United States
Environmental Protection
Agency
Office of
Research and
Development
Office of Solid Waste
and Emergency
Response
EPAJ540JS-99JD01
September 1 999
&EPA Ground Water Issue
Microbial Processes Affecting Monitored Natural
Attenuation of Contaminants in the Subsurface
Ann Azadpour-Keeley,1 Hugh H. Russell,2 and Guy W. Sewell3
Introduction
The EPA Regional Ground Water Forum is a group of EPA
professionals representing Regional Superfund and Resource
Conservation and Recovery Act (RCRA) Offices, committed to
the identification and resolution of ground-water issues impacting
the remediation of Superfund and RCRA sites. Innovative
technologies for subsurface remediation are being evaluated
more often for specific sites, as the limitations to conventional
technologies are recognized. The purpose of this Issue Paper
is to provide those involved in assessing remediation technologies
with some basic information regarding monitored natural
attenuation (MNA) processes, specifically in determining overall
contribution of microbial processes.
On April 21, 1999, the Office of Solid Waste and Emergency
Response (OSWER) issued Directive 9200.4-17P, titled "Use of
Monitored Natural Attenuation at Superfund, RCRA Corrective
Action, and Underground Storage Tank Sites." It specifies the
current EPA policy regarding the use of MNA for the remediation
of contaminated soil and ground water at OSWER sites. The
Directive's aim is to promote consistency in the manner in which
MNA remedies are proposed, evaluated, and approved.
Accordingly, "EPA does not consider MNA to be a presumptive
or default remedy - it is merely one option that should be
considered with other applicable remedies." Thus, during the
process of selecting a site remedy, MNA may be evaluated and
compared with other remedial technologies and chosen upon
attainment of each relevant remedy selection criteria, including
the full protection of human health and the environment, and
achieving the intended site remedial objectives within a time
frame that is comparable to the other remedial techniques. The
Directive expects that "source control and long-term performance
monitoring will be fundamental components of any MNA remedy"
(U.S. EPA, 1999).
It shouid be emphasized that this document is not intended to be
used in establishing protocol involved in natural attenuation
investigations or policies leading to the interpretation of the
results of those investigations. To that end, the reader is
referred to the EPA Protocol for chlorinated solvents (U.S. EPA,
1998a), and ASTM for petroleum hydrocarbons (ASTM, 1998).
EPA is also preparing a guidance for long term monitoring (LTM)
for MNA.
For further information contact Dr. Ann Azadpour-Keeley
(580-436-8890) atthe Subsurface Protection and Remediation
Division of the National Risk Management Research
Laboratory, Ada, Oklahoma.
Background
Interest in the natural attenuation of ground-water contaminants
has increased in recent years as the result of dealing with
complexities of subsurface systems and the inherent problems
and costs associated with more conventional remedial
technologies, such as pump-and-treat systems. There is a
growing perception in the environmental community that, under
favorable conditions, the selection of natural attenuation as a
remedy will result in significant savings in cost over more
intrusive remedial alternatives which are exacerbated by the
complex geochemical, biochemical, and hydrogeological
uncertainties which are dominant at most sites. While this
perception may or may not be true, depending upon site specific
National Risk Management Research Laboratory, ORD, U.S. EPA,
Ada, OK (formerly with Dynamac Corporation)
CHR2 Environmental Services, Inc., Oilton, OK
National Risk Management Research Laboratory, ORD, U.S. EPA,
Ada, OK
Superfund Technology Support Center for Ground Water
National Risk Management Research Laboratory
Subsurface Protection and Remediation Division
Robert S. Kerr Environmental Research Center
Ada, Oklahoma
< f '&»' -^ ~~,
Technology Innovation Office
r Offfce,of'Solid Waste and Emergency
' Re$ponse,AJS ERAt Washington,'DC
&*4 s ,« ~7 3*
Walter W.Kovalick, Jr.; Ph.D, '- ,
pir^ctor ,*- --* v " ^
Printed on Recycled Paper
-------
characteristics, like all relatively new technologies, actual cost
and performance data will be required as the remediation
alternative matures. Natural attenuation, which is also referred
to as natural assimilation, intrinsic remediation, intrinsic
bioremediation, natural recovery, or passive remediation, is the
use of natural processes to remove contaminants from soil or
ground water. Whilethe mechanisms of chemicaltransformation,
dispersion, dilution, sorption and volatilization have been
identified, aerobic and anaerobic degradation comprise the
major processes for the reduction of contaminant mass in the
subsurface. Some considerations needed for the evaluation of
these processes are depicted in Table 1.
There is little question that naturally occurring biodegradation
processes are taking place at many sites where sediments have
been contaminated (Davis et al., 1994; Lee, 1988). To those
who have examined subsurface sediment, the reduced
mineralogy (darkening) and unpleasant odors (anaerobic
metabolic products) are indicators of microbial degradation
processes. While there are limitations to natural attenuation
due to factors such as complex hydrogeology, microbial toxicity
of contaminants, and other physical, biological, chemical and
environmental factors, many of the organic compounds
introduced into the subsurface can be transformed by indigenous
microorganisms. The primary challenge in evaluating natural
attenuation is not in demonstrating that biodegradation is
occurring. This can be a relatively easy task accomplished by
determining that the production of metabolites and the loss of
electron acceptors are concomitant with the distribution of
contaminants. As with any other remedial technology, the
appropriate evaluation of natural attenuation as a remedial
alternative is to make the determination that the transformation
processes are taking place at a rate that is protective of human
health and the environment. The evaluation should include a
reasonable expectation that these processes will continue at an
acceptable rate for an acceptable period of time.
Chlorinated aliphatics are among the most widespread
contaminants in ground water (Figure 1) and soil due to their use
for degreasing, dry cleaning, and as solvents (Westrick et al.,
1984). At the same time, the impact of petroleum hydrocarbons
among the various environmental media is ubiquitous (Tiedje
and Stephens, 1988; Sharpies, 1992; Kennedy and Hutchins,
1992). Because of their prominence as environmental
contaminants, these two groups of chemicals will be used as
case examples in this issue paper.
Monitored Natural Attenuation Model
Contaminants in the subsurface partition into four phases
including sorption to the soil and aquifer solids, a free phase
which displaces water from the pore spaces, dissolved in the
Tebla 1. Soma Information Needed for Prediction of Organic Contaminant Movement and Transformation in Ground Water. (Reprinted from
Btotlcand Abiotic Transformation of HalogenatedAliphatic Compounds, T.M. Vogel, Ph.D. Thesis, Stanford University, Stanford, CA,
p. 42,1988, with permission of T.M. Vogel.)
