, -dichJoroethene (1,1 -DCE),
and VC. Significant levels of ethene and methane were measured at the site
(Kitanidis et al. 1993 and McCarty and Wilson 1992), confirming the natural
attenuation of TCE.
McCarty and Wilson (1992) delineated contours of the chemical oxygen
demand (COD, a surrogate for the capacity of a donor to supply electrons) and
correlated them with contours of chlorinated aliphatic compounds (see Figure 3).
The authors found a correlation between COD decrease and transformations of TCE
to VC. Also shown in Figure 3 are the locations of three transects where data were
collected for the detailed bioattenuation characterization. Table 3 summarizes the
electrons released by TCE reduction to different products, along with the equivalent
amount of COD decrease. Essentially, the reduction of 1 mole or 131 g of TCE to
ethene releases six electrons, and an equivalent decrease of 48 g of COD is needed.
Table 4 summarizes the CAHs, ethene, and methane found at some of the
monitoring locations. The average concentration of methane at depths of 25 m or
more was 6 mg/L. This concentration of methane corresponds to an equivalent
COD decrease of 24 mg/L, However, Wilson and McCarty (1992) measured COD
decrease, as high as 200 mg/L. They attribute this inconsistency to either dilution
effects between the lagoon and the detailed characterization location or to the
presence of other electron acceptors such as nitrate and sulfate.
Apparent Degradation Constants
Wilson et al. (1994b) studied the western plume at the site to estimate the
contaminant mass flux and to estimate apparent degradation constants. Data
collected in 1991 from three transects near the source of the western plume and data
collected in 1992 from two additional transects were used in the analysis (see Figure
4 for locations of the transects). The mass estimates combined with the flow
velocities were used to estimate the mass flux at each transect. As would be
expected, the mass fluxes decline toward the downgradient edge of the plume.
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The mass per unit thickness of TCE at transects 2,4, and 5 was used to
estimate first-order degradation constants. Table 5 lists the computed apparent loss
coefficients for three different estimates of the hydraulic conductivity at the site.
The rate for TCE degradation ranges from 0.0048 to 0.011 wk"1 between the
upgradient transects 2 and 4. This rate increases up to 0.023 wk"l between transects 4
and Lake Michigan.
DEMONSTRATING INTRINSIC BIOREMEDIATION IN THE FIELD
Initial Monitoring to Determine if Intrinsic Bioremediation is Feasible
Extensive field monitoring is initially conducted to determine if intrinsic
bioremediation is feasible as a remedial alternative at a site. Rifai (1990) and Borden
(1990) proposed preliminary sampling protocols for soils, groundwater, and soil gas
(see Table 6). These protocols were intended for underground storage tank sites and
mostly focused on defining the electron acceptor distribution in the groundwater
and on quantifying the by-products of biodegradation in groundwater, and soil gas.
Since then, the Air Force Center for Environmental Excellence (AFCEE) has
developed a detailed protocol for site characterization in support of the natural
attenuation alternative at Air Force facilities (Wiedemeier et al. 1994).
At a minimum, site characterization in an intrinsic bioattenuation protocol
should provide data on the location and extent of contaminant sources, the extent
and distribution of dissolved contaminants, groundwater geochernical data (i.e.,
concentrations of electron acceptors and by-products of biodegradation
mechanisms), geologic characterization data, and hydrogeologic parameters such as
hydraulic conductivity, gradients, and potential migration pathways.
The site characterization data are analyzed to quantify the extent of intrinsic
bioattenuation. Overall, three indicators of natural attenuation can be developed
from this information:
1. Compound disappearance: one of the most convincing arguments for
natural attenuation involves demonstrating the disappearance of a dissolved
organic chemical at the site relative to the persistence of another "conservative" or
"recalcitrant" organic (internal standard). In some cases, it is sufficient to
demonstrate that the extent of migration of the organic of concern is less than that
of the "conservative" tracer, and thus its transport downgradient is being limited by
natural attenuation. In cases where compound disappearance cannot be related to
an internal standard, it may be possible to demonstrate mass loss as a function of
time for the organic of interest. Another alternative involves analyzing the peak
concentrations of the organic at the different monitoring wells downgradient from
the source. This analysis should demonstrate the overall decline of these
concentrations as a function of time and distance.
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2. Loss of electron donors: measuring dissolved oxygen concentrations and
those of other electron acceptors can provide a key indicator of natural attenuation.
Reduced oxygen, nitrate, and sulfate concentrations within the plume relative to
their background concentrations are considered to be strong evidence of intrinsic
bioattenuatioru Many field studies, for instance, have correlated depressed oxygen
concentrations within the center of a plume with high concentrations of the
dissolved contaminant.
3. Degradation products: the accumulation of dissolved iron and the
production of carbon dioxide, hydrogen sulfide, and methane are additional
indicators of biological attenuation at intrinsic bioremediation sites.
One of the interesting questions that is currently being investigated by
researchers is whether it may be possible to complete a mass balance on the supply
of electron donors and electron acceptors at a given field site. This question is
complicated, because of sampling and field data collection limitations. Other
complicating factors involve the temporal nature of the distributions of the electron
acceptors and donors.
