Monitored Natural 

Attenuation of 

Tertiary Butyl Alcohol (TBA) 

in Ground Water at 

Gasoline Spill Sites
EPA/600/R-07/100

October 2007

Monitored Natural Attenuation of Tertiary Butyl 

Alcohol (TBA) in Ground Water at Gasoline 

Spill Sites

John T. Wilson and Cherri Adair

U.S. Environmental Protection Agency 

Office of Research and Development 

National Risk Management Laboratory

Ada, Oklahoma 74820

Project Officer

John T. Wilson

Ground Water and Ecosystems Restoration Division

National Risk Management Research Laboratory

Ada, Oklahoma 74820

National Risk Management Research Laboratory

Office of Research and Development

U.S. Environmental Protection Agency

Cincinnati, Ohio 45268
ii

Notice

This document is intended for internal Agency use only. The U.S. Environmental 

Protection Agency through its Office of Research and Development funded portions 

of the research described here. Mention of trade names and commercial products 

does not constitute endorsement or recommendation for use. All research projects making conclusions and recommendations based on environmentally related 

measurements and funded by the Environmental Protection Agency are required to 

participate in the Agency Quality Assurance Program. This project was conducted 

under a Quality Assurance Project Plan for Task 10013, Fate of Fuel Oxygenates 

in Aquifer Materials. Work performed by U.S. EPA employees or by the U.S. EPA 

on-site analytical contractor followed procedures specified in these plans without 

exception. Information on the plans and documentation of the quality assurance 

activities and results are available from Cherri Adair.
iii

Foreword

The U.S. Environmental Protection Agency is charged by Congress with protecting the Nation’s land, air, 

and water resources. Under a mandate of national environmental laws, the Agency strives to formulate 

and implement actions leading to a compatible balance between human activities and the ability of natural 

systems to support and nurture life. To meet this mandate, EPA’s research program is providing data 

and technical support for solving environmental problems today and building a science knowledge base 

necessary to manage our ecological resources wisely, understand how pollutants affect our health, and 

prevent or reduce environmental risks in the future.

The National Risk Management Research Laboratory is the Agency’s center for investigation of technological and management approaches for preventing and reducing risks from pollution that threatens human 

health and the environment. The focus of the Laboratory’s research program is on methods and their 

cost-effectiveness for prevention and control of pollution to air, land, water, and subsurface resources; 

protection of water quality in public water systems; remediation of contaminated sites, sediments and 

ground water; prevention and control of indoor air pollution; and restoration of ecosystems. NRMRL 

collaborates with both public and private sector partners to foster technologies that reduce the cost of 

compliance and to anticipate emerging problems. NRMRL’s research provides solutions to environmental 

problems by: developing and promoting technologies that protect and improve the environment; advancing scientific and engineering information to support regulatory and policy decisions; and providing the 

technical support and information transfer to ensure implementation of environmental regulations and 

strategies at the national, state, and community levels.

This publication has been produced as part of the Laboratory’s strategic long-term research plan. It is 

published and made available by EPA’s Office of Research and Development to assist the user community 

and to link researchers with their clients.

Tertiary butyl alcohol (TBA) is one of the most widely distributed organic contaminants in ground water 

at gasoline spill sites. The U.S. EPA does not have a Maximum Contaminant Level (MCL) for TBA in 

drinking water. Nevertheless, many states have set standards for TBA in drinking water and clean up 

goals for TBA at gasoline spill sites. Because other contaminants, such as benzene and methyl tertiary 

butyl ether, are often biologically degraded in anaerobic ground water, the state agencies that implement the Under Ground Storage Tank program rely heavily on monitored natural attenuation to clean up 

these contaminants at gasoline spill sites. This report reviews the prospects for using monitored natural 

attenuation to manage the risk from TBA in ground water at gasoline spill sites.

The report reviews the distribution of TBA in ground water at gasoline spill sites, the process that produces TBA from anaerobic biodegradation of MTBE, and the prospects for natural biodegradation of 

TBA in ground water. The report presents data from a microcosm study conducted by U.S. EPA on TBA 

degradation in sediment from six gasoline spill sites distributed around the United States. Finally the 

report reviews the limited knowledge on use of stable carbon and stable hydrogen isotopes to evaluate 

natural biodegradation of TBA at field scale. 

Stephen G. Schmelling, Director

 Ground Water and Ecosystems Restoration Division

National Risk Management Research Laboratory

v

Contents

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

Figures ......................................................................................................................................... vii

Acknowledgments ......................................................................................................................... ix

Abstract ......................................................................................................................................... xi

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

Distribution of TBA, MTBE, and Benzene at Gasoline Spill Sites...................................................1

Interpreting TBA Biodegradation at Field Scale ..............................................................................4

Attenuation with Distance from the Source ..............................................................................4

Attenuation with Time in Individual Monitoring Wells................................................................5

Biodegradation of TBA ...................................................................................................................5

Rates of Biodegradation of TBA in Ground Water .........................................................................9

Extent of Biodegradation of TBA ..................................................................................................11

EPA microcosm study of anaerobic TBA biodegradation .............................................................15

Construction of Microcosms ..................................................................................................15

Laboratory Analytical Procedures ..........................................................................................16

Data Quality ...........................................................................................................................16

Biodegradation of TBA in Microcosms ...................................................................................17

Use of Stable Isotope Ratios .................................................................................................20

Summary ......................................................................................................................................26

Recommendations ........................................................................................................................27

Research Needs .....................................................................................................................28

References ...................................................................................................................................29

Appendix ......................................................................................................................................32
vi
vii

Figures

Figure 1. Frequency distribution of the maximum concentrations of TBA, MTBE, and Benzene 

at gasoline spill sites in Los Angeles County, California.....................................................................2

Figure 2. Distribution of MTBE and TBA at gasoline spill sites in Orange County, California, in 2002 and in 

the Eastern United States in 1999 .....................................................................................................3

Figure 3. Relationship between the fractionation of carbon isotopes in MTBE produced by biodegradation 

of MTBE to TBA and the ratio of TBA to MTBE in ground water in monitoring wells at 13 gasoline 

spill sites in Orange County, California. .............................................................................................4

Figure 4. Removal of MTBE and production or removal of TBA in anaerobic sediment from a gasoline spill 

site in Fountain Valley, California (98UT010) and a gasoline spill site in Laguna Niguel, California 

(91UT086). .........................................................................................................................................9

Figure 5. Relative importance of electron acceptors at 25 fuel spill sites in North America. .........................12

Figure 6. Distribution of sulfate at 77 gasoline spill sites in the Eastern United States. .................................12

Figure 7. General distribution of terminal electron accepting processes (TEAPs) in ground water down 

gradient from a spill of gasoline. ......................................................................................................13

Figure 8. Distribution of the concentrations of TBA and sulfate in selected monitoring wells at gasoline spill 

sites in Orange County, California. ..................................................................................................14

Figure 9. Relationship between depth below the water table and the concentrations of TBA and sulfate in 

ground water at a gasoline spill site in Port Hueneme, California. ..................................................14

Figure 10. Removal of TBA in triplicate microcosms constructed with material for a gasoline spill site in 

Petaluma, California. ........................................................................................................................18

Figure 11. Behavior of TBA in triplicate microcosms constructed with material for gasoline spill sites in Deer 

Park, New York and Parsippany, New Jersey. ..................................................................................18

Figure 12. Behavior of TBA in microcosms constructed with material for gasoline spill sites in Boca Raton, 

Florida. .............................................................................................................................................19

Figure 13. Behavior of TBA in microcosms constructed with material from a gasoline spill site at Port 

Hueneme, California and a site at Vandenberg AFB, California. .....................................................19

Figure 14. Expected relationship between the ratio of stable isotopes of carbon in MTBE or TBA and the 

extent of biodegradation of MTBE or TBA. ......................................................................................21

Figure 15. Changes in concentrations of MTBE and TBA in monitoring wells over time at three sites in 

Orange County, California, data plotted on logarithmic scale. .........................................................22

Figure 16. Changes in concentrations of MTBE and TBA in monitoring wells over time at three sites in 

Orange County, California, data plotted on arithmetic scale. ...........................................................22

Figure 17. Changes in concentrations of MTBE and TBA in monitoring wells over time at a site in Orange 

County, California, where a decline in concentration of MTBE was followed at a later time by a 

decline in concentration of TBA. ......................................................................................................23
viii

Figure 18. Relationship between the isotopic ratio of carbon in TBA in ground water at a manufacturing 

facility in Pasadena, Texas, and the concentration of TBA. .............................................................24

Figure 19. Comparison of the range of variation in the isotope ratio of carbon (d 13C) in TBA at monitoring 

wells at sites in Orange County, California, (Kuder et al., 2005) and at a site in South America 

(Zwank et al., 2005) to the range of variation in the isotope ratio of carbon (d 13C) in MTBE in the 

same monitoring wells. .....................................................................................................................24

Figure 20. Comparison of the range of variation in the isotope ratio of hydrogen (δ D) in TBA at monitoring 

wells at a site in South America (Zwank et al., 2005) to the range of variation in the isotope ratio 

of carbon (δ 13C) in MTBE in the same monitoring wells. ..............................................................25

Tables

Table 1. Rates of Anaerobic Biodegradation of TBA in Aquifer Sediments or Field Scale Plumes at 

Gasoline Spill Sites. ..........................................................................................................................8

Table 2. Distribution of TBA, MTBE, Methane, and Sulfate in Monitoring Wells at the three Gasoline 

Service Stations in Orange County, California that Provided Sediment for the Microcosm 

Study of DeVaull et al. (2003). .........................................................................................................10

Table 3. Comparison of first order rates of anaerobic biodegradation of TBA, MTBE, and Benzene............11

Appendix: 

Table 4. Typical Quality Performance Data for Analysis of TBA in Water. .....................................................32

Table 5. Typical Quality Performance Data for Analysis of Sulfate in Water. .................................................35
ix

Acknowledgments

This document was peer reviewed by Dr. Matthew Caldwell of the National Risk Management Laboratory, 

U.S. EPA, by Dr. Kevin T. Finneran of the University of Illinois, and by Dr. Yue Rong, Senior Environmental 

Scientist, California Environmental Protection Agency, California Regional Water Quality Control Board 

Los Angeles Region.
x
xi

Abstract

The state agencies that implement the Underground Storage Tank program rely heavily on Monitored 

Natural Attenuation (MNA) to clean up contaminants such as benzene and methyl tertiary butyl ether 

(MTBE) at gasoline spill sites. This is possible because the contaminants are biologically degraded in 

anaerobic ground water at the site. Tertiary Butyl Alcohol (TBA) is generally considered to be more 

readily degradable than MTBE, and there is a danger that the state agencies will consider contamination 

from TBA a good prospect for MNA. A close examination of the available information indicates that a 

default presumption that TBA is readily degraded in anaerobic ground water is not justified. Anaerobic 

biodegradation of TBA will require a supply of an electron acceptor such as sulfate or biologically available 

Iron(III) or Manganese(IV). The available survey data indicate that ground water in the source area of 

the majority of known plumes is devoid of sulfate. Although a procedure to estimate biologically available 

Iron(III) is commercially available, it is not routinely applied to gasoline spill sites. There is no established 

procedure to estimate biologically available Manganese(IV). To date, the performance of available approaches to document anaerobic biodegradation of TBA at specific field sites has been disappointing. 

These include field monitoring to show a statistically significant attenuation in concentration with distance 

along the flow path, microcosm studies conducted with sediment from the site, and analysis of stable 

isotope ratios in TBA in the plume. 
xii
1

Introduction

Tertiary butyl alcohol (TBA) is one of the most widely distributed organic contaminants in ground water at gasoline spill 

sites. The U.S. EPA does not have a Maximum Contaminant 

Level (MCL) for TBA in drinking water. Nevertheless, many 

states have set standards for TBA in drinking water and 

clean up goals for TBA at gasoline spill sites. 

The state agencies that implement the Underground 

Storage Tank program rely heavily on monitored natural 

attenuation to clean up organic contaminants in ground 

water at gasoline spill sites. There are a variety of processes that attenuate the concentrations of contaminants 

in ground water, including sorption, hydrodynamic dispersion, and biodegradation. At many gasoline spill sites, 

sorption and dispersion alone are not adequate to prevent 

the contaminants from reaching a receptor. To protect the 

receptor, the contaminant must be degraded in ground 

water. Monitored natural attenuation has been selected as 

a remedy, or part of the remedy, at certain fuel spill sites 

because it has shown that the contaminants of concern, 

such as benzene, are biologically degraded in anaerobic 

ground water (Parsons Engineering Science, 1999; Wiedemeier et al., 1999a; Wiedemeier et al., 1999b; Wiedemeier 

et al., 1999c). 

Because TBA is very soluble in water, it dissolves readily 

out of spilled gasoline into ground water, and can reach 

high concentrations in ground water. Sorption of TBA to the 

solids in the aquifer matrix is negligible (Interstate Technology & Regulatory Council, 2005). To use monitored natural 

attenuation as a remedy for TBA contamination in ground 

water at many gasoline spill sites, it will be necessary to 

demonstrate that TBA is biologically degraded in the ground 

water at the site. 

This report reviews the distribution of TBA in ground water 

at gasoline spill sites, the process that produces TBA from 

anaerobic biodegradation of MTBE, and the prospects for 

natural biodegradation of TBA in ground water. Additionally, 

the report evaluates available information on attenuation 

of TBA with distance along a flow path in ground water. 

This report reviews available information on the rates and 

extent of biodegradation of TBA in ground water. Data is 

presented from a microcosm study conducted by U.S. EPA 

on TBA degradation in sediment from six gasoline spill sites 

distributed around the United States. This report reviews the 

limited knowledge on the use of stable carbon and stable 

hydrogen isotopes to evaluate natural biodegradation of 

TBA at field scale. Finally, the report makes recommendations for using monitored natural attenuation to manage the 

risk posed by TBA in ground water at gasoline spill sites.

Distribution of TBA, MTBE, and Benzene at 

Gasoline Spill Sites

Tertiary butyl alcohol (TBA) is widely distributed in ground 

water that is contaminated by spills of gasoline. Shih 

et al., (2004) compared the distribution of the maximum 

concentrations of TBA, methyl tertiary butyl ether (MTBE), 

di-isopropyl ether (DIPE), tertiary amyl methyl ether 

(TAME), ethyl tertiary butyl ether (ETBE) and benzene in 

ground water at 868 leaking fuel tank sites in Los Angeles, 

California. The geometric mean concentration of TBA was 

1,730 μg/L. The geometric mean concentrations of MTBE, 

benzene, DIPE, TAME, and ETBE were 900, 700, 31, 24, 

and 7 μg/L. The concentration of TBA was slightly higher 

than the concentrations of MTBE and benzene, and substantially higher than the concentrations of DIPE, TAME, 

and ETBE. The frequency distribution of TBA, MTBE, and 

benzene as reported by Shih et al., (2004) is presented in 

Panel A of Figure 1.