BIOLOGICAL Ground-Water Characteristics
Ionic strength
PH
temperature
nutrients
substrate
0* N03; S04-
macro (P,S, N)
trace
organism
CHEMICAL Ground-Water Characteristics
Ionic strength
pH
temperature
NO3; SO4-, Oz
toxicants
Aquifer Characteristics
grain size
active bacteria number
Monod rate-constants
Aquifer Characteristics
potential catalysts
metals, clays
Contaminant Characteristics
potential products
toxicity
concentration
Contaminant Characteristics
potential products
concentration
HYDRAULIC
SORPTION
Contaminant Source
location
amount
rate of release
Distribution Coefficient
characteristic of
concentration
Wells
location
amount
depth
pump rates
Hydrogeologic Environment
extent of aquifer and
aquitard
characteristics of aquifer
hydraulic gradient
ground-water flow rate
Characteristics of the Aquifer Solid Contaminant Characteristics
organic carbon content
clay content
octanol/water partition
coefficient
solubility
-------
Federal Republic of Germany
Tetrachloroethene
Trichloroethene
cis-1,2-Dichloroethene
Benzene
Vinyl chloride
Trichloromethane
1,1,1-Trichloroethane
Xylenes
trans-1,2-Dichloroethene
Toluene
Ethylbenzene
Dichloromethane
S Dichlorobenzenes
Chlorobenzene
Tetrachloromethane
USA
Trichloroethene
Tetrachloroethene
trans-1,2-Dichloroethene
Trichloromethane
1,1 -Dichloroethene
Dichloromethane
1,1,1 -Trichloroethane
1,1-Dichloroethane
1,2-Dichloroethane
Phenol
Acetone
Toluene
bis-(2-ethylhexyl)-phthalate
Benzene
Vinyl chloride
50 40 30
20 10 00 10 20
Frequency of Detection (%)
30 40 50
Aliphatic Chlorinated Hydrocarbons
Aromatic Chlorinated Hydrocarbons
Aromatic Hydrocarbons
Oxygen Containing Compounds
Figure 1. The 15 most frequentlydetected organic compounds in groundwaterat waste disposal sites in Germany and the U.S. (Reprinted from
The Landfill, Baccini, P., Ed., p. 399,1989, Arneth, J.-D., Milde, G., Kerndorff, H., and Schleyer, R. with permission of Springer-Verlag,
New York, Inc., New York, NY.)
water, and vapor (Figure 2). The degree to which contaminants
partition into these phases is determined by their physical/
chemical properties or notably the sorption coefficient, Henry's
Law Constant (gas partition coefficient), octanol-water coefficient,
Figure 2. Distribution of contaminants in the subsurface. (Modified
from In-Situ Bioremediation of Ground Water and
Geological Material: A Review of Technologies, p. 2,
1993, by Norris et al., EPA/600/R-93/124.)
and solubility (water partition coefficient). The mass in each
phase will therefore depend on the preference of the
contaminants of concern for each phase.
Plume movement is dependent on the same partition coefficients,
in addition to the hydrology of the site itself. In most instances,
chemicals will prefer to partition to organic matter within
sediments, to oily or free-phase material that may completely fill
pore spaces, or be dissolved in water at solubility levels. The
mass of the contaminants in the aqueous phase is usually lower
than that in the other phases.
As water moves through pore spaces, chemicals are desorbed
from sediments or dissolved from free-phase liquids. Once in
solution, these chemicals move with the aqueous phase. Their
movement downgradient is impeded by further sorption to
sediments and biodegradation. The basis of natural attenuation
is thatthe partition of chemicals into the aqueous phase reaches
equilibrium with the processes of biological transformation at an
acceptable time and distance from the source. To understand
these biological processes, some knowledge of microbiology is
required.
Microlbial Physiology
Bacteria, although unicellular, share characteristics with most
living things. Information is encoded in DNA (deoxyribonucleic
-------
acid) and transferred through RNA (ribonucleic acid) to ribosomes
to make proteins orenzymes which are usedto operate systems
within the organism. In regard to this discussion, enzymes are
responsible forthe degradation of organic carbon, which is used by
the bacterial cell to produce both the building blocks of life and
energy. The degradation of any organic molecule, including
contaminants, requires the production and efficient utilization of
enzymes. In most instances, degradation is merely a complex
oxidation/reduction reaction. The electrons or reducing equivalents
(hydrogen or electron-transferring molecules) produced must be
transferred to a terminal electron acceptor (TEA). During the
transfer process, energy is produced which is utilized by the cell.
In regard to TEAs, bacteria are "generally grouped into three
categories:
1. Aerobic bacteria Bacteria which can only utilize
molecular oxygen as aTEA. Without molecular oxygen,
these bacteria are not capable of degradation.
2. Facultative aerobes/anaerobes Bacteria which can
utilize molecular oxygen or when oxygen concentrations
are low or nonexistent, may switch to nitrate,
manganese oxides or iron oxides as electron acceptors.
3. Anaerobes Bacteria which cannot utilize oxygen as
- an electron acceptor and for which oxygen is toxic.
Though members may utilize nitrate or other electron
acceptors, it can be said that they generally utilize
sulfate or carbon dioxide as electron acceptors.
In this discussion, three modes of contaminant degradation are
addressed including aerobic, cometabolic, and anaerobic. The
first is the case in which the contaminants, for example, petroleum
hydrocarbons, are utilized by bacteria as a sole source of
carbon. Petroleum hydrocarbons are degraded through a
series of enzymatic reactions to produce needed cellular
constituents. Electrons or reducing equivalents must be
regenerated. If a contaminant serves as a sole source of carbon
and energy, conditions must be within acceptable pH, Eh, and
temperature limits and the appropriate TEA must be present. In
this case, the rate of degradation will be determined by the rate
of dissolution of toxic end products away from the microbial
population and the rate at which the TEA is replenished. Some
of the lesser chlorinated solvents, such as dichloroethene
(DCE), may also serve as sole sources of carbon; however,
tetrachloroethene (PCE) and trichloroethene (TCE) are not
thought to serve as sole sources of carbon.
In the case of TCE (and lesser chlorinated solvents), degradation
may occur through cometabolic processes. Under aerobic
conditions, the enzymes necessary forthe degradation, however,
must be induced. Inducible enzymes are those that are not
produced unless an inducer compound is present within the
bacterial cell. Pertinent to this discussion are the inducers for
methane monooxygenase and various mono- and dioxygenase
enzymes produced by aromatic degrading bacteria.
In the presence of oxygen and methane, methanotrophic bacteria
are known to produce the enzyme methane monooxygenase
(Hanson and Hanson, 1996; Patel et al., 1982). The substrate
for this enzyme is methane, but it has been shown to have a
broad substrate specificity including chlorinated solvents (Mayer
et a!., 1988). Methanotrophs downgradient from a chlorinated
solvent event may feed on methane produced within the
anaerobic portion of the plume and cometabolically degrade
some chlorinated solvents.
Numerous authors have shown that oxygenase enzymes
produced by bacteria capable of degrading aromatjc
hydrocarbons are capable of degrading chlorinated solvents.