Risk Assessment
The successful application of the natural attenuation alternative involves an
exposure and risk assessment analysis that considers the location of receptors at a
site in relation to the extent of the contaminant plume. While a detailed discussion
of risk assessment is beyond the scope of this paper, it is important to note that
intrinsic bioremediation will only be viable as a remedial alternative if it can be
demonstrated that little risk to human health and the environment will be incurred
as a consequence of this management strategy. To demonstrate the viability of
intrinsic bioremediation, it maybe necessary to conduct fate and transport
predictions of the future conditions at contaminated sites.
Prediction of Flume Migration — Modeling Approaches
Two key questions need to be answered when determining the viability of
natural attenuation as a remediation alternative: (1) How far will the dissolved
plume migrate before it is attenuated to below a predetermined cleanup standard?
and (2) How long will it take for the attenuation process to "cleanup" the plume?
Both questions can be readily answered using analytical and numerical models of
fate and transport. Analytical models are simpler to use, but they are limited in
their capabilities to simplified hydrogeologic scenarios. Numerical models are more
complicated, but can be used to simulate heterogeneous systems and more complex
hydrogeologic and contaminant scenarios.
Numerous fate and transport models have been developed over the years.
The majority of these models simulate advection, dispersion, and sorption and
some form of source representation. A smaller number of these models, however,
can actually simulate complex biological and chemical transformation processes. In
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the analytical modeling arena-, the most common method for simulating
biodegradation is through the use of a first-order decay coefficient. The contaminant
of concern is assumed to biodegrade exponentially, and the modeler specifies the
first-order decay constant for a given site and a given contaminant This leaves the
modeler with the dilemma of selecting a first-order decay constant.
While many laboratory and field studies have developed first-order decay
constants for a variety of contaminants under a number of hydrogeologic scenarios,
these constants are not readily transferable to other sites. Another problem noted by
Rifai (1994) is that the first-order decay model does not account for electron acceptor
limitations and thus can overestimate the effect of biodegradation on a given
system. Connor et al. (1994) proposed using an instantaneous reaction expression
similar to that used in BIOPLUME II (Rifai et al. 1988) as an alternative to the first-
order decay model. Additionally, Rifai and Hopkins (1994) have developed,
through modeling with BIOPLUME II, "electron-acceptor" limited decay coefficients
for scenarios with contaminant sources removed and continuous contaminant
sources (Tables 7a and 7b) to provide more "applicable" decay constants.
Finally, and for sites where a long history of monitoring exists, it may be
possible to estimate a first-order decay constant based on the observed mass loss in
the aquifer. The dissolved concentrations for different sampling events are used to
generate an estimate of the dissolved mass in the aquifer as a function of time. The
resulting data allow the modeler to estimate a first-order decay constant. It should
be mentioned, however, that this procedure is highly dependent on the density of
the sampling network and would not be very accurate for sites with a limited
number of monitoring wells.
Numerical models, as mentioned earlier, can provide more simulation
capabilities than analytical models. The BIOPLUME H model (Rifai et al. 1988)
allows the user to simulate a heterogeneous aquifer system with a variable flow
field. The BIOPLUME IT model is one of the few two-dimensional models in two
dimensions that can simulate the transport of an electron acceptor (oxygen in this
case) and its reactions with the aquifer contaminants. The model currently
simulates the instantaneous reaction between oxygen and aromatic hydrocarbons.
Rifai et al. (1995) are extending the BIOPLUME II model to allow the simulation of
multiple electron acceptors within a contaminant plume. A number of other
numerical biodegradation models exist in addition to BIOPLUME II (Table 8). The
majority of these models, however, are either one-dimensional or of a proprietary
nature.
One of the difficulties encountered in using numerical models is determining
what data are required and how to incorporate the field data into the modeling
process. Most fate and transport models require an estimate of the aquifer thickness,
matrix conductivity, porosity, and sorptive characteristics. Additionally, most
models require some description of the hydraulic and hydrologic stresses on the
system in the form of boundary conditions or recharge and discharge specification.
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One of the most complicated parameters to estimate for numerical models is the
source representation because, in most cases, the history of contamination at the site
is not known with any degree of certainty. Finally, biodegradation models require
input on the electron acceptor availability within the aquifer.
The process of simulating natural attenuation at a site using a numerical
model requires: (1) calibrating the model to the hydraulics at the site so that the
model can emulate the direction of flow and observed groundwater velocities in the
field; and (2) calibrating the model to simulate existing contamination conditions.
Once those two steps have been completed, the numerical model can be used to
determine the distribution of contaminants at the site as a function of time.
One of the problems faced in modeling intrinsic bioattenuation of organic
chemicals at sites is the fact that the observed data usually incorporate the effects of
advection, dispersion, sorption and biodegradation. Therefore, it may be difficult to
estimate the advective and dispersive components independently from these data.
A possible solution is to use a "conservative tracer" or an "internal standard" in the
calibration of the numerical model that does not sorb or biodegrade. For example,
when simulating gasoline spills, it may be possible to use MTBE concentrations in
the calibration process if the data exist Methyl terbutyl ether (MTBE) does not sorb or
biodegrade and thus would reflect the advective and dispersive characteristics of the
aquifer). The biodegradation and sorption of the aromatic hydrocarbons within the
gasoline plume can be readily estimated by comparing the BTEX plume to the MTBE
plume. Wiedemeier et al. (1995) have also suggested the use of tetramethylbenzene
as a more recalcitrant internal standard at fuel spiE sites.
