1

It should 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) at the 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

United States

Environmental Protection

Agency

Office of Solid Waste

and Emergency

Response

Office of

Research and

Development

Ground Water Issue

Microbial Processes Affecting Monitored Natural

Attenuation of Contaminants in the Subsurface

National Risk Management Research Laboratory

Subsurface Protection and Remediation Division

Robert S. Kerr Environmental Research Center

Ada, Oklahoma

Superfund Technology Support Center for Ground Water

Technology Innovation Office

Office of Solid Waste and Emergency

Response, US EPA, Washington, DC

Walter W. Kovalick, Jr., Ph.D.

Director

1 National Risk Management Research Laboratory, ORD, U.S. EPA,

Ada, OK (formerly with Dynamac Corporation) 2 CHR2 Environmental Services, Inc., Oilton, OK 3 National Risk Management Research Laboratory, ORD, U.S. EPA,

Ada, OK

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).

EPA/540/S-99/001

September 1999
2

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. While the mechanisms of chemical transformation,

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; Sharples, 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

Table 1. Some Information Needed for Prediction of Organic Contaminant Movement and Transformation in Ground Water. (Reprinted from

Biotic and Abiotic Transformation of Halogenated Aliphatic Compounds, T.M. Vogel, Ph.D. Thesis, Stanford University, Stanford, CA,

p. 42, 1988, with permission of T.M. Vogel.)

BIOLOGICAL Ground-Water Characteristics Aquifer Characteristics Contaminant Characteristics

ionic strength grain size potential products

pH active bacteria number toxicity

temperature Monod rate-constants concentration

nutrients

substrate

O2, NO3

-, SO4 =

 macro (P,S, N)

trace

organism

CHEMICAL Ground-Water Characteristics Aquifer Characteristics Contaminant Characteristics

ionic strength potential catalysts potential products

pH metals, clays concentration

temperature

NO3

-, SO4 =, O2

toxicants

HYDRAULIC Contaminant Source Wells Hydrogeologic Environment

location location extent of aquifer and

amount amount aquitard

rate of release depth characteristics of aquifer

pump rates hydraulic gradient

ground-water flow rate

SORPTION Distribution Coefficient Characteristics of the Aquifer Solid Contaminant Characteristics

characteristic of organic carbon content octanol/water partition

 concentration clay content coefficient

solubility
3

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,

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 that the 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.

Microbial Physiology

Bacteria, although unicellular, share characteristics with most

living things. Information is encoded in DNA (deoxyribonucleic

Figure 1. The 15 most frequently detected organic compounds in groundwater at 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.)

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.)

Source

Vapor

Phase

 Capillary

 Fringe

Water

Table

Ground

Water

Dissolved Contaminants Residual

Saturation

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

Tetrachloroethene

Trichloroethene

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

1

2

3

4

5

6

7

8

9

10

11

12

13

14

15

50 40 30 20 10 0 0 10 20 30 40 50

Aliphatic Chlorinated Hydrocarbons

Aromatic Chlorinated Hydrocarbons

Aromatic Hydrocarbons

Oxygen Containing Compounds

Federal Republic of Germany USA

Frequency of Detection (%)
4

acid) and transferred through RNA (ribonucleic acid) to ribosomes

to make proteins or enzymes which are used to operate systems

within the organism. In regard to this discussion, enzymes are

responsible for the 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 a TEA. 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 for the 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 al., 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 aromatic

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 et al., 1996). The anaerobic process

for degradation of chlorinated solvents is known as reductive

dechlorination (Bouwer et al., 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 their primary 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
5

Figure 3. Anaerobic degradation pathways for trichloroethylene and 1,1,2,2-tetrachloroethane . (Modified from Chen et al., 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 macroand 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

Hydrogenolysis Dichloroelimination Dehydrochlorination

Explanation

H Cl

H Cl

C = C

1,1-

dichloroethylene

H H

Cl Cl

C = C

cis-1,2-DCE

cis-1,2-

dichloroethylene

H Cl

Cl H

C = C

trans-1,2-

dichloroethylene

(1,1-DCE) trans-1,2-DCE

H H

H Cl

C = C

vinyl chloride

(VC)

H H

H H

C = C

ethene

H

Cl

Cl

C C Cl

H

Cl

1,1,2,2-tetrachloroethane

(PCA)