The frequency distribution of concentrations of TBA and 

MTBE in Orange County, California are presented in Panel 

B of Figure 1 (Wilson et al., 2005a; personal communication 

Seth Daugherty, Supervising Hazardous Waste Specialist-Retired, Orange County Health Care Agency). Orange 

County is on the Pacific Coast in Southern California, just 

south of Los Angeles County. As might be expected, the 

distribution of concentrations of TBA and MTBE in Orange 

County are almost identical to the distribution in Los Angeles County.

Other regions of the United States have a similar distribution 

of concentrations of TBA. Panel C of Figure 1 presents 

data from a survey of 74 gasoline stations in the Eastern 

United States (Kolhatkar et al., 2000). A total of 41 sites 

from Pennsylvania, 8 sites from New Jersey, 5 sites from 

New York, 5 sites from Florida, 7 sites from Indiana, 3 sites 

from Maryland, 2 sites from Washington DC, and 3 sites 

from Ohio were included in the survey. At least six wells 

were sampled from each site. Panel C of Figure 1 presents 

the frequency distribution of the maximum concentration of 

TBA, MTBE, and benzene in any well at the 74 gasoline 

spill sites. There was no significant difference between 

the frequency distributions in the spill sites in the Eastern 

United States and the distribution reported by Shih et al., 

(2004) for Los Angeles, California. The concentrations of 

TBA were equivalent to the concentrations of MTBE. The 

geometric mean concentration of TBA was 1,512 μg/L while 

the geometric mean concentrations of MTBE and benzene 

were 1,724 μg/L and 447 μg/L respectively.
2

Figure 1. Frequency distribution of the maximum concentrations of TBA, MTBE, and Benzene at gasoline 

spill sites. Panel A presents data from Los Angeles County, California. Panel B presents data from 

Orange County, California in 2002. Panel C presents data from the Eastern United States in 1999. 

Concentrations of TBA are represented by red diamonds, MTBE by blue triangles, and benzene by 

black squares.
3

Although the frequency distribution of concentrations of TBA 

and MTBE are very similar in Southern California and the 

Eastern United States, the distributions of concentrations 

at individual gasoline spill sites are not similar. Panel A of 

Figure 2 compares the maximum concentration of TBA in 

any well at a site to the maximum concentration of MTBE 

in any well at a site in Orange County, California, in 2002 

(Redrawn from Figure 1 in Wilson et al., 2005a). There 

was no correlation between concentrations of MTBE and 

TBA. 

Wilson et al. (2005a) and Wilson et al. (2005b) used equilibrium partitioning theory to predict the concentrations of 

TBA and MTBE in ground water in contact with gasoline in 

Figure 2. Distribution of MTBE and TBA at gasoline spill sites. Panel A presents data from Orange County, 

California, in 2002. Panel B presents data from the Eastern United States in 1999. The figure 

compares the maximum concentration of TBA in any well at a particular gasoline spill site in any 

sampling period in the year against the maximum concentration of MTBE in any well at any sampling 

period in the year. The dashed straight lines and solid curved lines bound the range of concentrations 

of TBA and MTBE expected in water in contact with gasoline containing 11% MTBE and 0.22% TBA. 

an aquifer at reasonable values for residual gasoline. They 

assumed that gasoline spilled in Orange County was 11% 

MTBE by weight, and that the technical grade of MTBE used 

in the gasoline was 2% TBA by weight. The solid curved 

lines in Figure 2 are the range of concentrations of TBA and 

MTBE that would be expected if the residual concentration 

of gasoline in the spill varied from 1000 to 40,000 mg/kg 

total petroleum hydrocarbons. The dashed lines bound 

the range of concentrations that would be expected if concentrations of TBA or MTBE in contact with gasoline were 

attenuated by dilution and dispersion in ground water, and 

the TBA and MTBE did not sorb and were not biologically 

degraded in the ground water.
4

Many of the data in Panel A of Figure 2 fall above the range 

of concentrations of TBA that would be expected for partitioning of TBA from gasoline. TBA is produced as the first 

biodegradation product during anaerobic biodegradation of 

MTBE (Kolhatkar et al., 2002); the most plausible source of 

the high concentrations of TBA is from natural anaerobic 

biodegradation of MTBE. Recent work makes it possible 

to recognize and track biodegradation of MTBE in ground 

water by a change in the ratio of the stable isotopes of 

carbon in MTBE (Hunkeler et al., 2001; Gray et al., 2002; 

Kolhatkar et al., 2002; Kuder et al., 2005; Wilson et al., 

2005b; Zwank et al., 2005). Wilson et al., (2005a) used 

this approach to show that MTBE was being degraded at 

many of the sites in Orange County, California, and that 

the sites where MTBE was extensively degraded had high 

ratios of TBA to MTBE (Figure 3).

The distribution of TBA and MTBE in the survey of MTBE 

gasoline spill sites in the Eastern United States is presented in Panel B of Figure 2 (redrawn from Figure 3.10 in 

Wilson et al., 2003). The data are sparse, but the general 

distribution of concentrations is similar to the distribution 

in Southern California. As was the case in Southern 

California, the concentrations of TBA at many stations was 

from one hundred to one thousand-fold higher than would 

be expected from partitioning of TBA from gasoline. The 

high concentrations of TBA suggest that the TBA is largely 

produced by biodegradation of MTBE. 

Figure 3. Relationship between the fractionation of carbon isotopes in MTBE produced by biodegradation of 

MTBE to TBA and the ratio of TBA to MTBE in ground water in monitoring wells at 13 gasoline spill 

sites in Orange County, California. (Redrawn from Figure 3 in Wilson et al., 2005a)

At this writing (June 2006), the standards set by individual 

states for TBA in ground water vary from 12 μg/L to 3,200 

μg/L. Based on the frequency distribution of TBA in Orange 

County, California, approximately 5% of sites have at least 

one well with concentrations of TBA greater than 110,000 

μg/L. In the data set reported by Shih et al. (2004) for Los 

Angeles, California, 5% of sites had TBA concentrations 

greater than 97,000 μg/L. A concentration of 110,000 μg/L 

would have to be diluted 10,000 fold to meet a standard of 

12 μg/L, but only 34 fold to meet a standard of 3,200 μg/L. If 

natural attenuation is to be used as a remedy in those states 

that have clean up goals for TBA in the range of 12 to 140 

μg/L, then natural biodegradation must make a significant 

contribution to natural attenuation. The remainder of this 

report evaluates the prospects for natural biodegradation 

of TBA in ground water at gasoline spill sites.

Interpreting TBA Biodegradation at Field 

Scale

Attenuation with Distance from the Source

The most direct approach to evaluate biodegradation of 

TBA at field scale is to compare the concentration of TBA in 

highly contaminated wells in the source area to concentrations in wells down gradient. Unfortunately, this approach 

rarely works well for TBA.
5

Buscheck and Alcantar (1995) developed an equation 

that could predict the first order rate of biodegradation of 

a compound along an inferred flow path in ground water 

from the changes in concentration in monitoring wells along 

the flow path and the seepage velocity of the contaminant 

along the flow path. The equation corrects for attenuation 

due to dispersion and sorption to aquifer solids. Kolhatkar 

et al., (2000) used the approach of Buscheck and Alcantar 

(1995) to evaluate TBA biodegradation in ground water 

at 74 gasoline spill sites. To correct for dilution effects 

in monitoring wells, Kolhatkar et al., (2000) divided the 

concentration of TBA by the concentration of methane in 

ground water. 

In this approach the first-order biodegradation rate constant 

(λ) in a steady-state plume is estimated from (1) a linear 

regression of the natural logarithm of TBA concentration on 

distance along the flow path, (2) the coefficient of longitudinal dispersivity and (3) the flow velocity of the contaminant 

(equal to the seepage velocity of ground water divided by 

the retardation factor). 

λ

α

= + α 















 



 −

















v k

v

c

x

x

x 4

1 2 1

2

Where: λ = first-order biological degradation constant 

(per year),

 νc = retarded contaminant velocity in the x-direction 

(meters per year),

 αx = longitudinal dispersivity (meters)

 k/vx = rate of attenuation in contaminant concentration 

with distance along the flow path (per meter).

The rate of attenuation in contaminant concentration with 

distance along the flow path can be estimated as the negative of the slope of a regression of the natural logarithm 

of contaminant concentration on distance along the flow 

path (meters). Their criterion for biodegradation of TBA at 

a site was a rate of attenuation in contaminant concentration with distance that was statistically greater than zero. 

Specifically, the slope of the regression had to be greater 

than zero at 95% confidence.

Of 74 gasoline spill sites that they examined in the Eastern United States, they found only three sites where the 

regression for TBA was significant at 95% confidence. It is 

possible that TBA biodegradation was rare in this population 

of field sites. However, these sites were not research sites. 

The wells were installed at the direction of the regulatory 

authorities for monitoring to support a risk assessment 

and to select appropriate remedies. It is also possible that 

conventional practice for locating monitoring wells failed to 

recognize TBA biodegradation. In either case, monitoring 

using conventional practice provided little interpretable 

evidence that biodegradation of TBA was important at 

gasoline spill sites.

As a practical matter, it is difficult to distinguish changes 

in concentration due to biodegradation from changes due 

to purely physical processes. The concentration data may 

be affected by uncertainties associated with hydrological 

factors. If the ground water flow velocity is slow, the concentrations in down gradient wells may not have reached 

the maximum concentration. The plume may change flow 

direction, and carry the contaminated ground water away 

from the monitoring well. The down gradient well may be 

askew of the predominant flow path in the aquifer, and concentrations are lower because the well failed to sample the 

centerline of the plume. If there is variation in the lithology 

at the site, and the plume is confined to one flow zone and 

other zones produce clean ground water, concentrations 

in a down gradient well may be lower because the plume 

was diluted to a greater extent in the down gradient well. 

The concentrations of TBA can also be attenuated by 

hydrodynamic dispersion in the aquifer. Finally, the concentrations of TBA are related to the distribution of MTBE 

and BTEX contamination, particularly if the MTBE is being 

biodegraded to produce TBA.

Attenuation over Time in Individual 

Monitoring Wells

Because it is often difficult to compare changes in concentration of TBA from one well to another, it is important 

to maintain a good continuous monitoring record of TBA 

concentrations over time in strategically located wells. If 

the general trend in concentrations over time is down, that 

is the best evidence available from conventional monitoring 

that natural attenuation processes are reducing concentrations. However, when the only data are TBA monitoring 

data, it is impossible to attribute natural attenuation to 

natural biodegradation. Very likely most of the reduction 

in concentration is due to physical weathering of TBA, or 

its parent compound MTBE from the source area of contamination at the spill site. If a transport and fate model is 

used as part of the risk evaluation, there is no justification to 

include biodegradation in the model. Including biodegradation in the model requires direct evidence that bacteria in 

the contaminated aquifer are capable of degrading TBA, or 

in fact, have degraded TBA. In the past, evidence that the 

aquifer harbored organisms that could degrade an organic 

contaminant was attained by conducting microcosm studies. 

In recent years, more direct evidence of biodegradation of 

benzene, other BTEX compounds, and MTBE has been 

attained from the analysis of stable isotope ratios (Schmidt 

et al., 2004b; Meckenstock et al., 2004). 

The following sections review the literature on biodegradation of TBA, and then specifically consider anaerobic 

biodegradation of TBA in microcosms constructed from contaminated aquifer material. The final section evaluates the 

use of stable isotopes to recognize TBA biodegradation.

Biodegradation of TBA

At many gasoline spill sites, TBA occurs at high concentrations in groundwater, and dilution and dispersion alone 

cannot be expected to bring concentrations of TBA to clean 

up goals before the ground water reaches a receptor. At 

these sites, Monitored Natural Attenuation will only be a 
6

viable option if the TBA is biologically degraded to harmless materials. 

Organic compounds in ground water can be biologically 

degraded through a variety of mechanisms. These mechanisms are biochemical reactions. In a chemical reaction, if 

an element or compound loses an electron it is said to be 

oxidized; if the element or compound receives an electron 

it is said to be reduced. This old nomenclature goes back 

to the chemistry of metal ores. Molecular oxygen tends 

to obtain additional electrons. The chemical that lost an 

electron to oxygen is said to be oxidized. When a metal 

ore is processed to recover the pure metal, the weight of 

the pure metal that is produced is less than the weight of 

the original ore. The ore was reduced to produce the pure 

metal. In the process the metal atoms in the ore received 

electrons.

Biological metabolism is a linked series of biochemical 

oxidation/reduction reactions where one compound loses 

electrons and is oxidized and the other compound receives 

electrons and is reduced. If the organic compound is entirely oxidized to carbon dioxide, the process is termed a 

respiration. If the organic compound is oxidized, something 

else must be reduced. That “something else” is described 

as the terminal electron acceptor. Oxygen, nitrate, sulfate, 

and Iron(III) minerals can serve as the terminal electron 

acceptors during respiration. When molecular oxygen is 

reduced to water, the process is termed aerobic respiration. 

When sulfate is reduced to sulfide, the process is termed 

sulfate reduction. When nitrate is reduced to ammonia or 

molecular nitrogen, the process is called nitrate reduction 

or denitrification. When insoluble Iron(III) minerals are reduced to form soluble Iron(II) salts, the process is termed 

iron reduction and when insoluble Manganese(IV) minerals are reduced to form soluble Manganese (II) salts, the 

process is termed manganese reduction. 

In some cases, organic compounds can also serve as 

both an electron donor and an electron acceptor. These 

reactions are called fermentations. In a fermentation reaction, an organic compound is the terminal electron acceptor. Sometimes it is the same compound that acts as 

the electron donor. The most straightforward example of 

fermentation is the anaerobic biodegradation of acetate to 

produce methane and carbon dioxide. One of the carbon 

atoms in acetate is oxidized to carbon dioxide; the other is 

reduced to form methane.

Bacteria that carry out aerobic respiration are widely distributed in soil and sediment and have a great metabolic 

diversity. However, oxygen has limited solubility in water and 

is usually unavailable in contaminated aquifer sediments. 

Ambient concentrations of nitrate are usually low. As a 

result, the most important electron acceptors in contaminated ground water are sulfate and Iron(III) minerals in the 

aquifer sediment (Wiedemeier et al., 1999a). 

Biodegradation of TBA has been reviewed by Schmidt et 

al., (2004a). Degradation of TBA has been reported with 

a variety of terminal electron acceptors including oxygen, 

nitrate, sulfate, Iron(III) and Manganese(IV). Biodegradation 

of TBA is also theoretically possible through fermentation 

where part of the TBA molecule is reduced to methane or 

hydrogen and part is oxidized to carbon dioxide. Schmidt 

et al. (2004a) calculated the free energy yield for biodegradation of TBA with oxygen, nitrate, sulfate, or Iron(III) as 

an electron acceptor and for the fermentation of TBA to 

methane. The free energy yield under environmental conditions, ∆rG0

(w), was -2659 kilojoules per mole for oxygen 

as electron acceptor, -2360 kilojoules per mole for nitrate 

as electron acceptor, -814 kilojoules per mole for Iron(III) 

as electron acceptor, -171 kilojoules per mole for sulfate 

as electron acceptor, and -72.06 kilojoules per mole during 

methanogenesis. 