SUMMARY OF INTRINSIC BIOREMEDIATION HELD SITES
Over the past decade, a large number of sites undergoing intrinsic
bioremediation have been studied in detail. Some of the major characteristics of
selected sites are provided in Table 9. Upgradient background concentrations of
oxygen, nitrate, and sulfate provide an indication of the concentrations potentially
+2
available for biodegradation. Elevated concentrations of Fe and CH4 in the plume
reflect the importance of iron reduction and methanogenic fermentation. At many
sites, significant concentrations of iron, nitrate and sulfate are available to support
hydrocarbon biodegradation. Although methane fermentation has been
documented at several sites, it appears to be less important than iron and sulfate
reduction.
Biodegradation results from field and laboratory studies are reported in Table
10. Laboratory results are reported only if the laboratory study was designed to
simulate field conditions. Effective decay rates have been estimated in the field
using several approaches. The most reliable approach is to calculate a mass balance
for a known mass of contaminant injected into an aquifer using a dense network of
monitoring points. This approach is feasible only when a pulse of contaminant is
injected. For continuous sources at steady-state, the degradation rates may be
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calculated from the change in total mass flux across several lines of monitoring
wells. Both the mass balance and mass flux approaches are expensive to implement
because of the high number of monitoring points required. At most field sites, the
only feasible approach is to estimate the degradation rate from a plot of peak
contaminant concentration versus travel time from the source. This general
approach has been modified by normalizing the contaminant concentrations to an
'internal standard' that is poorly biodegradable, has similar sorption and
volatilization properties as the contaminant, and is present in the waste source.
Ideally, the internal standard approach should correct for changes in concentration
due to dilution.
There is a wide range in reported biodegradation rates. Reported first order
degradation rates for benzene range from non detectable to approximately 1% per
day, with an average of approximately 0.2% per day. Degradation rates for other
hydrocarbons were typically somewhat higher, but in the same general range.
Higher biodegradation rates occur most frequently at sites containing higher
concentrations of sulfate in the background water and higher concentrations of
dissolved iron in the plume. Where high concentrations of methane are observed,
biodegradation rates are often lower.
SUMMARY
Past research has shown that intrinsic bioremediation can control the
migration of dissolved hydrocarbon plumes. Field biodegradation rates are often
lower than would be expected based on laboratory results, but are often sufficient to
contain the contaminant plumes within reasonable transport distances. Fewer data
are available on chlorinated hydrocarbon plumes, but ongoing studies suggest that
intrinsic bioremediation may also be technically feasible at these sites. At many sites
intrinsic bioremediation alone may be the best alternative available for risk
management.
Intrinsic bioremediation will be the preferred alternative when the costs of
conventional remediation are high, the problem compounds are easily
biodegradable, aquifer conditions are appropriate, there are no nearby groundwater
receptors, and/or there is a well-defined surface water discharge. Intrinsic
bioremediation alone may not be the best alternative when the costs of
conventional remediation are moderate to low and/or there is a large source of
poorly degradable compounds. Norris et al. (1994) argue that, at some sites, it may be
more cost effective to implement some low-level activities (e.g., limited air sparging
or venting) than to rely on intrinsic bioremediation as the only management
technique. If these limited remediation activities can shorten the monitoring
period by several years, they may more than pay for themselves in reduced long-
term monitoring costs.
-------
13
ACKNOWLEDGMENTS
Portions of the research described in this article were supported by the
American Petroleum Institute under Grant No. GW-25A-0400-38. The opinions,
findings, and conclusion expressed are those of the authors and do not necessarily
represent those of the American Petroleum Institute.
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-------
Sand
Decay Rates (per day)
Dissolved Oxygen
Gradient
12mg/L
9mg/L
6mg/L
3mg/L
0.01
3.8E-04
2.5E-04
1.8E-04
1.0E-04
0.001
1.0E-06
1.0E-06
7.5E-07
7.5E-07
Silt
Decay Rates (per day)
Dissolved Oxygen
Gradient
12mg/L
9mg/L
6mg/L
3mg/L
0.01
1.0E-05
1.0E-05
7.5E-06
5.0E-06
Clay
Decay Rates (per day)
Dissolved Oxygen
Gradient
12mg/L
9mg/L
6mg/L
3mg/L
0.01
1.0E-05
1.0E-05
1.0E-06
1.0E-06
Note:
The decay rates listed in this table were obtained through Eioplume II model simulations.
A continuous source scenario assuming 100 mgfl. was injected at a rate of 10 gal/day in the sand
hydrogeologic environment.
A lower injection rate, 1 gal/day, was used in the sift and clay hydrogeologic environment to minimize
mounding in the model.
Centerline concentrations at different receptor well locations modeled for oxygen-limited degradation were
matched to concentrations using first-order decay.
-------
Sand
Decay Rates (per day)
Dissolved Oxygen
Gradient
12mg/L
9mg/L
6mg/L
3mg/L
0.01
-3.2E-01
-2.1E-01
-1.3E-01
-6.8E-02
0.001
-3.3E-02
-2.4E-02
-1.7E-02
-1.0E-02
0.0001
-8.7E-03
-6.3E-03 .