H

H

Cl

C C Cl

H

Cl

1,1,2-trichloroethane

(1,1,2-TCA)

H

H

Cl

C C

H

Cl

1,2-dichloroethane

(1,2-DCA)

H

H

Cl

C C H

H

H

chloroethane

H

H

H

C C H

H

H

ethane

H Cl

Cl

C = C

trichloroethylene

(TCE) abiotic

Cl

H

*

*

 (biotic or abiotic)

 (biotic or abiotic)

*

Dominant Pathways
6

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 Once available oxygen is consumed,

to Nitrate-Reducing active aerobic populations begin to

Conditions shift to nitrate respiration.

Denitrification will continue until

available nitrate is depleted, or

usable carbon sources become

limiting.

Nitrate-Reducing to Once nitrate is depleted, populaManganese-Reducing tions which reduce manganese

Conditions may become active. Bacterial

metabolism of substrates utilized by

manganese-reducing populations

will continue until the concentration

of manganese oxide becomes

limiting.

Manganese-Reducing When manganese oxide becomes

to Iron-Reducing limiting, iron reduction becomes

Conditions the predominant reaction mechanism. Available evidence suggests

that 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-Reducing to Iron reduction continues until

Sulfate-Reducing substrate or carbon limitations allow

Conditions sulfate-reducing bacteria to become

active. Sulfate-reducing bacteria then

dominate until usable carbon or sulfate

limitations impede their activity.

Sulfate-Reducing to Once usable carbon or sulfate

Methanogenic limitations occur, methanogenic

Conditions bacteria are able to dominate.

Figure 4. Examples of redox reactions involved in degradation of organic matter (expressed as the model compound CH2O) in different redox

environments. (Reprinted from Attenuation of Landfill 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.)

Methanogenic, fermentative:

2CH2O CH3COOH CH4 + CO2 ∆G°(W) = -22 kcal/mol

Note: For organic matter deviating from the used model compound, the fermentation will lead to

generation of H2, which may be oxidized by CO2 reduction according to:

CO2 + 4H2 CH4 + 2H2O

Sulfate reduction:

 2CH2O + SO4

2- + H+ 2CO2 + HS- + 2H2O ∆G°(W) = -25 kcal/mol

Iron (ferric) reduction:

CH2O + 4Fe(OH)3 + 8H + CO2 + 4Fe2+ + 11 H2O ∆G°(W) = -28 kcal/mol

Manganese (manganic) reduction:

CH2O + 2MnO2 + 4H+ CO2 + 2Mn2++ 3H2O ∆G°(W) = -81 kcal/mol

Denitrification:

5CH2O + 4NO3

- + 4H+ 5CO2 + 2N2 + 7H2O ∆G°(W) = -114 kcal/mol

Aerobic respiration, oxygen reduction:

CH2O + O2 CO2 + H2O ∆G°(W)= -120 kcal/mol

These processes are microbially mediated.

be represented by the following stoichiometric reactions and the corresponding Gibbs free-energy changes at

pH equal to 7.

Conversion of organic matter, represented by the model compound CH2O, in different redox environments may
7

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 available from various sites indicate that the 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 et al., 1987; Salanitro, 1992; and McAllister

and Chiang, 1994). Although anaerobic biodegradation of

toluene and xylene under nitrate-reducing (Barbaro et al., 1992;

Schocher et al., 1991), iron-reducing (Lovley, et al., 1989;

Lovley and Lonergan, 1990), sulfate-reducing (Beller et al.,

1992a and b; Edwards and Grbic  ́ - Galic  ́ , 1992; Rabus et al.,

1993), and methanogenic (Vogel and Grbic  ́ - Galic  ́ , 1986; Bouwer

and McCarty, 1983; Edwards and Grbic  ́ - Galic  ́ , 1994) conditions

have been extensively reported. Until recently, unequivocal

biodegradation of benzene under strict anaerobiosis was not

demonstrated (Edwards and Grbic  ́ - Galic  ́ , 1992; Lovley et al.,

1994). According to Borden et al. (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

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 et al., 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 (AlvarezCohen and McCarty, 1991a; Broholm et al., 1990), and because

of the competitive inhibition between a cosubstrate and the

secondary substrate for oxygenase enzymes (Alvarez-Cohen

and McCarty, 1991b; 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 sulfatereducing 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

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/R93/124.)