Aromatic compounds, such as toluene and phenol, have been
shown to induce the responsible enzymes. In contaminated
aquifers which contain both aromatic hydrocarbons and
chlorinated solvents, degradation of both may occur.
It should be noted that reports on the degradation of PCE under
aerobic conditions do not exist in the peer-reviewed literature.
The structure and oxidative state of PCE may preclude its
aerobic degradation (Chen etal., 1996). The anaerobic process
for degradation of chlorinated solvents is known as reductive
dechlorination(Bouweretal., 1981; Bouwer, 1994). Chlorinated
solvents are not utilized as a carbon source, rather as an
acceptor for electrons produced during the metabolism of other
oxidizable carbon (electron donor). Thus, this process cannot
be termed cometabolism. Electrons or reducing equivalents
formed during metabolism are accepted by the chlorinated
solvent. As an example, PCE will accept electrons or reducing
equivalents formed during metabolism. This results in the reduction
of PCE, the concomitant release of a chloride ion and the formation
of TCE. While almost any degradable carbon appears capable of
driving reductive dechlorination, in most instances, a low percentage
of the electrons or reducing equivalents are utilized in the reductive
process. This is a function of other metabolic requirements for
reducing equivalents (substrate specific) and the presence or
absence of more suitable TEAs.
The behavior of the chlorinated solvent plumes can be classified
into three types based on theirprimary substrate source (U.S. EPA,
1998a).
Type 1 Behavior Primary substrate is adequate
amount of anthropogenic organic carbon, solvent
plume degrades
Type 2 Behavior Primary substrate is adequate
amount of native organic carbon, solvent plume
degrades
Type 3 Behavior Low native organic carbon
concentrations or low anthropogenic organic carbon
concentrations, PCE, TCE, and DCE do not degrade
For complete detoxification, the parent chlorinated solvent must
be dechlorinated in a stepwise fashion to the environmentally
benign ethene. This is illustrated in Figure 3. While PCE and
TCE are readily reduced as a result of their oxidative states, the
more reduced daughter products (DCE and VC) are less prone
to reductive processes. These intermediates tend to accumulate
in anaerobic aquifers where contaminants are allowed to naturally
attenuate (Lesage et al., 1990). While this may indeed be a
function of the oxidative state of the lesser-chlorinated
compounds, it may also be a function of the concentration of
degradable organic matter within the contaminated system.
The efficacy of an anaerobic microbial population for natural
degradation of chlorinated solvents is determined by the same
environmental factors as for the other systems (i.e., pH, Eh,
temperature, osmotic potential), the presence of a carbon
source which can be readily degraded (electron donor), and a
-------
trichloroethylene ^ abiotic 112 2-tetrachloroethane
H Cl " ' (PCA)
\ /
c=c
/ \
Cl Cl
f $* ^ *^^
H H
I I
I I
a c c ci
_^~-^^^ I I
_____ -~~^^ Cl Cl
^^ Jl
. ... 1'1' , cis-1,2- -^ trans-1,2- 1,1,2-trichloroethane
dichloroethylene dichloroethylene dichloroethylene (11 2-TCA)
H C! H H H Cl
\ / \ / \ /
c=c c=c c=c
/ \ / \ / \
H Cl Cl Cl Cl H
(1,1-DCE) cis-1,2-DCE trans-1,2-DCE
\ % t?
vinyl chloride
(VC)
H H
|
H _ C _ C _ Cl
I
CI Cl .
1 ,2-dichloroethane
(1,2-DCA)
H H
Hv /H ^ (biotic or abiotic) |
c = c *
/ \
H Cl
* *
ethene
H H ,,-'
\ /^.-'
C = C
/ \ .
H H
Explanation
Jl Hydrogenolysis J/ Dichloroelimination ! Dehydrochlorination
Y
^Dominant Pathways
, H C C H
,.-'-' I
..-' CI Cl
chloroethane
H H
I ' I
H C C H
I I
Cl H
ethane
H H
II
H C C H
I I '
H H
Figures. Anaerobic degradation pathways fortrichloroethylene and 1,1,2,2-tetrachloroethane. (Modifiedfromchenetal., 1996; McCarty, 1996;
Nyer and Duffin, 1997; and Vogel et al., 1987.)
TEA other than chlorinated solvents (Vogel et al., 1987). In
addition to organic carbon and a TEA, bacteria require macro-
and micronutrients, most notably nitrogen and phosphorous, for
the production of DNA, RNA and other needed cellular
constituents. In most instances within an aquifer, sufficient
macro- and micronutrients will be available for microbial
processes. On the other hand, the availability of organic carbon
can often be the limiting factor in the continuance of these
processes. For example, during the reductive dechlorination of
chlorinated solvents, the ratio of the mass of the electron donor
to that of the contaminant ranges between 100/1 and 1000/1
(Bouwer, 1994).
Mechanisms of Biodegradation
In theory, in any environment in which microbial activity occurs,
there is a progression from aerobic to anaerobic (methanogenic)
conditions. There is a definite sequence of electron acceptors
used in this progression through distinctly different redox states
(Figure 4).
The rate, type of active microbial population, and level of activity
under each of these environments are controlled by several
factors. These include the concentration of the electron
acceptors, substrates which can be utilized by the bacteria, and
specific microbial populations leading to the progression of an
aquifer from aerobic to methanogenic conditions (Salanitro,
1993). This results in a loss of organic carbon and various
electron acceptors from the system as well as a progression in
the types and physiological activity of the indigenous bacteria.
If microbial activity is high, the aquifer environment would be
expected to progress rapidly through these conditions. The
following scenario outlines a general sequence of events in which
-------
Conversion of organic matter, represented by the model compound CH2O, in different redox environments may
be represented by the following stoichiometric reactions and the corresponding Gibbs free-energy changes at
pH equal to 7.
Methanogenic, fermentative:
CH4 + CO2 AG°(W; = -22 kcal/mol
Note: For organic matter deviating from the used model compound, the fermentation will lead to
generation ofH2, which maybe oxidized by CO 2 reduction according to:
CO2 + 4H2-> CH4 + 2H2O
Sulfate reduction:
2CH2O + SO£- + H+-*2CO2 + HS' + 2H2O ACf(W) = -25 kcal/mol
Iron (ferric) reduction:
CHgO + 4Fe(OH)3 + 8H + ^ CO2 + 4Fe2+ + 1 1 H2O AGf(W) = -28 kcal/mol
Manganese (manganic) reduction:
CH2O + 2MnO2 + 4H+ -» CO2 + 2Mn2++ 3H2O AG°(W) = -81 kcal/mol
Denitrification:
5CH2O + 4NO3" + 4H+ > 5CO2 + 2N2 + 7H2O AG>(W) = -114 kcal/mol
Aerobic respiration, oxygen reduction:
CH2O + O2 -» CO2 + H2O AG°(W;= -120 kcal/mol
These processes are microbially mediated.