If the value for ∆rG0(w) is negative, then the reaction is 

theoretically possible. Based on the thermodynamics of 

the reactions, biodegradation of TBA should be theoretically 

possible under aerobic, nitrate-reducing, sulfate-reducing, 

Iron(III)-reducing, and methanogenic conditions. The more 

negative the value for ∆rG0(w), the greater the amount of 

energy available from biodegradation. Much more energy 

is available from aerobic biodegradation and nitrate reduction compared to iron reduction, sulfate reduction, and 

methanogenesis. 

When oxygen or nitrate is available, biodegradation of TBA 

in sediment is rapid and extensive. Bradley et al., (2002) 

compared biodegradation of radio-labeled TBA in surface 

water sediments. When sediment from Charleston, South 

Carolina, was amended with nitrate and incubated at 23°C 

for 198 days, 28 ± 5% of the label was recovered as carbon 

dioxide. When the sediment was amended with oxygen, 99 

± 3% of the label was recovered as carbon dioxide. When 

sediment from Laurens, South Carolina, was amended 

with nitrate or oxygen, the recovery was 70 ± 10% and 

99 ± 2% respectively.

Schirmer et al., (2003) constructed laboratory microcosms 

using sediment from the Borden field site in Ontario, Canada. The sediment was collected from a region in the aquifer 

where there was evidence of field scale biodegradation of 

MTBE. After 22 days of incubation at 10°C, the sediment 

degraded 650 μg/L TBA to a concentration of less than 

5 μg/L. Hunkeler et al., (2001) showed transitory accumulation of TBA in laboratory cultures derived from sediment 

from the Borden field site. During aerobic biodegradation 

of 10 mg/L MTBE, up to 4 mg/L of TBA accumulated. After 

the MTBE was completely degraded, the TBA began to 

degrade. The TBA was completely degraded within 15 days 

after the MTBE was no longer available. 

Kane et al., (2001) evaluated MTBE biodegradation under aerobic conditions in sediment from fuel spill sites in 

California. They added 4.5 mg/L of MTBE to sediment 

from a spill site in Palo Alto, California. When the MTBE 

was degraded, they respiked the sediment with 4.5 mg/L 

MTBE. As the MTBE was degraded, TBA accumulated to 

a maximum concentration near 2.0 mg/L. Within 25 days, 

the TBA was degraded. 

Aerobic conditions may exist at the down gradient margin of 

a plume, or in sediment where a plume discharges to aero-
7

bic surface water; however, ground water is often devoid of 

oxygen in the source area of plumes where concentrations 

of TBA are high. If Monitored Natural Attenuation is to be 

a viable remedy for TBA, then TBA must be biologically 

degraded in the absence of oxygen. 

White et al., (1986) evaluated biodegradation of TBA under 

anaerobic conditions in sediment from a refinery near the 

Schuykill River in Philadelphia, Pennsylvania. They incubated 12 grams of sediment and 4 ml of sterile ground water 

without headspace in test tube microcosms at 12°C. After a 

period of acclimation, the TBA was completely removed. 

Zenker et al., (1999) constructed microcosms using sediment from a site in North Carolina that had been exposed to 

MTBE. The microcosms were constructed under anaerobic 

conditions. After 200 days of incubation, TBA was completely consumed in each of triplicate microcosms.

Bradley et al., (2001) added radio-labeled [14C] TBA to microcosms prepared using stream and lake-bed sediments 

from sites in Charleston, South Carolina; Jacksonville, 

Florida; and Picatinny Arsenal, New Jersey. When the 

microcosms were prepared to simulate anaerobic conditions, 8%± 1%, 11% ± 1%, and 8% ± 1% of the label was 

recovered as carbon dioxide after 166 days of incubation. 

Finneran and Lovley (2001) reported rapid and extensive 

degradation of TBA under mixed iron(III)-reducing, sulfatereducing, and methanogenic conditions in sediment from 

the Potomac River. They added radio-labeled [14C] TBA to 

the sediment. After 62 days of incubation, approximately 

7% of the label was recovered as methane and 25% of the 

label was recovered as carbon dioxide. If the methane had 

been produced by fermentation of TBA, the molar ratio of 

methane produced to carbon dioxide produced should have 

been three to one. If TBA is degraded by sulfate-reducing 

bacteria or Iron(III)-reducing bacteria, all the label should 

go to carbon dioxide. The actual ratio was near 0.3 to one. 

Finneran and Lovley (2003) concluded that TBA in the 

sediment from the Potomac River was being degraded by 

a variety of terminal electron accepting processes, which 

included sulfate reduction and Iron(III) reduction. 

In a subsequent experiment, Finneran and Lovly (2003) 

added additional sulfate to the sediment from the Potomac 

River. If the addition of sulfate allowed the sulfate-reducing 

bacteria to compete with methanogenic bacteria for the 

TBA, the fraction of radio-label going to methane should 

be smaller. When sulfate was not added, approximately 

24% of the label was recovered as carbon dioxide and 

10% of the label was recovered as methane. The ratio of 

label in methane to label in carbon dioxide was near 0.4 

to one. When 960 mg/L of sulfate was added to sediment 

from the Potomac River, approximately 26% of the label 

was recovered as carbon dioxide and approximately 3% of 

the label was recovered as methane, and the ratio of label 

in methane to label in carbon dioxide was near 0.1 to one. 

The addition of sulfate inhibited the production of methane. 

Molybdate is a specific inhibitor of sulfate reduction. When 

Finneran and Lovley (2003) added sulfate and molybdate to 

sediment from the Potomac River, approximately 23% of the 

label was recovered as carbon dioxide and approximately 

3% of the label was recovered as methane. Because the 

molybdate did not produce a substantial reduction in the 

amount of label in carbon dioxide, Finneran and Lovley 

(2003) concluded These results indicate that two processes 

are probably responsible for the [14C]-TBA mineralization 

– sulfate reduction and Fe(III) reduction (because nitrate 

was not detected in these sediments).

Bradley et al., (2002) compared anaerobic biodegradation of radio-labeled [14C] TBA in surface water sediments 

under various electron accepting conditions. In sediment 

from Charleston, South Carolina, the sediment as collected 

was methanogenic. After 200 days of incubation, none of 

the 14C from TBA was recovered as carbon dioxide. They 

should have been able to detect 2% of the original label 

as carbon dioxide. When the sediment was amended with 

Iron(III) or Manganese(IV), no 14C from TBA was detected as 

carbon dioxide. However, when the sediment was amended 

with sulfate, 4 ± 1% of the 14C from TBA was recovered as 

carbon dioxide. 

In sediment from a site in Laurens, South Carolina, the 

unamended sediment was also methanogenic. After 200 

days of incubation, none of the labeled TBA was recovered 

as carbon dioxide. When the sediment was amended 

with Iron(III), none of the labeled TBA was recovered as 

carbon dioxide. When the sediment was amended with 

Manganese(IV), 75 ± 20% of the label was recovered as 

carbon dioxide. When the sediment was amended with 

sulfate, 5 ± 1% of the label was recovered as carbon 

dioxide. 

Finneran and Lovley (2003) added radio-labeled [14C] acetate to material from a contaminated aquifer near Bemidji, 

Minnesota. Based on the distribution of label between 

methane and carbon dioxide, they concluded that methanogenesis was the only important electron accepting process 

in the sediment. They added radio-label [14C] TBA to this 

sediment. After 23 days of incubation, approximately 5% 

of the 14C from TBA was recovered as methane and approximately 9% was recovered as carbon dioxide.

The interpretation of their experiments on TBA biodegradation under sulfate reducing and methanogenic conditions is 

not straightforward. Some portion of the 14C from TBA is 

recovered as methane, indicating that at a minimum some 

transformation product of TBA can be degraded under 

methanogenic conditions. However, the yield of methane 

is much lower than would be expected from the complete 

fermentation of TBA under methanogenic conditions. 

C4H10O + H2O → 3CH4 + CO2

If TBA is fermented to methane according to the stoichiometry above, 75% of label in [14C] TBA should be recovered 

as methane. In sediment where TBA is degraded in the 

absence of oxygen or nitrate, approximately 36% to 10% 

of the label is recovered as methane. This would suggest 

that some portion of the TBA is being oxidized by sulfate, 

Iron(III), or Manganese(IV). 
8

There is not unequivocal evidence that TBA can be metabolized under strictly fermentative conditions. In their 

review, Schmidt et al., (2004a) concluded Thus, in contrast 

to MTBE degradation, there is general consensus that 

TBA is recalcitrant under methanogenic conditions. If this 

is the case, either sulfate or biologically available Iron(III) 

or Manganese(IV) is required for biodegradation of TBA in 

the absence of oxygen or nitrate. 

As part of a study to explain the distribution and behavior 

of TBA in ground water at gasoline spill sites in Orange 

County, California, DeVaull et al., (2003) collected sediment from three locations at each of three gasoline spill 

sites. The sediment was used to construct microcosms. 

One experimental treatment in their report simulated the 

natural anaerobic conditions in the aquifer at their study 

sites. A 160 ml glass vial was filled with 75 ml of water 

and 50 g of wet sediment. The headspace was purged 

with nitrogen for 30 minutes and then the vial was sealed 

with a septum. The microcosms were spiked with MTBE 

and TBA to bring the concentrations in the pore water to 

approximately 10 mg/L. The ratio of water to solids in the 

experimental system was approximately 2.1 to 1 (gm/gm). 

The ratio of water to solids in a sandy aquifer is near 0.25 

to 1 (gm/gm). Because the experimental system had a high 

water to solids ratio, and because both the TBA and soluble 

electron acceptors such as sulfate were supplied in water, 

the microcosm system was more sensitive than a natural 

aquifer to the contribution of soluble electron acceptors and 

less sensitive to the contribution of Iron(III). 

The concentrations of TBA, MTBE, methane, and sulfate in 

monitoring wells at the sites that provided sediment for the 

microcosm study of DeVaull et al., (2003) are presented 

in Table 1. The water samples were collected in 2002 

and 2003. All the wells at all three sites had much higher 

concentrations of TBA than MTBE. There were high concentrations of methane in at least one well from each site. 

As will be discussed later, the oxidation of TBA by sulfate 

reduction requires 3.9 mg/L sulfate for each 1.0 mg/L of 

TBA. All the wells at the site at Fountain Valley, California, 

and Laguna Niguel, California, had adequate concentrations 

of sulfate to support complete metabolism of TBA by sulfate 

reduction. The site at San Juan Capistrano, California, did 

not have adequate sulfate to meet the demand associated 

with the TBA.

The terminal electron accepting processes (TEAP) were not 

defined in the treatment that simulated natural anaerobic 

conditions in the aquifer. Sulfate and nitrate were available 

at the beginning of the experiment, but DeVaull et al., (2003) 

did not present data on depletion of nitrate or sulfate, or 

accumulation of Iron(II).

In microcosms constructed with three different sediment 

samples from a site in Fountain Valley, California, MTBE 

was degraded in all three microcosms after 91 days of incubation. In two of the three microcosms, TBA was either 

completely degraded or substantially degraded within 161 

days of incubation (Panel A of Figure 4 presents data from 

one of the microcosms). 

In microcosms constructed with three different sediment 

samples from a site in San Juan Capistrano, California, TBA 

was completely degraded in two of the three microcosms 

within 114 days, but MTBE did not degrade within 154 days. 

In the third microcosm, degradation followed the opposite 

pattern; MTBE degraded within 70 days, but TBA did not 

degrade within 154 days. 

In microcosms constructed with three different sediment 

samples from a site in Laguna Niguel, California, TBA 

degraded within 90 days in one microcosm and 145 days 

Table 1. Distribution of TBA, MTBE, Methane, and Sulfate in Monitoring Wells at the three Gasoline Service 

Stations in Orange County, California that Provided Sediment for the Microcosm Study of DeVaull et 

al. (2003). 

Station Well TBA MTBE Methane Sulfate

ml/L ml/L ml/L ml/L

Fountain Valley 16969 mw-6 28.0 0.079 0.0982 183

Fountain Valley 16969 mw-15 2.0 0.015 0.0293 128

Fountain Valley 16969 mw-13 14.8 0.012 0.535 79

Laguna Niguel 30011 mw-7 45.9 0.042 1.02 1290

Laguna Niguel 30011 mw-10 33.2 0.028 0.953 2640

Laguna Niguel 30011 mw-11 40.3 0.229 0.0176 2680

San Juan Capistrano 27101 w-2 10.8 0.230 4 11

San Juan Capistrano 27101 mw-8 12.0 0.002 1.56 17

San Juan Capistrano 27101 mw-11 69.9 0.084 0.835 6
9

Rates of Biodegradation of TBA in Ground 

Water

The rates of attenuation of TBA in sediment under aerobic 

conditions are rapid. Schirmer et al., (2003) reported a 

first order rate of removal of TBA in sediment from the 

Borden field site of 0.12 per day. The initial concentration 

of TBA was 0.7 mg/L. Wilson et al., (2002) conducted a 

microcosm study with sediment from a gasoline spill site 

at Vandenberg AFB, California. The sediment was spiked 

with MTBE, and then spiked again six more times after the 

MTBE was degraded. The spiked concentrations of MTBE 

were as high as 16 mg/L. The first order rates of removal of 

MTBE were 0.05, 0.15, 0.03, 0.07, and 0.08 per day. TBA 

was produced by metabolism of MTBE. In each cycle of 

spike and removal of MTBE, the TBA never accumulated 

to any significant extent. The rate of removal of MTBE was 

the rate of production of TBA. TBA would be expected to 

accumulate unless the rate of removal of TBA was faster 

than the rate of removal of MTBE. 

Kane et al., (2001) spiked and respiked MTBE into sediment 

from a gasoline spill site in Palo Alto, California. Each spike 

of MTBE was near 5 mg/L. As the MTBE degraded, TBA 

accumulated to concentrations near 2 mg/L. Once MTBE 

was entirely degraded, the concentrations of TBA started to 

decline. The rate of removal of TBA corresponded to first 

order rates of degradation of 0.10 and 0.13 per day.

The rates of aerobic biodegradation are so fast that they 

are effectively instantaneous in the context of ground water 

movement at field scale. Shih et al., (2004) reported the distribution of the maximum concentration of TBA in monitoring 

wells at gasoline spill sites in Los Angeles, California. The 

median of the TBA concentrations was 1,880 μg/L. A first 

order rate of aerobic biodegradation of 0.1 per day would 

bring the median concentration of TBA to the California 

Action Limit of 12 μg/L in fifty days. 

