U.S. Department of the Interior

U.S. Geological Survey

Circular 1303

Toxic Substances Hydrology Program 

Prepared in cooperation with the

Strategic Environmental Research and Development Program

A Framework for Assessing the Sustainability of 

Monitored Natural Attenuation
Cover: Cyclic biodegradation of chlorinated ethenes (adapted from figure 4).
A Framework for Assessing the Sustainability of 

Monitored Natural Attenuation

By Francis H. Chapelle, John Novak, John Parker, Bruce G. Campbell, and 

Mark A. Widdowson

Prepared in cooperation with the 

Strategic Environmental Research and Development Program

Circular 1303

U.S. Department of the Interior

U.S. Geological Survey
U.S. Department of the Interior

DIRK KEMPTHORNE, Secretary

U.S. Geological Survey

Mark D. Myers, Director

U.S. Geological Survey, Reston, Virginia: 2007

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Suggested citation:

Chapelle, F.H., Novak, John, Parker, John, Campbell, B.G., and Widdowson, M.A., 2007, A framework for assessing 

the sustainability of monitored natural attenuation: U.S. Geological Survey Circular 1303, 35 p.

Library of Congress Cataloging-in-Publication Data

A framework for assessing the sustainability of monitored natural attenuation / by Francis H. Chapelle ... [et al.].

 p. cm. -- (Circular ; 1303)

 Includes bibliographical references and index.

 ISBN 978-1-4113-1774-1

 1. Hazardous wastes--Natural attenuation--Evaluation. 2. Hazardous waste site remediation--Evaluation. 3. In situ remediation--

Evaluation. 4. Groundwater--Pollution. I. Chapelle, Frank. 

 TD1052.F73 2007

 628.5--dc22

 2007007476
Foreword

Natural attenuation, a series of naturally occurring processes that transform potentially dangerous environmental contaminants into innocuous byproducts, occurs to some degree in all 

hydrologic systems. The efficiency of natural attenuation, however, varies widely for different 

hydrologic environments and for different environmental contaminants. Monitored Natural 

Attenuation (MNA) is the regulatory term for the use of these naturally occurring processes to 

help manage human-generated wastes. Good environmental stewardship requires, however, 

that MNA be applied in a technically sound manner that reliably protects humans and wildlife 

from chemical contaminants. 

Several years ago, the Strategic Environmental Research and Development Program (SERDP), a 

cooperative program of the Department of Defense, the Department of Energy, and the Environmental Protection Agency, initiated research designed to improve the scientific basis for 

evaluating MNA. Because MNA is often considered as a part of remediation systems that need 

to operate for decades or centuries, methods for understanding the long-term sustainability of 

MNA were of particular priority. SERDP, in partnership with the U.S. Geological Survey, Virginia 

Polytechnic Institute and State University, and Oak Ridge National Laboratory, initiated a project 

to assess the long-term sustainability of MNA. This report summarizes some of the findings of 

that study. 

The results of this joint study highlight opportunities and potential dangers associated with 

applying MNA to contaminants such as chlorinated solvents. The principal opportunity is that 

microorganisms capable of cycling chlorinated solvents to innocuous byproducts (carbon dioxide 

and chloride) are ubiquitous and are present in all hydrologic systems. This, in turn, shows that 

MNA can be a useful tool for managing chlorinated solvents that have been released to the 

environment. The potential danger is that chemical energy, principally in the form of organic 

matter, is required to drive these degradation processes to completion—a requirement that is 

not met in some hydrologic systems. 

This report outlines methodology for assessing the efficiency and sustainability of MNA. As 

such, it provides environmental managers a sound scientific basis for identifying sites where 

MNA is sustainable and an appropriate part of overall site remediation. Just as importantly, 

it also helps identify those sites where MNA may not be sustainable and not appropriate for 

site remediation. The principal mission of SERDP is to provide Federal facilities managers and 

regulators with sound, scientifically based information for implementing cost-effective environmental remediation. This report advances that mission.

 

Robert M. Hirsch

U.S. Geological Survey, Associate Director for Water

Bradley Smith

Strategic Environmental Research and Development Program, Executive Director

iii 

Contents

Foreword ........................................................................................................................................................iii

Abstract...........................................................................................................................................................1

Introduction.....................................................................................................................................................1

Biological Cycles and the Nature of Sustainability..................................................................................2

DDT—An Incomplete Waste-Substrate Cycle ................................................................................3

Chlorinated Ethenes in an Oxygenated Aquifer—An Incomplete 

Waste-Substrate Cycle ..........................................................................................................4

Chlorinated Ethenes in an Anoxic Aquifer—A Completed Waste-Substrate Cycle..................5

Mass Balance, Energy Balance, and the Sustainability of Natural Attenuation ................................6

Mass and Energy Balance ..................................................................................................................7

Mass Balance in Contaminated Ground-Water Systems ..............................................................8

Energy Balance in Contaminated Ground-Water Systems............................................................8

Quantifying Mass and Energy Balance......................................................................................................9

An Empirical Approach to Mass and Energy Balance ...................................................................9

A Deterministic Approach to Mass and Energy Balance ............................................................10

Conceptual Model of the Kings Bay Site ........................................................................................11

Constructing the Deterministic Model ............................................................................................13

Constraining the Mass-Balance Model..........................................................................................14

Rates of Ground-Water Flow ...................................................................................................14

Areal Recharge to the Semiconfined Aquifer.......................................................................16

Mass Balance of NAPL.............................................................................................................16

Mass Balance of Electron Donors and Acceptors ..............................................................17

Mass Balance of Dissolved Chlorinated Ethenes ................................................................17

Assessing the Sustainability of Natural Attenuation....................................................................17

NAPL Dissolution and Time of Remediation..........................................................................19

NAPL Removal and Time of Remediation ..............................................................................19

Dissolved Oxygen/Dissolved Organic Carbon Flux: Short-Term Sustainability...............19

Available Organic Carbon and Dissolved Organic Carbon Flux: 

Long-Term Sustainability.............................................................................................20

Short-Term and Long-Term Sustainability .............................................................................22

Electron Acceptor Depletion and Sustainability ..................................................................22

Conclusions...................................................................................................................................................24

References....................................................................................................................................................24

Appendix 1. Description of the Deterministic Model .........................................................................27

Appendix 2. Parameters Used to Simulate the Kings Bay Site.........................................................29
Figures

1–4. Diagrams showing­—

 1. The carbon cycle...................................................................................................................2

 2. Role of biochemical pathways such as glycolosis and the citric acid cycle 

in the carbon cycle................................................................................................................3

 3. DDT as an example of an incomplete waste-substrate cycle.......................................4

 4. Biochemical cycling of chlorinated ethenes....................................................................5

5. Photograph showing the organic-matter-rich semiconfining bed overlying the 

semiconfined aquifer ...................................................................................................................6

6–9. Diagrams showing­—

 6. Effect of mass and energy balance on the sustainability of natural attenuation.......7

 7. Interaction between a contaminant source area, a dissolved plume, 

and natural attenuation ........................................................................................................8

 8. Ways in which energy flow in ground-water systems affects the 

transformation of chlorinated ethenes ..............................................................................9

 9. The empirical approach to describing the mass balance of a system 

depends on site-specific monitoring data.......................................................................10

10. Conceptual model of the Kings Bay site in southeastern Georgia.....................................12

11. Map showing concentrations of vinyl chloride in the Kings Bay, Georgia, 

plume, November 1998...............................................................................................................13

12–14. Graphs showing—

 12. Concentrations of chlorinated ethenes along the flowpath 

of the Kings Bay plume.......................................................................................................13

 13. Concentrations of redox indicator parameters along the flowpath 

of the Kings Bay plume.......................................................................................................14

 14. Concentration changes of sulfate and chlorinated ethenes 

over time at well KBA-11-13A from 1998 through 2004 .................................................15

15–18. Maps showing­—

 15. Observed chloride concentrations in ground water at the 

Kings Bay site, November 1998.........................................................................................16

 16. Simulated and observed 1998 concentrations of chlorinated ethenes......................18

 17. The simulated and measured 2004 concentrations of chlorinated 

ethenes following source-area removal in 1998 ............................................................20

 18. Simulated 1998 PCE concentrations for DO/DOC ratios of 1/5, 5/5, and 5/1...............21

19. Diagram showing the short- and long-term components of the sustainability 

of monitored natural attenuation .............................................................................................22

20. Maps showing simulated 2050 VC concentrations for DO/DOC ratios 

of 1/5, 5/5, and 5/1 .......................................................................................................................23

Tables

1. Observed sensitivity of simulated chloride concentrations to a 

non-aqueous phase liquid (NAPL) mass.................................................................................16

2. Estimated times of remediation for different non-aqueous 

phase liquid (NAPL) masses .....................................................................................................19

3. Organic carbon content of Kings Bay sediments..................................................................23

vi
Conversion Factors

Inch/Pound to Metric

Multiply By To obtain

Length

foot (ft) 0.3048 meter (m)

mile (mi) 1.609 kilometer (km)

Volume

gallon (gal) 3.785 liter (L) 

gallon (gal) 0.003785 cubic meter (m3) 

Flow rate

foot per second (ft/s) 0.3048 meter per second (m/s)

foot per day (ft/d) 0.3048 meter per day (m/d)

foot per year (ft/yr) 0.3048 meter per year (m/yr)

cubic foot per second (ft3/s) 0.02832 cubic meter per second (m3/s)

Metric to Inch/Pound

Multiply By To obtain

Length

meter (m) 3.281 foot (ft) 

Area

square meter (m2) 0.0002471 acre 

Volume

cubic meter (m3) 264.2 gallon (gal) 

liter (L) 61.02 cubic inch (in3) 

cubic meter (m3) 35.31 cubic foot (ft3)

Flow rate

meter per day (m/d) 3.281 foot per day (ft/d)

Mass

gram (g) 0.03527 ounce, avoirdupois (oz)

kilogram (kg) 2.205 pound avoirdupois (lb)

Density

gram per cubic centimeter (g/cm3) 62.4220 pound per cubic foot (lb/ft3) 

Hydraulic conductivity

meter per day (m/d) 3.281 foot per day (ft/d) 

Temperature in degrees Celsius (°C) may be converted to degrees Fahrenheit (°F) as follows:

°F=(1.8×°C)+32

Temperature in degrees Fahrenheit (°F) may be converted to degrees Celsius (°C) as follows:

°C=(°F-32)/1.8

vii
Abbreviations and Acronyms

AOC Available organic carbon

Ar Advection

ATP Adenosine triphosphate

Br Biodegradation

CH4 Methane

CH2O Organic carbon

(CH2O)6 Glucose

Cl Chloride

CN NAPL mass in grams of NAPL per gram of sediment

CO2 Carbon dioxide

DCE Dichloroethene

DDD Dichlorodiphenyldichloroethane

DDE Dichlorodiphenyldichloroethene

DDT Dichlorodiphenyltrichloroethane

DO Dissolved oxygen

DOC Dissolved organic carbon

Dr Dispersion

EA Electron acceptors

Fe(II) Ferrous iron

Fe(III) Ferric iron

H2O Water

LOI Weight loss on ignition

LTSR Long-term sustainability ratio

mg/L Milligrams per liter

mM/L Millimoles per liter

Mn(IV) Solid-phase manganese

MNA Monitored natural attenuation

μg/L Micrograms per liter

NAC Natural attenuation capacity

NAPL Non-aqueous phase liquid

nM/L Nanomoles per liter

NO3 Nitrate

O2 Molecular oxygen

PCE Perchloroethene

Pr Plant transpiration

SEAM3D 3-dimensional sequential electron acceptor model

SO4 Sulfate

Sr Sorption

STSR Short-term sustainability ratio

TCE Trichloroethene

TOR Time of remediation

VC Vinyl chloride

Vr Volatilization

viii
Abstract

The sustainability of monitored natural attenuation (MNA) over time depends upon (1) the presence of chemical/

biochemical processes that transform wastes to innocuous byproducts, and (2) the availability of energy to drive 

these processes to completion. The presence or absence of 

contaminant-transforming chemical/biochemical processes 

can be determined by observing contaminant mass loss over 

time and space (mass balance). The energy available to drive 

these processes to completion can be assessed by measuring 

the pool of metabolizable organic carbon available in a system, 

and by tracing the flow of this energy to available electron 

acceptors (energy balance). For the special case of chlorinated ethenes in ground-water systems, for which a variety 

of contaminant-transforming biochemical processes exist, 

natural attenuation is sustainable when the pool of bioavailable 

organic carbon is large relative to the carbon flux needed to 

drive biodegradation to completion.

These principles are illustrated by assessing the sustainability of MNA at a chlorinated ethene-contaminated site in 

Kings Bay, Georgia. Approximately 1,000 kilograms of perchloroethene (PCE) was released to a municipal landfill in the 

1978–1980 timeframe, and the resulting plume of chlorinated 

ethenes migrated toward a nearby housing development. A 

numerical model, built using the sequential electron acceptor 

model code (SEAM3D), was used to quantify mass and energy 

balance in this system. The model considered the dissolution 

of non-aqueous phase liquid (NAPL) as the source of the PCE, 

and was designed to trace energy flow from dissolved organic 

carbon to available electron acceptors in the sequence oxygen > chlorinated ethenes > ferric iron > sulfate > carbon dioxide. The model was constrained by (1) comparing simulated 

and measured rates of ground-water flow, (2) reproducing the 

observed distribution of electron-accepting processes in the 

aquifer, (3) comparing observed and measured concentrations 

of chlorinated ethenes, and (4) reproducing the observed production and subsequent dilution of dissolved chloride, a final 

degradation product of chloroethene biodegradation.

Simulations using the constrained model indicated that 

an average flux of 5 milligrams per liter per day of organic 

carbon (CH2O) per model cell (25 square meters) is required 

to support the short-term sustainability of MNA. Because this 

flux is small relative to the pool of renewable organic carbon 

(about 4.7 x 107 milligrams [mg] per model cell) present in the 

soil zone and nonrenewable carbon (about 6.9 x 108 mg per 

model cell) in an organic-matter-rich sediment layer overlying 

the aquifer, the long-term sustainability of MNA is similarly 

large. This study illustrates that the short- and long-term sustainability of MNA can be assessed by:

Estimating the time required for contaminants to dissolve/

disperse/degrade under ambient hydrologic conditions 

(time of remediation).

Quantifying the organic carbon flux to the system needed 

to consume competing electron acceptors (oxygen) and 

direct electron flow toward chloroethene degradation 

(short-term sustainability).

Comparing the required flux of organic carbon to the pool 

of renewable and nonrenewable organic carbon given the 

estimated time of remediation (long-term sustainability).

These are general principles that can be used to assess the 

sustainability of MNA in any hydrologic system.

Introduction

Monitored natural attenuation (MNA) is defined as the use 

of naturally occurring contaminant degradation, dispersion, and immobilization processes to reach site-specific 

remediation goals (Wiedemeier and others, 1996; U.S. Environmental Protection Agency, 1998, 1999). In current engineering 

practice, the effectiveness of MNA is evaluated on a site-by-site 

basis by considering three lines of evidence: (1) historical monitoring data showing decreasing concentrations and/or contaminant mass over time, (2) geochemical data showing that site 

conditions favor contaminant transformation or immobilization, 

or (3) site-specific laboratory studies documenting ongoing biodegradation processes (U.S. Environmental Protection Agency, 

1998). Various field and laboratory methods for assessing 

these three lines of evidence have been developed and are currently in use (­Wiedemeier and others, 1999).

1.

2.

3.

A Framework for Assessing the Sustainability of 

Monitored Natural Attenuation

By Francis H. Chapelle,1 John Novak,2 John Parker,3 Bruce G. Campbell,1 and Mark A. Widdowson2

1

 U.S. Geological Survey.

2

 Virginia Polytechnic Institute and State University.

3

 U.S. Department of Energy.
For most sites, the effectiveness of chlorinated ethene 

natural attenuation is typically assessed over relatively short 

periods of time, often only a few months or years. When MNA 

becomes part of a long-term remediation strategy, however, it 

must be assumed that processes observed during site assessment will remain intact over the system’s operational lifetime. 

This operational lifetime depends on the length of time that 

contaminants are released from source areas to ground water 

(National Research Council, 2000), a period of time that may 

encompass decades or even centuries. This, in turn, raises 

an important question. Will the natural attenuation processes 

observed during site assessment continue with the same efficiency in the future? In other words, will MNA be sustainable 

throughout the operational life of the remediation system?