Figure 4. Examples of redox reactions Involved in degradation of organic matter (expressed as the model compound CHZO) in different redox
environments. (Reprinted from Attenuation of Land fill Leachate Pollutants in Aquifers, box 1, p. 138,1994, by Christensen, T.H., P. Kjeldsen,
H-J. Albrechtsen, G. Heron, P.H. Nielsen, P.L. Bjerg, and P.E. Holm, with permission of CRC Press, Inc., Boca Raton, FL.)
aerobic metabolism of preferential carbon sources would occur
first. The carbon source may be contaminants of interest or other
more readily degradable carbon which has entered the system
previously or simultaneously with the contamination event.
Oxygen-Reducing
to Nitrate-Reducing
Conditions
Nitrate-Reducing to
Manganese-Reducing
Conditions
Once available oxygen is consumed,
active aerobic populations begin to
shift to nitrate respiration.
Denitrification will continue until
available nitrate is depleted, or
usable carbon sources become
limiting.
Once nitrate is depleted, popula-
tions which reduce manganese
may become active. Bacterial
metabolism of substrates utilized by
manganese-reducing populations
will continue until the concentration
of manganese oxide becomes
limiting.
Manganese-Reducing
to Iron-Reducing
Conditions
Iron-Reducing to
Sulfate-Reducing
Conditions
Sulfate-Reducing to
Methanogenic
Conditions
When manganese oxide becomes
limiting, iron reduction becomes
the predominant reaction mechan-
ism. Available evidence suggests
thiat iron reduction does not occur
until all manganese IV oxides are
depleted. In addition, bacterial
Mn(IV) respiration appears to be
restricted to areas where sulfate is
nearly or completely absent.
Iron reduction continues until
substrate or carbon limitations allow
sulfate-reducing bacteria to become
active. Sulfate-reducing bacteria then
dominate until usable carbon orsulfate
limitations impede their activity.
Once usable carbon or sulfate
limitations occur, methanogenic
bacteria are able to dominate.
-------
The ambient redox condition of the aquifer is important when
determining the contribution of microbial degradation to MNA
mechanisms. In the case of petroleum hydrocarbons, because
of their highly reduced condition, the preferred TEA for microbial
processes would be oxygen (Brown et al., 1996; Clark et al.,
1997). From a thermodynamic standpoint, this is the most
favorable reaction mechanism. When the soluble portion of
petroleum hydrocarbons, BTEX (benzene, toluene,
ethylbenzene, xylenes), are the contaminants of concern, an
inverse relationship between BTEX and dissolved oxygen
concentrations within a plume is indicative of the microbial
metabolism of these compounds as well as other hydrocarbons
in the mixture (Donaldson et al., 1992; Huesemann and Truex,
1996). Data availablef rom various sites indicate thatthe natural
attenuation of BTEX proceeds at higher rates under oxic
conditions than normally achieved in anoxic environments
(Figure 5), with rate constants ranging from 0.3 to 1.3 percent
per day when modeled as a first-order process (Chiang et al.,
1989; Kemblowski etal., 1987; Salanitro, 1992; and McAllister
and Chiang, 1994). Although anaerobic biodegradation of
toluene and xylene under nitrate-reducing (Barbara et al., 1992;
Schocher et al., 1991), iron-reducing (Lovley, et al., 1989;
Lovley and Lonergan, 1990), sulfate-reducing (Seller et al.,
1992a and b; Edwards and Grbid-Galic', 1992; Rabus et al.,
1993), and methanogenic (Vogel and Grbid -Galic", 1986; Bouwer
and McCarty, 1983; Edwards and Grbk5 - Galid, 1994) conditions
have been extensively reported. Until recently, unequivocal
biodegradation of benzene under strict anaerobiosis was not
demonstrated (Edwards and Grbid-Galic", 1992; Lovley et al.,
1994). According to Borden etal. (1997), even though accurate
description of anaerobic biodegradation of individual BTEX
constituents may not follow a simple first-order decay function,
biodegradation of total BTEX seems to more closely approximate
a first-order decay function.
Biodegradation of chlorinated solvents, depending on the degree
of halogenation (Figure 6), is fundamentally different from that
of petroleum hydrocarbons and other oxidized chemicals
(Wiedemeier et al., 1995). The preferred redox conditions for
the effective degradation of these chemicals is anaerobic
(exception is vinyl chloride, VC). Effective degradation of these
compounds may occur only when redox conditions are below
nitrate reducing.
Although under aerobic conditions, cometabolism of TCE by
autotrophic bacterial populations, obtained from soil and ground
water, have been demonstrated, it is generally accepted that
Oxygenated - Uncontamlnated Ground Water
Oxygenated - Uncontamlnated Ground Water
Figure 5. Plan view of a typical hydrocarbon plume undergoing
natural attenuation. (Modified from In-Situ Bioremediation
of Ground Water and Geological Material: A Review of
Technologies, p. 9-8, 1993, by Norris et al., EPA/600/R-
93/124.)
Reductive
Dechlorinatiori
Rate
cm
Carbontf
Figure 6.
Degree of Chlorination ^^^^-
Monochlorinated ^^-^-^^- Polychlorinated
0.25 > 4
Relationships between degree of chlorination and anaerobic
reductive dechlorination, aerobic degradation andsorption
onto subsurface material. (Reprinted from In-Situ
Bioremedlation of Ground Water and Geological Material:
A Review of Technologies, p. 10-19,1993, byNorrisetal.,
EPA/600/R-93/124.)
this route of removal is limited only to low concentrations of TCE.
The cometabolism of TCE proceeds in the presence of methane
(Fogel, et al., 1986; Henson et al., 1988; Wilson and Wilson,
1985), ammonia (Arciero etal., 1989), or toluene (Ensley, 1991;
Mu and Scow, 1994; McCarty and Semprini, 1994) as
cosubstrate. Due to the inherent toxicity of TCE to
microorganisms responsible for degradative process (Alvarez-
Cohen and McCarty, 1991 a; Broholm et al., 1990), and because
of the competitive inhibition between a cosubstrate and the
secondary substrate for oxygenase enzymes (Alvarez-Cohen
and McCarty, 1991 b; Hopkins et al., 1993), special attention to
concentrations of TCE and its cosubstrate is warranted. Toxic
or inhibitory effects of TCE are more serious than those of 1,1,1 -
TCA (Broholm et al., 1990).
Microcosm studies involving anaerobic biotransformation of
PCE/TCE from environmental samples including sediments
(Parsons et al., 1984), ground water (Wilson et al., 1983; Sewell
and Gibson, 1991), and soil (Kleopfer et al., 1985) have been
documented. Also, the reductive dechlorination of PCE and
other chlorinated compounds under methanogenic conditions
has been reported (Vogel and McCarty, 1985 and 1987b; Vogel
et al., 1987; Freedman and Gossett, 1989; McCarty, 1988).