-4.1E-03
-2.5E-03
Silt
Decay Rates (per day)
Dissolved Oxygen
Gradient
12mg/L
9mg/L
6mg/L
3 mg/L
0.01
-5.8E-03
-4.1E-03
-3.0E-03
-1.6E-03
0.001
-2.8E-04
-2.4E-04
-2.1E-04
-1.8E-04
0.0001
-2.5E-05
-2.4E-04
-2.1E-05
-2.3E-05
Clay
Decay Rates (per day)
Dissolved Oxygen
Gradient
12mg/L
9mg/L
6mg/L
3 mg/L
0.01
-4.0E-04
-3.4E-04
-3.0E-04
-2.6E-04
0.001
-2.1E-04
-3.5E-05
-3.0E-05
-2.7E-05
0.0001
-4.3E-06
-3.7E-06
-3.2E-06
-2.9E-06
Note:
The decay rates listed in this table were obtained through Eioplume II model simulations,
A removed source scenario assuming a plume 45 ft wide and 50 ft long with an initial concentration of 10
mg/L was simulated. Log of present mass dissolved at 1 month to 5 to 10 years was plotted over time. The
decay rate is the slope of the plotted line.
-------
Name
Description
Authors)
1-D, aerobic, microcolony, Monod
BIOPLUME 1-D, aerobic, Monod
1-D, analytical first order
B1Q1D 1-D, aerobic & anaerobic, Monod
1-D, comeiabolic, Monod
1 -D, aerobic anaerobic, nutrient limitations,
microcolony, Monod
1-D, aerobic, cometabofic, multiple
substrates, fermentative, Monod
BIOPLUME II 2-D, aerobic, instantaneous
2-D, Monod
BIOPLUS 2-D, aerobic, Monod
ULTRA 2-D, first order
2-D, denitriflcation
2-D, Monod, Biofilm
Molz et al. (1986)
Borden et al. (1986)
Domenico (1987)
Srinivasan and Mercer (1988)
Semprini and McCarty (1991)
Ceiiaetal. (1989)
Rifai et al. (1988)
MacQuarrie et al. (1990)
Wheeler etai. (1987)
Tucker et al. (1986)
Kinzelbach et al. (1991)
Odencrantz et al. (1990)
(From Bedient et al.. Ground Water Contamination: Transport and Remediation. Prentice Hall. NJ. 1994)
-------
Background
Plume
Site
Aquifer Type
pH
02
(mq/L)
N03-N
(mq/L)
S04
(mq/L)
Fa*2
(mg/L)
ch4
(mq/L)
Borden, Ontario
{Barker et al. 1987)
glaciolacustrine,
medium to fine sand
Borden, Ontario
(Barbara etal. 1992)
glaciolacustrine,
medium to fine sand
7.4
<0.07
<0.05
220
Rocky Point, NC
(Borden et al. 1995)
marine, fine sand
4.7
2.5
6
30
102
0.4
Kalkaska, Ml
(Chiang et al. 1989)
medium-course sand,
qravel interbeds
5.5-7.0
9
0.8
1-6
Columbus, MS
(Maclntyre et al.
1993)
fluvial,
heterogeneous sands
and clays
2.6-3.8
Sleeping Bear, Ml
(Wilson et al. 1994a)
glacial outwash,
course sand/caravel
6.4
2.4
15
20
17-28
24-30
Eglin AFB, FL
(Wilson et al. 1994a)
sands and sitty peats
5.6-6.7
8.
17
Hill AF8-1.UT
(Wiedemeier et al.
1994)
thin channel sands in
deltaic deposits
-7
6
8
100
50
2
Patrick AFB. FL
(Wiedemeier et al.
1994)
tine to course marine
sand with shell
fraaments
-7
3.7
0.3
86
1.9
14.6
Fairfax, VA
(Buscheck et al.
1993)
5
2.5-4
43
Sampson Co.. NC
(Borden personal
communication 1995)
clayey and silty
sands
4-5
8
17
8
0.4
<0.01
Traverse City
(Wilson et al. 1990)
glaciolacustrine,
medium sand with
qravel
7
9
3
10
12
17
Broward Co., FL
(Caldwell et al. 1992)
7.2
2-3
0.3
<5
1.8
Pensacola, FL
(Godsy et al. 1992;
Bekins et al. 1993)
poorly sorted fine to
course deltaic sands
6.9
0.04
2
6
33
13
Bermidji, MN
(Baedecker et al.
1993)
glacial outwash,
course sand with silt
layers
7.6
7.7
0.001
2
16
12
Galloway, NJ
(Cozzarelli and
Baedecker1992)
fine to course sand,
perched water table
on clay lense
4.7
6
9
24
37
.007
Perth, Australia
(Thierrin et al. 1993)
medium to fine
aeolian sand
5.8
0.4
0.1
20-100
1.4
Manufacturing Plant
(Davis et al. 1994)
glacial silty sand over
bedrock
6.9
1.5
<0.005
Cliffs-Dow
(Klecka et al. 1990)
course sand and
qravels
6.1
1.2
1.7
48
Hill AFB-2, UT
(Dupont et al. 1994)
4
6.9
-------
Site
Contaminant
V (m/d) Fiold Results
Laboratory Results
Borden, Ontario
(Barker et al, 1987)
BIX stock solution injected
into unconlaminaled aquifer
0.09 Zero-order decay rales from
mass balance method:
benz. = 30 mg d"1;
tol. = 37 mgd-1;
m-xyl. = 47 mg d1;
p-xyl. = 55 mg d*1;
o-xyl, = 33 mg d'1.