A variety of circumstances affect the sustainability of 

MNA for chlorinated ethenes (Newell and Aziz, 2004), and 

thus, its sustainability can be expected to vary from site to site. 

In spite of this variability, the sustainability of MNA at any 

particular site should reflect the same set of general principles. 

The purpose of this report is to identify these general principles, arrange them into a framework suitable for assessing 

the sustainability of MNA, and illustrate these principles using 

a field example. This framework, in turn, leads directly to 

methods for characterizing and quantifying the sustainability 

of MNA that can be applied at any site.

Biological Cycles and the 

Nature of Sustainability

Life on Earth has been sustained for approximately 

3.4 billion years. In considering the concept of sustainability, therefore, it is useful to begin with those characteristics 

of living systems that have made them sustainable for such 

long periods of time.

One characteristic of life systems is their cyclic nature. 

The use of carbon by living organisms, for example, is based 

on the alternate reduction of carbon dioxide (CO2) and water 

(H2O) to organic carbon (CH2O) with the release of molecular 

oxygen (O2) as a waste product. This oxygen, in turn, becomes 

a substrate for heterotrophic organisms that oxidize organic 

carbon back to carbon dioxide. Thus, the waste product (oxygen) of one biochemical process (photosynthesis) becomes the 

substrate (oxygen) for another biochemical process (heterotrophic metabolism). When the waste from one process becomes 

the substrate for another process, the result may be termed a 

waste-substrate cycle. Continuous, ongoing waste-substrate 

cycles, which are analogous to the action of a mechanical 

engine, are driven by the energy of the sun (fig. 1).

But while the overall stoichiometry of the carbon cycle 

is simple,

CO2 + H2O  CH2O + O2 CO2 + H2O , (1)

the underlying biochemistry is enormously complex. Thousands of biochemical pathways are involved in the production of organic matter from carbon dioxide and solar energy. 

Similarly, there are just as many biochemical pathways that 

transform organic carbon back to carbon dioxide in order to 

complete the cycle.

Two well-known examples of heterotrophic biochemical 

pathways are glycolosis and the citric acid cycle (fig. 2). Glycolosis begins with the simple sugar glucose (CH2

O)6, which is 

sequentially fermented to pyruvate and carbon dioxide, yielding chemical energy in the form of adenosine triphosphate 

(ATP). At least nine enzymatically catalyzed steps are involved 

in glycolosis. Pyruvate formed by glycolosis can then enter 

the citric acid cycle, again undergoing numerous intermediate 

steps, with the final production of additional chemical energy 

(ATP) and CO2. The combination of glycolosis and the citric 

acid cycle results in the complete mineralization of glucose to 

carbon dioxide, and the conversion of chemical energy stored 

in glucose to ATP. Numerous biochemical pathways, of which 

glycolosis and the citric acid cycle are just two examples, have 

evolved over geologic time. In the absence of these enzymatically catalyzed biochemical processes, oxygen and organic 

carbon can coexist together indefinitely at room temperature 

and not be converted to carbon dioxide.

Figure 1. The carbon cycle.

 A Framework for Assessing the Sustainability of Monitored Natural Attenuation
The carbon cycle in its present form, which has been 

sustained since the advent of biochemical photosynthesis 

2.2 billion years ago, illustrates two important characteristics 

of sustainable biological systems. The first is that biochemical 

mechanisms must be in place (glycolosis, the citric acid cycle) 

to transform the various carbon substrates that are available. 

This is not a trivial requirement. All of these biochemical 

processes took long periods of time to evolve. The presence 

of biochemical mechanisms to perform certain functions, 

therefore, is not a given. Those that are present have evolved 

and developed over time, and they typically fill a particular 

available ecological niche.

The second characteristic of these biochemical pathways is that energy must be available to drive the processes.

Biochemical pathways perform work in which organic 

molecules are assembled or disassembled, and work requires 

the utilization of energy. In the case of photosynthesis, it is 

the continuous delivery of solar energy that drives the carbonoxygen cycle. Importantly, however, chemical energy stored in 

reduced carbon compounds also is available to drive biochemical pathways such as glycolosis and the citric acid cycle 

(fig. 2) that serve to complete the carbon cycle. These considerations suggest a working definition of sustainability:

A sustainable biological process is cyclic in nature, 

is based on existing biochemical mechanisms, and 

has a source of energy available to drive it over time.

This broad definition can be applied to the natural attenuation of human-generated contaminants. If particular contaminants (chlorinated ethenes) are actively transformed to innocuous products (carbon dioxide and chloride) by biochemical 

processes that are present in natural systems (reductive dechlorination, oxidation), and if energy (organic carbon) is available 

to drive these processes to completion, then the natural attenuation of chlorinated ethenes is inherently sustainable.

The converse of this definition also is true and is equally 

important. If biochemical processes are not present to transform contaminants, or if the available energy is not sufficient 

to complete this transformation, then the natural attenuation 

of chlorinated ethenes is inherently unsustainable. These definitions of sustainable and nonsustainable natural attenuation 

are central to this report.

DDT—An Incomplete Waste-Substrate Cycle

In 1962, Rachel Carson published the book “Silent 

Spring” that described environmental damage caused by the 

unregulated use of pesticides such as DDT. This book, which 

marks the beginning of the modern environmental movement, 

showed how DDT was entering the food chain by agricultural 

spraying and was accumulating to harmful levels in some 

organisms. One example was the bioaccumulation of DDT in 

birds near the top of the food chain. In the case of bald eagles, 

DDT contributed to the production of fragile eggshells that 

tended to break during incubation of the eggs. This, in turn, 

nearly led to the extinction of bald eagles in the continental 

United States. If the use of DDT continued, Carson argued, 

then more bird species would be adversely affected and 

humans would be faced with the prospect of springtime in 

which there were no singing birds—a silent spring. Because 

“Silent Spring” has been so influential in the modern environmental movement, because DDT is a chlorinated compound, 

and because DDT was the first human-produced chemical to 

be banned for environmental reasons in the United States, it 

is useful to consider DDT in the context of natural attenuation 

and sustainability.

Dichlorodiphenyltrichloroethane (DDT) was originally 

developed during World War II as a treatment for body lice 

on humans. The bacterium Rickettsia prowazekii is part of 

the normal flora of the common body louse, and causes the 

disease typhus in humans. During World War II, unsanitary 

conditions led to the proliferation of lice in human populations, and millions of people were threatened by typhus. 

DDT in powdered form is highly toxic to body lice and other 

insects. Because it is virtually insoluble in water, however, 

DDT can be applied externally to the skin in controlled 

amounts that are not toxic to humans. The introduction of 

DDT as a de-lousing agent in 1943 and later its use to control malaria-spreading mosquitoes is estimated to have saved 

as many as 500 million human lives (National Academy of 

Science, 1970).

Figure 2. Role of biochemical pathways such as glycolosis and 

the citric acid cycle in the carbon cycle.

Biological Cycles and the Nature of Sustainability 
The later use of DDT as an agricultural pesticide, 

however, a practice that began in the late 1940s, initiated the 

adverse environmental effects described by Rachel Carson and 

others. Specifically, DDT was not observed to degrade once 

it was released into the environment, and it accumulated in 

soils and aquatic sediments. In addition, fat-soluble DDT also 

tended to bioaccumulate as it was passed up the food chain, 

eventually reaching levels that could be harmful to birds and 

mammals. These adverse effects led to bans on DDT use, first 

in Sweden (1970) and then in the United States (1972).

One way to consider the observed behavior of DDT in the 

environment is in the context of sustainable natural attenuation. Even though DDT has not been used in the United States 

in 35 years, it remains a common contaminant of aquatic sediments (Van Metre and Mahler, 2004). The reason DDT persists in the environment has to do with its chemical structure. 

The combination of two linked chlorinated phenols with a 

chlorinated ethane is a chemical structure not commonly found 

in nature. Consequently, biochemical mechanisms capable of 

cycling DDT to carbon dioxide and chloride have not evolved 

in natural ecosystems. Soil and sediment microorganisms are 

able to partially dechlorinate DDT to dichlorodiphenyldichloroethene (DDE) and then to dichlorodiphenyldichloroethane 

(DDD). But this process attacks only the chlorinated ethane 

portion of the DDT molecule and happens only under highly 

reducing conditions. Under the aerobic conditions characteristic of many agricultural soils, biochemical mechanisms 

are not present to transform DDT to innocuous products such 

as carbon dioxide and chloride (fig. 3). Consequently, DDT, 

DDE, and DDD commonly accumulate in the environment, in 

some cases reaching toxic levels. Because of the lack of DDTtransforming biochemical pathways, the natural attenuation of 

DDT in aerobic soils is by definition unsustainable.

The lack of sustainable DDT natural attenuation is due to 

the lack of specific chemical/biochemical mechanisms available for transforming the waste. Equally important, however, 

is the case where transforming biochemical mechanisms are 

present, but where sufficient energy to complete the cycle is 

not available. An example of this behavior is given by highly 

chlorinated ethenes such as perchloroethene (PCE) and trichloroethene (TCE).

Chlorinated Ethenes in an Oxygenated Aquifer—

An Incomplete Waste-Substrate Cycle

In the late 1970s, an unusually large number of children 

living in Woburn, Massachusetts, developed childhood leukemia, and several of these children eventually died. Painstaking 

investigation finally showed that the incidence of leukemia 

was associated with drinking water produced from two municipal wells drilled into the underlying buried-valley aquifer 

(Harr, 1995). Analysis of the well water showed that it was 

contaminated by PCE and TCE. While both of these chemicals had been used for years as solvents and for dry cleaning, 

few Americans had ever heard of them. Furthermore, the fact 

that these contaminants had migrated to Woburn’s municipal 

wells led many people to the logical assumption that PCE and 

TCE, like DDT, did not biodegrade in the environment. After 

all, PCE and TCE, like DDT, did not occur naturally in the 

environment and were manufactured exclusively by human 

chemical technology. Why would naturally occurring microorganisms have the enzymatic capability to degrade chemicals 

like PCE and TCE?

But as the extent of ground-water contamination by 

chlorinated ethenes gradually was recognized in the early 

1980s, it became apparent that other chlorinated ethenes such 

as dichloroethene (DCE) and vinyl chloride (VC) were often 

present in contaminated water. In many cases, these compounds were not present in the original contaminants, suggesting that DCE and VC were degradation products of PCE 

and TCE. Subsequent research established that PCE and TCE 

were subject to a variety of biodegradation processes including aerobic cometabolism (Wilson and Wilson, 1985) and 

anaerobic reductive dechlorination (Vogel and others, 1987). 

Later, it was also discovered that DCE and VC, the byproducts 

of reductive dechlorination, also were subject to both aerobic 

(Hartmans, 1995) and anoxic mineralization (Bradley and 

Chapelle, 1996).

Human-generated

 chemical wastes

Biochemical

 reduction

 No

 biochemical

pathway

Incomplete

waste–substrate

cycle

p c

Figure 3. DDT as an example of an incomplete 

waste-substrate cycle.

 A Framework for Assessing the Sustainability of Monitored Natural Attenuation
Numerous naturally occurring biochemical pathways 

are now recognized to exist, and these pathways are capable 

of completely transforming chlorinated ethenes to innocuous 

byproducts such as carbon dioxide and chloride (fig. 4). Furthermore, these biochemical pathways appear to have evolved 

so that microorganisms can utilize natural chlorinated organic 

compounds that are ubiquitous in aquatic environments (Grimvall, 1995; Bradley, 2000, 2003). While these naturally occurring biochemical pathways are not fully transferable to all 

human-produced chlorinated compounds (such as DDT), they 

are transferable to chlorinated ethenes. Thus, the first requirement for a completed waste-substrate cycle—­biochemical 

processes capable of transforming chlorinated ethenes to 

innocuous byproducts—has been met.

Establishing a waste-substrate cycle that is inherently 

sustainable, however, also requires the availability of energy. 

In the case of chlorinated ethenes such as PCE and TCE, this 

energy must be present in the form of organic carbon. Organic 

carbon acts as an electron donor in the metabolism of PCE 

and TCE, degrading microorganisms such as Dehalococcoides

(DiStefano and others, 1991), in which PCE and TCE act as 

electron acceptors (fig. 4). The establishment of completed, 

sustainable waste-substrate pathways, therefore, requires the 

presence of sufficient chemical energy in the form of organic 

matter to drive the reductive dechlorination of PCE and 

TCE. When sufficient organic matter is available, chlorinated 

ethenes are subject to a completed waste-substrate cycle. 

Conversely, when sufficient organic matter is not available, the 

waste-substrate cycle cannot be completed, and chlorinated 

ethenes will persist.

This explains the persistence of PCE and TCE in the 

aquifer underlying Woburn, Massachusetts, and the subsequent 

transport of these chemicals to the municipal wells. The underlying buried-valley aquifer consists of sands and gravels that 

were deposited by retreating glaciers at the end of the ice ages 

about 15,000 years ago. Partly because of the cold climatic 

conditions that prevailed at that time, and partly because the 

sediments have been extensively reworked by flowing water, 

these sands and gravels contain little natural organic carbon. 

The amount of available organic carbon present in this glacial 

aquifer, and thus the energy that is available, is not sufficient 

to drive reductive dechlorination. The most obvious indication 

of this is that ground water in the glacial outwash aquifer contains relatively high concentrations of dissolved oxygen. Thus, 

the flux of dissolved oxygen into the aquifer with recharging 

ground water exceeds oxygen consumption by carbon-utilizing 

heterotrophic bacteria. This, in turn, diverts available chemical 

energy away from chlororespiration resulting in the observed 

accumulation of PCE and TCE. Even though contaminanttransforming biochemical mechanisms may be present, the 

waste-substrate cycle for PCE and TCE is not completed in 

oxygenated systems because of the lack of available energy.

Chlorinated Ethenes in an Anoxic Aquifer—

A Completed Waste-Substrate Cycle

Cases of environmental contamination such as Woburn, 

Massachusetts, where an incomplete waste-substrate cycle 

results in the exposure of humans to contaminants, are those 

that most often lead to adverse health affects in humans. Many 

other hydrologic systems, however, do contain sufficient 

organic matter to complete the waste-substrate cycle for PCE 

and TCE.

An example of this behavior was described at the Kings 

Bay Naval Submarine Base in southeastern Georgia (Chapelle 

and Bradley, 1998). This particular aquifer was deposited 

in a temperate marine environment that supported abundant 

photosynthesis, resulting in the deposition of large amounts 

of organic carbon. As is common in marginal marine sediments, however, the distribution of this sedimentary organic 

carbon varies depending upon the sedimentary depositional 

environment. Beach sands, for example, typically are organicmatter poor since the action of wind and surf tends to remove 

sedimentary organic matter from them. Back-bay deposits, 

on the other hand, which are deposited in relatively quiet 

Human-generated

 chemical wastes

Biochemical

 reduction

 Biochemical

reduction

 Biochemical

oxidation

Complete

waste–substrate

cycle

Figure 4. Biochemical cycling of chlorinated ethenes.

Biological Cycles and the Nature of Sustainability 
lagoons between barrier islands (beaches) and the mainland, 

commonly contain large amounts of organic matter. As these 

sedimentary units are deposited by prograding shorelines, 

organic-matter-poor barrier island sands are typically overlain 

by organic-­matter-rich back-bay sediments. A photograph of 

an outcrop of the organic-matter-rich semiconfining bed, taken 

2 miles from the Kings Bay site, is shown in figure 5. As rainwater percolates through the organic-matter-rich sediments and 

recharges the underlying beach-sand aquifer, dissolved oxygen 

is consumed. In addition, sedimentary organic matter can be 

partially dissolved and transported to the aquifer. The presence 

of dissolved organic matter and the absence of oxygen result 

in conditions favorable to reductive dechlorination and leads to 

efficient biodegradation of PCE at the Kings Bay site.

The Kings Bay site is an example of a system where the 

availability of energy in the form of sedimentary organic carbon is sufficient to drive the chlorinated ethene waste-­substrate 

cycle to completion (fig. 4). This completed waste-substrate 

cycle stands in marked contrast to organic-matter-poor systems 

such as Woburn, Massachusetts, where the amount of available 

organic matter is insufficient to drive the waste-­substrate cycle 

to completion.

Figure 5. Photograph of the organic-matter-rich semiconfining bed overlying the semiconfined aquifer.