Bagley and Gossett (1990) suggested that the ability of sulfate-
reducing enrichment cultures for PCE dechlorination is
apparently less than that of mixed methanogenic cultures. The
main by-product of anaerobic biodegradation of chlorinated
ethene is VC which is more toxic than the parent compounds
PCE, TCE, and DCE (Chu and Jewell, 1993). It is noted that
anaerobic reduction of VC to ethylene is a slow and inefficient
process (Freedman and Gossett, 1989). Gantzer and Wackett
(1991) determined that dechlorination of chlorinated ethenes
proceed via first-order rate constants.
The oxidation-reduction (redox) potential is a relatively simple
and inexpensive indicator of the redox state of an aquifer. If the
redox is positive, one can assume that dissolved oxygen is
present and the system has not been stressed by biological
L
-------
activity (Border) etal., 1995). If the redox potential is significantly
negative, it can be assumed that processes favored under
aerobic conditions, such as BTEX degradation, are occurring at
a substantially reduced rate. Figure 7 suggests that the redox
should be -400 mV or less and the dissolved oxygen should be
below 0.25 mg/1 before anaerobic microbial reactions could take
place for the more highly chlorinated compounds (i.e., PCE,
PCA). It should be stressed that one normally does not attempt
to determine the actual redox conditions for comparison between
different sites, rather the differences between points within a
plume. Furthermore, due to the lack of internal equilibrium (Morris
and Stumm, 1967) and the mixed Eh potentials of natural aqueous
systems, the use of any measured master Eh as an input in
equilibrium hydrogeochemical model for predicting the equilibrium
chemistry of redox reactions is misleading. Instead, measuring
certain sensitive species such as oxygen, Fe(ll), hydrogen sulfide,
or methane as qualitative guides to the redox status of the water
may generate better results (Lindberg and Runnells, 1984).
In anoxic waters, where low pH and Eh exist, the reduced form
of manganese, Mn(II) is favored (Stumm and Morgan, 1981).
Reduction of Fe(lll) and Mn(IV), due to chemical processes or
microbial metabolic reactions that couple the oxidation of organic
matters to the reduction of these chemical species, has a major
influence on the distribution of Fe(II) and Mn(IV) in aquatic
sediments, submerged soil, and ground waters (Stone and
Morgan, 1984; Burdige and Nealson, 1986; Ehrlich, 1987; Di-
Ruggieroand Gounot, 1990; Lovley, 1991). Thus, measurable
Fe(ll) or Mn(ll) may indicate suboxic conditions in the absence
of detectable oxygen concentration (Higgo et al., 1996).
In addition to establishing background conditions away from the
plume, dissolved oxygen, nitrate, manganese, iron, sulfate, and
sulfide should be measured along the axis of the plume, as well
as transverse to it, in orderto characterize biological activity with
respect to the redox state at those locations. This information
will allow an estimation of the current redox state at various
parts of the contaminated plume, thereby defining the types of
reactions that may take place.
The rate of change in the concentration of these parameters can
be useful as input to predictive models. This set of data will also
characterize the abundance of the principal electron acceptors,
oxygen, nitrate, and sulfate, to allow an estimate of how long
natural attenuation will remain a viable remedial alternative.
Another approach that may be used to indicate the terminal
electron acceptor process (TEAR) predominant in the areas
of contamination is the hydrogen (H2) concentration (Lovley
and Goodwin, 1988). Hydrogen concentrations forthe various
terminal electron acceptors are shown in Table 2 (Chapelle et
al., 1995).
1.0
2H2O
2NO3 + 12H+ + 108- N2 + 6H2O
,5 Mn02(s) + HC03 + 3H+ + 2e" -
g MnC03(s)+2H20 0.5'
FeOOH(s)+ HCO3
FeC03(s)+2H20
804 + 9H+ + 8e- MS" + 4H2O
CO2 + 8H''- + 8e- CH4 + 2H2O
2CO2 + 8H+ + 8e' CH3COOH + 2H2O
e- H2
-0.5
Aerobic
-} (Oxygen as
Electron Acceptor)
II
I
Typical Primary
:} Substrates
(Electron Donors)
Figure 7. Important electron donors and acceptors in biotransformation processes. Redox potentials data were obtained from Stumm and
Morgan (1981). (Modified from In-Situ Bioremediation of Ground Water and Geological Material: A Review of Technologies, p. 8-3,
1993, by Norris et al., EPA/600/R-93/124.)
-------
Table 2. Range of Hydrogen Concentrations fora Given Terminal
Electron-Accepting Process. (Data from Chapelle et al.,
1995.)
Terminal Electron
Accepting Process
Hydrogen (HJ Concentration
(nanomoles per liter)
Denitrification
Iron (III) Reduction
Sulfate Reduction
Methanogenesis
0.2 to 0.8
1to4
5-20
Parameters that investigators should analyze for petroleum
hydrocarbons include dissolved oxygen, nitrate, Fe(ll), sulfate,
redox potential, pH, Mn(IV), dissolved methane, and total
petroleum hydrocarbons (ASTM, 1998). The majority of these
parameters can be determined using field measurements,
Hach kits, and/or CHEMetrics test kits. Since methane is
produced after other TEAs (nitrate, iron, sulfate) are depleted,
dissolved methane data is superior to contaminant data
(Underground Storage Tank Technology Update, 1998). The
parameters for chlorinated solvents may include: temperature,
redox potential, DO, sulfide, Fe(ll), methane, ethane/ethene,
alkalinity, pH, sulfate, nitrate, chloride, dissolved organiccarbon,
and hydrogen. Since Fe(lll) may be dissolved from aquifer
matrix, Fe(ll) is measured as proxy for biodegradation due to
iron reduction. Although there may be a correlation between
sediment redness and the hematite content of soil, when a soil
sample represents a mixture of several iron species, the color
is not a useful indicator (Heron et al., 1994).
The microbial activities of a site are thus determined by the
dissolved organiccarbon, presence of macro- and micronutrients
and the TEA (Semprini et al., 1995). The presence and
concentration of each will determine not only the activity, but the
predominant population.
Different levels of QA/QC may be required for those analyses
determined in the field versus those performed under laboratory
conditions (U.S. EPA, 1996; Klusman, 1980; Shampine et al.,
1992; Koterba et al., 1996 ). For example, dissolved iron and
oxygen, redox, and temperature must be determined on-site
(Shelton, 1994; Wood, 1981) using field test kits because of the
deterioration that would normally occur between the time, of
sample collection and that of arrival at the laboratory. On the
other hand, parameters such as metals, organics (Shelton and
Capel, 1994; Fishman and Friedman, 1985), and bacteria can
be properly preserved by cooling, capping, or chemical fixation,
and thereby subjected to a higher level of QA/QC.
Factors Affecting the Demonstration of
Natural Attenuation
Hydrogeology
State and federal agencies are increasingly relying on risk
based corrective action (RBCA) and/or MNA for cleanup of
contaminated sites (Brady et al., 1998). The American Society
for Testing and Materials released an RBCA protocol two years
ago (ASTM, 1998) and recently finalized an MNA protocol. The
U.S. EPA has recently published a directive on MNA (U.S. EPA,
1999) and a protocol for chlorinated solvents (U.S. EPA, 1998a).