Zero-order decay rates (per
1,800 L) from aerobic
microcosms; benz. = 58 d*1;
tol. = 61 mgd"1;
m-xyl. = 50 mg d ';
p-xyl. = 65 mg d*1;
o-xyl.« 54 mg d'1.
Borden, Ontario
(Barbara et al. 1992)
stock solution contacted wilh
gasoline then injecled into
leachate plume
0.09
% loss over 4 rn travel,
Rocky Point, NC
(BorcJen et al. 1995)
residual gasoline from UST 0.08
Rates from conc. vs. travel
time: benz. = 0.0002 d1;
tol. = 0.0021 d*>;
e-benz. = 0.0015 d"1;
mp-xyl. = 0.0013 d"1;
o-xyl. = 0.0021 d"1.
Rales from Fe/S04 reducing
microcosms;
benz. = 0.024 d"1;
tol. = 0.045 d1;
e-benz. = 0.002 d"';
mp-xyl. = 0.02 d*1;
o-xyl. a 0.056 d"1.
Kalkaska, Ml
(Chiang et al. 1989)
natural gas condensate-BTEX 0.2
Rales from mass balance:
benz. = 0.0095 d"1.
Rales from aerobic
microcosms: BTX » 0.01 to
0.1 d'1.
Columbus, MS
(Maclntyre et al. 1993)
stock solution of benzene,
p-xylene, naphthalene,
o-dichlorobenzene
Tritium used as non-reactive
tracer. Mineralization proven
using ,4C - p-xyl. Rales from
mass balance:
benz. = 0.0070 d"';
p-xyl. = 0.0107 d"1;
naphthalene = 0.0064 d*1;
o-DCB = 0.0046 d"1.
Sleeping Bear, Ml
(Wilson et al. 1994; Schafer,
1994)
residual gasoline from UST
release - BTEX
0-0.4 Rates from conc. vs. travel
lime using 2,3-
dimethylpenlane as an internal
standard: benz. = N.S.;
tol. = 0.02 - 0.07 d*1;
e-benz.. = 0.03-0.011 d"1;
m-xyl. = 0.004 - 0.014 d*1;
p-xyl. = 0.002 -0.010 d"';
o-xyl. = 0.004 - 0.011 d'1.
Rates from melhanogenic
microcosms: benz. = N.S.;
tol. = 0.007 -0.04 d*1;
e-benz. = N.S.;
mp-xyl. = N.S.;
o-xyl. = N.S.
12 to 16 mg/L CH4 produced
in lab microcosms.
Indian River, FL
(Kemblowski el al 1987)
gasoline from UST - BTEX 0.06
Conc. vs. travel lime:
benz. = 0.0085 d*1.
1st-onder rates from aerobic
microcosms: benz, = 0.02 to
0.2 d'1.
-------
Site
Contaminant
V (m/d)
Field Results Laboratory Results
Morgan Hill, CA
(Kemblowski et al 1987)
gasoline - BTEX
0.05
Rates from conc. vs. travel
time: benz. = 0.0035 d'1.
Eglin AFB, FL
(Wilson et al, 1994)
JP-4 from POL depot
1.3
Rates from conc. vs. travel
time using 1,2,4-
trimelhylbcnzeno as internal
slandard: benz. = B.O.;
tot. = 0.05 lo 0.013 d"1;
e-benz. = 0.03 to 0.05 d"1;
m-xyl, = 0.02 to 0.1 d'1;
p-xyl. = 0.02 to 0.00 d"1;
o-xyl. = 0.21 d"1.
Hill AFB, UT
(Wiedemeier ©1 al, 1994)
JP-4 from POL depot
0.5
Rates Irom conc, vs. travel
time using total
trimelhylbenzene as internal
standard:
benz. =0.03 to 0.09 d"1;
e-benz. =0.01 to 0.08 d"<;
p-xyl. = 0.01 to 0.03 d"1;
m-xyl. = 0 lo 0.03 d"1;
o-xyl. = 0 lo 0.02 d"1. Toluene
rate not calculable.
Patrick AFB, FL
(Wiedemeier et al. 1994)
700 gallons unleaded gasoline
from UST
0.13
Rates from conc. vs. travel
time using tolal methane as
internal standard;
benz. =0 to 0.004 d"1;
tol. = 0.0006 to 0,004 d"1;
e-benz. = 0.0001 to 0.004 d'1;
p-xyl. = 0.001 to 0.003 d"1;
m-xyl. = 0.001 to 0.004 d*1;
o-xyl. = 0.004 to 0.02 d*1.