Mass Balance, Energy Balance, 

and the Sustainability of 

Natural Attenuation

The different biodegradation patterns exhibited by chlorinated ethenes at the Woburn, Massachusetts, and the 

Kings Bay, Georgia, sites illustrate how the sustainability 

of natural attenuation processes can differ in ground-water 

systems. In both systems, microorganisms with the capability 

of degrading chlorinated ethenes were present, although probably in differing abundances. At one site (Woburn) the lack of 

energy to drive reductive dechlorination caused the preservation and accumulation of TCE and PCE in ground water. In 

the other system (Kings Bay), sufficient energy was available 

to drive reductive dechlorination, and thus, natural attenuation 

has been sustained for at least 30 years. This, in turn, helped 

to restrict the migration of environmental contaminants and 

contributed to overall site remediation (Chapelle, Bradley, and 

Casey, 2005). Clearly, it is useful to distinguish hydrologic 

systems characterized by sustainable or unsustainable natural 

attenuation for chlorinated ethenes. Methods for making such 

distinctions can be derived from the twin principles of mass 

and energy balance.

 A Framework for Assessing the Sustainability of Monitored Natural Attenuation
Mass and Energy Balance

Some of the ways in which mass and energy balance 

affect the sustainability of MNA are illustrated in figure 6.

The reality is that human activities lead to the production 

of wastes, and these wastes may be released to the environment. Once released, wastes typically partition between 

soils, surface-water, and ground-water systems where they 

are subject to chemical and/or biochemical processes that (in 

some cases) transform these wastes to innocuous byproducts. 

If the rate of transformation equals or exceeds the rates of 

waste loading, then the natural attenuation processes have met 

the short-term criterion of sustainability (presence of chemical/biochemical processes capable of detoxifying chemical wastes). Conversely, if the rate of contaminant loading 

exceeds transformation rates, then the natural attenuation 

processes are by definition unsustainable. These different 

outcomes can be recognized by the degree of accumulation 

or non­‐accumulation of wastes in the environment. Thus, the 

short-term requirement for sustainable natural attenuation can 

be recognized by mass balance considerations.

The sustainability of natural attenuation processes, 

however, does not end here. In addition to having chemical/biochemical processes that transform wastes to innocuous 

byproducts, sufficient energy must be available in the system 

to drive these processes to completion. For the case of chlorinated ethenes, this energy must be provided by an external 

source of carbon. Furthermore, the energy provided by the 

organic carbon must not be diverted by competing biochemical processes (such as oxygen-based respiration) away from 

contaminant-degrading processes. For petroleum hydrocarbons, the waste itself provides the source of energy driving 

degradation. In this case, a supply of electron acceptors must 

be available to facilitate energy flow. In both cases, however, 

the long-term requirement for sustainable natural attenuation 

can be recognized by energy balance considerations (fig. 6).

Short-term

sustainability

Long-term

sustainability

Human activity

Chemical wastes released to environment

Figure 6. Effect of mass and energy balance on the sustainability of natural attenuation.

Mass Balance, Energy Balance, and the Sustainability of Natural Attenuation 
Defining the sustainability of MNA in terms of mass balance (short-term) and energy balance (long-term) immediately 

suggests ways for assessing the sustainability of MNA in different hydrologic systems. First, documenting the accumulation 

or non-accumulation of particular wastes (is the contaminant 

plume stable, shrinking, or enlarging) in a hydrologic system 

addresses short-term sustainability. Secondly, it suggests that 

evaluating energy sources available to drive waste transformation processes for the length of time required to fully transform 

contaminants addresses long-term sustainability.

Mass Balance in Contaminated 

Ground-Water Systems

The efficiency of natural attenuation can be defined in 

terms of a balance between the delivery to, and removal of, 

contaminants from ground-water systems by naturally occurring processes. This is most readily illustrated by considering 

the classic interaction between a contaminant source area, 

a dissolved plume emanating from the source area, and the 

attenuation of that plume (fig. 7). In 

most systems, contaminant mass in 

source areas is present in a variety of 

forms that include non-aqueous phase 

liquids (NAPL), chemically sorbed 

mass, or contaminant mass that has 

diffused into low-­permeability clays, 

shales, or other rocks. Regardless of 

how it is stored, however, the contaminant mass tends to dissolve from 

these sources, enters aqueous solution, and forms plumes. The sum of 

the processes contributing dissolved 

contaminant mass to water (NAPL 

dissolution, desorption, diffusion) 

represents the contaminant loading component of the mass balance. 

Once formed, contaminant masses in 

plumes are subject to transport and 

natural attenuation processes that 

include advection (Ar), dispersion 

(Dr), biodegradation (Br), sorption 

(Sr), volatilization (Vr), and plant 

transpiration (Pr) processes. The 

sum of these dispersive and removal 

processes represents the natural 

attenuation capacity component of 

the mass balance.

This conceptualization illustrates 

how the balance between contaminant loading and natural attenuation 

capacity determines the size of 

plumes, and thus their environmental 

impact. It may seem intuitively obvious that low contaminant loading 

will lead to the development of a small plume; however, high 

contaminant loading combined with a high natural attenuation capacity also can result in a relatively small plume. Thus, 

the overall size of a plume does not necessarily distinguish 

between cases of small contaminant loading and cases of 

large natural attenuation capacity. When a site is sufficiently 

well-instrumented so that the contaminant loading and natural 

attenuation capacity can be assessed by direct observation, 

these differences become evident.

Energy Balance in Contaminated 

Ground-Water Systems

The efficiency of natural attenuation processes also 

depends upon the energy balance of ground-water systems. 

Reductive dechlorination requires the transfer of electrons 

from electron donors such as organic carbon, to chlorinated 

ethenes or other chlorinated organic compounds. If sources 

of electrons are lacking in a system, or if available electrons 

Figure 7. Interaction between a contaminant source area, a dissolved plume, and 

natural attenuation.

 A Framework for Assessing the Sustainability of Monitored Natural Attenuation
are diverted away from chlorinated ethene reduction, natural 

attenuation will not be sustainable.

Examples of ways that energy flow in ground-water 

systems affects the transformation of chlorinated ethenes are 

illustrated in figure 8. As organic matter oxidizes, electrons 

are transferred to available electron acceptors such as oxygen, 

nitrate, chlorinated ethenes, manganese [Mn(IV)], ferric iron 

[Fe(III)], sulfate, or carbon dioxide. Because the electronegativity of these compounds differs, their efficiency as electron 

acceptors differs as well. Dissolved oxygen typically is the 

most efficient electron acceptor present in the environment, 

so microorganisms gain more energy coupling organic-matter 

oxidation to oxygen reduction than other electron acceptors. 

If oxygen is available, ambient microorganisms will preferentially use oxygen.

Once oxygen is consumed, a condition that is common 

in subsurface environments isolated from the atmosphere, 

nitrate is the next most efficient electron acceptor commonly 

available. Nitrate is followed, in succession, by solid manganese oxides, solid ferric oxyhydroxides, dissolved sulfate, 

and finally carbon dioxide (fig. 8). Highly chlorinated ethenes 

such as PCE and TCE, which contain highly electronegative 

chlorine, also are potential electron acceptors in contaminated 

ground-water systems. This chlororespiration produces DCE, 

VC, and ethene as daughter products.

The ecological succession of electron-accepting processes 

(fig. 8) channels the flow of electrons, and thus the flow of 

energy, in predictable ways. For example, if excess dissolved 

oxygen is available in an aquifer, oxygen will sequester much 

of the electron flow from organic-matter degradation, effectively short-circuiting reductive dechlorination and leaving 

PCE and TCE largely untransformed. Conversely, if Fe(III), 

sulfate, or carbon dioxide are the predominant electron acceptors present, PCE and TCE can intercept a portion of the 

overall electron flow. This, in turn, transforms PCE and TCE 

to DCE, VC, or ethene. The potential for transformation of 

PCE and TCE, therefore, depends both on a source of electron 

donors capable of supporting microbial metabolism, and conditions that allow chlorinated ethenes to intercept part of the 

overall electron flow. In other words, the sustainable transformation of chlorinated ethenes depends upon the energy balance 

of the system.

Quantifying Mass and Energy Balance

There are two philosophically distinct, but equally useful, approaches to quantifying mass and energy balance 

in ground-water systems. The first approach is empirical in 

nature and uses monitoring data to directly calculate the mass 

loading of contaminants to a system and the system’s natural attenuation capacity. Similarly, the energy flow through 

a ground-water system can be empirically determined by 

evaluating the succession of electron-accepting processes. The 

second approach is deterministic in nature, quantifying mass 

and energy balance in terms of equations that describe the 

physics and biochemistry of a system. Both approaches have 

advantages and disadvantages. In practice, it is often useful 

to combine these approaches in order to address site-specific 

questions. These approaches can be illustrated using the Kings 

Bay site in Georgia as an example.

An Empirical Approach to Mass and 

Energy Balance

The empirical approach to contaminant mass balance 

(Chapelle, Campbell, and others, 2005) is illustrated in figure 9. Given a well-characterized plume, the flux of contaminants across a given cross-sectional transect of an aquifer can 

be estimated from the equation:

mass fluxtransect = (Cave) x (Q), (2)

where Cave, is the average contaminant concentration (in 

micrograms per liter), and Q is the ground-water flux in units 

of length per unit time (LT-1). When this flux is summed over 

a given transect of the aquifer, it gives mass fluxtransect in units 

of mass per unit time (MT-1). The ambient natural attenuation 

capacity (NAC) of the system can then be quantified as the difference in contaminant mass flux between two transects:

NAC = (Mass fluxtransect 1) – (Mass fluxtransect 2) . (3)

In this formulation, NAC is expressed in units of contaminant 

mass attenuated per unit time (MT-1).

Figure 8. Ways in which energy flow in ground-water systems 

affects the transformation of chlorinated ethenes.

Quantifying Mass and Energy Balance 
 In principle, this empirical approach also can be used 

to assess the energy balance of a system. The large number 

of potential electron donors and electron acceptors present 

in ground-water systems makes this a difficult problem. The 

observed succession of electron-accepting processes (fig. 8), 

however, provides a framework for distributing energy flow in 

a predictable way. Empirically, energy flow can be assessed by 

observing the consumption of different electron acceptors as 

ground water flows downgradient (fig. 9). The overall flux of 

energy to particular electron-accepting processes then can be 

assessed using equations 1 and 2.

The chief advantage of using the empirical mass-balance 

approach is that no a priori assumptions need be made about 

the processes contributing to contaminant loading, contaminant attenuation, or the consumption of electron donors and 

acceptors. As a description of the present behavior of a system, 

therefore, the empirical approach is highly useful. Because the 

individual processes contributing to contaminant loading and 

natural attenuation are not explicitly considered, however, the 

empirical approach cannot predict how a system will respond if 

conditions change at some point in the future. Thus, the utility 

of the empirical approach for assessing long-term questions, 

such as the sustainability of natural attenuation, is limited.

A Deterministic Approach to Mass and 

Energy Balance

Mass and energy balance in ground-water systems also 

can be quantified by considering the individual processes that 

deliver, transport, transform, partition, or sequester mass and 

energy in ground-water systems. Numerous processes including advection, hydrodynamic dispersion, sorption, and biodegradation affect the mass and energy balance of solutes being 

transported by ground-water flow (fig. 6). The deterministic 

approach uses mathematical equations to represent each individual process, and then sums the processes to obtain a set of 

coupled mass and energy balance equations. These equations 

Figure 9. The empirical approach to describing the mass balance of a system depends on site-specific 

monitoring data.

10 A Framework for Assessing the Sustainability of Monitored Natural Attenuation
then can be solved simultaneously to evaluate the interactions of contaminants, electron donors, and electron acceptors 

over time for a variety of specified environmental conditions. 

The chief advantage of this deterministic approach is that it 

provides time-dependent mass and energy balances which then 

can be used to assess how a system may respond if conditions 

change over time. This approach is well-suited, therefore, for 

assessing the long-term sustainability of natural attenuation.

In this report, the 3-dimensional Sequential Electron 

Acceptor Model (SEAM3D) was used to quantify mass and

energy balance (Waddill and Widdowson, 1997). This model 

was selected because it explicitly calculates a mass balance 

for the amount of contaminant NAPL present in a system, 

and because it tracks the flow of energy from electron donors 

(organic carbon) and the various competing electron accepting 

processes. This, in turn, makes it possible to simultaneously 

track the dissolution of NAPL over time, and to couple the 

biotransformation of dissolved contaminant mass to the organic 

carbon and electron acceptors available in an aquifer. A detailed 

description of the SEAM3D code is provided in Appendix 1.

The description of SEAM3D given in Appendix 1 shows 

that, in order to use a deterministic approach for quantifying 

mass and energy balance at any given site, a large number 

of parameters describing NAPL dissolution, advective and 

dispersive transport of contaminants, biotransformation of 

contaminants, and the delivery and utilization of electron 

donors and acceptors are required. The model-building process 

begins with conceptualizing the hydrologic system, moves to 

estimating values for the various parameters required, and then 

constrains these estimated values by using available hydrologic, geochemical, and microbiologic data. Once the model 

has been adequately constrained, it can be used to evaluate 

how the sustainability of natural attenuation in a particular 

hydrologic system responds to various environmental conditions over time.

Conceptual Model of the Kings Bay Site

A conceptual model of the Kings Bay site, in southeastern Georgia, was developed from drilling logs and long-term 

monitoring data (fig. 10). The sediments underlying the site 

are of marginal marine origin and reflect the progradation of 

the shoreline over geologic time. Sediments exposed at land 

surface consist of fine- to medium-grained sands of wind-blown 

dune (aeolian) origin and range from 2 to 3 meters in thickness. Below these surface sands is a layer of organic-matterrich sands and clays that record sedimentation in a back-bay 

lagoon environment. This organic-matter-rich layer is typically 

3 meters thick and is areally extensive in the vicinity of the site. 

A photograph of this unit in outcrop, taken 2 miles from the site 

is shown in figure 5. Underlying the layer of organic-matter-rich 

sediments is another layer of fine- to medium-grained aeolian 

sand that ranges from 5 to 7 meters thick. This sandy layer is 

underlain by clayey sediments of marginal marine origin.

The hydrology of this site reflects the interaction of these 

geologic units. The sands exposed at land surface are relatively 

permeable and absorb most of the atmospheric precipitation 

that falls on the site. During precipitation, water saturates the 

sandy soil and percolates downward into the subsurface. For 

the first 2 or 3 meters of vertical flow, percolating recharge 

encounters soils and fine-grained quartz sands until it reaches 

the water table. The depth to the water table varies as much as 

1 to 2 meters, depending on the amount of recent rainfall. Typically, the water table resides in the sandy surface sediments 

above the organic-matter-rich layer. The organic-matter-rich 

layer is characterized by a lower hydraulic conductivity than 

the surface sands, but during and after rainfall, water moves 

downward through the organic-matter-rich layer, recharging 

the semiconfined aquifer beneath. The semiconfined aquifer 

has a higher hydraulic conductivity than the overlying organicmatter-rich layer, and ground-water flow changes from being 

predominantly vertical (in the organic-matter-rich layer) to 

horizontal (in the semiconfined layer). The semiconfined aquifer, therefore, acts as a regional “drain” moving recharge water 

from topographic highs near the landfill to discharge areas to 

the west (fig. 10).

During operation of the landfill (1978–1985), trenches 

were dug to a depth of about 8 feet, filled with municipal 

waste, and covered. Traces of these trenches are still visible 

at land surface (fig. 10). At some time during landfill operation, PCE, probably waste from a dry-­cleaning operation, was 

disposed of in the trenches. The relatively dense PCE moved 

downward through the soil, the organic-matter-rich confining bed, and the semiconfined aquifer, finally pooling on the 

underlying confining bed. Horizontally flowing ground water 

dissolved the PCE, carrying it downgradient in the direction of 

a nearby housing subdivision. Because of the ambient anoxic 

conditions present in the semiconfined aquifer, PCE was 

subject to reductive dechlorination, transforming it to TCE, 

DCE, and VC. Each of these contaminants was subject to 

natural attenuation processes including advection, dispersion, 

sorption, and biodegradation, which tended to restrict expansion of the plume. In 1998, a series of observation wells were 

installed to delineate the areal extent of the plume (fig. 11).

At that time, VC was observed 100 meters downgradient of 

the contaminant source area and impacted the nearby subdivision. Concentrations of chlorinated ethenes along the axis 

of the plume show that PCE and TCE predominated near the 

source area, and they were transformed to DCE and VC along 

the flowpath (fig. 12). Importantly, the VC concentration was 

observed to drop below detectable levels at the most downgradient well (KBA-11-37) although DCE was still detectable.