Since RBCA and MNA incorporate no safety factors to reduce
contaminant concentrations compared to active remediation
technologies, these strategies rely solely on accurate and high
quality hydrogeological site characterizations (Boulding, 1993a
and b; U.S. EPA, 1997a) to demonstrate adequatepublicprotection.
Adequate monitoring is one of the most important facets of
proving that natural bioremediation is occurring in the subsurface.
Many factors other than natural attenuation or bioremediation
can have an effect on the observed concentration of contaminants
at a monitoring well (Black and Hall, 1984). The infiltrating
precipitation into a system may have a profound effect on
contaminant concentrations, especially if the contaminants are
light nonaqueous phase liquids (LNAPLs) such as petroleum
hydrocarbons (Kemblowski and Chiang, 1990; Lenhard and
Parker, 1990). High rates of infiltration may lower the apparent
concentration of contaminants due to dilution. Pettyjohn (1982)
has shown that there are actually two time periods after a
precipitation event when dilution may have an effect on
monitoring. The first is a few hours after a precipitation event
where flow is through macropores or "wormholes" to the water
table, and the second is a few days after with flow through the
vadose zone. The time required for both of these events to
occurwill beafunction of the number and size of the macropores,
overall permeability of the unsaturated zone and depth to the
water table. Monitoring during these events can result in an
apparent decrease in the concentration of contaminants by
dilution, or if contaminants are present in the vadose zone, an
increase in concentrations due to their infiltration. Seasonal
variations may also occur in flow paths within an aquifer (Schmidt,
1977). These variations may cause an apparent increase or
decrease in concentration of contaminants due to dilution or
plume movement. Although dilution is considered to be a part
of natural attenuation, this reduction in concentration should not
be attributed to degradation.
A major problem with monitoring wells is that a sample from the
uncontaminated portion of the aquifer may be a composite of
contaminated water from the plume drawn into the cone of
depression along with clean water from the aquifer. This will result
in an apparent increase in contamination. The reverse is also often
possible if the monitoring well is pumped for an extended period
prior to sampling, the amount of clean water coming into the well
in relation to contaminated water will result in an apparent decrease
in the concentration of the contaminants of interest. These
problems are often exacerbated by using well screens that are too
long (greater than 2 meters), inconsistent screened intervals, and
inappropriate sampling methods (Church and Granato, 1996).
Pumping wells other than those designed for monitoring (such
as those in an interdiction field, or water supply wells for
municipalities or irrigation) may influence the movement of a
plume as well as flow lines within an aquifer. Depending on
when these wells were designed and constructed, how they are
pumped, and how they affect plume movement, apparent
decreases in concentration may be observed in monitoring
wells (Martin-Hayden and Robbins, 1997; Martin-Hayden et al.,
1991; Robbins, 1989; Robbins and Martin-Hayden, 1991).
Monitoring programs should be designed such that these
concerns are taken into account (Zeeb et al., 1999). A good
L
-------
monitoring program will require sampling (Puls and Barcelona,
1996) of not only monitoring wells which are completed into the
plume, but also monitoring wells outside of the contaminated
zone In order to establish background conditions (Figure 8).
Data from monitoring wells in the contaminated portion are then
compared to background wells. The number and location of
monitoring wells are not only determined by plume geometry
and ground-water flow (ASTM, 1991), but by the degree of
confidence required to statistically demonstrate that natural
attenuation is taking place, to estimate the rate of attenuation
processes, and to predict the time required to meet established
remediation goals. These issues will be addressed in detail in
the EPA guidance for long term monitoring (LTM) for MN A which
is in preparation.
MNA Monitoring Well Network Considerations
RBCA and MNA both rely on sentinel wells for early warning
signals of plume migration. The installation of monitoring wells
(Aller et al., 1991) to adequately identify contaminant
concentrations is of paramount importance in determining the
overall contribution of biological processes to a reduction in
eitherconcentration or mass of contamination. Monitoring wells
which are to be used to determine the contribution of natural
bioremediation to site cleanup cannot be located until sufficient
knowledge of the aquifer system is obtained (Zeeb et al., 1999).
Information that must be obtained before installation include
depth to water table, hydraulic conductivity (Molz et al., 1994)
andgradient, direction of ground-water flow, storage coefficient
orspecific yield, vertical and horizontal conductivity distribution,
direction of plume movement and the effects of any man-made
or natural influences (i.e., lagoons or seeps) on the aquifer
system. It is also important to determine if the hydraulic gradient
is affected seasonally. Also, sentinel well screen depth and
length are important. Often it is advantageous to use short
screensto minimize averaging ofverticalwaterquality differences
(Martin-Hayden and Bobbins, 1997).
The location, number (ASTM, 1995) and other pertinent data
regarding monitoring wells (U.S. EPA, 1998b) for the evaluation
of MNA should be determined on a site-specific basis. The
design of the monitoring network will be determined by the size
of the plume, site complexity, source strength, ground-water/
Contaminant
Source
MW-I
Grovndwatar
Flow
MW-7
MW-6
Plume
Boundary
| Monitoring Waff
Figure 8. Recommended groundwater monitoring well network for
demonstrating natural attenuation. (Reprinted from A
Practical Approach to Evaluating Natural Attenuation of
Contaminants In Ground Water, p. 166, 1994, by P.M.
McAllister and C.Y. Chiang. Reprinted by permission of
Ground Water Monitoring & Remediation, Westerville,
OH, Copyright 1994. All rights reserved.)
subsurface water interactions, distance to receptors and the
confidence limits each party involved wishes to place in the data
obtained. By necessity, the denser the monitoring network, the
greaterthe degree of confidence one may place in the data. The
wells should be capable of monitoring singular flow paths within
a plume's course and subsequent movement of contaminants
along these flow paths. One way to determine natural attenuation
is to determine the concentration of appropriate parameters at
one location and sample the same volume of waterforthe same
parameters at some distance downgradient. It is generally
impractical to monitorflow paths within a plume with the exception
of the plume axis which is the only flow line that can be located
with any reasonable level of certainty.
A second way to monitor for natural attenuation assumes that
the plume is in equilibrium. While less costly and less time-
consuming than monitoring single flow paths for contaminants
where a mass balance cannot be performed, the confidence
level for data obtained is significantly less when plume equilibrium
is assumed. Problems arise because plumes are never in
complete equilibrium. Monitoring wells even a few feet apart
differ significantly in observed concentrations of contaminants.
It has been suggested that internal tracers such as
trimethylbenzene (TMB), a biologically recalcitrant compound,
can increase the confidence one might place in this method.
One simply uses the difference between TMB concentrations at
the upgradient and downgradient points to measure the observed
loss which can be attributed to other factors such as dispersion.