The reported first order rates of biodegradation of TBA 

under anaerobic conditions (as reported in Table 2) are 

slower than under aerobic conditions. In their study of TBA 

degradation in sediment from a refinery near Philadelphia, 

Pennsylvania, White et al., (1986) reported their data as 

a zero order rate of utilization. At initial concentrations of 

1.7, 13.5, and 150 mg/L, the utilization rates were 0.057, 

0.52 and 2.00 mg/L/day respectively. A first order rate 

constant can be thought of as a zero order rate of change 

in concentration (∆C/∆t) normalized to the concentration at 

the particular instance in time over which the rate is operating. By definition of a first order rate constant (k), first order 

rate = ∆C/∆t = k*C; therefore k = (∆C/∆t)/C). We estimated 

the first order rate constants for attenuation of TBA at the 

beginning of the experiment by dividing the reported zero 

order rates of utilization by the initial concentration of TBA 

in the pore space of the microcosms. At an initial concentration of 1.7 mg/L and a zero order rate of utilization of 

0.057 mg/L/day, the first order rate constant for degradation 

was 0.057 mg/L/day divided by 1.7 mg/L or 0.0335 per day, 

equivalent to a rate of 12 per year. At an initial concentration 

of 13.5 mg/L, the first order rate constant was 14 per year. 

Figure 4. Removal of MTBE and production or 

removal of TBA in anaerobic sediment 

from a gasoline spill site in Fountain 

Valley, California (98UT010) and a 

gasoline spill site in Laguna Niguel, 

California (91UT086). Plotted from data 

in DeVaull et al. (2003). Data are plotted 

from one of three microcosm studies 

constructed with sediment from each of 

the sites. 

in a second microcosm, but MTBE did not degrade within 

196 days in either of the two microcosms. In the third microcosm, degradation followed the opposite pattern; TBA 

did not degrade within 196 days, but MTBE completely 

degraded within 90 days (Panel B of Figure 4).

The results of the microcosm studies reported by DeVaull 

et al., (2003) suggest that conditions that are favorable for 

anaerobic biodegradation of TBA are widespread in Orange 

County, California. At least one microcosm from each of 

the three sites degraded TBA under anaerobic conditions. 

However, the microcosm studies also suggest that the capacity to degrade TBA may be heterogeneously distributed 

at particular sites; TBA failed to degrade in at least one of 

the microcosms constructed with one of the three different 

sediment samples from each of the three sites. 
10

Table 2. Rates of Anaerobic Biodegradation of TBA in Aquifer Sediments or Field Scale Plumes at Gasoline 

Spill Sites. If More than one Rate is Reported from Microcosm Experiments Constructed from 

Material from One Location at a Site, all the Rates are Reported in the Table.

Source of Material Field 

Rate Lab Rate

First Order Rate 

of Biodegradation 

(per year)

Reference

Fountain Valley, 

California X 26, 17 DeVaull et al. (2003)

Laguna Niguel, 

California X 18, 5.1 DeVaull et al. (2003)

Potomac River X 15 Finneran and Lovley 

(2001)

Philadelphia, 

Pennsylvania X

14, 

12, 

4.9

White et al. (1986)

New York X 8.8 Kolhatkar et al.

(2000)

Florida X 7.3 Kolhatkar et al.

(2000)

Pennsylvania X 7.2 Kolhatkar et al.

(2000)

San Juan Capistrano, 

California X 5.1 

3.5 DeVaull et al. (2003)

Pasadena, Texas

flow path from location 150 X 1.1 Day and Gulliver (2003)

Pasadena, Texas

flow path from location 57 to X 0.97 Day and Gulliver (2003)

Pasadena, Texas

flow path from location 165 X 0.26 Day and Gulliver (2003)

At an initial concentration of 150 mg/L, the first order rate 

constant was 4.9 per year. As would be expected for a first 

order process, the estimated first order rates of attenuation 

were essentially independent of concentration.

Finneran and Lovley (2001) conducted spike and respike 

experiments with TBA in sediment from the Potomac River. 

The spiked concentrations were approximately 50 mg/L. 

The time course of TBA degradation reported in Figure 3 

of Finneran and Lovley (2001) corresponds to a first order 

rate constant of 15 per year. 

In the microcosm experiment conducted by DeVaull et al., 

(2003) the rates of TBA biodegradation in the anaerobic 

microcosms where TBA degraded were in the range of 26 

to 3.5 per year (Table 2). The rates in sediment samples 

from Fountain Valley, California, were 26 and 17 per year. 

The rates in sediment samples from San Juan Capistrano, 

California, were 3.5 and 5.1 per year. The rates in sediments from Laguna Niguel, California were 18 and 5.1 per 

year.

There are also a few reports on the rate of TBA biodegradation under anaerobic conditions at field scale (Table 2). 

Kolhatkar et al., (2000) evaluated the distribution of TBA at 

74 gasoline spill sites in the Eastern United States. They 

used the approach of Buscheck and Alcantar (1995) to 

extract first order rate constants for biodegradation of TBA 

along the flow path in the plume. They were able to extract 

rate constants that were statistically significant at 95% 

confidence from three of the 74 plumes. The rate of TBA 

biodegradation in a plume in New York was 8.8 per year, 

the rate in a plume in Pennsylvania was 7.26 per year, and 

the rate in a plume in Florida was 7.3 per year.

Day and Gulliver (2003) evaluated the natural attenuation of a large plume of TBA at a chemical manufacturing 

plant at Pasadena, Texas. They compared the reduction in concentration of TBA with distance along the flow 

path to the reduction in concentration of the co-occurring 

contaminants 1,1-dichloroethene and 1,1-dichloroethane. 

The attenuation of TBA was significantly faster than that 
11

of 1,1-dichloroethene and 1,1-dichloroethane. They used 

the approach of Buscheck and Alcantar (1995) to extract 

first order rate constants for natural biodegradation of TBA 

along three flow paths in the aquifer. The rate constants 

were 0.97, 0.26, and 1.1 per year.

The rates of TBA biodegradation under anaerobic conditions (when biodegradation occurs) vary over two orders 

of magnitude (Table 2). The rates are faster than the rates 

reported for anaerobic biodegradation of MTBE or for benzene, which is the usual “risk-driver” at gasoline spill sites 

(Table 3). If the microbial community at a particular spill 

site acclimates to anaerobic biodegradation of TBA, and if 

there is an adequate supply of the metabolic requirements 

for anaerobic degradation, then natural anaerobic biodegradation of TBA can provide a substantial contribution to 

natural attenuation of TBA.

Extent of Biodegradation of TBA 

The extent of TBA biodegradation by aerobic respiration, 

nitrate reduction, Iron(III) reduction or sulfate reduction will 

depend on the supply of these electron acceptors in the 

ground water or aquifer sediment. Wiedemeier et al., (1995) 

compared the concentrations of oxygen, nitrate, Iron(II) 

and sulfate in ground water in the interior of 25 fuel spills 

to the concentrations in the surrounding ground water that 

had not been affected by the fuel spill. They calculated the 

consumption of each of the soluble electron acceptors and 

the production of soluble Iron(II) in the ground water from 

insoluble Iron(III) compounds in the aquifer solids. Then 

they compared the concentrations of electrons that were 

transferred by each of the electron accepting processes on 

a milliequivalent per liter basis. Their results are presented 

in Figure 5. Sulfate-reduction was the dominant electron 

accepting process.

Figure 6 presents data on the concentration of sulfate in 

monitoring wells at 77 gasoline spill sites in the Eastern 

United States (unpublished data from study published in 

Kolhatkar et al., 2000). The figure compares the maximum 

sulfate concentration in any well at each site and the minimum sulfate concentration in any well at each site. The 

maximum concentration is an estimate of the ambient 

concentration in ground water in the aquifer. The minimum 

concentration is an estimate of the sulfate concentration 

in the LNAPL source area of the gasoline spill. At 75% 

of sites, the ambient concentration of sulfate was at least 

56 mg/L, at 50% of sites the concentration was at least 

107 mg/L, and at 25% of sites the concentration was at 

least 304 mg/L. 

Oxidation of TBA by sulfate reduction requires 3.9 mg/L 

sulfate for each 1.0 mg/L of TBA:

C4H10O + 3SO4

-2 → 3S-2 + 4CO2

 + 5H2O

Ambient concentrations of 56, 107 and 304 mg/L sulfate 

can support degradation of 14, 27, and 78 mg/L of TBA.

In the most typical scenario at a gasoline spill site, a plume 

of MTBE in ground water is produced and sustained by 

continual dissolution of MTBE from the residual gasoline to 

ground water. The major portion of the TBA is produced by 

biodegradation of the MTBE once it is dissolved in ground 

water. Other soluble components of gasoline, such as 

the BTEX compounds also dissolve in ground water. The 

demand for oxygen, nitrate, and sulfate for biodegradation 

of the other components of the gasoline depletes sulfate 

and other soluble electron acceptors from the ground water 

that are in contact with the residual gasoline. 

Based on the supply of electron acceptors, and the free 

energy available from the reaction, anaerobic regions 

in aquifers tend to develop distinct areas dominated by 

different electron acceptors (Finneran and Lovley, 2003; 

Wiedemeier et al., 1999a). When gasoline is spilled into 

an aquifer, oxygen is depleted first, then nitrate is depleted, 

then sulfate is depleted, and finally the terminal electron 

accepting process transitions to methanogenesis. As a 

consequence, the ground water immediately adjacent to 

a gasoline spill is often methanogenic. This region is surrounded by ground water that is sulfate reducing, which in 

turn is surrounded by ground water that is Iron(III)-reducing, 

which in turn is surrounded by ground water that is nitrate 

reducing. Although one electron accepting process tends to 

dominate, they often proceed concomitantly. If the supply of 

biologically available Iron(III) is adequate, Iron(III) reduction 

Table 3. Comparison of First Order Rates of Anaerobic Biodegradation of TBA, MTBE, and Benzene.

TBA MTBE Benzene

number of rates 14 10 20

mean (per year) 9.2 1.0 3.7

median (per year) 7.3 0.41 Not Provided

Reference this report Wilson 

(2003)

Suarez and Rifai 

(1999)
12

Figure 5. Relative importance of electron acceptors at 25 fuel spill sites in North America. Adapted from 

Wiedemeier et al., 1995.

Figure 6. Distribution of sulfate at 77 gasoline spill sites in the Eastern United States. The maximum 

concentration represents the likely concentration in ambient ground water that was not impacted 

by the spill. The minimum concentration represents the concentration in the LNAPL source area of 

ground water contamination. Unpublished data from Kolhatkar et al. (2000).
13

Figure 7. General distribution of terminal electron accepting processes (TEAPs) in ground water down gradient 

from a spill of gasoline.

can occur in ground water that is also sulfate reducing or 

is methanogenic. Sulfate reduction and methanogenesis 

can occur together, particularly at lower concentrations of 

sulfate. These relationships are depicted diagrammatically 

in Figure 7.

The concentration of sulfate at which sulfate becomes 

limiting for sulfate reduction varies widely from one organism to the next and with different environmental conditions. Ingvorsen and Jørgensen (1984) found that the half 

saturation constant for sulfate reduction in four strains of 

bacteria varied from 0.5 mg/L to 7 mg/L. Fukui and Takii 

(1994) showed that the half saturation constant for sulfate 

reduction of a Desulfovibrio desulfuricans was 0.8 mg/L 

when the cells were associated with particles of FeS, and 

24 mg/L when the cells were free living. Concentrations of 

sulfate less than 5 mg/L are considered sulfate poor (Fukui 

and Takii, 1994), and flow of electrons can be expected to 

transition to methanogenesis at sulfate concentrations less 

than 40 mg/L (Personal Communication Kevin T. Finneran, 

University of Illinois, August, 2006). 

If sulfate has been depleted in contaminated ground water, 

it is reasonable to presume that nitrate and oxygen are also 

depleted (Finneran and Lovley, 2003; Lovley, et al., 1994; 

Wiedemeier et al., 1999a). In 62 of the 77 sites in the 

survey of gasoline spills in the Eastern United States, the 

concentration of sulfate was less than 1.0 mg/L in the most 

contaminated wells in the LNAPL source area (Figure 6). 

Based on the available knowledge of anaerobic biodegradation of TBA, biodegradation of TBA supported by aerobic 

respiration, nitrate reduction, or sulfate reduction would not 

be expected in the source area of these plumes. 

If the biodegradation of TBA in ground water is limited by 

the supply of sulfate, there should be an inverse relationship between the concentration of TBA in ground water 

and the concentration of sulfate. Figure 8 compares the 

concentration of sulfate to the concentration of TBA in 58 

wells at 13 gasoline spill sites in Orange County, California. 

These are the same wells sampled by Wilson et al., (2005c) 

to determine whether stable isotope ratios in MTBE could 

be used to predict MTBE biodegradation in ground water. 

In general, an inverse correlation did apply to the data set 

from these wells (Figure 8), but there is more scatter than 

would be expected. 

Part of the scatter of the data may be an artifact caused by 

mixing of ground water from different plumes in monitoring 

wells. Most conventional monitoring wells are screened 

over ten or fifteen vertical feet (3.05 to 4.57 meters) in the 

aquifer. Many plumes at gasoline spill sites are vertically 

heterogeneous over ten to fifteen feet. This relationship is 

illustrated in Figure 9 using data from a gasoline spill site 

at Port Hueneme, California. The plume was in a layer of 

sands and gravels. Water samples were taken with temporary push wells with a vertical screened interval of 1.5 feet 

(0.46 meters). Samples started just below the water table 

and extended to a clay confining layer. The ground water 

in the shallow intervals had low concentrations of sulfate 

and high concentrations of TBA while water in the deeper 

interval had high concentrations of sulfate and low concentrations of TBA. However, water from a well screened 

across this aquifer would have intermediate concentrations 

of both TBA and sulfate. The concentrations of TBA and 

sulfate in the water produced by the monitoring well would 
14

Figure 9. Relationship between depth below the water table and the concentrations of TBA and sulfate in 

ground water at a gasoline spill site in Port Hueneme, California.

Figure 8. Distribution of the concentrations of TBA and sulfate in selected monitoring wells at gasoline spill sites 

in Orange County, California.
15

suggest that sulfate was available for anaerobic biodegradation of TBA. 

The TBA at this site was produced through biodegradation 

of MTBE that dissolved out of residual gasoline near the 

water table (data not shown). Biodegradation of alkylbenzenes (the BTEX compounds) associated with the residual 

gasoline is probably responsible for the lower concentrations of sulfate near the water table. The mixed water 

sample produced by a conventional monitoring well in this 

aquifer would not have produced a water sample that was 

representative of the water surrounding a particular sulfate 

reducing bacterium. 

In general, if water produced from a monitoring well is 

devoid of sulfate, then sulfate would not be available to 

bacteria in the ground water sampled by the well. The inverse is not necessarily true. The presence of both sulfate 

and TBA in water produced by a monitoring well does not 

mean that sulfate is available for organisms to degrade the 

TBA. The sulfate and TBA may have come from different 

depth intervals.