Fairfax, VA
0.015
Rales from conc. vs. travel
(Buscheck et al. 1993)
time: benz. = 0.00055 d"1;
tol. = 0.00045 d"1;
e-benz. = 0.00045 d1;
mpo-xyt. = 0.00040 d"'.
San Francisco, CA
(Buscheck et al. 1993)
0.03
Rales from conc. vs. travel
time: benz. = 0.0028 d"1;
tol. = 0.0022 d'1;
e-benz. = 0.0033 d*1;
mpo-xyl. = 0,0023 d"1.
-------
Site
Contaminant
V (m/d)
Field Results
Laboratory Results
Alameda County, CA
(Buscheck el ai, 1993)
gasoline - BTEX
.01
Rates from conc. vs. travel
time: benz. = 0.0020 d"1;
tol. = 0.0017 d'1;
e-benz. = 0.0020 d"1;
mpoxyl. = 0.0017 d"1.
Elko County, NV
(Buscheck et al. 1993)
gasoline - BTEX
0.04
Rates from conc. vs. travel
time; benz. = 0.001 d"'.
Sampson Co., NC
(Borden personal
communication 1995)
gasoline from UST -
BTEX/MTBE
0.04
High nitrate concentrations in
groundwater may onhnnco
biodegradation 1st-order rates
from mass flux:
MTBE = 0.0006 d*1;
benz. = 0.0000 d"1;
tol. = 0.0021 d'1;
e-benz. = 0.0023 d*1;
mp-xyl. a 0,0016 d*';
o-xyl. = 0.0009 d"1.
Toluene and ethylbenzeno
rapidly degraded in denitrifying
microcosms after a 56-day lag
period.
Traverse City
(Wilson et al. 1990)
aviation gasoline from UST -
BTEX
1.5
Rates from conc. vs. travel
time; benz. = 0.001 d-';
tol, =0.2 d1;
mpoxyl. = 0.004 d"'.
Anaerobic microcosm rates:
benz. = 0.07 d"';
tol. = 0.04 d*1;
mp-xyl. = 0.06 d*1;
o-xyl. =0.07 d"1.
Methane produced in
microcosms.
Broward Co., FL
(Caldwell el al. 1992)
gasoline from UST - BTEX
and MTBE
0.1
Anaerobic decay rate Irom
matching BIOPLUME for total
BTEX = 0.00012 d*1. Aerobic
decay will increase net
biodegradation.
Pensacola, FL
(Bekins et al. 1993)
creosote - phenols
0.3 to 1.2
Selected phenols were
completely degraded over a
100-
-------
Site
Contaminant
V(m/d)
Field Results
Laboratory Results
Perth, Australia
(Thierrin et al. 1993)
gasoline from UST - BTEX
0.4
Rates from conc. vs. Iravel
time: benz. = N.S.;
tol. = 0.006 d*1;
e-benz. = 0.003 d*1;
mp-xyl. = 0.004 d*1;
o-xyl. = 0.006 d*1;
naphthalene = 0.004 d"1.
Field (plume scale) rates
closely match rales from
tracer test using deuteraled
compounds.
Anaerobic columns with 14
mg/L S04
benz.« N.S.
tol.» 2.3 d"'
e-benz. = N.S.
o-xyl. - N.S.
Manufacturing Plant
(Davis et a. 1994)
benzene only
0.16
BIOID match to field data
showed benzene decay rate
>0.01 d"1.
Over 90% benzene loss over
77 days in methanogenic and
sulfate-reducinq microcosms.
Cliffs-Dow
(Klecka et al. 1990)
charcoal wastes, phenols,
naphthalene
0.2 to
0.46
All organics degraded within
100 rn of source.
All organics degraded in
aerobic microcosms within 30
to CO d.
Hill AFB-2, UT
(Dupont et al. 1994)
18,000 gallon UST
0.14
Rates from mass balance: 1st
order for TPH = 0.005 d'1.
Zero order for:
benz. = 0.02 kg/d;
e-benz. = 0.06 kg/d;
p-xyl. = 0.06 kq/d.
Gas Plant
(Piontek et al. 1994)
NAPL released from nalurai
gas plant - BTEX
105 reduction in BTEX over
100 m.
Picatinny Arsenal, NJ
(Martin and Imbrigiotta 1994)
TCE, 1,1,1-trichloroelhane
and metals from plating
wastewater
0.3 lo 1.0
Spatial distribution of TCE,
DCE, and VC indicate
reductive dechlorination.
Anaerobic Microcosms: TCE =
0.0001 to 0.003 d"1.
St. Joseph, Ml
(Wilson et al. 1994)
TCE from lagoons/dry wells
0.1
Rates from mass flux:
TCE = 0.001 -0.003 d*1.
Finger Lakes, NY
(Major et al. 1994)
TCE, acetone, methanol
Spatial distribution of TCE,
DCE, VC, and ethene were
indicative of reductive
dechlorination.
benz. = benzene, tol. a toluene, e-benz. = ethylbenzene, xyl. = xylenes.
N.S. = not significant.
-------
Concentration (mg/L)
NJ t>- 01 OJ O f>J
O O Q O O O O
at
Concentration (meq/L)
Concentration (mg/L)
O M O) CO O M
1111111111II11111II11111j if 11
>
Oi Oi
OM JZO
(p>pjsj4
58-28 3
-------
Transform
product
Mol. Wt.