The observed patterns of chlorinated ethene biodegradation in the semiconfined aquifer reflects the redox chemistry 

of the site (fig. 13). Ground water in the semiconfined aquifer, 

including ground water outside of the contaminant plume, is 

predominantly anoxic, and contains moderate concentrations 

of ferrous iron [Fe(II)], sulfide, and methane. This chemistry 

reflects the transport of recharge water through the organicmatter-rich confining bed (fig. 10), consumption of dissolved 

oxygen by aerobic microorganisms, and the establishment of 

Quantifying Mass and Energy Balance 11
Figure 10. Conceptual model of the Kings Bay site in southeastern Georgia.

12  A Framework for Assessing the Sustainability of Monitored Natural Attenuation
anoxic conditions. Observed concentrations of dissolved iron 

and sulfide, as well as concentrations of hydrogen in the 0.3 

to 2.0 nanomoles per liter (nM/L) range (fig. 13), indicate predominantly sulfate-reducing and Fe(III)-reducing conditions 

in the semiconfined aquifer. The observed increase in methane 

concentrations may reflect methane generation in the organicmatter-rich semiconfined layer rather than active methanogenesis in the semiconfined aquifer itself. Near the landfill, 

sulfate-reducing conditions appear to predominate, and as 

ground water flows downgradient, Fe(III) reduction becomes 

increasingly predominant in the semiconfined aquifer.

The sulfate-reducing conditions near the source area 

facilitate the initial reduction of PCE and TCE to DCE and VC 

(fig. 12). Farther downgradient, however, the lack of measurable 

ethene concentrations in ground water suggests a net mineralization of DCE and VC. One possibility for this difference is that VC 

is reduced to ethene, which then mineralizes under anoxic conditions to CO2

. Another possible explanation is that VC is directly 

mineralized to CO2 with no intermediate reductive step. Because 

VC is less likely than DCE to undergo reduction, because VC is 

absent from the most downgradient ground water, and because 

DCE is present at low concentrations (about 10 micrograms per 

liter [μg/L]) in downgradient ground water, anoxic mineralization 

of VC is a possible mechanism at this site.

Constructing the Deterministic Model

The conceptual model of the Kings Bay site described 

previously (fig. 10) was used to construct a deterministic 

model of the site for the purpose of quantifying the mass and 

energy balance over time. The goals of this exercise included:

Quantifying the flux of electron donors and acceptors, 

including chlorinated ethenes, to the semiconfined aquifer 

and evaluating how this flux affects redox conditions and 

biodegradation processes.

Estimating the time required for the NAPL source to dissolve and disperse, eliminating the contaminant plume. This 

provides an estimate of how long the plume can be expected 

to persist, and thus the length of time natural attenuation 

processes will need to be sustained into the future.

Quantifying the flux of dissolved organic carbon to the 

semiconfined aquifer, quantifying the mass of renewable 

(soil zone) and nonrenewable (organic-matter-rich confining bed) sources of organic carbon in order to assess the 

long-term sustainability of natural attenuation at this site.

The deterministic mass- and energy-balance model was 

built in three steps. First, estimates of parameters describing the geologic, hydrologic, and microbiologic conditions 

of the site were estimated and entered as initial input to the 

model. Secondly, these model parameters were constrained to 

reproduce observed rates of ground-water flow and observed 

redox conditions at the site. Thirdly, the mass of NAPL present in the semiconfined aquifer was constrained by comparing 

observed and simulated concentrations of dissolved chloride, 

1.

2.

3.

Figure 11. Concentrations of vinyl chloride in the Kings Bay, 

Georgia, plume, November 1998.

0

500

1,000

1,500

2,000

2,500

3,000

3,500

0 25 50 100 125 150 175 200

DISTANCE ALONG FLOW PATH, IN METERS

PCE

TCE

DCE

VC

CONCENTRATIONS OF CHLORINATED

ETHENES, IN MICROGRAMS PER LITER

Figure 12. Concentrations of chlorinated ethenes along the 

flowpath of the Kings Bay plume.

17

465

683

110

76

95

29

208

792

888

196 77

314

219

327 KBA-34

Approximate extent

of pre-1998

contaminant source

area

888

Vinyl chloride plume

Observation well and vinyl chloride

 concentration, in micrograms per liter

EXPLANATION

0 25 50 METERS

0 50 100 150 200 FEET

GEORGIA STATE ROAD SPUR 40

FENCE LINE

FENCE

LINE

Atlanta

Study

area

Base modified from U.S. Geological Survey

1:24,000-scale digital data

Quantifying Mass and Energy Balance 13
Figure 13. Concentrations of redox indicator parameters along 

the flowpath of the Kings Bay plume.

a realistic representation of the natural system. The large 

number of parameters involved makes it impossible to select 

a unique combination of parameters that describe the system. 

Given this challenge, the approach taken in this report was 

to constrain the parameter values using available hydrologic, 

microbiologic, and geochemical field data. The model was 

constrained in four steps:

Given observed water levels and aquifer-test estimates 

of hydraulic conductivity and porosity, the MODFLOW 

portion of the code was used to simulate rates of groundwater flow. Simulated rates of ground-water flow were 

then compared to observed rates of sulfate transport from 

a Fenton’s Reagent injection monitored for 6 years. The 

hydraulic conductivity and porosity values specified in 

the model were chosen to reproduce observed rates of 

ground-water flow.

The mass of NAPL in the contaminant source area 

initially was estimated by considering how much PCE 

might have been disposed of at the site. These estimates 

were refined by considering the observed distance of 

PCE transport downgradient from the source area, the 

distribution of chlorinated ethene daughter products, and 

observed concentrations of dissolved chloride produced 

from within the source area.

The areal recharge to the semiconfined aquifer was 

constrained by comparing the observed and simulated 

dilution of the chloride plume as it is transported downgradient from the source area.

The transport of dissolved oxygen and organic carbon 

substrate to the semiconfined aquifer was constrained by 

comparing observed and simulated distribution of redox 

conditions, and by comparing observed and simulated 

concentrations of dissolved organic carbon.

Rates of Ground-Water Flow

The first step in obtaining a realistic mass balance of 

a ground-water system is to have an accurate estimate of 

ground-water flow rates. These estimates are obtained by solving the ground-water flow equation for appropriate boundary 

and initial conditions, and then applying Darcy’s equation to 

obtain ground-water velocity. In the SEAM3D code, these 

calculations are made by using the MODFLOW ground-water 

flow model. The hydrologic parameters used to simulate 

ground-water flow at this site are provided in Appendix 2. 

Initial estimates of these parameters were made from aquifer 

tests conducted at the site (Leeth, 1999), and were adjusted so 

that the observed head distribution at the site was reproduced 

by MODFLOW simulation. The observed head distribution could be simulated by using constant-head nodes at the 

upgradient and downgradient boundaries to simulate lateral 

flow into and out of the system, assuming an areally constant 

value for hydraulic conductivity, and assuming a constant flux 

1.

2.

3.

4.

14  A Framework for Assessing the Sustainability of Monitored Natural Attenuation

the final degradation product of chloroethene transformation. 

Finally, rates of dissolved chlorinated ethene biotransformation were constrained by comparing observed and simulated 

concentrations of dissolved PCE, TCE, DCE, and VC.

The boundary conditions, hydrologic parameters, NAPL 

dissolution parameters, geochemical, redox, and microbial 

population parameters, and chlorinated ethene biodegradation 

parameters used in the SEAM3D simulations of this system 

are provided in Appendix 2.

Constraining the Mass-Balance Model

The large number of hydrologic, microbiologic, and 

geochemical parameters needed for the coupled set of 

mass-­balance equations (Appendixes 1 and 2) solved by the 

SEAM3D code presents a significant challenge. Specifically, 

the parameter values must be adjusted so that the model gives 

-0.5

0.0

0.5

1.0

0.2

0.4

0.6

0.8

1.0

1.2

0

0.5

1.0

1.5

2.0

2.5

0

0

1

2

3

4

5

6

7

DISTANCE ALONG THE FLOWPATH, IN METERS

0 20 40 60 80 100 120 140 160

0.5

1.0

1.5

2.0

2.5

DISSOLVED OXYGEN DISSOLVED IRON DISSOLVED SULFIDE

(milligrams per liter)

DISSOLVED METHANE DISSOLVED HYDROGEN

(nanomoles per liter)
of vertical recharge to the aquifer from rainfall. While these 

hydrologic parameters yield a simulated head distribution 

that closely matched the observed head distribution, it does 

not follow that the simulated flux of water is representative of 

the system.

One way to constrain ground-water flow rates simulated 

by the model, which have a large affect on the simulated mass 

balance of the system, is to compare simulated and observed 

transport rates of a solute through the system. Because 

Fenton’s Reagent, a mixture of ferrous sulfate and hydrogen 

peroxide, was injected into the source area of this site in 1998, 

the transport of sulfate downgradient provides an estimate 

of ground-water travel time in the aquifer. Sulfate, which is 

subject to retardation, is not the ideal solute for this approach. 

Nevertheless, the observed breakthrough of sulfate in the 

downgradient wells can be used to constrain the simulated 

flux of water through the model if sorption is considered. The 

breakthrough of sulfate at well KBA-11-13A, which occurred 

approximately 1.5 years after the injection, indicates a sulfate 

velocity of about 15 meters per year (fig. 14). The transport 

of a tracer, with a sorption coefficient appropriate for sulfate, 

was simulated with the model, and a hydraulic conductivity 

needed to match the observed sulfate velocity was identified 

(Appendix 2). The model, therefore, simulates a ground-water 

velocity comparable to what is observed in the field. 

In addition to helping constrain rates of ground-water 

flow at the site, the arrival of Fenton’s Reagent-treated highconcentration sulfate water at well KBA-11-13A coincided 

with large declines in chlorinated ethene concentrations 

(fig. 14). This, in turn, shows that the Fenton’s Reagent treatment was effective in lowering contaminant concentrations at 

the source area.

Figure 14. Concentration changes of (A) sulfate and (B) chlorinated ethenes over time at 

well KBA-11-13A from 1998 through 2004.

Quantifying Mass and Energy Balance 15

1998 1999 2000 2001 2002 2003 2004 2005

TRICHLOROETHENE A

ND PERCHLOROETHENE

(micrograms per liter)

0

10

20

30

40

50

60

70

DCE (micrograms per liter)

0

100

200

300

400

500

TCE

DCE

PCE

B.

SULFA

TE (milligrams per liter)

0

200

400

600

800

1,000

1,200

1,400

1,600

1,800

VIN

YL CHLORIDE (micrograms per liter)

0

20

40

60

80

100

Sulfate

VC

A. Initial

Fenton's Reagent

Injection
Areal Recharge to the Semiconfined Aquifer

The amount of areal recharge to the semiconfined 

aquifer was constrained by observed concentration changes 

of dissolved chloride in the plume. Near the contaminant 

source area at the first line of USGS observation wells, where 

chloroethene biodegradation is relatively incomplete, average 

chloride concentrations were 10.6 mg/L (fig. 15). Chloride 

concentrations increase at the second line of wells (to the 

west) to an average of 17.6 mg/L, reflecting the biotransformation of chlorinated ethenes. Significantly, chloride concentrations decrease slightly at the third line of wells to an average 

of 14.5 mg/L. Model simulations that did not include areal 

recharge through the semiconfining bed indicated that chloride 

concentrations would continue to increase between the second 

and third line of wells. When areal recharge of low-chloride 

water to the aquifer was included, the observed downgradient 

dilution of the chloride plume could be reproduced. An areally 

constant value of recharge to the aquifer of 0.001 meter per 

day (m/d) per model cell (1 liter per day per 25 square meters 

[L/d/25 m2]) captured the observed downgradient dilution of 

the chloride plume.

Mass Balance of NAPL

The NAPL Dissolution Package of SEAM3D was 

used to simulate the dissolution of NAPL present at the site 

(Appendixes 1 and 2). Several lines of evidence were used 

to refine NAPL mass estimates. First, the distance that PCE 

was observed to move downgradient from the source area was 

captured over time at well KBA-11-13A. PCE was observed to 

reach well KBA-11-13A by the early 1980s at concentrations 

of about 20 μg/L, but by 1998, the PCE plume had regressed 

slightly back toward the source area (Southern Division, Naval 

Facilities Engineering Command, 2004). These data indicate 

that the dissolution of PCE over time has decreased and the 

size of the PCE plume was shrinking by 1998. The amount 

of NAPL placed in the modeled source area was constrained 

to reproduce the observed initial transport of PCE to well 

KBA­‐11-13A after 20 years of simulation, and to reproduce 

the subsequent recession of the PCE plume from that well.

Another line of evidence available to constrain the mass 

of NAPL initially present in the source area is observed 

concentrations of chloride near the source area. Since chloride 

is a final product of chloroethene degradation, the amount of 

chloride delivered to ground water is proportional to the mass 

of PCE dissolving and being transformed. At the Kings Bay 

site in 1998, observed concentrations of chloride at the first 

(most eastern) line of observation wells (near the plume source 

area) averaged 10.6 mg/L. At the second line of observation 

wells, chloride concentrations increased to 17.6 mg/L, reflecting the overall biodegradation of chlorinated ethenes (fig. 15). 

The final step in constraining the amount of NAPL present in 

the source area was to adjust the parameters to reproduce the 

observed chloride plume.

The mode of landfill operation suggests that between 

one and ten 55-gallon barrels of NAPL were released at 

the site. Table 1 shows the sensitivity of simulated chloride 

concentrations at the second line of wells to different assumed 

masses of NAPL. When two-thirds of a 55-gallon barrel of 

Figure 15. Observed chloride concentrations in ground water at 

the Kings Bay site, November 1998.

Table 1. Observed sensitivity of simulated chloride 

concentrations to a non-aqueous phase liquid (NAPL) mass.

[kg, kilogram; g/g, gram per gram; mg/L, milligrams per liter]

Simulated NAPL mass 

in each of three source 

area model cells 

(kg)

Simulated NAPL 

mass in each of 

three source area 

model cells 

(g/g)1

Simulated chloride 

concentration 

increase at the 

second line of 

observation wells 

(mg/L)

477 0.010 9

373 0.008 8

187 0.004 7

94 0.002 6

1

Assumes cell volume of 25 cubic meters.

16  A Framework for Assessing the Sustainability of Monitored Natural Attenuation

15.6

15.0

KBA-37

9.4

14.2

17.5

15.0

17.2

10.6

10.7

10.6

10.5

20

14.8

18.5

10.5

KBA-37

Observation well and chloride concentration,

 in milligrams per liter

Monitoring well and number

EXPLANATION

0 25 50 METERS

0 50 100 150 200 FEET

GEORGIA STATE ROAD SPUR 40

FENCE LINE

FENCE

LINE

Atlanta

Study

area

Base modified from U.S. Geological Survey

1:24,000-scale digital data
PCE (187 kg) was simulated in each of the three source area 

model cells (CN= 0.004 g/g), the resulting increase in chloride 

was approximately 7 mg/L. 

For a specified CN (NAPL mass in grams of NAPL 

per gram of sediment) of 0.008 g/g, the simulated chloride 

increase was 8 mg/L, and for a CN of 0.010 g/g, the chloride 

concentration increase was 9 mg/L. Because the observed 

increase in chloride concentration was approximately 8 mg/L, 

a value of 0.008 g/g was assigned to each source area node in 

the constrained model. This simulated NAPL mass is equivalent to about four 55-gallon barrels of PCE in the source area. 

This was simulated by assigning a NAPL mass of 373 kg of 

PCE to each of three source-area model cells. Because of the 

uncertainty associated with NAPL mass, however, simulations 

designed to estimate times of remediation used a range of 

NAPL estimates.

Mass Balance of Electron Donors and Acceptors

SEAM3D is designed specifically to track the sequential 

utilization of electron acceptors in ground-water systems. The 

code uses inhibition functions (Appendix 1) to specify the 

sequential utilization of electron acceptors in the order oxygen > chlorinated ethenes > Fe(III) > sulfate > carbon dioxide 

coupled to the oxidation of dissolved organic carbon. If dissolved oxygen is available (greater than 1 mg/L), the oxidation of organic carbon is coupled stoichiometrically to oxygen 

reduction. SEAM3D also considers solid ferric oxyhydroxides 

present in aquifer sediments as an electron acceptor. Thus, 

once oxygen has been consumed, oxidation of organic carbon 

is coupled to solid ferric hydroxides. The product of this reaction, dissolved ferrous iron [Fe(II)], is then tracked as a solute. 

Once solid ferric hydroxides have been depleted, oxidation of 

organic carbon switches to sulfate, and finally to carbon dioxide (fig. 8). These features enable the user to simulate electron 

flow between electron donors (organic carbon) and available 

electron acceptors, and thus, to quantify the energy balance of 

a system.