Anaerobic Margin

F L O W

F L O W

Aerobic Margin

Aerobic Margin

Oxygenated - Uncontaminated Ground Water

Oxygenated - Uncontaminated Ground Water

Figure 6. Relationships between degree of chlorination and anaerobic

reductive dechlorination, aerobic degradation and sorption

onto subsurface material. (Reprinted from In-Situ

Bioremediation of Ground Water and Geological Material:

A Review of Technologies, p. 10-19, 1993, by Norris et al.,

EPA/600/R-93/124.)

Degree of Chlorination

Sorption

Reductive

Dechlorination

Rate

Monochlorinated Polychlorinated

0.25 4

Aerobic

Degradation

Cl#

Carbon#

Degradation Rate

Sorption onto Subsurface Material
8

activity (Borden et al., 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/l 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(II), 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(III) 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; DiRuggiero and Gounot, 1990; Lovley, 1991). Thus, measurable

Fe(II) or Mn(II) 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 order to 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 (TEAP) predominant in the areas

of contamination is the hydrogen (H2) concentration (Lovley

and Goodwin, 1988). Hydrogen concentrations for the various

terminal electron acceptors are shown in Table 2 (Chapelle et

al., 1995).

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.)

1.0

0.5

0

-0.5

{

{

Redox Potential (pH = 7) in Volts

Typical Primary

Substrates

(Electron Donors)

Aerobic

(Oxygen as

Electron Acceptor)

Anaerobic (Alternate Electron Acceptors)

Decreasing Energy Yield During Electron Transfer (Respiration)

O2 + 4H+ + 4e- 2H2O

2NO3 + 12H+ + 10e- N2 + 6H2O -

MnO2(s) + HCO3 + 3H+ + 2eMnCO3(s)+ 2H2O

-

FeOOH(s) + HCO3 + 2H+ + eFeCO3(s)+ 2H2O

-

SO4 + 9H+ + 8e- HS- + 4H2O =

CO2 + 8H+ + 8e- CH4 + 2H2O

2CO2 + 8H+ + 8e- CH3COOH + 2H2O

2H+ + 2e- H2
9

Parameters that investigators should analyze for petroleum

hydrocarbons include dissolved oxygen, nitrate, Fe(II), 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,

HachTM kits, and/or CHEMetricsTM 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(II), methane, ethane/ethene,

alkalinity, pH, sulfate, nitrate, chloride, dissolved organic carbon,

and hydrogen. Since Fe(III) may be dissolved from aquifer

matrix, Fe(II) 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 organic carbon, 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 adequate public protection.

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

occur will be a function 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

Table 2. Range of Hydrogen Concentrations for a Given Terminal

Electron-Accepting Process. (Data from Chapelle et al.,

1995.)

Terminal Electron Hydrogen (H2) Concentration

Accepting Process (nanomoles per liter)

Denitrification <0.1

Iron (III) Reduction 0.2 to 0.8

Sulfate Reduction 1 to 4

Methanogenesis 5-20
10

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 MNA 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

either concentration 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)

and gradient, direction of ground-water flow, storage coefficient

or specific 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

screens to minimize averaging of vertical water quality differences

(Martin-Hayden and Robbins, 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/

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

greater the 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 water for the same

parameters at some distance downgradient. It is generally

impractical to monitor flow 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 timeconsuming 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 as a conservative

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, TCE, 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

conservative tracer, to account for these 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

various chlorinated solvents. They will not arrive at the 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.

Contaminant

Source

Plume

Boundary

Groundwater

Flow

Upgradient

Well

Monitoring Well

MW-1

MW-2 MW-3 MW-4 MW-5 MW-6

MW-7

MW-8

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.)
11

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 for MNA

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

degradative rate(s) under the 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 part of any MNA 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. Only when sufficient quantitative

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
12

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

model.

Once a model is chosen, it may be applied using the site data

(Hunt et al., 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 at the 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(II), Mn(II),

HCO3

-

, CO3

=

,

 and methane.

The use of MNA or passive remediation at contaminated groundwater 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

lower than 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, longterm 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
13

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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