*
Half Reaction
Equiv.
COD
decrease
g COD/g
product*3)
Methane
16
C02+8H++8e*—>CH4+2H20
4
DCE
97
CHCI=CCl2+H+2e-->CHCI=CHCI+CI-
0.16
VC
62.5
CHCI=CCI2+2H++4e--»CH2=CHCI+2CI-
0.51
Ethene
28
CHCI=CCI2+3H++6e"—~CH2=CH2+3CI"
1.71
(a) From (02+4H"+4e*-»2H20), the half-reaction for oxygen, one electron equivalent of COD equals one-
fourth mole of molecular oxygen or 8 grams, thus equiv. COD Decrease = (8n/Mol. Wt), where n is the
number of electrons in the half-reaction.
-------
Samp let3)
TCE
1.1 DCE
cDCE
tDCE
vc
Ethene
CH4
Total
1-2-70
mg/L
Equiv. COD
4.03
0.09
4.14
0.81
3.19
6.62
4.61
Decrease, mg/L
0.00
0.01
0.66
0.13'
0.41
11.32
18.44
30.97
% of Equiv. COD
0.0
0.0
2.1
0.4
1.3
36.6
59.5
1-3-75
mg/L
Equiv. COD
12.80
0.27
16.90
0.67
56.40
2.25
6.62
Decrease, mg/L
0.00
0.04
2.70
0.11
28.80
3.84
26.00
61.5
% of Equiv. COD
0.0
0.1
4.4
0.2
46.8
6.2
42.2
22-1-75
mg/L
Equiv. COD
0.44
0.09
13.40
0.21
1.46
3.15
7.43
Decrease, mg/L
0.00
0.01
2.14
0.03
0.74
5.39
29.72
38.03
% of Equiv. COD
0.0
0.0
5.6
0.1
1.9
14.2
78.1
2-6-65
mg/L
Equiv. COD
0.51
4.70
0.03
2.66
4.27
11.72
Decrease, mg/L
0.00
0.00
0.75
0.00
1.36
5.98
46.88
54.97
% of Equiv. COD
0.0
0.0
1.4
0.0
2.5
10.9
85.3
3-2-80
mg/L
Equiv. COD
2.53
0.01
0.90
0.03
0.27
4.87
11.59
Decrease, mg/L
0.00
0.00
0.14
0.00
0.14
8.32
46.36
54.96
% of Eauiv. COD
0.0
0.0
0.3
0.0
0.3
15.1
84.4
(aJ Firs! value is transect number, second value is borehole number, and third value is depth of sample
below ground surface in feet.
-------
Apparent Loss Coefficients at St. Joseph, Ml
Transect 2 to 4
Low Conductivity Avg, Conductivity High Conductivity
Estimate Estimate'' Estimate
(1/wk) (1/wk) (1/wk)
TCE 0.0048 0.0074 0.011
cDCE 0.0064 0.0097 0.013
tDCE 0.0076 0.012 0.0016
1,1-DCE 0.0066 0.010 0.0135
VC 0.0023 0.0035 0.0047
Apparent Loss Coefficients at St. Joseph, Ml
Transect 4 to 5
Low Conductivity Avg. Conductivity High Conductivity
Estimate Estimate Estimate
(1/wk) (1/wk) (1/wk)
TCE 0.016 0.025 0.033
cDCE 0.010 0.016 0.021
tDCE 0.010 0.016 0.021
1,1-DCE 0.012 0.018 0.024
VC 0.011 0.017 0.023
Apparent Coefficients at St. Joseph, Ml
Transect 5 to Lake
Low Conductivity
Estimate
(1/wk)
Avg. Conductivity
Estimate
(1/wk)
High Conductivity
Estimate
(1/wk)
TCE
0.011
0.018
0.023
cDCE
0.038
0.059
0.079
tDCE
0.0092
0.014
0.019
1,1-DCE
-
-
.
VC
0.053
0.081
0.11
-------
TABLE 1.
TABLE 2.
TABLE 3,
TABLE 4.
TABLE 5.
TABLE 6.
TABLE 7a.
TABLE 7b.
TABLE 8.
TABLE 9.
TABLE 10.
Environmental conditions generally associated with reductive
transformations of chlorinated solvents.
Comparison of intrinsic bioremediation rates from field monitoring,
laboratory microcosms, and in situ test chambers.
Half reactions indicating electron equivalents of change and associated
equivalent COD decrease associated with change.
Concentrations of CAHs, ethene, and methane found at selected
sampling locations along detailed characterization transects and the
equivalent COD decrease associated with the products.
Apparent loss coefficients at St Joseph, Michigan.
Proposed parameters for field measurements.
Aerobic decay rates for removed source scenario.
Aerobic decay rates for continuous source scenario.
Biodegradation models.
Characteristics of intrinsic bioremediation field sites.
Biodegradation results from field and laboratory studies at intrinsic
bioremediation sites.
FIGURE 1. Variations in (A) BTEX components, (B) electron acceptors, and (C)
indicator parameters in a profile along the plume centerline.
FIGURE 2. Vertical distribution of dissolved hydrocarbons in iron-reducing zone
170 m downgradient from the former UST.