SEAM3D simulations then were used to evaluate the 

energy balance for the Kings Bay site. Once estimates for the 

amount of areal recharge entering the aquifer were constrained 

by the observed dilution of chloride along the flowpath, different concentrations of dissolved organic carbon were assigned to 

vertically percolating recharge, simulating the transport of carbon from the soil zone and the organic-matter-rich beds overlying the aquifer. In addition, recharging water was specified to 

contain various concentrations of electron acceptors in order to 

simulate the distribution of energy between available electron 

acceptors. Concentrations of solid ferric hydroxides present in 

the aquifer matrix were varied between 50 and 100 micrograms 

per gram (μg/g). The model was constrained by reproducing 

the observed distribution of redox processes at the site. This 

included simulated dissolved oxygen concentrations less than 

0.2 mg/L, sulfate-reducing conditions near the landfill grading to Fe(III)-reducing conditions downgradient, and dissolved 

organic carbon concentrations of approximately 2 mg/L.

Mass Balance of Dissolved Chlorinated Ethenes

The last step in constraining the mass-balance model was 

to adjust the Monod kinetic parameters for reductive dechlorination of chlorinated ethenes, and the direct oxidation of VC, 

in order to reproduce the observed 1998 concentrations of dissolved chlorinated ethenes. This was done on a trial and error 

basis, and began with comparing the observed and simulated 

distribution of the PCE plume near the source area. Once the 

PCE plume was adequately reproduced, attention turned to the 

TCE plume, then to DCE, and finally to VC. The most difficult 

part of this procedure was to reproduce the observed behavior 

of the DCE and VC plumes. Monitoring data show that the 

DCE plume extended farther downgradient then the VC plume. 

This observed pattern could not be reproduced by considering 

reductive dechlorination to be the only ongoing biodegradation process. Because VC is produced from DCE by reductive 

dechlorination, it was not possible to adjust the parameters so 

that the DCE plume was longer than the VC plume. Microcosm 

experiments conducted with aquifer material from the Kings 

Bay site indicate that VC is subject to a combination of reductive dechlorination and anoxic mineralization (Bradley, 2003). 

For this reason, kinetic parameters for the anoxic mineralization of VC were included in the model. This, in turn, allowed 

the reproduction of the observed DCE and VC plume configurations. The final simulated and observed 1998 distributions of 

PCE, TCE, DCE, and VC are shown in figure 16.

Assessing the Sustainability of 

Natural Attenuation

After the model was constrained, SEAM3D was used to 

assess the sustainability of natural attenuation at the Kings 

Bay site. This sustainability depends upon the mass balance 

between contaminant loading and attenuation, the delivery of 

dissolved organic carbon needed to transform the contaminants, and the sequestration of energy from organic carbon 

between competing electron-accepting processes. These issues 

of mass and energy balance can be quantitatively assessed 

using the deterministic model.

One approach for assessing sustainability is to use the 

model to illustrate how possible future changes in site conditions may affect contaminant behavior. This approach utilizes 

the time-dependent nature of the mass and energy balance 

equations that comprise the model, and can suggest ways of 

managing the site over time so as to maximize the sustainability of natural attenuation. The sustainability of MNA at the 

Kings Bay site was assessed in several ways.

The dissolution of NAPL and the size of the resulting chlorinated ethene plume were simulated over time to estimate 

the length of time needed for natural processes to deplete 

the contaminant source and remediate the plume. This, in 

turn, provided an estimate of the time frame over which 

natural attenuation processes would need to be sustained.

1.

Quantifying Mass and Energy Balance 17
The removal of the NAPL source was simulated to 

estimate the length of time needed for the contaminant 

plume to degrade once its source was gone. This provided 

estimates of the minimum time frame over which natural 

attenuation processes would need to be sustained.

The flux of dissolved oxygen and dissolved organic 

carbon to the aquifer was varied to estimate the minimum 

delivery of organic carbon needed to sustain reductive 

dechlorination over time.

2.

3.

The minimum flux of dissolved organic carbon required 

to sustain reductive dechlorination was compared to 

renewable and nonrenewable sources of organic carbon to 

assess the long-term sustainability of MNA at the site.

The effects of ferric oxyhydroxide concentrations in 

aquifer sediments on the sustainability of VC-degrading 

processes over time.

4.

5.

KBA-11-18 KBA-11-18

KBA-11-18

KBA-PS-03

KBA-PS-03

KBA-PS-03

KBA-PS-03

KBA-11-10B

KBA-11-10B

KBA-11-10B

KBA-11-10B

KBA-11-03B

KBA-11-03B

KBA-11-03B

KBA-11-03B

5.7500

4.7500

3.7500

2.7500

1.7500

0.7500

5.7500

4.7500

3.7500

2.7500

1.7500

0.7500

5.7500

4.7500

3.7500

2.7500

1.7500

0.7500

5.7500

4.7500

3.7500

2.7500

1.7500

0.7500

GS-14 GS-14

GS-13

GS-12

GS-11

GS-10

GS-9

GS-8

KBA-11-18

GS-14

GS-13

GS-12

GS-11

GS-10

GS-9

GS-8

GS-13

GS-12

GS-11

GS-10

GS-9

GS-8

GS-14

GS-13

GS-12

GS-11

GS-10

GS-9

GS-8

GS-7

GS-6

GS-5

GS-4

GS-7

GS-6

GS-5

GS-4

KBA-PDA-08

GS-7

GS-6

GS-5

GS-4

KBA-PDA-08

KBA-PDA-08

GS-7

GS-6

GS-5

GS-4

KBA-PDA-08

GS-3

GS-2

GS-1

GS-15

GS-3

GS-2

GS-1

GS-15

GS-3

GS-2

GS-1

GS-15

GS-3

GS-2

GS-1

GS-15

Source1

Source2

Source3

Source1

Source2

Source3

Source1

Source2

Source3

Source1

Source2

Source3

PCE TCE

DCE VC

MT3D BC Symbols

EXPLANATION

Point source/sink

MODFLOW BC Symbols

Constant head

0 25 50 METERS

0 50 100 150 200 FEET

Base modified from U.S. Geological Survey

1:24,000-scale digital data

Figure 16. Simulated and observed 1998 concentrations of chlorinated ethenes.

18  A Framework for Assessing the Sustainability of Monitored Natural Attenuation
NAPL Dissolution and Time of Remediation

Assessing the sustainability of MNA requires an estimate 

of how long reductive chlorination processes will need to be 

sustained over time. One way to estimate this is to consider 

the length of time needed for a NAPL source to dissolve and 

for the resulting plume to disperse. Simulations allowing the 

estimated mass of NAPL present in the system, as constrained 

by available hydrologic and geochemical data, were performed 

in order to make these estimates. Because the mass of NAPL 

emplaced in the aquifer is subject to uncertainty, a range of 

NAPL masses were simulated in each of three source-area 

cells ranging from 187 kg (two-thirds of a 55-gallon barrel) 

to 467 kg (one and two-thirds of a 55-gallon barrel). For these 

simulations, the time of remediation (TOR) was defined when 

concentrations of VC dropped below 5 μg/L throughout the 

model domain. The results of these simulations are summarized in table 2.

The estimated TORs from these simulations ranged from 

a minimum of 100 years to more than 500 years. These results 

indicate that TOR at this site—assuming that the NAPL is left 

in place and allowed to dissolve over time—will be greater 

than 100 years. Thus, in the absence of engineered NAPL 

removal actions, it is reasonable to conclude that MNA will 

have to be sustained at this site for approximately 300 years.

NAPL Removal and Time of Remediation

Guidance from the U.S. Environmental Protection 

Agency (1999) states that “monitored natural attenuation can 

be selected as a remediation alternative when the timeframe is 

reasonable compared to other remediation technologies.” The 

Georgia Department of Environmental Protection, the State 

agency responsible for environmental regulation at the Kings 

Bay site, concluded that any remedial strategy that would take 

longer than 50 years was not “reasonable” and could not be 

accepted. As a result, the U.S. Navy began considering other 

strategies for remediating the Kings Bay site. The technology 

selected was in situ oxidation using Fenton’s Reagent injection. Accordingly, in December 1998, a series of Fenton’s 

Reagent injections were performed and a monitoring program 

was initiated to assess the behavior of the contaminant plume 

after treatment with Fenton’s Reagent.

The waste-removal action was simulated in SEAM3D by 

instantly removing the NAPL after 20 years of simulation (at 

1998) and simulating the subsequent collapse of the contaminant plume. Because monitoring data following source area 

removal are available at this site, it is possible to compare 

simulated and measured chloroethene concentrations. Simulated and measured VC concentrations for 2004, which are 

shown for comparison in figure 17, are generally similar. The 

model was then used to simulate the length of time needed 

for VC concentrations to decline below 2 μg/L throughout the 

plume. Based on this approach, the estimated time of remediation was approximately 14 years following Fenton’s Reagent 

treatment which corresponds to the year 2012.

These two TOR estimates—for the case of no NAPL 

removal and for the case of NAPL removal—indicate that 

natural attenuation processes would need to be sustained for at 

least 100 years if NAPL were not removed, and approximately 

14 years if the NAPL were removed. In practice, the option to 

not remove the NAPL is moot since Fenton’s Reagent treatment has already occurred. Nevertheless, the model can be 

used to illustrate how different environmental conditions may 

affect the sustainability of MNA in this system.

Dissolved Oxygen/Dissolved Organic 

Carbon Flux: Short-Term Sustainability

Qualitatively, the sustained reductive dechlorination of 

chlorinated ethenes depends upon the establishment of anoxic 

conditions, at least to Fe(III) reduction, allowing part of the 

electron flow to be captured by chlororespiration (fig. 8). The 

SEAM3D model described in this report can be used to illustrate how the interaction of dissolved oxygen and dissolved 

organic carbon present in recharging ground water affects the 

establishment of anoxic conditions.

The conceptual model of the Kings Bay site (fig. 10) 

assumes that dissolved oxygen and dissolved organic carbon 

are delivered to the semiconfined aquifer by areal recharge. 

The actual delivered concentrations of these species, however, 

is uncertain. By varying the concentrations of dissolved oxygen and organic carbon being delivered to the aquifer in the 

SEAM3D model, it is possible to illustrate the effects of these 

species on redox conditions and chlorinated ethene biodegradation in the system.

During the model-constraining procedure, the most realistic comparison between simulated and measured redox conditions and chlorinated ethene biodegradation was achieved 

using concentrations of 5 mg/L dissolved oxygen [0.156 millimoles per liter (mM/L)] and 5 mg/L dissolved organic carbon 

(0.167 mM/L) in recharging water, which reflect the stoichiometry of oxygen consumption

O2 + CH2O → CO2 + H2O

Table 2. Estimated times of remediation for different 

non-aqueous phase liquid (NAPL) masses.

[kg, kilogram; g/g, gram per gram; >, greater than]

Simulated NAPL mass 

in each of three source 

area model cells 

(kg)

Simulated NAPL 

mass in each of 

three source area 

model cells 

(g/g)1

Estimated time 

of remediation 

(years)

467 0.010 >500

373 0.008 500

280 0.006 160

187 0.004 100

1

Assumes cell volume of 25 cubic meters.

Quantifying Mass and Energy Balance 19
and the slightly greater molar concentration of dissolved 

organic carbon relative to oxygen. The short-term requirement 

for the sustainability of MNA for chlorinated ethenes is that 

the molar ratio of dissolved oxygen (DO) to organic carbon 

(both dissolved and particulate; DOC) delivered to the aquifer 

be less than 1. As a result,

DO	(mol/L) STSR	(dimensionless)	=	 1

DOC	(mol/L)

< , (4)

where STSR may be called the “short-term sustainability 

ratio” and concentrations of DO and DOC are expressed in 

molar units. When STSR is less than 1, anoxic conditions 

develop and the flow of energy proceeds toward the biodegradation of chlorinated ethenes. Conversely, when STSR is 

greater than 1, oxic conditions develop and the flow of energy 

is diverted away from chloroethene biodegradation.

The effects of the dissolved oxygen/dissolved organic 

carbon (DO/DOC) ratio on the sustainability of MNA at Kings 

Bay can be illustrated using model simulations. When the DO/

DOC ratio (in units of milligrams per liter) in areal recharge 

was low (1/5 mg/L), then redox conditions were more reducing and the size of the simulated PCE plume was correspondingly small (fig. 18), reflecting active chlororespiration. When 

the DO/DOC ratio was increased to 5/5, conditions remained 

anoxic, but more electron flow was sequestered by sulfate and 

Fe(III) reduction relative to chlororespiration and the simulated 

size of the PCE plume increased. When oxic conditions were 

established by simulating a DO/DOC ratio of 5/1, simulated 

DOC concentrations declined below 1 mg/L. This further 

decreased the energy flow toward chlororespiration and caused 

the size of the simulated PCE plume to expand (fig. 18).

The most basic, short-term requirement for sustaining 

MNA for chlorinated ethenes, therefore, is that the molar flux 

of organic carbon to a system exceeds the molar flux of dissolved oxygen. At the Kings Bay site, this requirement for the 

short-term sustainability of MNA is met.

Available Organic Carbon and Dissolved Organic 

Carbon Flux: Long-Term Sustainability

Given estimates of DOC flux to the semiconfined aquifer, 

a flux sufficient to maintain short-term MNA for chlorinated 

ethenes, it becomes possible to address the long-term sustainability of MNA as well. Qualitatively, if the amount of available 

organic carbon (AOC) present in the system is large relative 

to the flux of DOC needed to sustain anoxic conditions, then 

MNA will be more sustainable in the long term. Conversely, 

if the amount of AOC is small relative to the DOC flux, then 

MNA will be less sustainable. The results of the SEAM3D 

simulations, which give estimates of the DOC flux, make it possible to evaluate the ratio between DOC flux and AOC.

For each model grid block (5 m x 5 m), the areal recharge 

rate is 0.001 m/d (1 L/d). Assuming this recharge contains 

5 mg/L of dissolved organic carbon, the DOC flux to the aquifer 

from each model cell is 5 mg/d. This flux can be compared to 

the pool of organic carbon present in the soil zone and organicmatter-rich confining bed overlying the aquifer (fig. 10).

3

35

127

<1

128

37

3

<1

61

<1

<1 <1

<1

15

1

111

49

0 25 50 METERS

0 50 100 150 200 FEET

GEORGIA STATE ROAD SPUR 40

FENCE LINE

FENCE

LINE

Atlanta

Study

area

26.7

26.3

<1

<1

<1

4.4

<1

<1

106

<1

<1

<1

6.6

25.5

2.8

<1 <1

GEORGIA STATE ROAD SPUR 40

FENCE LINE

FENCE

LINE

Approximate extent of the

Fenton’s Reagent Treatment

Approximate extent of the

Fenton’s Reagent Treatment

A. B.

1

Vinyl chloride plume

Observation well and

 vinyl chloride con-

 centration, in micro-

 grams per liter

EXPLANATION

Base modified from U.S. Geological Survey

1:24,000-scale digital data

Figure 17. The (A) simulated and (B) measured 2004 concentrations of chlorinated ethenes following source-area removal in 1998.

20 A Framework for Assessing the Sustainability of Monitored Natural Attenuation
The most basic measure of the organic carbon content of 

soils and sediments is weight loss on ignition (LOI). Because 

much of the combustible organic carbon indicated by LOI is 

not bioavailable, however, it is not an appropriate measure 

of available organic carbon. Methods for directly measuring 

organic carbon content and extractable bioavailable organic 

carbon are available (Rectanus and others, 2005); these methods 

were used to estimate AOC at the Kings Bay site. The results of 

these analyses indicate considerable variability in the amount of 

organic carbon present (table 3). In general, soil-zone sediments 

have the least amount of organic carbon, and the organic-matter-rich confining bed has the most. These estimates for the 

amount of extractable bioavailable organic carbon present in the 

sediments overlying the semiconfined aquifer make it possible 

to evaluate the ratio of DOC flux to the amount of AOC.