One must be cautioned, however, when using TMB asaconservative
tracer due to its degradation under anaerobic conditions and the
resulting production of aromatic acid intermediates (Fang et al.,
1997).
To monitor the anaerobic degradation of chlorinated solvents,
monitoring of singular flow paths may not be necessary, if a
pseudo-mass balance can be performed. This method assumes
that at each monitoring point either the parent chlorinated
solvent or daughter product should be observed. Using PCE as
an example at each monitoring point, samples are analyzed for
PCE, f CE, DCE, VC, and ethene. The total concentration ;of
analytes is compared on a molar basis. Changes in total molar
concentrations are assumed to be losses due to flaws in the
monitoring system. The pseudo-mass balance method does not
account for physical loses due to dispersion, sorption, or dilution,
therefore, one would also have to utilize some method, such as a
conservativetracer, to accountforthese other processes. To place
even greater confidence in data, the time frame for analysis must
be sufficient to allow for differences in subsurface mobility of the
variouschlorinatedsolvents. They will not arrive atthe downgradient
monitoring point at the same time.
Often hazardous waste sites are monitored for natural attenuation
using monitoring wells which have previously been installed,
especially if plume equilibrium is assumed. This in itself presents
a number of problems, especially if the wells were installed without
adequate knowledge of the subsurface and plume movement. If
it can be demonstrated that contamination exists between these
wells, by using tracer study for example, they can be used as
appropriate monitoring points for assessing natural attenuation.
Reinhard and Goodman (1984) used chloride as a reference to
investigate the behavior of the trace organic compounds in leachate
plumes of two sanitary landfills. Kampbell et al. (1995) and
Wiedemeier et al. (1996) described the use of TMB to normalize for
changes in BTEX concentration due to abiotic processes of
dispersion, dilution, sorption, and volatilization.
10
-------
It is also necessary to determine what contribution physical
processes between sequential monitoring wells have on apparent
reduction of chemicals. It may also be necessary to construct
new monitoring wells which are offset to present wells or to
perform borings near existing wells.
Statistical Validity of Data forMNA
Adequate monitoring is critical (Reinhard and Goodman, 1984),
especially when considering natural attenuation as a remedial
alternative. The number of sampling points (Barcelona et al.,
1994 and 1985) and sampling rounds are often insufficient to
establish statistically valid trends, given the natural variability in
ground water quality, the variability due to pumping and sampling,
and differences between analytical laboratories. Schmidt (1977)
lists a number of conditions in which large fluctuations in water
quality may be noted, indicating that minor changes in water
chemistry may be related to sampling procedures. In order to
minimize the effects of natural seasonal variations, sampling,
and subsurface heterogeneity on ground water quality, any
natural attenuation monitoring program (Gibbons, 1994; Gilbert,
1987) should be based on a detailed statistical evaluation of
pertinent data (Hardin and Gilbert, 1993; McDonald and Erickson,
1994; O'Brien et al., 1991; O'Brien, 1997). Although it is
recognized that intensive monitoring is expensive, increasing
sampling points and frequencies, along with acceptable QA/QC
procedures, will give more statistically reliable information.
Caution should be exercised when drawing conclusions from
limited data sets, especially when attempting to model complex
situations.
The number of samples required for evaluating natural
attenuation is intrinsically a function of a preselected confidence
error and the variance of the data. Quarterly samples collected
for a year or two, for example, may not be adequate for
evaluating an overall reduction in the mass of contaminants
from a particular monitoring well. Sampling frequencies such as
this offer small windows for viewing contaminant reduction
rates. Contaminant fronts may or may not have reached the
monitoring points at the time of sampling, or as stated previously,
any number of processes such as infiltration, dilution, and
sorption may bias results when addressing only the
bioremediation component of natural attenuation.
If the reduction of contaminant mass is to be determined by
temporal trends (least squares analysis), for example, the
statistical confidence is based on the variance of the data and
the square root of the number of samples. In a least square
analysis, the correlation coefficient (r) is a function of the
degrees of freedom (df) which, in this case, is the number of
observations minus two (n-2). For example, if a well were
sampled quarterly for two years, the degrees of freedom would
be 6 (8-2) which requires a correlation coefficient of about 0.8
to demonstrate statistical significance at a 95 percent level of
confidence (an r of 1.0 denotes a perfect correlation). The point
is that one must sample with enough frequency, over a protracted
period in order to obtain a statistically meaningful correlation
between the reduction of contaminant mass and time.
MNA Degradation Rate Constant Considerations
Precision and accuracy in estimation of rate constants are
essential to conclude how quickly the ground-water plume will
be cleaned up following the source control (McNabb and Dooher,
1998). Reviews by Vogel et al. (1987a) and Howard et al. (1991)
contain a compilation of chlorinated solvents rate constants.
Experimental data on the neutral and base-catalyzed abiotic
hydrolysis rates of chlorinated ethanes and ethenes was
determined by Jeffers et al. (1989). Chapelle et al. (1996)
integrated field and laboratory data to estimate rates of petroleum
hydrocarbon biodegradation.
In the derivation of the rate constants, it is of considerable
significance to calculate concentrations at the point of
compliance, compare rates at the site to those in the literature
to determine if the site is behaving like other sites, and predict
changes caused by fluctuations in flow (Weaver et al., 1996).
Although microcosms are used as an effective tool to determine
the biodegradation potentials, the use of the microbiological
laboratory data for calculation of rate constants may be
inappropriate since they are not always representative of the
degradativerate(s)underthe field conditions (U.S. EPA, 1997b).
Results from laboratory studies may significantly over- or
underestimate biodegradation rates if environmental conditions
in the laboratory differ from conditions in the field (Borden,
1994). Rifai (1997) also points out that, although useful in
evaluating the biodegradation potential, microcosms can disrupt
the normal structure of ecosystems and prevent the direct
extrapolation of microcosm-determined biodegradation rates to
field scale. If rate constants for attenuation of chlorinated
contaminants are to be used for exposure assessments, it is
necessary to estimate the residence time of the contaminants in
the aquifer as accurately as possible (Molz and Boman, 1997).
One should also be cautioned to not substitute literature
biodegradation rates in the place of site specific values.
Often, first-order kinetics obtained from field studies are used to
approximate the degradation mechanism. Wiedemeier et al.
(1996) described two methods to estimate first-order rate
constants: (1) the use of a conservative tracer, a biologically
recalcitrant compound found in the dissolved contaminant plume;
and (2) the interpretation of a steady state contaminant plume
as proposed by Buscheck and Alcantar (1995). The later
method is founded on a one-dimensional steady state analytical
solution to the advection-dispersion equation presented by
Bear (1979).