If sulfate is depleted in the source area of a gasoline spill, 

anaerobic biodegradation of TBA will require admixture of 

TBA in the contaminated ground water in the plume with 

sulfate in the receiving ground water in the aquifer down gradient. Mixing by dispersion is controlled by flow of ground 

water and the geometry of the plume. The mixing occurs 

across the interface between the plume and the ambient 

ground water. If a plume has a large volume relative to the 

surface area presented to the ambient ground water, then 

mixing by dispersion will be a slow process, and adequate 

admixture of sulfate to meet the stoichiometric demand for 

biodegradation of TBA will likely occur at some distance 

away from the source area. 

At most gasoline spill sites, the monitoring wells are on 

the property of the gasoline station or on the property of 

the immediate neighbors. The existing monitoring wells 

may not be located in the portion of the plume where 

sulfate is available to support biological degradation of 

TBA. At many gasoline spill sites, it may be impossible to 

use monitoring wells that are in or near the source area 

to document natural biodegradation of TBA carried out by 

sulfate reducing bacteria. 

If TBA is not biologically degraded under methanogenic 

conditions, the only plausible agents for biodegradation of 

TBA in ground water that has been depleted of sulfate and 

the other soluble electron acceptors are Iron(III) reducing 

bacteria or Manganese(IV) reducing bacteria. If sulfate 

and the other soluble electron acceptors are depleted, the 

extent of biodegradation of TBA will be limited by the supply of biologically available Iron(III) or Manganese(IV) in 

the aquifer sediments. At present there is a commercially 

available assay for biologically available Iron(III) in sediment, 

(Bioavailable Ferric Iron (BAFe III) Assay, New Horizons 

Diagnostics Corp., 9110 Red Branch Road, Columbia, 

Maryland USA 21045, 1-800-888-5015 ext. 0 or 235, fax: 

410-992-0328, NHDiag@aol.com ). The performance of 

the assay to predict biologically available Iron(III) has been 

evaluated by the Environmental Security and Technology 

Certification Program (ESTCP, 2005). However, to the 

authors’ knowledge, the assay has never been used to 

evaluate natural attenuation of TBA in ground water. There 

is no commercially available assay for biologically available 

Manganese(IV). Until techniques to directly evaluate the 

supply of biologically available Iron(III) and Manganese(IV) 

in aquifer sediment become standard practice, the potential 

contribution of Iron(III)-reducing and Manganese(IV)-reducing bacteria to the natural attenuation of TBA cannot be 

characterized. It is inappropriate to attribute apparent disappearance of TBA from ground water to Iron(III)-reducing or 

Manganese(IV)-reducing bacteria simply because Iron(II) 

or Manganese(II) accumulates in ground water.

EPA microcosm study of anaerobic TBA biodegradation

To provide an independent evaluation of the natural anaerobic biodegradation of TBA in ground water at gasoline 

spill sites, EPA/ORD conducted microcosm studies using 

material from several gasoline spill sites around the United 

States. Sediment was collected from BP gasoline stations 

at Petaluma, California; Deer Park, New York; Parsippany, 

New Jersey; and Boca Raton, Florida. The microcosms 

were part of a larger study that examined the anaerobic 

biodegradation of MTBE and ethanol. Sediment was also 

collected from a motor gasoline spill site at a U.S. Navy 

Base at Port Hueneme, California, and a motor gasoline 

spill site at Vandenberg Air Force Base, California. All the 

samples were from shallow aquifers in sandy unconsolidated sediments. Sediment from Deer Park, Parsippany, 

and Boca Raton were selected for the microcosm study 

because field data collected in 1999 indicated that the 

rates of degradation of MTBE and TBA in ground water, 

using the method of Buscheck and Alcantar (1995), were 

statistically significant at 80% confidence. Sediment from 

Parsippany was selected because the field data indicated 

that MTBE was being degraded, and TBA was not accumulating. The sediments from Port Hueneme and Vandenberg 

AFB were selected as negative control sites. At the time 

the microcosms were constructed, there was no evidence 

of biodegradation of MTBE or TBA in the plumes at Port 

Hueneme and Vandenberg AFB.

Construction of Microcosms

The sediment was collected and stored in 1-L glass jars. 

To protect the anaerobic microorganisms that might be 

present in the samples from oxygen in the atmosphere, the 

head space above the sediment was replaced with ground 

water. The sediment samples were shipped by air freight 

and were stored at 4 °C until they were used to construct 

microcosms.

All manipulations to prepare the microcosms were carried 

out aseptically in an anaerobic glove box. An oxygen meter 

indicated that the concentration of oxygen in the atmosphere of the glove box was less than 1 ppm by volume. 

Microcosms were prepared in sterile glass serum bottles 

with a volume of 25 mL. When available, ground water from 

monitoring wells at the sites was added to the sediment to 
16

make a slurry, and the sediment samples were stirred to 

blend well. If ground water was not available, the slurry was 

made with autoclaved reverse osmosis water. The added 

water was 5% or less of the final volume of the slurry. The 

slurry was transferred to the serum bottles with a sterile 

scoop. Each microcosm received approximately 45 gm of 

slurry and 1.0 mL of an aqueous dosing solution containing 

a sterile aqueous solution of TBA. The concentration of 

TBA in the dose solution for microcosms constructed with 

sediment from Deer Park, New York; Petaluma, California; 

Vandenberg AFB, California; Parsippany, New Jersey; 

and Boca Raton, Florida, varied from 13 to 15 mg/L. The 

microcosms constructed with sediment from Port Hueneme, California, had 80 mg/L TBA. The microcosms were 

sealed with a sterile Teflon-faced gray butyl rubber septum 

and a crimp cap. The microcosms were stored at room 

temperature (20 to 22 °C) in the same glove box, under an 

atmosphere containing 2% to 5% v/v hydrogen. 

They were incubated from eighteen months to two years. 

Every two to three months, triplicate microcosms were 

selected for analysis. Prior to sampling, the contents of 

each microcosm were mixed with a vortex mixer while the 

microcosms were still sealed, and then the sediment was 

allowed to settle. The septum was removed and approximately 1 mL of the standing water was taken and diluted in 

14 ml of distilled water containing 1% trisodium phosphate 

as a preservative. The diluted samples contained approximately 15 ml of diluted water and 6.0 ml of head space. 

The diluted samples were sealed with a septum and crimp 

cap, and then shaken to bring the water and head space 

to equilibrium. 

Laboratory Analytical Procedures

The concentrations of TBA were determined by head space 

gas chromatography/mass spectrometry (GC/MS) using 

a modification of EPA Method 5021A, “Volatile Organic 

Compounds in Various Sample Matrices using Equilibrium 

Headspace Analysis,” June 2003. Samples were collected 

for analysis with an automated static headspace sampler. 

Analytes were determined by gas chromatography/mass 

spectrometry using an Ion Trap Detector. The lowest calibration standard was 10 μg/L; the method detection limit 

was 2.4 μg/L. Concentrations of sulfate were determined 

with a Waters Quanta 4000 Capillary Ion Analyzer, using 

a modification of EPA Method 6500, “Dissolved Inorganic 

Anions in Aqueous Matrices by Capillary Ion Electrophoresis,” January 1998. The method detection limit for sulfate 

was 0.172 mg/L. 

Depending on the amount of standing water that was 

sampled from each microcosm, the pore water in the microcosms was diluted in a range between 15:1 and 30:1 before 

analysis. In the samples with concentrations of TBA below 

the lower calibration limit, the pore water in the microcosms 

was diluted 15:1 before analysis. As a result, the effective 

lower limit for calibration of TBA in the original pore water of 

the microcosms was 150 μg/L, and the effective detection 

limit was 36 μg/L. The effective method detection limits for 

sulfate in diluted samples were 1.2 mg/L.

Data Quality

Laboratory analyses for data presented in Panel C of 

Figure 1, in Panel B of Figure 2, Figure 6, Figures 8 

through 13, and Table 1 were conducted at the R.S. Kerr 

Environmental Research Center in accordance with a 

Quality Assurance Project Plan prepared for in-house task 

10013 (Fate of Fuel Oxygenates in Aquifer Material). 

Major quality assurance and quality control (QA/QC) 

evaluations for the analyses included method blank (MB), 

continuing calibration check (CCC), second source check 

(QC) using a sample obtained from the second source as 

identified by their designated names, laboratory duplicates 

(LD), and matrix spike (MS). A method blank was analyzed 

in the beginning and end of a sample set. Continuing calibration check standards (CCC) were analyzed every ten 

samples as well as in the beginning and end of a sample 

set. QC checks were analyzed every ten samples. Lab 

duplicates were analyzed every ten samples. Matrix spikes 

were analyzed every twenty samples. 

The data quality objectives for TBA were as follows: 

The target analyte in the method blank would be below 

method detection limit. The reported concentration of the 

continuing calibration check standard (CCC) would agree 

with the expected concentration plus or minus 20% of the 

known concentration: the matrix spike would agree with the 

expected concentration plus or minus 30% of the known 

concentration (i.e., Recovery of the expected value would 

be in the range of 70-130%). Laboratory duplicates would 

agree with each other with a relative percent difference of 

± 25%. 

The data quality objectives for sulfate were as follows: The 

target analyte in the method blank would be below method 

detection limit. The reported concentration of the continuing 

calibration check standard (CCC) and the QC check would 

agree with the expected concentration plus or minus 10% 

of the known concentration, the matrix spike would agree 

with the expected concentration plus or minus 20% of the 

known concentration (i.e., Recovery of the expected value 

would be in the range of 80-120%). Laboratory duplicates 

would agree with each other with a relative percent difference of ± 10%.

Performance of the Quality Checks is presented in Tables 

4 and 5 in the Appendix. For analysis of TBA, 80 of the 81 

continuing calibration check standards met the goal. One 

of the standards was 126% of the nominal concentration. 

For analysis of TBA, 16 of 16 matrix spikes met the goal. 

For the analysis of TBA, 16 of 18 laboratory duplicates met 

the goal; the relative percent difference of one duplicate 

was ± 25.2% and the relative percent difference of another 

duplicate was ± 29.5%. In the 38 blanks for analysis of TBA, 

reported concentrations were less than the lower calibration 

limit, or less than the method detection limit, depending on 

which concentration was reported by the analyst. 

For analysis of sulfate, 92 of 93 continuing calibration check 

standards met the goal. For the analysis of sulfate, 19 of 

19 matrix spikes met the goal. For the analysis of sulfate 
17

19 of 19 laboratory duplicates met the goal. The reported 

concentration of all 19 blanks was less than the method 

detection limit. 

There were 19 sample sets for analysis of sulfate; 10 of 

the sets were analyzed within 30 days, and the maximum 

holding time for any set was 165 days. There were 20 

sample sets for analysis of TBA; 12 of the sets were analyzed within 30 days, and the maximum holding time for any 

set was 145 days. Based on the reproducibility of data in 

the container controls (as presented in Figures 10, 11, 12 

and 13) the excess holding times do not appear to have 

compromised the data quality for TBA. All the data were 

determined to be of acceptable quality, and the data were 

used in the report. 

Biodegradation of TBA in Microcosms

Of the six sites in the survey, TBA only degraded in microcosms constructed from sediment from Petaluma, California 

(Figure 10). There were three experimental treatments in 

the study: microcosms constructed with sediment as collected, microcosms constructed with sediment that had 

been autoclaved to kill organisms that could biodegrade 

TBA, and container controls that did not contain sediment. 

Live microcosms were spiked to initial concentrations of 

approximately 1,400 μg/L. Reductions in concentrations 

were apparent after 91 days of incubation. After 182 days 

of incubation, the concentration was down to the effective 

detection limit in two of the three replicate microcosms 

sampled. After 273, 314 and 456 days of incubation, the 

concentration of TBA was below the detection limit in all 

the microcosms sampled. Disregarding the lag, the rate of 

removal of TBA was 6.5 ± 4.1 per year at 90% confidence. 

This rate is in good agreement with the rates presented 

in Table 2.

In one of the treatments, the sediment was autoclaved at 

121 °C overnight in an attempt to sterilize the sediment 

and then dosed. There was not significant removal of 

TBA over 734 days of incubation in the autoclaved control 

microcosms. One of the experimental treatments was a 

control for loss from the microcosm container. This control 

consisted of the sterile glass serum bottle filled with sterile 

water and dose solution, and sealed with a sterile septum 

and crimp cap. There was no sediment in the container 

control. As expected, there was no loss of TBA from the 

container control.

The pore water of the live microcosms contained 11.7 ± 

0.2 mg/L sulfate at the start of the experiment. At the time 

when the TBA was entirely consumed, the sulfate concentration was 15.9 ± 0.1 mg/L. As was discussed earlier, oxidation of TBA by sulfate reduction requires 3.9 mg/L sulfate 

for each 1.0 mg/L of TBA. The sulfate demand to oxidize 

1.4 mg/L TBA in the live microcosms would be 5.5 mg/L. 

It is not entirely clear why sulfate concentrations appeared 

to increase over time. The method for analysis of sulfate 

is recalibrated when calibration check standards are off by 

10%. It is possible that sulfate desorbed from the anion 

exchange complex in the sediment. In any case, there were 

adequate concentrations of sulfate to meet the theoretical 

demand for sulfate reduction of the TBA. However, the 

sulfate concentrations were low. They were near the half 

saturation constant for sulfate reduction. It is equally plausible that TBA in the sediment from Petaluma was degraded 

by Iron(III) or Manganese(IV) reducing bacteria. 

Panel A of Figure 11 presents data for the sediment from 

Deer Park, New York. In contrast to the behavior of TBA in 

sediment from Petaluma, California, there was no sustained 

or significant removal of TBA in the live microcosms, in the 

control microcosms, or in the container controls. The estimated rate of MTBE biodegradation in the plume at Deer 

Park was 5.3 ± 3 per year at 95% confidence (Kolhatkar 

et al., 2001). However, the companion microcosm study 

on anaerobic biodegradation of MTBE also failed to show 

any degradation of MTBE (Wilson et al., 2005b).

The maximum concentration of TBA in the live microcosms 

was 350 μg/L. If 10% of the TBA is converted to microbial 

biomass, and if the average dry weight of a microbial cell 

is 10-12 g, then complete consumption of 350 μg/L would 

produce 3.8 x 107

 cells per liter. The microcosms were 

incubated for 740 days. If initially, there was only a single 

TBA degrading bacterium in each microcosm, and the 

growth rate of the organism was at least 3% per day, the 

organisms would have grown to consume the TBA in the 

incubation period.

The concentration of sulfate in the pore water of the Deer 

Park microcosms was less than 1.0 mg/L at the beginning 

of the incubation period and at the end of the incubation 

period. However, the theoretical demand for sulfate for TBA 

biodegradation would have only been 1.4 mg/L. It is highly 

likely, but not definitely established, that concentrations 

of sulfate limited TBA biodegradation in the microcosms 

constructed with sediment from the Deer Park site. 

In microcosms constructed from sediment from Parsippany, 

New Jersey, the concentration of TBA in the live microcosms increased approximately tenfold during 744 days 

of incubation (Panel B of Figure 11). This was probably 

due to TBA formed by anaerobic biodegradation of MTBE 

in residual gasoline in the sediment used to construct the 

microcosms (Wilson et al., 2005c). If TBA was degraded 

in the sediment, the rate did not exceed the rate of TBA 

production from biodegradation of MTBE. As expected, 

there was no loss of TBA from the killed control microcosms 

or the container controls.