The 2-m thickness of sandy soil and sediments at land 

surface (fig. 10) has a bioavailable organic carbon content 

of approximately 500 mg/kg (table 3). Given the volume of 

each model grid block (5 m x 5 m x 2 m) and a bulk density 

of 1.85 x 103 kg/m3, the resulting sediment mass in each block 

is 9.3 x 104 kg. Furthermore, given the measured concentration of bioavailable organic carbon in the sediment (500 mg/

kg), the sediment contains on the order of 4.7 x 107 mg of 

potentially metabolizable organic carbon. The 3-meter-thick 

organic-matter-rich confining bed overlying the aquifer can 

be treated in the same manner. Given a grid block volume of 

5 m x 5 m x 3 m and a measured concentration of bioavailable 

organic matter of 5,000 mg/kg, the resulting mass of potentially metabolizable organic carbon is 6.9 x 108

 mg. In other 

words, the organic-matter-rich confining bed contains approximately an order of magnitude more bioavailable organic 

carbon than the soil zone. The sum of potentially available 

organic carbon in the soil zone and in the organic-matter-rich 

confining bed at the Kings Bay site is on the order of 7.4 x 

108mg per 125 m3, or 5.9 x 106 mg/m3.

If it is assumed that DOC is continually delivered to the 

aquifer by vertically percolating recharge at a concentration of 

5 mg/d per model cell of aquifer, the ratio of potentially available carbon to the flux of DOC to the aquifer is:

LTSR	(yr/m AOC	(mg/m

Flux	of	DOC	(mg/yr)

3

3

) ) = , (5)

where LTSR (in units of years per meter cubed of sediment) 

may be termed the “long-term sustainability ratio” for the 

MNA of the system. Ideally, the LTSR would quantify the 

length of time organic carbon in a given volume of aquifer can 

sustain the DOC flux needed to maintain MNA. In practice, 

however, the physics and chemistry of DOC mobilization are 

far too complex to make this estimate quantitative. On the 

other hand, it is clear that when the LTSR is large, the longterm sustainability of MNA at a site will be proportionally 

large. Conversely, when the LTSR is small, the sustainability 

of MNA will be small.

Figure 18. Simulated 1998 PCE concentrations for DO/DOC 

ratios of 1/5, 5/5, and 5/1. All ratios are in milligrams per liter.

Quantifying Mass and Energy Balance  21

1/5 DO/DOC ratio

5/5 DO/DOC ratio

5/1 DO/DOC ratio

GS-14

GS-13

GS-12

GS-11

GS-10

GS-9

GS-8

GS-14

GS-13

GS-12

GS-11

GS-10

GS-9

GS-8

GS-14

GS-13

GS-12

GS-11

GS-10

GS-9

GS-8

GS-7

GS-6

GS-5

GS-4

KBA-PDA-08

GS-3

GS-2

GS-1

GS-15

GS-7

GS-6

GS-5

GS-4

KBA-PDA-08

GS-7

GS-6

GS-5

GS-4

KBA-PDA-08

GS-3

GS-2

GS-1

GS-15

Source1

Source2

Source3

Source1

Source2

Source3

Source1

Source2

Source3

KBA-PS-03

KBA-PS-03

KBA-PS-03

KBA-11-10B

KBA-11-10B

KBA-11-10B

KBA-11-03B

KBA-11-03B

KBA-11-03B

KBA-11-18

KBA-11-18

KBA-11-18

5.7500

4.7500

3.7500

2.7500

1.7500

0.7500

5.7500

4.7500

3.7500

2.7500

1.7500

0.7500

5.7500

4.7500

3.7500

2.7500

1.7500

0.7500

MT3D BC Symbols

EXPLANATION

Point source/sink

MODFLOW BC Symbols

Constant head

0 25 50 METERS

0 50 100 150 200 FEET

Base modified from U.S. Geological Survey

1:24,000-scale digital data
In the case of the Kings Bay site, the LTSR is:

LTSR	= 5.9	 	10 	mg/m

365	mg/yr

	=	1.6	 	10 	 	yr/m

6 3  ́ 4 3  ́  ́ . (6)

This relatively large LTSR value illustrates that the amount of 

AOC in the system is large relative to the flux of DOC needed 

to sustain MNA. This, in turn, provides an explanation for 

the observed sustainability of MNA at the Kings Bay site for 

at least 20 years, and indicates that MNA could be sustained 

virtually indefinitely in this system.

Short-Term and Long-Term Sustainability

This analysis, which is based on simulations using a 

deterministic mass-balance model, indicates that the “sustainability” of MNA for chlorinated ethenes includes distinct 

short- and long-term components. The interaction of these 

components is illustrated in figure 19. When the short-term 

sustainability ratio (DO/DOC) is large, and the long-term 

sustainability ratio (AOC/DOC flux) is small, oxic conditions 

develop and MNA for chlorinated ethenes will be inherently 

unsustainable. When the STSR decreases below 1, anoxic conditions will develop and MNA will be sustainable for at least 

the short term. If the LTSR is too small, however, the amount 

of organic carbon may be insufficient to sustain MNA over 

the long term. When the STSR is small and the LTSR is large, 

MNA for chlorinated ethenes will be sustained over the long 

term. The threshold value of the LTSR needed to insure longterm (greater than 300 years) sustainability is not presently 

known. On the basis of this study, however, a value on the 

order of 104 yr/m3 may be sufficient to meet the requirements 

of long-term sustainability.

Electron Acceptor Depletion and Sustainability

The simulations of PCE transport shown on figure 18 

indicate that a high DO/DOC ratio (relatively reducing conditions) serves to limit the extent of the PCE plume as would 

be intuitively expected. Sometimes, however, model simulations can reveal behavior that is less intuitively obvious. For 

example, model simulations indicate that over relatively long 

timeframes (about 80 years), VC concentrations do not follow 

the pattern observed for PCE (fig. 18). Figure 20 shows simulated VC concentrations for different DO/DOC ratios. In this 

case, however, both highly reducing (DO/DOC ratio of 1/5) 

and oxic (DO/DOC ratio of 5/1) conditions were smaller than 

the intermediate (5/5 DO/DOC ratio) condition. For the highly 

reducing case, reductive dechlorination was capable of restricting VC transport, and thus the smaller VC plume. Similarly, 

under oxic conditions, oxic mineralization of VC restricted VC 

transport. For the intermediate condition, however, the amount 

of Fe(III) supporting anoxic mineralization of VC became 

limiting, which caused the VC plume to expand.

The simulations shown in figure 20 are probably not a 

realistic representation of the system as it now exists. Because 

of the Fenton’s Reagent injection, a large amount of Fe(III) 

has been added to the system. Much of the previous sulfatereducing portion of the plume (near the source area) is in fact 

now dominated by Fe(III) reduction as a result of the added 

ferric iron. Model simulations that increased the amount of 

Fe(III) present in the system prevented the increased size of 

the VC plume for the 5/5 ratio of DO/DOC shown in figure 20. 

Nevertheless, these simulations illustrate that the sustainability of MNA for chlorinated ethenes in general, and for VC 

in particular, may be subject to electron-acceptor limitations 

in addition to electron-donor limitations. As a result, these 

electron-acceptor limitations also should be considered in 

evaluations of sustainability.

Short-term

sustainability ratio

Long-term

sustainability ratio

Dissolved oxygen

Unsustainable

MNA

Short-term

sustainable,

and

long-term

unsustainable

MNA

Short-term

and

long-term

sustainable

Large MNA Small

Small Large

Dissolved organic carbon

Available organic carbon

Dissolved organic carbon flux

Ratio

Ratio

Figure 19. Diagram showing the short- and long-term components of the sustainability of monitored natural attenuation (MNA).

22  A Framework for Assessing the Sustainability of Monitored Natural Attenuation
Table 3. Organic carbon content of Kings Bay sediments.

[mg/kg, milligrams per kilogram; —, not determined]

Sample

horizon

Replicate 

analysis 

number

Weight loss on 

ignition1

(mg/kg)

Carbon 

content2

(mg/kg)

Extractable 

bioavailable 

organic 

carbon3

(mg/kg)

Soil-zone sample, 

cored at well 

KBA-11-13A

1 460 940 80 ± 2

2 450 — 260 ± 5

3 700 — 20 ± 2

4 — — 80 ± 33

5 — — 60 ± 4

Mean 540 ± 140 940 500 ± 20

Organic-matterrich confining 

bed, outcrop 

samples

1 20,300 16,000 1,500 ± 68

2 23,100 — 1,100 ± 73

3 22,700 — 600 ± 74

4 — — 9,900 ± 830

5 — — 1,200 ± 61

Mean 22,000 ± 1,500 16,000 14,300 ± 490

Organic-matterrich confining 

bed, cored 

samples at well 

USGS-9

1 5,800 3,600 400 ± 170

2 4,500 — 400 ± 11

3 4,600 — 130 ± 1

4 — — 1,100 ± 115

5 — — 300 ± 19

Mean 5,000 ± 700 3,600 2,300 ± 120

1

 Loss in weight of pre-dried samples after 8 hours at 400°C.

2

 Elemental carbon analysis by catalytic combustion at 900°C.

3Extractable bioavailable organic carbon, mean is sum of five extractions shown. 

Quantifying Mass and Energy Balance  23

Figure 20. Simulated 2050 VC concentrations for DO/DOC 

ratios of 1/5, 5/5, and 5/1. All ratios are in milligrams per liter.

1/5 DO/DOC ratio

5/5 DO/DOC ratio

5/1 DO/DOC ratio

GS-14

GS-13

GS-12

GS-11

GS-10

GS-9

GS-8

GS-14

GS-13

GS-12

GS-11

GS-10

GS-9

GS-8

GS-14

GS-13

GS-12

GS-11

GS-10

GS-9

GS-8

GS-7

GS-6

GS-5

GS-4

KBA-PDA-08

GS-3

GS-2

GS-1

GS-15

GS-7

GS-6

GS-5

GS-4

KBA-PDA-08

GS-7

GS-6

GS-5

GS-4

KBA-PDA-08

GS-3

GS-2

GS-1

GS-15

GS-3

GS-2

GS-1

GS-15

Source1

Source2

Source3

Source1

Source2

Source3

KBA-PS-03

KBA-PS-03

KBA-PS-03

KBA-11-10B

KBA-11-10B

KBA-11-10B

KBA-11-03B

KBA-11-03B

KBA-11-03B

KBA-11-18

KBA-11-18

KBA-11-18

5.7500

4.7500

3.7500

2.7500

1.7500

0.7500

5.7500

4.7500

3.7500

2.7500

1.7500

0.7500

5.7500

4.7500

3.7500

2.7500

1.7500

0.7500

X

XX

Source1

Source2

Source3

0 25 50 METERS

0 50 100 150 200 FEET

MT3D BC Symbols

EXPLANATION

Point source/sink

NAPL point source/sink

MODFLOW BC Symbols

Constant head

Base modified from U.S. Geological Survey

1:24,000-scale digital data
Conclusions

The sustainability of monitored natural attenuation (MNA) 

for any site and for any contaminant depends upon (1) the 

presence of chemical/biochemical processes capable of 

cycling contaminant(s) to innocuous byproducts, and (2) the 

availability of sufficient energy to drive these transformations 

to completion (fig. 6). By this definition, the example of the 

Kings Bay site meets all of the criteria for both short- and 

long-term sustainability of MNA for chlorinated ethenes. First, 

biochemical processes are in place (reductive dechlorination, 

VC oxidation) that transform chlorinated ethenes to innocuous byproducts. Secondly, sufficient energy exists in the form 

of organic carbon capable of driving these transformations to 

completion over the timeframe required (about 14 years).

The sustainability of natural attenuation for chlorinated 

ethenes can be quantitatively assessed by considering the mass 

and energy balance of ground-water systems. Such mass and 

energy balances can be made by using either empirical or 

deterministic methods. But regardless of the methods used, it 

is necessary to:

Quantify the flux of ground water moving through the system. Any assessment of mass and energy balance is highly 

sensitive to ambient rates of ground-water flow, and it is 

important for this parameter to be highly constrained.

Delineate the mass and distribution of chlorinated ethenes 

and/or NAPL present in the system. A true mass balance 

cannot be achieved without estimates of contaminant 

mass. While such estimates are difficult to make and 

are subject to considerable uncertainty, the sustainability of MNA cannot be assessed without them. A variety 

of methods are available for making these estimates. 

Given an estimated mass of NAPL present, it is possible 

to estimate the time of remediation (TOR) needed for a 

contamination to completely disperse.

Quantify the mass and flux of available electron acceptors, especially dissolved oxygen, being delivered to the 

system. Oxygen can divert electron flow from chlorinated 

ethene reduction; understanding the dynamics of electronacceptor delivery to the system is fundamental to quantifying MNA sustainability.

Quantify the mass and flux of available organic carbon. 

Microbially utilizable organic carbon in soil and aquifer 

sediments can be estimated using sediment-extraction 

procedures. Similarly, measurements of dissolved organic 

carbon provide an estimate of carbon mass present in the 

aqueous phase.

Assess the interaction of electron donors and acceptors. 

The short-term sustainability of MNA for chlorinated 

ethenes requires that the molar ratio of dissolved oxygen 

to dissolved organic carbon to a system be less than 1. 

The long-term sustainability of MNA requires that the 

1.

2.

3.

4.

5.

ratio of available organic carbon in a system be large 

relative to the flux of organic carbon needed to maintain 

anoxic conditions.

The importance of sustainability when assessing MNA 

has only recently been identified as an important issue in environmental remediation (Newell and Aziz, 2004). The framework outlined in this report provides methods for assessing and 

quantifying the sustainability of MNA for chlorinated ethenes.

References

Alexander, Martin, 1994, Biodegradation and bioremediation: 

San Diego, Academic Press, 302 p.

Bradley, P.M., 2000, Microbial degradation of chloroethenes 

in groundwater systems: Hydrogeology Journal, v. 8, no. 1, 

p. 104–111.

Bradley, P.M., 2003, History and ecology of chloroethene 

biodegradation­—A review: Bioremediation Journal, v. 7, 

no. 2, p. 81–109.

Bradley, P.M., and Chapelle, F.H., 1996, Anaerobic mineralization of vinyl chloride in Fe(III)-reducing, aquifer sediments: Environmental Science and Technology, v. 30, no. 6, 

p. 2084–2086.

Carson, Rachel, 1962, Silent spring: Boston, Riverside Press, 

368 p.

Chapelle, F.H., and Bradley, P.M., 1998, Selecting remediation goals by assessing the natural attenuation capacity of 

groundwater systems: Bioremediation Journal, v. 2, no. 3, 

p. 227–238.

Chapelle, F.H., Bradley, P.M., and Casey, C.C., 2005, Behavior 

of a chlorinated ethene plume following source-area treatment with Fenton’s Reagent: Ground Water Monitoring and 

Remediation, v. 25, no. 2, p. 131–141.

Chapelle, F.H., Bradley, P.M., Lovley, D.R., O’Neill, Kathleen, 

and Landmeyer, J.E., 2002, Rapid evolution of redox processes in a petroleum hydrocarbon-contaminated aquifer: 

Ground Water, v. 40, no. 4, p. 353–360.

Chapelle, F.H., Campbell, B.G., Looney, B.B., Wiedemeier, 

T.H., and Sink, C.H., 2005, Empirical and deterministic 

methods for assessing monitored natural attenuation, in

Proceedings of the In Situ and On-Site Bioremediation 

Eighth International Symposium, Baltimore, Md., 

June 6–9, 2005: Battelle Press.

Davis, J.W., and Carpenter, C.L., 1990, Aerobic biodegradation of vinyl chloride in groundwater samples: Applied and 

Environmental Microbiology, v. 56, no. 12, p. 3878–3880.

24  A Framework for Assessing the Sustainability of Monitored Natural Attenuation
References  25

DiStefano, T.D., Gossett, J.M., and Zinder, S.H., 1991, Reductive dechlorination of high concentrations of tetrachloroethene to ethene by an anaerobic enrichment culture in the 

absence of methanogenesis: Applied and Environmental 

Microbiology, v. 57, no. 8, p. 2287–2292.

Fetter, C.W., 1988, Applied hydrogeology (2d ed.): New York, 

Merrill Publishing Company, p. 68.

Grimvall, Anders, 1995, Evidence of naturally produced and 

man-made organohalogens in water and sediments, in Grimvall, Anders, and de Leer, E.W.B., eds., Naturally-produced 

organohalogens: Boston, Kluwer Academic Publishers, 

p. 3–20.

Harr, Jonathon, 1995, A civil action: New York, Random 

House, 500 p.

Hartmans, S., 1995, Microbial degradation of vinyl chloride, 

in Singh, V.P., ed., Biotransformations—microbial degradation of health-risk compounds: New York, Elsevier Science, 

p. 239–248.

Leeth, D.C., 1999, Hydrogeology of the surficial aquifer in the 

vicinity of a former landfill, Naval Submarine Base Kings 

Bay, Camden County, Georgia: U.S. Geological Survey 

Water-Resources Investigations Report 98-4246, 28 p.