Modeling as a Predictive Tool for MNA
MNA requires two types of models: (1) conceptual and (2) fate
and transport. In order to understand ground water-flow and
contaminant movement (Bear et al., 1992), the construction of
a three-dimensional conceptual model of the site must be an
integral partof anyMNA Work Plan. A comprehensive conceptual
model should be used as a clear and concise aid for the general
understanding of the nature of the site, the acquired sampling
results, and to indicate where additional sampling efforts should
be directed. Once a conceptual model has been accepted, a
period of monitoring is required to verify that the forecast of the
conceptual model is adequate. Onlywhensufficientquantitative
site characterization data are generated and the conceptual
model is well developed, can an appropriate analytical or
numerical fate and transport model be chosen for the site.
Care should be exercised when choosing the models to predict
the fate and transport of contaminants in the subsurface. This
may also include the use of screening models such as
BIOSCREEN and BIOCHLOR (Newell et al.; 1996 and 1998).
Appropriateness of the model to the actual hydrogeologic
11
-------
situation, assumptions, limitations, and manner of application
are all considerations (Corapcioglu and Baehr, 1987; Carey et
al., 1988). The overall effectiveness of the model for predicting
fate and transport of contaminants at a particular site depends
on all of these factors. Caution should also be exercised when
drawing conclusions from limited data sets, especially when
modeling complex situations. The validity of the input data is
critical in determining the accuracy of predictions made with the
mode).
Once a model is chosen, it may be applied using the site data
(Hunt et a!., 1988) and calibrated. Calibration is a process of
careful modification of site hydrogeologic or contaminant
transport parameters over numerous simulations to identify a
set of parameters which generate simulation results which
closely match field measured values of hydraulic head for the
flow model and contaminant concentrations for the transport
model. Calibration of the model results to observed values
requires that an acceptable range of error be identified for each
calibration target. This range will depend on the model purpose
and also on the amount and reliability of the field data (Kresec,
1997). Once calibrated, the fate and transport model can then
be used to predict the future extent and concentration of a
dissolved contaminant plume by simulating the combined effects
of advection, dispersion, sorption, and biodegradation (Rifai et
al., 1989).
Many times during calibration, if a model does not fit observed
concentrations, it is assumed that the biodegradation coefficient
Is the proper parameter to be adjusted. Using biodegradation
to adjust a model without supporting field data should not be
done until all abiotic mechanisms for reduction are explored.
When using a model which incorporates a biodegradation term,
care should be taken to verify that assumptions made about
degradation rates and the amount and activity of biomass are
valid for the site in question. Degradation rates are sensitive to
a wide array of field conditions which have been discussed
previously. Extrapolation of laboratory derived rates to a site
can also lead to significant errors. Likewise, using models to
derive degradation rates from limited field data where abiotic
variables are not well defined can be misleading. Models can
be useful tools in determining plume movement (Mercer, 1998)
and the contribution of natural attenuation (Rifai et al., 1995) to
reductions in contaminant mass provided that all model inputs
are correct, particularly those associated with biodegradation
rates. Kinetic constants derived from laboratory microcosms or
other sites are generally not useful on a wide scale to predict
overall removal rates. Site specific degradation rates should be
developed and incorporated into a model.
Summary
The behavior of a contamination plume, whether stable,
shrinking, or expanding, is the primary evidence for the
occurrence of natural attenuation. In the majority of the cases,
historical data to indicate the status of a plume are not available.
In these events, there are at least four basic conditions which
must be present to confirm that natural attenuation processes
are taking place. These include, but are probably not limited to:
1. The points of sampling must be on flow lines from the
source of contamination or an upgradient point of
observation. It must be demonstrated that the
downgradient observations accurately reflect the abiotic
and biotic processes which have occurred between
the two points. Ideally, one would sample the same
volume of water at the downstream point that was
sampled earlier atthe source or upgradient observation
well. Since this is rarely practical, it must be assumed
that the plume is in equilibrium with respect to natural
attenuation processes between the two points of
observation. If these conditions are not satisfied, any
downgradient measurement of contaminant
concentration must be lower than the true value, and
therefore, exaggerate the effectiveness of natural
attenuation. Confidence levels may be enhanced by 1)
increasing the number of observation points and times at
fixed frequencies, and 2) use of a conservative tracer.
2. There must be a reduction in contaminant mass or
concentration. One could argue that natural attenuation
results in a reduction in contaminant concentration by
sorption, volatilization, or dilution, with the only loss of
mass being that of volatilization to the atmosphere.
Natural bioremediation, on the other hand, must result
in a reduction of mass of the contaminants of concern
by the eventual conversion to environmentally
acceptable compounds.
3. Site geochemistry must assure that conditions are
right for reduction of contaminant concentration, such
as the presence of mineral nutrients and electron
acceptors, the state of redox, temperature, and pH.
4. Daughter products of contaminants must be present,
perhaps with indicators of mineralization. For example,
claims of intrinsic or natural bioremediation need to be
supported by data including the relationship between
the mass loss and the loss of oxygen. In addition to the
biological utilization of oxygen, nitrate, and sulfate,
these natural attenuation processes often result in the
creation of by-products such as dissolved Fe(ll), Mn(ll),
HCO3-, CO3= and mesthane.
The use of MN A or passive remediation at contaminated ground-
water sites is gaining attention both within the scientific and
regulatory communities. The OSWER directive recommends
that MNA be applied concurrently with or subsequently to active
measures such as source control or active remedial technologies.
Selection of MNA as a remedy or part of a remedy can be
advantageous since it may minimize the transfer of contaminants
to other media, is less intrusive to the environment, may be
applied to all or part of a site, and overall remedial costs may be
lowerthan for an active remedy. There are several factors which
may limit the application of natural attenuation as a remedial
alternative. They are: (1) the longer time frame that may be
required to achieve remedial goals, (2) site characterization
investigations that may be more extensive and costly, (3) the
added responsibility for long-term monitoring and costs, (4)
toxicity of byproducts, (5) potential for continued contaminant
migration, and (6) required alternatives if natural attenuation
fails to meet established goals.
Under proper conditions, MNA along with source removal, long-
term monitoring, and land use restrictions might be selected
over more expensive conventional technologies. There will be
other sites where natural attenuation will not be acceptable as
a remedial alternative because of regulatory constraints or the
site conditions are not favorable for its application. In the end,
the selection of a remedial technology at a specific site will be
12
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determined by time constraints in obtaining remediation
objectives, the hydrogeology and geochemistry at the site, the
contaminants of concern, regulatory constraints, and considerations
of environmental exposure and cost.
Disclaimer
The U.S. Environmental Protection Agency through its Office of
Research and Development partially funded and collaborated
in the research described here under Contract No. 68-C4-0031
to Dynamac Corporation. It has been subjected to the Agency's
peer and administrative review and has been approved for
publication as an EPA document. Mention of trade names or
commercial products does not constitute endorsement or
recommendation for use.
Quality Assurance Statement
All research projects making conclusions or recommendations
based on environmentally related measurements and funded by
the Environmental Protection Agency are required to participate in
the Agency Quality Assurance Program. This project did not involve
physical measurements and as such did not require a QA plan.
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18
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