Similarly, there was no evidence of removal of TBA over 

730 days of incubation in microcosms constructed with 

sediment from the Boca Raton site (Figure 12). The initial 

concentration of TBA was 1.6 mg/L. The theoretical demand for sulfate was 6.2 mg/L. The initial concentration 

of sulfate was 4.8 ± 1.9 mg/L. The final concentration of 

sulfate was 3.4, 1.3, and <0.3 mg/L respectively in the 

triplicate microcosms. It is likely that sulfate was limiting 

for TBA biodegradation in the microcosms constructed with 

sediment from the Boca Raton site. 

Finally, there was no evidence of TBA biodegradation in 

microcosms constructed with sediment from the Port Hueneme site (Panel A of Figure 13) or the Vandenberg AFB site 
18

Figure 11. Behavior of TBA in microcosms constructed with material from gasoline spill sites in Deer Park, New 

York and Parsippany, New Jersey.

Figure 10. Removal of TBA in microcosms constructed with material from a gasoline spill site in Petaluma, 

California.
19

Figure 12. Behavior of TBA in microcosms constructed with material from a gasoline spill site in Boca Raton, 

Florida.

Figure 13. Behavior of TBA in microcosms constructed with material from a gasoline spill site at Port Hueneme, 

California and a site at Vandenberg AFB, California. These are sites where natural anaerobic 

biodegradation of TBA was not expected.
20

(Panel B of Figure 13). Both of these aquifers are naturally 

anaerobic and sulfate reduction is the dominant electron 

accepting process. Unfortunately, data on concentrations of 

sulfate were not collected in the microcosm experiments. 

In summary, the microcosm experiments did not present 

compelling evidence that anaerobic TBA biodegradation 

was widespread at gasoline spill sites. At three of the sites, 

the microcosm studies failed to confirm initial expectations 

that TBA was being degraded at the site based on field 

monitoring data.

Use of Stable Isotope Ratios

In recent years an alternate approach has been developed 

to recognize natural biodegradation of MTBE in ground 

water (Hunkeler et al., 2001; Gray et al., 2002; Kolhatkar 

et al., 2002; Kuder et al., 2005; Wilson et al., 2005b; Zwank 

et al., 2005). Organic compounds contain two stable isotopes of carbon. Carbon with a weight of twelve daltons 

(

12C) is approximately one hundred times more abundant 

than carbon with a weight of thirteen daltons (13C). During 

biodegradation, MTBE molecules with 12C in the methyl 

group are degraded more rapidly than MTBE molecules with 

13C in the methyl group. Over time 13C accumulates in the 

residual MTBE molecules that have not been degraded, and 

the extent of biodegradation of MTBE can be inferred by an 

increase in the ratio of 13C to 12C in the residual MTBE.

The use of stable isotopes has two important advantages. It 

uses the compound of interest as its own tracer. Molecules 

composed with 13C and molecules with 12C share a common 

source and should have the same behavior with respect to 

sorption, hydrodynamic dispersion, or dilution in a monitoring well. A second and more important advantage is that the 

stable isotope ratios recognize and validate biodegradation 

that has actually occurred in the aquifer. Microcosm studies 

done in the laboratory with sediment samples only establish 

a potential for biodegradation in the aquifer. 

The behavior of stable isotopes is described by two parameters: δ13C and ε. The δ13C of a compound is a measure of 

the ratio of 13C atoms to 12C atoms in the molecule. It will 

be defined formally later in the text. The usual pronunciation of δ13C is “delta thirteen sea.” During the course of 

biodegradation, the compound that is still remaining will 

have more of the heavy isotope 13C, and the value of δ13C 

will become more positive. The value of ε (epsilon) relates 

the change in δ13C to the fraction of the original contaminant 

that is still remaining. As a consequence, ε is often called 

the isotopic enrichment factor.

The ratio of isotopes is determined with an isotope ratio 

mass spectrometer. The mass spectrometer does not 

measure the ratio of the stable carbon isotopes to each 

other. Rather, it measures the deviation of the ratio in 

the sample from the ratio of a standard used to calibrate 

the instrument. The substance used as the international 

standard for stable carbon isotopes has a ratio of 13C to 12C of 0.0112372. 

The conventional notation for the ratio of 13C to 12C in a 

sample (δ13C) reports the ratio in terms of its deviation from 

the ratio in the standard.

δ13

13 12 13 12

13 12 C

C C C C

C C

sample standard

standard

= ( ) − ( )

( )

















×1000

The units for δ13C are parts per thousand, often represented 

as ‰, or per mil, or per mill.

The extent of isotopic fractionation is typically determined by 

a linear regression of the δ13C in MTBE on the natural logarithm of the fraction of MTBE remaining after biodegradation. The slope of the regression line is termed the isotopic 

enrichment factor (ε). Fractionation of carbon in MTBE is 

much greater during anaerobic biodegradation compared 

to biodegradation during aerobic respiration. Somsamak 

et al., (2006) reported values for ε during biodegradation of 

MTBE under methanogenic conditions of -13.3‰ to -14.6‰, 

and values under sulfate reducing conditions of -13.4‰ 

to -14.6‰. In contrast, Hunkeler et al., (2001) reported 

enrichment factors for aerobic biodegradation that varied 

from –1.52‰ ± 0.06‰ to –1.97‰ ± 0.05‰, and Gray et al. 

(2002) determined the enrichment factors that varied from 

–1.4‰ ± 0.1‰ to –2.4‰ ± 0.3‰.

Figure 14 compares the extent of fractionation of MTBE 

that would be expected during biodegradation under aerobic and anaerobic conditions. The most negative value 

for ε was plotted because this value will predict the least 

biodegradation of MTBE with an increase in the value of 

δ13C in MTBE. The most negative value for ε is the most 

conservative value. 

Any change in the value of δ13C in MTBE caused by biodegradation must be compared to the normal variation of 

δ13C in MTBE used to make gasoline. Smallwood et al., 

(2001) reported that the normal range of δ13C for MTBE in 

gasoline is from –28.3‰ to –31.6‰; more recent surveys 

indicate that the normal range extends between –27.5‰ 

and –33‰ (O’Sullivan et al., 2003). This natural variation 

is represented as a filled arrow in Figure 14. The natural 

variation in δ13C in MTBE is of the same order as the fractionation that would be expected after 99% of the original 

amount of MTBE had been degraded under aerobic conditions. As a consequence of this weak fractionation, it has 

been difficult to document natural aerobic biodegradation of 

MTBE at field scale. However, the variation in δ13C during 

anaerobic biodegradation is much larger than the variation 

in δ13C in MTBE in gasoline, and determinations of δ13C 

in MTBE have been used to document natural anaerobic 

biodegradation of MTBE at field scale (Kolhatkar et al., 

2002; Kuder et al., 2005; Wilson et al., 2005a; Wilson et 

al., 2005b; Zwank et al., 2005).

Wilson et al., (2005a) used stable isotope ratios to evaluate 

production of TBA from natural biodegradation of MTBE at 

thirteen gasoline spill sites in Orange County, California. At 
21

these thirteen sites in Orange County, the concentrations 

of MTBE and TBA are not constant over time. Figures 15 

and 16 present data from wells at three separate sites that 

illustrate common patterns. At some wells, at some sites, 

concentrations of MTBE are greater than concentrations 

of TBA at the beginning of the monitoring record. Over 

time, the concentrations of MTBE go down by orders of 

magnitude and the concentrations of TBA increase by 

orders of magnitude (Figure 15). Occasionally there was 

an almost stoichiometric replacement of MTBE with TBA 

(Figure 16 and 17). An evaluation of the ratio of stable 

isotopes of carbon in the residual MTBE demonstrated that 

biodegradation of MTBE to TBA could explain the increase 

in concentrations of TBA (Wilson et al., 2005a).

In some wells, at four of the thirteen sites, the high concentrations of TBA persisted for a period and then the 

concentrations of TBA also declined over time (illustrated 

by Figure 17). The pattern was very similar to the pattern 

seen in the microcosm studies conducted with sediment 

from sites in Orange County, California (Figure 4, presenting data from DeVaull et al., 2003). 

Studies with enrichment cultures typically follow the same 

pattern (Bradley et al., 1997; Suflita et al., 1997; Walter, 

1997). The initial density of active organisms in the inoculum is low and the activity of the organisms has no 

perceptible effect on the concentration of their substrate. 

As the organisms grow and increase in number, they eventually reach a density where their activity has a perceptible 

effect on the concentration of substrate. As they continue 

to grow and increase in numbers they eventually exhaust 

the substrate.

Figure 14. Expected relationship between the ratio of stable isotopes of carbon in MTBE or TBA and the extent 

of biodegradation of MTBE or TBA. 

The pattern shown in Figure 4 for a laboratory microcosm 

and in Figure 17 for a monitoring well at field scale can be 

explained as a lag period of MTBE biodegradation during 

microbial acclimation to MTBE, followed by biodegradation 

of MTBE to TBA, then a lag during microbial acclimation to 

TBA, followed by biodegradation of TBA. However, there 

are other explanations for the pattern at field scale. The 

supply of MTBE may have been exhausted in the source 

area and as a result there was no continuing supply of 

MTBE to produce TBA. As a result, the concentrations 

of TBA decreased because the TBA that was previously 

produced was swept away by advective flow of ground 

water in the aquifer. It is also possible that the direction 

of ground water flow in the aquifer changed and the new 

flow path to monitoring well MW-4S at site 88UT138 did 

not contain MTBE or TBA. Because there are alternative 

explanations for the pattern illustrated in Figure 17, field 

monitoring data cannot provide unequivocal evidence that 

TBA is being biologically degraded in ground water.

It would greatly simplify the evaluation of natural biodegradation of TBA at field scale if the ratios of stable isotopes 

could be used to recognize and quantify TBA biodegradation 

in ground water. However, the authors are not aware of 

any report comparing fractionation of TBA under anaerobic conditions in either microcosm studies or enrichment 

cultures. The only data available on the fractionation of 

carbon isotopes during biodegradation of TBA are from a 

study of a field-scale plume and the data suggest that the 

fractionation coefficient is small. Day et al., (2002) and 

Day and Gulliver (2003) describe a detailed evaluation of 

natural TBA biodegradation in ground water at a chemical 
22

Figure 16. Changes in concentrations of MTBE 

and TBA in monitoring wells over time at 

three sites in Orange County, California, 

data plotted on arithmetic scale.

manufacturing plant at Pasadena, Texas. They compared 

the concentrations of TBA in the plume to the δ13C of TBA. 

Figure 18 in this report plots the data provided in Figure 

30-7 of Day and Gulliver (2003). Following the approach 

of Kolhatkar et al., (2002), a regression of δ13C of TBA on 

the natural logarithm of the concentrations of TBA in the 

plume was used to estimate the fractionation coefficient 

(ε) for biodegradation of TBA at field scale. The value of 

ε was -0.73‰.

Figure 15. Changes in concentrations of MTBE 

and TBA in monitoring wells over time at 

three sites in Orange County, California, 

data plotted on logarithmic scale.

Figure 14 compares the extent of fractionation of TBA that 

would be expected during biodegradation under anaerobic 

conditions to the extent of fractionation of MTBE that would 

be expected under aerobic and anaerobic conditions. If 

the value of the enrichment coefficient during anaerobic 

biodegradation of TBA is -0.73‰, then minimal change 

can be expected in the value of δ13C for TBA, even when 

the mass of TBA is reduced by two orders of magnitude 

through biodegradation of TBA. 
23

The prediction of the enrichment factor ε for anaerobic biodegradation of TBA in Figure 18, as projected in Figure 14, 

is consistent with field observations of δ13C for MTBE and 

δ13C for TBA at field sites in Orange County, California, 

(Kuder et al., 2005) and a large field site in South America 

(Zwank et al., 2005). Figure 19 compares the total variation 

in δ13C in TBA in ground water samples in monitoring wells 

to the variation in δ13C in MTBE in the same well. In the 

data reported by Kuder et al., (2005) for Orange County, 

the total variation of δ13C for TBA was 6.76‰ while the 

total variation of δ13C for MTBE was 89.5‰ (Panel A of 

Figure 19). In the data reported by Zwank et al., (2005) 

for South America, the total variation of δ13C for TBA was 

5.6‰ while the total variation of δ13C for MTBE was 66.7‰ 

(Panel B of Figure 19). There was substantial anaerobic 

biodegradation of MTBE in these studies, as documented 

by extensive fractionation of carbon isotopes in MTBE. 

There was little fractionation of TBA. This may have been 

a result of little biodegradation of TBA. It may also have 

been the result of the much weaker fractionation of TBA 

during anaerobic biodegradation. 

To properly interpret the variation of δ13C for TBA in ground 

water, it will be necessary to compare the measured value 

of δ13C for TBA to the natural variation of δ13C for TBA. To 

the authors’ knowledge, data on the natural variation in δ13C 

for TBA has not been published. In fact, two distributions 

on natural variation would be necessary to interpret δ13C 

for TBA in ground water, (1) the natural variation of δ13C in 

TBA in gasoline, and (2) the natural variation of δ13C in the 

tertiary butyl functional group of MTBE in gasoline. Zwank 

et al., (2005) determined the δ13C of a sample of MTBE, 

then hydrolyzed the MTBE to TBA and determined the δ13C 

of the TBA produced. The δ13C of the MTBE was -28.13 ± 

0.15‰, while the δ13C of the TBA was -25.49 ± 0.10‰. The 

difference between the δ13C of MTBE and the δ13C of the 

tertiary butyl functional group in MTBE is one half as wide 

as the entire reported range of variation of δ13C of MTBE 

in gasoline (–27.5‰ to –33‰; reported in O’Sullivan et 

al., 2003). It will not be appropriate to use the variation of 

δ13C of MTBE in gasoline as a surrogate for the variation 

of δ13C of TBA. 

It is possible that future research will show that the true 

coefficient of fractionation of TBA during anaerobic biodegradation is considerably more negative than -0.73‰. 

Until that work is published, and until the distribution of 

the natural variation of δ13C of TBA is published, it will be 

difficult or impossible to use measurements of δ13C of TBA 

in ground water to estimate the extent of biodegradation 

of TBA at field sites. 

In contrast to the limited variation in δ13C of the TBA, there 

was much wider variation of isotope ratios of hydrogen in 

TBA (expressed as δ D, the ratio of deuterium to hydrogen 

one as measured against a standard reference material for 

hydrogen isotopes). Data reported by Zwank et al., (2005) 

are presented in Figure 20. Although values of δ13C in TBA 

varied by 3.9‰ in the population of wells, values of δ D in 

TBA varied by 67.9‰. Future research may establish a 

relationship between δ D and the extent of biodegradation 

of TBA in ground water; however, there are concerns that 

isotopic exchange of the hydrogen in the alcohol functional 

group with hydrogen in water may alter δ D in TBA, making a straightforward interpretation of δ D in TBA difficult 

(Zwank et al., 2005).