Maymo-Gatell, Xavier, Chien, Y.T., Gossett, J.M., and Zinder, 

S.H., 1997, Isolation of a bacterium that reductively 

dechlorinates tetrachloroethene to ethene: Science, v. 276, 

p. 1568–1571.

National Academy of Science, 1970, Committee on Research 

in the Life Sciences of the Committee on Science and Public Policy.

National Research Council, 2000, Natural attenuation for 

groundwater remediation: Washington, D.C., National 

Academy Press, 274 p.

Newell, C.J., and Aziz, C.E., 2004, Long-term sustainability 

of reductive dechlorination reactions at chlorinated solvent 

sites: Biodegradation, v. 15, no. 6, p. 387–394.

Rectanus, H.V., Widdowson, M.A., Novak, J.T., and Chapelle, 

F.H., 2005, Method development for quantifying potential 

bioavailable organic carbon in sediments, in Proceedings 

of the Third European Bioremediation Conference, Chania, 

Crete, Greece, July 4–7, 2005.

Southern Division, Naval Facilities Engineering Command, 

2004, Electronic Administrative Record and Index, Southern 

Division Naval Facilities Engineering Command, Charleston, South Carolina, 45 p.

U.S. Department of Energy, 2002, Alternative project plan for 

technology development, accelerated risk reduction through 

innovative remediation of chlorinated ethenes using monitored natural attenuation: U.S. Department of Energy, Office 

of Environmental Management.

U.S. Environmental Protection Agency, 1998, Technical protocol for evaluating natural attenuation of chlorinated solvents 

in ground water: Washington, D.C., U.S. Environmental 

Protection Agency, Office of Research and Development, 

EPA/600/R-98/128, p. 232.

U.S. Environmental Protection Agency, 1999, Use of monitored natural attenuation at Superfund, RCRA corrective 

action, and underground storage tank sites, Final: U.S. Environmental Protection Agency, Office of Solid Waste and 

Emergency Response, Directive Number 9200.4-17P, 32 p.

Van Metre, P.C., and Mahler, B.J., 2004, Contaminant trends 

in reservoir sediment cores as records of influent stream 

quality: Environmental Science and Technology, v. 38, 

no. 11, p. 2978–2986.

 Vogel, T.M., Criddle, C.S., and McCarty, P.L., 1987, Transformations of halogenated aliphatic compounds: Environmental Science and Technology, v. 21, no. 8, p. 722–736.

 Vogel, T.M., and McCarty, P.L., 1985, Biotransformation of 

tetrachloroethylene to trichloroethylene, dichloroethylene, 

vinyl chloride, and carbon dioxide under methanogenic 

conditions: Applied and Environmental Microbiology, v. 49, 

no. 5, p. 1080–1083.

 Waddill, D.W., and Widdowson, M.A., 1997, SEAM3D—A 

numerical model for three-dimensional solute transport 

and sequential electron acceptor-based bioremediation in 

groundwater: Virginia Polytechnic Institute and State University Final Technical Report, prepared for the U.S. Army 

Corps of Engineers.

Wiedemeier, T.H., Rifai, H.S., Newell, C.J., and Wilson, J.T., 

1999, Natural attenuation of fuels and chlorinated solvents 

in the subsurface: New York, John Wiley and Sons, 617 p.

Wiedemeier, T.H., Swanson, M.A., Moutoux, D.E., Wilson, 

J.T., Kampbell, D.H., Hansen, J.E., and Haas, P., 1996, 

Overview of the technical protocol for natural attenuation 

of chlorinated aliphatic hydrocarbons under development 

for the U.S. Air Force Center for Environmental Excellence: 

EPA/540/R-96/509, p. 35–59.

Wilson, J.T., and Wilson, B.H., 1985, Biotransformation of 

trichloroethylene in soil: Applied and Environmental Microbiology, v. 49, no. 1, p. 242–243.

The SEAM3D code uses mathematical representations of 

contaminant loading and attenuation processes, and sums 

them in order to provide a quantitative mass and energy balance. Advection refers to dissolved mass carried by flowing 

ground water, and is proportional to the velocity of groundwater flow v [LT-1] and the change of concentration mass 

[ML-3] over a particular distance, x [L-1]:

- ¶

¶

v

C

x .

This advective term, therefore, has units of mass concentration per unit time [ML-3T-1]. The dispersion of dissolved 

mass in flowing ground water is proportional to a scaledependent constant, the coefficient of dispersion (D) [L2

T-1], 

and the second derivative of the concentration gradient:

2

2

C D

x

¶

¶ .

Again, the units of this term are mass concentration per 

unit time [ML-3T-1]. Sorption refers to the distribution of a 

mass between the aqueous and adsorbed phases and can be 

described by the term:

b d K C

n t

r ¶ - ¶ ,

where ρb is the average bulk density of aquifer solids [ML-3]; 

n is aquifer porosity (dimensionless); and Kd is a distribution 

coefficient describing solute partition between the adsorbed 

and dissolved state [L3M-1]. The units of this term are also in 

mass concentration per unit time [ML-3T-1].

The microbial transformation of electron donors and 

acceptors is evaluated by summing the effects of microbial 

colonies inhabiting the ground-water system on substrates 

and electron acceptors (EA). For each substrate degraded by 

microbial population x, the sink term is

, ,

bio x

sink ls x ls

x

M R r

q =å , (1.1)

where θ is the effective porosity (dimensionless); Mx is the 

microbial biomass concentration [MbLpm

-3] for x = 1,2....NM 

(number of microcolonies); and rx,ls is the utilization rate of 

substrate ls in microbial population x where the units of this 

term are in mass concentration per unit time [MlsMb

-1T-1]. For 

each electron acceptor, the sink term is

, ,

bio x

sink le x le

x

M R r

q =å . (1.2)

The utilization of each substrate within microbial population x is described by the equation

x ls, x ls le , ,

le

r =ån , (1.3)

where x ls le , , n is the specific rate of substrate utilization for 

microbial population x growing on substrate ls and electron 

acceptor le. Similarly, utilization of each electron acceptor 

follows

rx le x ls le x ls le

ls

, =åg n , , , , , (1.4)

where gx ls le , , is the electron-acceptor use coefficient 

[MleMls

-1]. Finally, Monod kinetics are used to describe 

substrate and electron-acceptor utilization

max

, , , , ,

_ _ _

, , ,

ls le

x ls le x ls le x le li

s e

x ls le x le le ls

S E N I

K S K E

n n

- -

é ù é ù ê ú ê ú ê ú ê ú = ê ú ê ú ê ú ê ú ê ú ê ú + ê ú ê ú + ë û ë û

, (1.5)

where max

x ls le , , n is the maximum specific rate of substrate utilization for microbial population x growing on substrate ls and 

electron acceptor le [MlsMb

-1T-1]; , ,

s Kx ls le is the half-saturation 

constant for substrate ls for utilization of electron acceptor 

le [MlsL-3]; ,

e Kx le is the effective half-saturation constant for 

electron acceptor le [MleL-3]; and 

_

Sls and 

_

Ele are the effective concentrations of substrate ls [MlsL-3] and electron acceptor le [MleL-3], respectively, (defined as the difference between 

the aqueous phase concentrations and minimum concentrations below which biodegradation ceases).

The model code, SEAM3D, uses inhibition functions to 

simulate the preferential utilization of electron acceptors in the 

order O2 > NO3 > Mn(IV) > Fe(III) > SO4 > CO2, an order that 

corresponds to the decreasing free energy available to microorganisms from each reaction. When concentrations of oxygen 

are above a given threshold (about 0.5 mg/L), the model 

inhibits the oxidation of competing electron acceptors using a 

function le li , I defined by

, 1 le li I = for le=1, (1.6)

and

1 ,

, _ 1

,

le

le li

le li li

le li li

I

E

k

k

-

=

é ù ê ú

= P ê ú ê ú ê ú + ë û for le = 2, 3, 4, 5, or 6, (1.7)

Appendix 1. Description of the Deterministic Model  27

Appendix 1. Description of the Deterministic Model
where kle li , is the electron-inhibition coefficient [MleL-3], representing inhibition of electron acceptor le by electron acceptor li. Thus, the model is able to simulate the order of electronacceptor utilization observed in natural systems (fig. 8).

Finally, the rate of dissolution of the NAPL mass 

,

NAPL RSource ls can be formulated in terms of the maximum concentration of a solute in equilibrium with a NAPL of given 

composition ( eq Cls ) as given by Raoult’s law:

( ) eq sol C f C ls ls ls = , (1.8)

where ls f is the mole fraction of the solute and sol Cls is the 

solubility of the solute, and the actual concentration at any 

given point in space (Cls) is:

, ( ) NAPL NAPL eq R k C C Source ls = -ls ls , (1.9)

where kNAPL is the NAPL dissolution rate constant (T-1). This 

gives ,

NAPL RSource ls in units of contaminant concentration per unit 

time [ML-3T-1].

The individual processes that affect any particular solute 

in ground-water systems, including hydrodynamic dispersion, 

advection, sorption, and biodegradation, and NAPL dissolution 

then can be summed giving an explicit mass balance governing 

the transport of any particular solute. In the case of chlorinated 

ethenes, the general mass-balance equation may be written:

where

Sls is the aqueous phase chlorinated ethene 

concentration [MlsL-3] for ls=1,2,...,NS 

(number of individual chlorinated ethenes); *

ls S is the chlorinated ethene point-source 

concentration [MlsL-3];

vi is the average pore-water velocity [LT-1];

xi is distance [L];

Dij is the tensor for the hydrodynamic dispersion 

coefficient [L2T-1];

,

bio Rsink ls is a biodegradation sink term dependent on 

the mode of respiration [MlsL-3T-1]; 

,

NAPL Rsource ls is a term associated with non-aqueous phase 

liquid (NAPL) dissolution [MlsL-3T-1];

Rls is the retardation factor for chlorinated ethene 

ls [Lo];

t is time [T];

	 θ is aquifer porosity [Lo]; and

qs is the volumetric flux of water per unit 

volume of aquifer [T-1] with qs > 0 for 

sources and qs < 0 for sinks.

In the same manner, mass-balance equations can be 

written for dissolved organic carbon serving as electron 

donors in the system, and for the various soluble and insoluble 

compounds serving as electron acceptors. This set of coupled 

equations then can be solved simultaneously to assess electron 

donor oxidation coupled to electron acceptor reduction, and 

thus the flow of energy through the system.

28  A Framework for Assessing the Sustainability of Monitored Natural Attenuation

*

, ,

ls ls s bio NAPL ls

i ij ls sink ls source ls ls

i i j

S S q S

v D S R R R

x x x t

¶ ¶ ¶ ¶

¶ ¶ ¶ q ¶

æ ö ç ÷

- + ç

ç

÷

÷+ - + =

ç ÷÷ è ø

, (1.10)
Figure

2.1. Map showing location of SEAM3D model area and boundary conditions........................31

Appendix 2. Parameters Used to Simulate the Kings Bay Site

Contents

Boundary Conditions and Hydrologic Parameters.................................................................................31

Solute Transport Parameters: The Dispersion Package.......................................................................31

Sources of Dissolved Chlorinated Ethenes: The NAPL Dissolution Package ...................................32

Geochemical and Redox Parameters: The Biodegradation Package ................................................32

Chlorinated Ethene Transformation Parameters: The Reductive Dechlorination Package............33

Sorption Parameters: The Chemical Reaction Package.......................................................................34

Tables

2.1. Hydrologic and dispersion parameters used in the SEAM3D model.................................31

2.2. Initial conditions, concentrations at the landfill boundary, and recharge 

concentrations for electron donors, electron acceptors, tracer, and 

biodegradation end products ...................................................................................................32

2.3. Initial concentration of microbial populations and biodegradation parameters 

for four terminal electron-accepting processes ...................................................................33

2.4. Initial concentration of PCE/TCE and DCE/VC degraders and kinetic reductive 

dechlorination parameters for the simulated chlorinated ethenes ...................................34

2.5. Biodegradation parameters for net mineralization of DCE and VC....................................34

2.6. Stoichiometric conversion factors for reductive dechlorination .......................................35

2.7. Sorption parameters for mobile constituents ........................................................................35

Boundary Conditions and Hydrologic Parameters

The site was modeled with one layer representing a semiconfined aquifer to which a non-aqueous phase liquid 

(NAPL) source had been delivered (fig. 10). The conceptualization shown in figure 10 was translated into a deterministic 

model by applying the Basic Transport, Advection, Dispersion, NAPL Dissolution, Source-Sink, Chemical Reaction, 

Biodegradation, and Reductive Dechlorination Packages 

of the SEAM3D model code. The steady-state flow field is 

constructed to represent the long-term average flow conditions 

at the site of the contamination in order to test the concept 

of monitored natural attenuation with SEAM3D over multidecade time scales. Thus, the finite-difference grid is site 

specific with 20 columns and 40 rows with cell dimensions of 

5 m by 5 m (an area of 20,000 m2, shown on figure 2.1). The 

upgradient and downgradient boundaries were simulated as 

constant-head boundaries to match the observed head gradient 

across the plume in the semiconfined aquifer. The upgradient 

boundary is near a ground-water divide identified previously 

at the site (Leeth, 1999) and is treated as a constant head 

based on the average head at this boundary. The downgradient boundary was placed in the housing subdivision where 

water-level fluctuations are buffered by the presence of several 

small ponds. The lateral boundaries were placed perpendicular to observed lines of equal hydraulic head at the site, 

which is consistent with their treatment as no-flow boundaries. The water-table aquifer above the semiconfined aquifer 

was not actively simulated; however, the effect of recharge to 

the underlying semiconfined aquifer was simulated using the 

Source-Sink Package of SEAM3D.

SEAM3D uses the steady-state flow field created by 

MODFLOW, given concentration initial conditions and boundary conditions for the reactive transport model. A constant 

value of hydraulic conductivity (10 m/d) was assigned for the 

MODFLOW model based on estimates derived from aquifer 

tests conducted at the site (Leeth, 1999). The site-specific 

flow model is simplified; the average flow through the site is 

constrained by using the average known gradient across the site 

(commonly referred to as Darcy velocity); however, the pore 

velocity is determined by dividing the Darcy velocity determined by MODFLOW by the porosity. A constant porosity of 

0.3 was applied in the model. The site-specific effective porosity is unknown (Leeth, 1999), so the value of 0.3 was selected 

as representative of the type of sediments encountered at the site 

(Fetter, 1988). A specific storage (0.0001 1/meter) and specific 

yield (0.1 dimensionless) were assigned (table 2.1) based on 

estimates made from aquifer tests at the site (Leeth, 1999).

 In the Source-Sink Package of SEAM3D, recharge to the 

semiconfined aquifer was simulated using areally constant values, and recharge was considered to contain constant concentrations of the electron acceptors dissolved oxygen (5 mg/L) 

and sulfate (10 mg/L). In addition, areal recharge was considered to contain dissolved organic carbon mobilized from 

the soil zone and from the organic-matter-rich layer overlying 

the semiconfined aquifer (table 2.2). Because the amount of 

areal recharge delivered to the semiconfining aquifer was not 

directly measured, rates of electron acceptor-bearing recharge 

to the aquifer were adjusted as the model was constrained.

Solute Transport Parameters: 

The Dispersion Package

Given the ground-water velocity flow field generated by 

the MODFLOW-computed hydraulic gradient and an 

assigned porosity, the dispersion coefficient is calculated from 

the assigned dispersivity (5.0 m) and the ratio of transverse 

to longitudinal dispersivity (0.01). The effective molecular 

diffusion coefficient, which for sites dominated by advective 

ground-water flow is typically much less than the dispersion, 

was set to 5.0 x 10-5 m2/d (table 2.1).

Figure 2.1. Location of SEAM3D model area and 

boundary conditions.

Table 2.1. Hydrologic and dispersion parameters used in the 

SEAM3D model.