FIGURE 3. COD and CAH (10 mg/L) contours at the St. Joseph, Michigan, NPL site.
FIGURE 4. Locations of the 1991 and 1992 sampling transects at the St Joseph,
Michigan, NPL site.
-------
KEYWORD LIST
Indicator parameters
intrinsic bioattenuation
risk assessment
numerical models
hydrocarbon degradation
-------
Redox Environment
Chlorinated Solvent
AH Denitri-
fication
Sulfate
Reduction
Methanogenesis
Carbon tetrachloride
CT-*CF
CT-»CXD2+CI
1,1,1-T richloroethane
TCA-»1,1-DCE
+CH3COOH
TCA-*1,1-DCA
TCA-»C02+CI
Tetrachloroethylene
PCE-»1,2-DCE
PCE-»ethene
Trichloroethylene
TCE-+1.2-DCE
TCE-+ethene
-------
Compound
Field Rate
Laboratory Rate
In Situ Rate
(d-1)
(d*1)
(d-1)
Benzene
0.0002
0.024
0.004
Toluene
0.0021
0.045
0.012
Ethylbenzene
0.0015
0.002
N.S.(a)
o-Xylene
0.0021
0.056
N.S.
m-.p-Xylene
0.0013
0.02(b)
0.014
(a) Not significant at 95% level.
(h) Only m-xylene in laboratory microcosms.
-------
2.13
2.43
2.74
_ 3.05
— 3.35
xz
g* 3.66
Q
3.96
4.27
4.57
4.88
2.43
2.74
3.05
E
3.35
jr.
Q.
d>
3.66
Q
3.96
4.27
4.57
4.88
Mesitylene
Pseudocumene
0 200 400 600 800
Concentration iigIL)
-------
Lake
Michigan
100 Meters
Chemical Oxygen Demand
(mg/l)
• Monitoring well
— Transect
Red Arrow
Highway
Transects
-------
Lake
Michigan
A
N
I
VC
DCE
H
100 Meiers
10 yg/L Contours
Red Arrow
Highway
Railroad
-1
DCE
-------
588 *
590
592
0 200 400
1 i i
N
598
-------
U.S. DEPARTMENT OF COMMERCE
Technology Administration
National Technical Information Service
Springfield, VA 22161 (703) 487-4650
-------
TECHNICAL REPORT DATA
'v (Please read Instructions on the reverse before com
1. REPORT NO, 2.
EPA/600/A-95/112
4. TITLE ANO SUBTITLE
intrinsic bioattenuation for- subsurface
RESTORATION
5. REPORT OATE
6. PERFORMING ORGANIZATION CODE
7. AUTHOR(S)
JOHN T. WILSON (1) HANADI S. & CALVIN H. WARD (2!
ROBERT C. BOREN (3)
8. PERFORMING ORGANIZATION REPORT NO.
9. PERFORMING ORGANIZATION NAME ANO ADORESS
U.S./EPA, NRMRL-ADA, P.O. BOX 1198, ADA. OK (1)
RICE UNIVERSITY, HOUSTON TX (2)
NORTH CAROLINA STATE UNIVERSITY, RALEIGH, NC (3)
10. PROGRAM ELEMENT NO.
11. CONTRACT/GRANT NO.
IN-HOUSE RPJW9
12. SPONSORING AGENCY NAME AND ADDRESS
U.S./EPA, NRMRL-ADA
SUBSURFACES PROTECTION & REMEDIATION DIVISION
P.O. BOX 1198
ADA, OK 74820
13. TYPE OF REPORT ANO PERIOD COVERED
BOOK CHAPTER
14. SPONSORING AGENCY CODE
EPA/600/15
15. SUPPLEMENTARY NOTES
16, ABSTRACT
Intrinsic bioattenuation has recently evolved as a viable remediation
alternative at a number of sites where the risk of exposure to contaminants is
within acceptable standards. Important mechanisms controlling the instrinsic
j bioattenuation include advection, dispersion, sorption, dissolution from a residual
source, and abiotic and biological transformations. Since intrinsic bioattenuation is
a plume management strategy, it requires characterizing and monitoring these
processes. Intrinsic bioattenuation involves an assessment of risks to public health
and the environment, and consequently requires prediction of the fate and transport
of contaminants at the candidate sites. This paper reviews the processes controlling
intrinsic bioremediation and summarizes case histories where intrinsic
bioattenuation has been observed at sites contaminated with petroleum
hydrocarbons and chlorinated solvents. The key steps in evaluating natural
attenuation as a remedial alternative are summarized.
17. KEY WORDS AND DOCUMENT ANALYSIS
a. DESCRIPTORS
b. IDENTIFIERS/OPEN ENDED TERMS
c. COSATi Field,Group
BIOREMEDIATION
GROUND-WATER
INTRINSIC REMEDIATION
INTRINSIC BIOATTENUATI
ON
18. DISTRIBUTION STATEMENT
RELEASE TO PUBLIC
19, SECURITY CLASS /This Repor,}
TTKFnT,A R.^TFTEP
21 NO. OP PAGES
20. SECURITY CLASS {This page-
UNCLASSIFIED
22. PRICE
EPA Form 2220—1 (R«». 4-77) previous edition is obsolete
------- |