In summary, there is not a good technique to determine 

biodegradation of TBA at field scale. At most field sites 

there will not be enough monitoring wells, or the wells will 

not be in the right place to document anaerobic biodegradation along a flow path in the aquifer. The application of 

stable isotope ratios to TBA in ground water (at the present 

level of development) cannot resolve TBA biodegradation 

at field scale.

Figure 17. Changes in concentrations of MTBE and TBA in monitoring wells over time at a site in Orange 

County, California, where a decline in concentration of MTBE was followed at a later time by a decline 

in concentration of TBA.
24

Figure 19. Comparison of the range of variation in the isotope ratio of carbon (d13C) in TBA at monitoring wells at 

sites in Orange County, California, (Kuder et al., 2005) and at a site in South America (Zwank et al., 

2005) to the range of variation in the isotope ratio of carbon (d13C) in MTBE in the same monitoring 

wells.

Figure 18. Relationship between the isotopic ratio of carbon in TBA in ground water at a manufacturing facility 

in Pasadena, Texas, and the concentration of TBA. The slope of the line is an indirect estimate of the 

isotopic fractionation factor (ε) for anaerobic biodegradation of TBA at the site. Adapted from Day and 

Gulliver (2003).
25

Figure 20. Comparison of the range of variation in the isotope ratio of hydrogen (dD) in TBA at monitoring wells 

at a site in South America (Zwank et al., 2005) to the range of variation in the isotope ratio of carbon 

(d13C) in MTBE in the same monitoring wells. 
26

Tertiary butyl alcohol (TBA) is widely distributed at gasoline 

spill sites, and is present at high concentrations. As an example, Shih et al., (2004) reported that TBA was detected 

in ground water at 61.1% of sites in Los Angeles County, 

California. The concentration of TBA was equivalent to the 

concentrations of methyl tertiary butyl ether (MTBE) and 

benzene. The mean and median concentration of TBA was 

30,100 and 1,880 μg/L, compared to 44,800 and 1,200 μg/L 

for MTBE and 83,800 and 1,370 μg/L for benzene.

At a major fraction of sites, the concentration of TBA in 

ground water is greater than can plausibly be expected 

from the TBA that was a constituent in the gasoline that 

was originally spilled (Wilson, 2003; Wilson et al., 2005a; 

this report Figure 2). Based on the ratio of TBA to MTBE 

at gasoline spill sites in Orange County, California, Wilson 

et al., (2005a) determined that TBA resulting from the biodegradation of MTBE could explain the concentrations of 

TBA at 85% of the sites. 

A review of available literature indicated that microorganisms can degrade TBA using oxygen, nitrate, Iron(III), 

Manganese(IV), or sulfate as a terminal electron acceptor 

(Bradley et al., 2002). The current consensus opinion is 

that an electron acceptor is necessary, and that TBA cannot 

be directly fermented to produce methane (Schmidt et al., 

2004a). Although TBA does not degrade under anaerobic 

conditions in many laboratory experimental systems, the 

average rate of anaerobic biodegradation of TBA (when it 

does occur) is faster than the average rate of anaerobic 

biodegradation of MTBE. The average first order rate of 

anaerobic biodegradation of TBA in the studies reported in 

Table 2 was 9.2 per year compared to an average rate of 

anaerobic biodegradation of MTBE of 1.0 per year (Wilson, 

2003; Table 3).

Data on the availability of electron acceptors in ambient 

ground water and the concentrations of electron acceptors 

in the source area of plumes indicates that sulfate is the 

most important electron acceptor at fuel spill sites (Wiedemeier et al., 1995). Data collected for a study published by 

Kolhatkar et al., (2000), but presented for the first time in 

this report, indicates that sulfate, and presumably oxygen 

and nitrate, are entirely depleted in the source area of approximately 80% of gasoline spill sites.

As a consequence, anaerobic biodegradation of TBA at 

many particular field sites will be limited by the availability of 

sulfate and the rate of TBA biodegradation will be limited by 

the rate that sulfate is supplied to the plume by diffusion and 

dispersion. This rate of supply can be orders of magnitude 

slower than the rate of TBA biodegradation in laboratory 

microcosm studies where sulfate is not limiting. 

There is a strong possibility that Iron(III) reducing and 

Manganese(IV) reducing bacteria degrade TBA in ground 

water that is depleted of soluble electron acceptors. However, at the current state of practice, it is impossible to evaluate the contribution of Iron(III) reducing and Manganese(IV) 

reducing bacteria to the natural biodegradation of TBA in 

ground water at gasoline spill sites. 

The OSWER Directive on MNA (U.S. Environmental Protection Agency, 1999) identifies three lines of evidence that 

can be used to support the selection of MNA as a remedy. 

The first line of evidence is Historical groundwater and/or 

soil chemistry data that demonstrate a clear and meaningful 

trend of decreasing contaminant mass and/or concentration 

over time at appropriate monitoring or sampling points. The 

second line of evidence is Hydrogeologic and geochemical 

data that can be used to demonstrate indirectly the type(s) 

of natural attenuation processes active at the site, and the 

rate at which such processes will reduce contaminant concentrations to required levels. Until techniques are applied 

that can estimate the supply of biologically available Iron(III) 

or Manganese(IV) in aquifer sediment, it is not possible to 

compare the supply of these insoluble electron acceptors 

to the demand for electron acceptors provided by TBA or 

other organic materials in contaminated ground water. It 

is not possible to provide the second line of evidence for 

natural biodegradation of TBA by iron-reducing or manganese-reducing bacteria.

Data on the fractionation of carbon isotopes in TBA during 

anaerobic biodegradation are indirect data from a field study 

instead of direct measurements from controlled laboratory 

studies and the data are available from only one site (Day 

et al., 2002; Day and Gulliver, 2003). The fractionation 

reported at the one site available in the literature is weak; 

suggesting that fractionation of carbon isotopes will not be 

generally useful to recognize anaerobic biodegradation of 

TBA at field sites. Fractionation of hydrogen isotopes in 

TBA is much stronger (Zwank et al., 2005) and may be 

useful in the future. However, there is a concern about 

isotopic exchange of the hydrogen in the –OH function of 

TBA with hydrogen in water. At this writing, there is not 

a consensus on the appropriate interpretation of shifts in 

the ratio of hydrogen isotopes of TBA. However, isotopic 

fractionation during biodegradation is an active area of 

research and advances in the state of knowledge can be 

expected in the future. 

Summary
27

If it is necessary to evaluate natural biodegradation of TBA 

at a gasoline spill site, do not rely on the data from conventional monitoring wells. It is necessary to obtain data 

on the vertical distribution of the concentrations of TBA, 

sulfate, benzene, and methane in ground water. This can 

be accomplished using the following steps. Collect ground 

water with push tools that sample a narrow vertical interval (six inches to two feet; 0.15 m to 0.61 m). Extend the 

vertical profile from the water table into clean water below 

the plume. Use the concentrations of benzene as a tracer 

for the contaminated ground water that might contain TBA. 

If the benzene has been biologically degraded and is not 

present in the ground water, use concentrations of methane 

as a tracer for the plume. The strongest and most direct 

evidence for natural biodegradation of TBA is a series of 

sampling locations or monitoring wells down gradient of 

the source area that have high concentrations of tracer 

compounds but are devoid of TBA. 

If unacceptable concentrations of TBA are still present in 

ground water from the down gradient wells, compare the 

concentrations of TBA to the supply of sulfate as an electron acceptor for biodegradation of TBA. As a loose rule of 

thumb, the stoichiometric demand for sulfate as an electron 

acceptor for complete oxidation of a fuel component is four 

times the concentration of the fuel component. There is a 

reasonable prospect for natural attenuation of TBA through 

natural anaerobic biodegradation if the concentration of 

sulfate exceeds four multiplied by the sum of the concentrations of TBA, MTBE, benzene, toluene, ethylbenzene, 

xylenes, and trimethylbenzenes. 

Many risk evaluations at gasoline spill sites use a simple 

transport and fate model such as BIOSCREEN (Newell et 

al., 1996) or the calculations in ASTM E-1739, Risk-Based 

Corrective Action (RBCA) at Petroleum Release Sites, 

issued by the American Society for Testing and Materials 

Standards (2002). Unless or until it has been shown that 

a sufficient supply of electron acceptor is available to meet 

the stoichiometric demand for TBA and the other organic 

compounds in the plume, assume that TBA will not be 

biologically degraded in the ground water plume and that 

natural attenuation of TBA will be due exclusively to hydrodynamic dispersion.

Do not assume that TBA is being biologically degraded 

under anaerobic conditions in ground water because there 

is evidence that natural anaerobic biodegradation of MTBE 

is occurring. The same organisms are not degrading MTBE 

and TBA. The initial step in anaerobic MTBE biodegradation is carried out by acetogenic bacteria that use MTBE 

to consume molecular hydrogen as part of their energy 

metabolism. The anaerobic biodegradation of TBA is most 

commonly carried out by sulfate reducing bacteria. 

A risk evaluation for TBA may indicate that it is necessary 

to remediate TBA in ground water. Long term monitoring 

data may indicate that the plume of TBA is continuing to 

expand down gradient. Farther expansion of the TBA 

plume may put a down gradient receptor at risk, or exceed 

some concentration-based goal at a down gradient point 

of compliance. Depending on the remedial goals, on the 

nature of the source area for TBA in ground water, and on 

the geochemistry of the ground water, Monitored Natural 

Attenuation may have a role in the overall strategy for risk 

management. Small and Weaver (1999) reasoned that 

plumes of MTBE and benzene expand because the MTBE 

or benzene is transferred from the source area to flowing 

ground water faster than natural biodegradation can remove 

MTBE or benzene from the flowing ground water. Plumes 

of TBA should follow the same pattern. 

The capacity for biodegradation of TBA is limited by the 

concentration of sulfate in ground water entering the source 

area and the amount of biologically available Iron(III) or 

Manganese(IV) associated with the aquifer sediments. The 

only remedial approach that can tip balance between the 

release of TBA to ground water and the degradation of TBA 

in ground water is to reduce the transfer of TBA from the 

source to flowing ground water in the aquifer. 

There are two common situations that lead to a continuing 

source of TBA contamination. In the first situation, MTBE 

in residual gasoline is slowly released to flowing ground 

water over time, and the MTBE is biologically degraded to 

TBA once it comes into solution in ground water. In the 

second situation, TBA that is present in residual gasoline, 

as an oxygenate, is slowly released to flowing ground water 

over time because the residual gasoline is held in clay or 

silt by capillary attraction. As a result the TBA must escape 

the clay or silt by diffusion through pore water before it can 

enter the major channel of ground water flow in the aquifer. 

In either situation, the first remedial response should be 

removal of the residual gasoline by excavation, or some 

effective technique for in situ remediation. At some sites, 

air sparging has proved effective to remove sources of 

MTBE in residual gasoline (Hattan et al., 2003). Because 

air sparging can effectively supply oxygen to ground water, 

air sparging should effectively remove TBA as well.

Recommendations
28

After the source has been remediated, it may be necessary to pump and treat the ground water that still contains 

high concentrations of TBA. It is not reasonable to expect 

natural biodegradation of TBA to remediate TBA in a plume 

unless the concentration of sulfate in the plume can meet 

the stoichiometric demand for complete metabolism of TBA 

and other organic compounds in the plume. If a pump 

and treat remedy is put in place, it should continue until 

the concentrations of sulfate in the plume are adequate to 

degrade TBA and all the other organic compounds in the 

ground water.

Research needs

More research and field studies are needed on the contribution of Iron(III) reducing and Manganese(IV) reducing 

bacteria to the natural anaerobic biodegradation of TBA at 

gasoline spill sites. In particular, techniques are needed 

to evaluate the supply of biologically available Iron(III) and 

Manganese(IV). 

More research is also needed on techniques to determine 

the presence and activity of naturally occurring microbes 

in ground water that can degrade TBA. Recently, new 

techniques have been developed to sample and evaluate 

the microorganisms in water in monitoring wells (Biggerstaff, 2007; Geyer et al., 2005; Sublette et al., 2006). 

These techniques are built around the use of Bio-Sep® 

beads (Microbial Insights, Rockford, TN). These beads 

are constructed from a composite of 25% aramid polymer 

(Nomex®, DuPont, Wilmington, DE) and 75% powdered 

activated carbon. The beads are from 2 to 4 mm in diameter. They have a high porosity (74%) and high specific 

surface area (600 m2 /g). The Bio-Sep® beads provide a 

surface for the microorganisms to colonize and grow. After 

a period of incubation, in ground water in a monitoring well, 

the beads are retrieved and the microorganisms that grew 

in the beads are extracted and analyzed. 

One particularly compelling approach is to amend Bio-Sep® 

beads with an organic compound that is mass labeled with 

the stable carbon isotope 13C. If the compound is biologically degraded, some portion of the mass label should find 

its way into the biomass that develops in the bead. Geyer 

et al., (2005) amended the beads with 13C labeled benzene 

or toluene by sorbing vapors of benzene or toluene to the 

powdered activated carbon. The beads were installed in a 

monitoring well at a contaminated site for 32 days and then 

recovered. The phospholipid fatty acids in the biomass were 

extracted and the concentration of 13C in the fatty acids was 

determined using compound specific isotope ratio mass 

spectrometry. Selected fatty acids were highly enriched in 

13C and the mass label could only have come from metabolism of the mass labeled benzene or toluene incorporated 

into the Bio-Sep® beads before they were deployed to the 

well. As of this writing, EPA funded research is applying 

the same approach to evaluate the biodegradation of TBA 

in contaminated ground water.

Finally, more effort is needed to monitor the lifecycle of 

TBA plumes in ground water at gasoline spill sites and 

to document the contribution of natural attenuation processes, including natural biodegradation and dilution and 

dispersion.
29

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Wiedemeier, T., J.T. Wilson, D.H. Kampbell, R.N. Miller, 

and J.E. Hansen. Technical Protocol for Implementing 

Intrinsic Remediation with Long-Term Monitoring for 

Natural Attenuation of Fuel Contamination Dissolved 

in Ground Water Volume II, Air Force Center for 

Environmental Excellence, Technology Transfer 

Division, Brooks AFB, San Antonio, Texas. 1999c. 

available http://www.afcee.brooks.af.mil/products/

techtrans/download/A324247a.pdf. and http://www.

afcee.brooks.af.mil/products/techtrans/download/

A324247b.pdf.

Wilson, J.T. “Fate and transport of MTBE and other gasoline 

components.” In MTBE Remediation Handbook, by 

E.E. Moyer and P.T. Kostecki, Amherst, Massachusetts: 

Amherst Scientific Publishers, 2003, 19-61.
32

Appendix

Table 4 part 1. Typical Quality Performance Data for Analysis of TBA in Water. All Values are μg/L Unless Otherwise 

Indicated.