[m/d, meter per day; m, meter; m2/d, square meter per day]

 Model parameter Model input

Hydraulic conductivity 10 m/d

Porosity 0.3

Specific storage 0.0001 1/meter

Specific yield 0.1 (dimensionless)

Longitudinal dispersivity 5.0 m

Transverse dispersivity 0.05 m

Molecular diffusion coefficient 5.0 x 10-5 m2/d

Appendix 2. Parameters Used to Simulate the Kings Bay Site  31

KBA-11-18

GS-14 GS-7

GS-3

GS-2

GS-1

GS-15

GS-6

GS-5

GS-4

GS-13

GS-12

GS-11

GS-10

5.7500

4.7500

3.7500

2.7500

1.7500

0.7500

GS-9

GS-8

KBA-PDA-08

KBA-PS-03 KBA-11-10B

KBA-11-03B

Source1

Source2

Source3

MT3D BC Symbols

EXPLANATION

Point source/sink

MODFLOW BC Symbols

Constant head

0 25 50 METERS

0 50 100 150 200 FEET

Base modified from U.S. Geological Survey

1:24,000-scale digital data
Sources of Dissolved Chlorinated Ethenes: 

The NAPL Dissolution Package

An important capability of the SEAM3D code is that it 

includes a NAPL Package designed to simulate the dissolution and mobilization of chlorinated ethenes from the NAPL 

present in the contaminant source area. This feature, when 

combined with the Biodegradation and Reductive Dechlorination Packages, allows simulation of a global mass balance that 

includes mineral and dissolved electron acceptors, dissolved 

organic carbon substrate, and the dissolution and natural 

attenuation of chlorinated ethenes. This capability, in turn, 

makes it possible to estimate times of remediation associated 

with NAPL dissolution, and thus, to assess the sustainability 

of natural attenuation.

The SEAM3D model allows NAPL dissolution to be conceptualized by assigning a given NAPL mass ( N C in grams 

per gram) to model cells according to the equation

/ N N C M V = cell b cell r , (2.1)

where N Mcell mass of NAPL is assigned to each model cell 

(g); Vcell is the volume of the cell (∆x x ∆y x NAPL thickness); and rb (g/m3

) is the sediment bulk density. SEAM3D 

then partitions the soluble components of the NAPL into the 

aqueous phase at each time step depending on the contaminant 

mass fraction of the NAPL, the molecular weight, solubility, 

rates of ground-water flow, and a NAPL dissolution constant. 

Depending on the amount of dissolution, the model then recalculates the amount of remaining NAPL for the next time step.

NAPL at the site was considered to contain both PCE 

(molecular weight of 166; solubility of 150 mg/L; initial mass 

fraction of 0.15) and TCE (molecular weight of 131; solubility of 1,100 mg/L; initial mass fraction of 0.01). These two 

contaminants were considered to coexist with an inert fraction 

of NAPL with a molecular weight of 150, and an initial NAPL 

dissolution rate constant of 0.01 d-1. This rate constant was chosen primarily because it is sufficiently fast so that hydrologic 

(rate of ground-water flow) and chemical (solubility) conditions would determine the overall rate of NAPL dissolution.

Estimating the mass of NAPL originally emplaced in an 

aquifer is necessary to make a true mass balance of the system. 

At the Kings Bay site, eyewitness accounts of waste-disposal 

practices are available, although written records of how much 

waste was disposed are not available. One eyewitness account 

recalled that the waste-disposal trenches were dug with a 

backhoe to a depth of between 6 and 8 feet, trucks delivering 

municipal waste filled the trenches, and the backhoe then covered the trenches with sand. Apparently, drums of PCE waste 

were disposed of in this fashion. Given the waste-disposal 

practices employed, the measured size of the PCE source areas 

(Chapelle and Bradley, 1998) and the approximate concentrations of PCE measured in the source area place some constraints on the mass of NAPL emplaced. Nevertheless, because 

of the inherent uncertainties associated with estimating NAPL 

mass, a range of values was simulated.

Based on an approximate 5- by 15-m area of the source 

zone (75 m2), a thickness of affected aquifer of 1 m, measured 

concentrations of PCE ranging between 1 and 10 milligrams 

per gram (mg/g) of sediment, and a sediment bulk density of 

1.85 x 106 g/m3 suggest that between 280 and 2,800 kg of PCE 

were disposed of at the site. Because one 55-gallon barrel of 

PCE weighs about 280 kg, one could assume that between 

1 and 10 barrels of waste PCE were disposed of at this site. 

Assuming that one barrel of PCE is assigned to each sourcearea model cell ( N Mcell = 280,000 g), that a cell area is 25 m2

, 

a NAPL thickness is 1 m, and the rb is 1.85 x 106 g/m3, the 

N C calculated from equation 2.1 would be 0.006 g/g. These 

estimates were used as the starting point for assigning an 

amount of PCE NAPL in the modeled source area, and then 

were refined as the model was constrained with available 

hydrologic and geochemical data.

Geochemical and Redox Parameters: 

The Biodegradation Package

Redox conditions at the Kings Bay site have previously 

been shown to be dominated by Fe(III) reduction and 

sulfate reduction. Four redox processes (oxygen reduction, 

Fe(III) reduction, sulfate reduction, and methanogenesis) were 

incorporated into the SEAM3D simulation of the site. All of 

these electron-accepting processes were coupled to the oxidation of a soluble electron donor (dissolved organic carbon of 

oxidation state zero and a molecular weight of 32). To activate 

these processes in the model, the species oxygen, sulfate, 

Fe(II), and methane were defined so that SEAM3D would 

write a mass-balance equation for each species. SEAM3D is 

designed to consider Fe(III) reduction occurring with solidphase Fe(III) oxyhydroxides. Near the landfill, where sulfatereducing conditions were observed, the Fe(III) content of 

aquifer material was set to zero. Downgradient, where Fe(III)-

reducing conditions were observed, the Fe(III) content of 

Table 2.2. Initial conditions, concentrations at the landfill 

boundary, and recharge concentrations for electron donors, 

electron acceptors, tracer, and biodegradation end products.

[mg/L, milligram per liter; O2, dissolved oxygen; Fe(III), ferric iron; SO4, 

sulfate; Fe(II), ferrous iron; CH4, methane; Cl, chloride; N/A, not applicable]

Constituent Concentrations (mg/L)

Initial Boundary Recharge

Electron donor 0.1 0.1 5.0

O2 0.0 0.0 5.0

(1)Fe(III) 50.0 N/A N/A

SO4 1.0 1.0 10.0

Tracer 0.0 5.0 0.0

Fe(II) 0.0 1.0 0.0

CH4 0.0 2.0 0.0

Cl 0.0 0.0 0.0

(1) Unit for Fe(III) is milligrams per gram.

32  A Framework for Assessing the Sustainability of Monitored Natural Attenuation
aquifer materials was set to an initial value of 50 micrograms 

per kilogram (μg/kg). This value was based on measurements 

of Fe(III) made in a shallow coastal-plain aquifer with similar 

hydrology to that of Kings Bay (Chapelle and others, 2002). 

SEAM3D computes a mass balance on Fe(III) for each time 

step as Fe(III) reduction occurs over time, allowing the amount 

of Fe(III) present in the system to decrease as Fe(III) reduction proceeds. This feature makes it possible to assess how the 

consumption of Fe(III) in the system affects the sustainability 

of natural attenuation.

A large number of parameters are required by SEAM3D 

in order to simulate the biodegradation of dissolved organic 

carbon coupled to the sequential reduction of oxygen, Fe(III), 

sulfate, and carbon dioxide. These include specifying the 

density of aerobic, Fe(III)-reducing, sulfate-reducing, and 

methanogenic microorganisms (grams per cubic centimeter); 

specifying the inhibition coefficients directing sequential utilization; specifying product generation coefficients for Fe(II) 

and methane; specifying Monod half-saturation constants for 

each electron-accepting process coupled to dissolved organic 

carbon (substrate) utilization; and specifying Monod maximum utilization rate constants for each electron-­accepting 

process. The problem was simplified by assuming that the 

microbial population is at steady state (rate of death equals 

rate of growth). The parameter values used to build the biodegradation package in the model are shown in table 2.3.

Chlorinated Ethene Transformation Parameters: 

The Reductive Dechlorination Package

The transformation of chlorinated ethenes was simulated 

within SEAM3D by assigning Monod kinetic parameters 

(equation 1.5), which describe reductive dechlorination for 

each species. In addition, the anaerobic oxidation of VC was 

simulated. At the Kings Bay site, ethene, the final degradation product of reductive dechlorination, is generally below 

detectable levels (0.01 mg/L), although occasional detections 

(about 0.1 mg/L) have been noted. If reductive dechlorination 

to ethene were the only biodegradation process occurring, then 

ethene concentrations should be on the order of 1,000 μg/L for 

much of the plume. There are two possible explanations for 

this discrepancy. First, VC may be reduced to ethene, which 

is then rapidly mineralized to carbon dioxide, preventing the 

accumulation of ethene in solution. Secondly, VC may be 

directly mineralized to carbon dioxide (Bradley, 2000). The 

occasional detections of ethene show that a component of VC 

reduction to ethene occurs at the site; however, the observation 

that DCE (the source of VC) is transported farther downgradient than VC indicates that VC is subject to anoxic mineralization as well. The SEAM3D code is capable of simulating 

both reductive and oxidative transformation of all chlorinated 

ethenes. At the Kings Bay site, reductive dechlorination was 

simulated for all four chlorinated ethenes, and anoxic mineralization was simulated only for VC.

The Monod kinetic parameters used to simulate chlorinated ethene transformation at the Kings Bay site are shown in 

table 2.4. Half-saturation constants (K) and maximum rates of 

substrate uptake (v) for each reductive process were assigned. 

Only VC was considered subject to anoxic mineralization. 

Thus, half-saturation constants for direct mineralization of 

PCE, TCE, and DCE were set to zero, and a finite value was 

assigned for VC mineralization coupled to Fe(III) reduction.

The sequential reduction of PCE, TCE, DCE, and VC 

was simulated by assigning inhibition factors for each process. 

Since oxygen reduction strongly inhibits the reduction of all 

chlorinated ethenes, the inhibition factors for oxygen were 

set relatively low (table 2.5). Conversely, because electronaccepting processes such as Fe(III) reduction, sulfate reduction, and methanogenesis are not observed to inhibit reductive 

dechlorination (Bradley, 2000), the inhibition concentrations 

are assigned larger values. Finally, the stoichiometric conversion factors for the various products of chlorinated ethene 

reduction, which allow model output in units of milligrams per 

liter, are shown in table 2.6.

Table 2.3. Initial concentration of microbial populations and biodegradation parameters for four terminal electron-accepting processes.

[O2, oxygen; Fe(III), ferric iron; SO4, sulfate; g m-3, grams per cubic meter; d-1, 1/day; g g-1, gram per gram]

Microbial

population

Mx nx s le

max

, , Kx s le

s

, , Kx le

e

, Yx l, , s le gx l, , s le  x kle,li

(1)

(g m-3) (d-1) (g m-3) (g m-3) (g g-1) (g g-1) (g g-1) O2 Fe(III) SO4

Aerobes 0.25 0.05 5.0 1.0 0.0 3.0 N/A N/A N/A N/A

Iron reducers 0.025 0.05 10.0 N/A 0.0 33.0 0.10 10.0 N/A N/A

Sulfate reducers 0.025 0.005 10.0 1.0 0.0 4.5 N/A 10.0 50.0 N/A

Methanogens 0.025 0.01 10.0 N/A 0.0 N/A 0.20 10.0 25.0 1.0

(1) Units for inhibition coefficients are in milligrams per liter, except for Fe(III), which is in milligrams per gram.

Appendix 2. Parameters Used to Simulate the Kings Bay Site  33
Sorption Parameters: The Chemical 

Reaction Package

I

n addition to hydrologic and microbiologic transformations, 

the mass balance of solutes in the semiconfined aquifer is 

subject to geochemical processes as well. All solutes are subject to chemical and electrical interactions with aquifer materials. These interactions, which are highly complex and not 

well-understood, are typically lumped together and modeled as 

sorption processes. The SEAM3D code allows consideration 

of sorption processes on the transport of solutes in the Chemical Reaction Package.

SEAM3D enables the user to use different approaches to 

modeling sorption including specifying linear and Freundlich 

adsorption isotherms. For the Kings Bay model, a linear 

Table 2.4. Initial concentration of PCE/TCE and DCE/VC degraders and kinetic reductive dechlorination parameters for the simulated 

chlorinated ethenes. 

[O2, oxygen; Fe(III), ferric iron; SO4, sulfate; PCE, perchloroethene; TCE, trichloroethene; DCE, dichloroethene; VC, vinyl chloride; g m-3, grams per cubic 

meter; d-1, 1/day; N/A, not applicable] 

Chlorinated

ethene

M y max,EA

lc n e Klc kle,li

(1)

(g m-3) (d-1) (g m-3) O2 Fe(III) SO4 PCE TCE DCE

PCE

0.025

0.60 5.0 0.5 100 50 N/A N/A N/A

TCE 1.0 10.0 0.5 100 50 50 N/A N/A

DCE

0.025

1.2 20.0 1.0 25 50 50 10 N/A

VC 0.5 15.0 1.0 25 50 50 10 1.0

(1) Units for inhibition coefficients are in milligrams per liter except for Fe(III), which is in milligrams per gram.

Table 2.5. Biodegradation parameters for net mineralization of DCE and VC.

[DCE, dichloroethene; VC, vinyl chloride; d-1, 1/day; g m-3, grams per cubic meter; g g-1, gram per gram; N/A, not applicable]

Microbial

population

DCE VC

max

x lc le , , n Kx lc le

ed

, , Kx le

e

, Yx l, , c le gx l, , c le

max

x lc le , , n , ,

ed Kx lc le Kx le

e

, Yx l, , c le gx l, , c le

(d-1) (g m-3) (g m-3) (g g-1) (g g-1) (d-1) (g m-3) (g m-3) (g g-1) (g g-1)

Aerobes 0.0 0.0 0.0 0.0 0.0 0.0 0.0 0.0 0.0 0.0

Iron reducers 0.0 0.0 0.0 0.0 0.0 0.08 1.0 N/A 0.0 0.0

34  A Framework for Assessing the Sustainability of Monitored Natural Attenuation

isotherm coefficient was specified for each mobile chlorinated 

ethene, substrate, and product considered in the mass balance. 

In addition, a tracer was specified and assigned a sorption 

coefficient to help constrain rates of ground-water flow. The 

retardation factors for chlorinated ethenes were estimated from 

the approximate organic matter content of aquifer material and 

appropriate octanol/water partition coefficients. The solutes 

and adsorption isotherm coefficients are shown in table 2.7.

Most of the biochemically active solutes such as dissolved 

oxygen and substrate were considered to be nonsorbing. 

Sulfate, however, is known to be slightly retarded relative 

to ground-water flow. Because sulfate transport was used to 

constrain rates of ground-water flow in the model, a sorption 

coefficient was specified. In addition, a tracer was specified 

with the same sorption coefficient as sulfate.
Table 2.6. Stoichiometric conversion 

factors for reductive dechlorination.

[PCE, perchloroethene; TCE, trichloroethene; Cl, 

chloride; DCE, dichloroethene; VC, vinyl chloride]

Reaction Stoichiometric factor

PCE→TCE 0.7923

PCE→Cl 0.2141

TCE→DCE 0.7379

TCE→Cl 0.270

DCE→VC 0.6447

DCE→Cl 0.3662

VC→Cl 0.568

VC→Ethene 0.4498

Table 2.7. Sorption parameters for mobile constituents.

[m3/d, cubic meters per day; O2, oxygen; SO4, sulfate; Fe(II), ferrous iron; 

CH4, methane; PCE, perchloroethene; TCE, trichloroethene; DCE, dichloroethene; VC, vinyl chloride; Cl, chloride]

Constituent Distribution 

coefficient (m3/d) Retardation factor1

Electron donor 0.0 1.0

O2 0.0 1.0

SO4 8.11 x 10–8 1.50

Tracer 8.11 x 10–8 1.50

Fe(II) 0.0 1.0

CH4 0.0 1.0

PCE 6.08 x 10–7 4.75

TCE 3.08 x 10–7 2.90

DCE 1.59 x 10–7 1.98

VC 1.95 x 10–7 2.20

Ethene 0.0 1.0

Cl 0.0 1.0

1

 Bulk density = 1.85 x 106 grams per cubic meter; total porosity = 0.30.

Appendix 2. Parameters Used to Simulate the Kings Bay Site  35
Prepared by the U.S. Geological Survey Enterprise Publishing Network

Report edited and designed by Kimberly A. Waltenbaugh and James R. Douglas

Technical Reviewers:

James Davis, U.S. Geological Survey

Lashun King, Department of Civil and Environmental Engineering, Virginia Polytechnic Institute and State University

Eve Kuniansky, U.S. Geological Survey

Dan Waddill, Naval Facilities Engineering Command

For more information concerning the research in this report, please visit the U.S. Geological Survey 

Toxic Substances Hydrology Program Web site at http://toxics.usgs.gov/

Chapelle, Novak, Parker, Campbell, and Widdowson—A Framework for Assessing the Sustainability of Monitored Natural Attenuation—Circular 1303