Date Collected 02/01/01 04/06/01 05/31/01 07/06/01 09/05/01

Date Analyzed 03/22/01 & 

03/26/01

06/07-14/01 09/27/01,

10/18-23/01

08/28-31/01 

& 09/19/01

11/02-06/01

CCC Standard Nominal 200 200 200 200 20.0

CCC Standard Measured 194 188 217 198 19.8

Percent of Check Standard 97.0% 94.0% 109% 99.0% 99.0%

CCC Standard Nominal 200 20.0 20.0 20.0 200

CCC Standard Measured 177 18.0 22.4 20.5 171

Percent of Check Standard 89.0% 90.0% 112% 103% 85.5%

CCC Standard Nominal 20.0 200 200 200

CCC Standard Measured 19.0 177 204 182

Percent of Check Standard 95.0% 88.5% 102% 91.0%

QC Standard Nominal 20.0 20.0 20.0 20.0 20.0

QC Standard Measured 17.4 16.6 18.9 21.1 19.0

Percent of Check Standard 87.0% 83.0% 94.5% 106% 95.0%

QC Standard Nominal 200 200 200 200 200

QC Standard Measured 194 190 189 209 210

Percent of Check Standard 97.0% 95.0% 94.5% 105% 105%

Blank 1 <10* <10* <10* <10* <10*

Blank 2 <10* <10* <10* <10*

Sample Analysis 31.2 34.2 20.8 91.6

Laboratory Duplicate 33.8 33.4 19.8 118

Relative Percent Difference 8.00% 2.37% 4.93% 25.2%

Spike Concentration NP NP NP 200

Sample Concentration 112 <1.00 17.1 31.4

Spike Recovery (Percent) 83.0% 89.0% 100% 99.0%

CCC: Continuing Calibration Check; QC: Second Source Check; NP: Not Provided

* Lower Calibration Limit
33

Table 4 part 2. Typical Quality Performance Data for Analysis of TBA in Water. All Values are μg /L Unless Otherwise 

Indicated.

Date Collected 10/11/01 10/25/01 11/16/01 12/05/01 01/04/02

Date Analyzed 11/28-30/01 11/30/01-

12/04/01

11/26

- 27/01

01/04-09/02 01/18/02

CCC Standard Nozminal 20.0 200 200 200 200

CCC Standard Measured 19.7 226 241 219 213

Percent of Check Standard 98.5% 113% 121% 110% 107%

CCC Standard Nominal 20.0 20.0 20.0 20.0

CCC Standard Measured 19.5 24.3 21.5 24.0

Percent of Check Standard 97.5% 122% 108% 120%

CCC Standard Nominal 200 20.0 200

CCC Standard Measured 194 20.9 201

Percent of Check Standard 97.0% 105% 101%

QC Standard Nominal 200 200 200 20.0 20.0

QC Standard Measured 219 237 252 20.3 22.2

Percent of Check Standard 110% 119% 126% 102% 111%

QC Standard Nominal 200

QC Standard Measured 198

Percent of Check Standard 99.0%

Blank 1 <10* <10* <20* <10* <10*

Blank 2 <10* <10* <10* <10*

Sample Analysis <20.0 55.2 113 57.9

Laboratory Duplicate <20.0 53.6 110 67.3

Relative Percent Difference 0.00% 2.94% 3.05% 15.0%

Spike Concentration NP NP NP NP

Sample Concentration <20.0 42.4 71.8 53.9

Spike Recovery (Percent) 97.0% 126% 121% 104%

CCC: Continuing Calibration Check; QC: Second Source Check; NP: Not Provided

Table 4 part 3. Typical Quality Performance Data for Analysis of TBA in Water. All Values are μg /L Unless Otherwise 

Indicated.

Date Collected 1/15/02 2/12/02 3/12/02 4/16/02 4/25/02

Date Analyzed 1/26/02 3/6/02 4/8/02 5/7/02 4/25-26/02

CCC Standard Nominal 100 20.0 200 200 200

CCC Standard Measured 88.7 19.3 189 213 223

Percent of Check Standard 88.7% 96.5% 94.5% 107% 112%

CCC Standard Nominal 200 100 200 20.0 20.0
34

Date Collected 1/15/02 2/12/02 3/12/02 4/16/02 4/25/02

CCC Standard Measured 160 93.7 200 21.9 22.5

Percent of Check Standard 80.0% 93.7% 100% 110% 113%

CCC Standard Nominal 20.0 100 100 200

CCC Standard Measured 17.9 93.3 116 205

Percent of Check Standard 89.5% 93.3% 116% 103%

QC Standard Nominal 200 200 20.0 100 200

QC Standard Measured 162 208 17.5 117 239

Percent of Check Standard 81.0% 104% 87.5% 117% 120%

QC Standard Nominal 100 20.0 1000

QC Standard Measured 84.9 19.1 962

Percent of Check Standard 84.9% 95.5% 96.2%

Blank 1 <10* <2.4** <2.4** <2.4** <2.4**

Blank 2 <10* <2.4** <2.4** <2.4** <2.4**

Sample Analysis 19.8 <2.40 <2.40 88.1 118

Laboratory Duplicate 20.3 <2.40 <2.40 90.0 105

Relative Percent Difference 2.49% 0.00% 0.00% 2.13% 11.7%

Spike Concentration 92.6 98.8 94.9 114 290

Sample Concentration 4.20 <2.40 <2.40 <2.40 276

Spike Recovery (Percent) 88.0% 98.8% 94.9% 114% 109%

CCC: Continuing Calibration Check; QC: Second Source Check

Table 4 part 4. Typical Quality Performance Data for Analysis of TBA in Water. All Values are μg /L Unless Otherwise 

Indicated.

Date Collected 5/29/02 7/16/02 12/04/02 12/11/02 5/15/03

Date Analyzed 6/6/02 to 8/7/02 12/12/02 12/18/02 5/16/03

CCC Standard Nominal 200 200 100 200

CCC Standard Measured 225 240 112 200

Percent of Check Standard 113% 120% 112% 100%

CCC Standard Nominal 100 100 20.0

CCC Standard Measured 116 101 22.5

Percent of Check Standard 116% 101% 113%

CCC Standard Nominal 100 50.0

CCC Standard Measured 106 54.5

Percent of Check Standard 106% 109%

QC Standard Nominal 20.0 200 20.0 100 100

QC Standard Measured 20.8 224 20.9 92.5 98.6

Percent of Check Standard 104% 112% 105% 93.0% 98.6%
35

Date Collected 5/29/02 7/16/02 12/04/02 12/11/02 5/15/03

QC Standard Nominal 200 100 200 50.0 200

QC Standard Measured 220 102 238 49.4 200

Percent of Check Standard 110% 102% 119% 99.0% 100%

Blank 1 <2.4** <2.4** <2.4** <2.4** <2.4**

Blank 2 <2.4** <2.4** <2.4** <2.4** <2.4**

Sample Analysis 38.8 <2.40 259 39.5 270

Laboratory Duplicate 52.2 <2.40 283 43.5 270

Relative Percent Difference 29.5% 0.00% 8.90% 9.64% 0.00%

Spike Concentration 200 109 256

Sample Concentration 40.4 <2.40 251

Spike Recovery (Percent) 94.0% 109% 103%

CCC: Continuing Calibration Check; QC: Second Source Check

Table 5 part 1. Typical Quality Performance Data for Analysis of Sulfate in Water. All Values are mg/L Unless 

Otherwise Indicated.

Date Collected 01/05/01 5/31/01 5/31/01 6/28/01 07/06/01

Date Analyzed 4/15-26/01 11/8-12/01 10/3-4/01 10/4-5/01 9/26-27/01

CCC Standard Nominal 29.8 25.0 28.3 28.3 28.3

CCC Standard Measured 30.8 25.5 28.1 27.1 28.2

Percent of Check Standard 103% 102% 99.3% 95.8% 99.6%

CCC Standard Nominal 1.00 5.00 5.00 5.00 1.00

CCC Standard Measured 1.08 4.87 4.93 4.67 0.95

Percent of Check Standard 108% 97.4% 98.6% 93.4% 95.0%

CCC Standard Nominal 25.0 25.0 5.00 5.00 50.0

CCC Standard Measured 25.2 26.4 4.93 4.78 47.9

Percent of Check Standard 101% 106% 98.6% 95.6% 95.8%

QC Standard Nominal 29.8 5.00 5.00 5.00 28.3

QC Standard Measured 29.5 5.24 4.95 4.92 29.0

Percent of Check Standard 99.0% 105% 99.0% 98.4% 103%

QC Standard Nominal 5.00 25.0 28.3 5.00 50.0

QC Standard Measured 5.25 27.4 28.1 5.45 48.3

Percent of Check Standard 105% 110% 99.3% 109% 96.6%

Blank 1 <0.50 <1.00 <1.00 <1.00 <1.00

Blank 2 <0.50 <1.00

Sample Analysis 2.85 16.0 2.82 2.50 2.04

Laboratory Duplicate 2.82 15.9 2.84 2.53 2.01

Relative Percent Difference 1.06% 0.63% 0.71% 1.19% 1.48%

Spike Concentration NP NP NP NP 50.0
36

Date Collected 01/05/01 5/31/01 5/31/01 6/28/01 07/06/01

Sample Concentration 3.49 12.8 2.77 2.39 2.28

Spike Recovery (Percent) 100% 101% 95.0% 96.0% 98.6%

CCC: Continuing Calibration Check; QC: Second Source Check; NP: Not Provided

Table 5 part 2. Typical Quality Performance Data for Analysis of Sulfate in Water. All Values are mg/L Unless 

Otherwise Indicated.

Date Collected 7/26/01 9/05/01 10/11/01 10/25/01 12/05/01

Date Analyzed 10/10/01 12/12-14/01 10/25/01 11/9-12/01 1/7/02

CCC Standard Nominal 28.3 9.87 28.3 28.3 9.87

CCC Standard Measured 28.2 9.09 27.6 28.6 9.32

Percent of Check Standard 99.6% 92.1% 97.5% 101% 94.4%

CCC Standard Nominal 5.00 25.0 25.0 5.00 25.0

CCC Standard Measured 4.83 24.9 24.8 5.18 24.4

Percent of Check Standard 96.6% 99.6% 99.2% 104% 97.6%

CCC Standard Nominal 5.00 5.00 25.0 25.0 5.00

CCC Standard Measured 5.00 4.97 24.9 25.8 4.93

Percent of Check Standard 100% 99.4% 99.6% 103% 98.6%

QC Standard Nominal 5.00 25.0 5.00 5.00 25.0

QC Standard Measured 5.31 24.8 4.85 5.24 25.1

Percent of Check Standard 106% 99.2% 97.0% 105% 100%

QC Standard Nominal 5.00 5.00 25.0 25.0 5.00

QC Standard Measured 5.27 5.11 25.9 26.1 5.01

Percent of Check Standard 105% 102% 104% 104% 100%

Blank 1 <1.00 <1.00 <1.00 <1.00 <1.00

Blank 2

Sample Analysis 2.70 3.25 2.00 2.69 4.49

Laboratory Duplicate 2.75 3.34 1.99 2.61 4.49

Relative Percent Difference 1.83% 2.73% 0.50% 3.02% 0.00%

Spike Concentration NP NP NP NP NP

Sample Concentration 2.81 4.97 1.99 2.78 2.27

Spike Recovery (Percent) 98.0% 102% 95.0% 97.0% 94.0%

CCC: Continuing Calibration Check; QC: Second Source Check; NP: Not Provided

Table 5 part 3. Typical Quality Performance Data for Analysis of Sulfate in Water. All Values are mg/L Unless 

Otherwise Indicated.

Date Collected 01/04/02 1/15/02 02/07/02 2/12/02 3/12/02

Date Analyzed 02/06/02 2/4-5/02 03/01/02 3/2/02 4/11/02

CCC Standard Nominal 9.87 9.87 9.87 9.87 9.87

CCC Standard Measured 9.18 8.91 8.95 8.32 9.24
37

Date Collected 01/04/02 1/15/02 02/07/02 2/12/02 3/12/02

Percent of Check Standard 93.0% 90.3% 90.7% 84.3% 93.6%

CCC Standard Nominal 25.0 25.0 25.0 25.0

CCC Standard Measured 24.5 25.1 25.1 25.3

Percent of Check Standard 98.0% 100% 100% 101%

CCC Standard Nominal 5.00 5.00 5.00 5.00

CCC Standard Measured 4.84 4.83 5.00 5.04

Percent of Check Standard 96.8% 96.6% 100% 101%

QC Standard Nominal 25.0 25.0 5.00 25.0 25.0

QC Standard Measured 25.2 25.4 5.14 25.4 26.7

Percent of Check Standard 101% 102% 103% 102% 107%

QC Standard Nominal 5.00 5.00 25.0 5.00 5.00

QC Standard Measured 5.05 5.16 25.8 5.25 5.20

Percent of Check Standard 101% 103% 103% 105% 104%

Blank 1 <1.00 <1.00 <1.00 <1.00 <1.00

Blank 2

Sample Analysis 2.65 12.4 6.60 10.6 8.16

Laboratory Duplicate 2.59 12.3 6.64 10.7 8.15

Relative Percent Difference 2.29% 0.81% 0.60% 0.94% 1.23%

Spike Concentration 3.58 3.67 13.4 16.8 8.49

Sample Concentration 2.54 2.74 1.16 9.16 12.1

Spike Recovery (Percent) 92.0% 92.0% 103% 98.0% 98.0%

CCC: Continuing Calibration Check; QC: Second Source Check

Table 5 part 4. Typical Quality Performance Data for Analysis of Sulfate in Water. All Values are mg/L Unless 

Otherwise Indicated.

Date Collected 4/16/02 4/25/02 5/29/02 7/16/02

Date Analyzed 5/14-15/02 5/23/02 6/25/02 8/15/02

CCC Standard Nominal 9.87 18.6 18.6 18.6

CCC Standard Measured 9.04 17.0 17.6 17.9

Percent of Check Standard 91.6% 91.4% 94.6% 96.2%

CCC Standard Nominal 25.0 25.0 25.0 5.00

CCC Standard Measured 25.0 24.6 24.7 5.09

Percent of Check Standard 100% 98.4% 98.8% 102%

CCC Standard Nominal 5.00 5.00 5.00 25.0

CCC Standard Measured 4.97 4.91 4.85 25.7

Percent of Check Standard 99.4% 98.2% 97.0% 103%

QC Standard Nominal 25.0 25.0 25.0 5.00

QC Standard Measured 25.7 25.8 24.5 5.04
38

Date Collected 4/16/02 4/25/02 5/29/02 7/16/02

Percent of Check Standard 103% 103% 98.0% 101%

QC Standard Nominal 5.00 5.00 5.00 25.0

QC Standard Measured 5.05 4.91 4.91 25.5

Percent of Check Standard 101% 98.2% 98.2% 102%

Blank 1 <1.00 <1.00 <0.33 <1.00

Blank 2

Sample Analysis 10.7 0.53 9.78 13.4

Laboratory Duplicate 10.7 0.51 9.78 13.5

Relative Percent Difference 0.00% 3.85% 0.00% 0.74%

Spike Concentration 18.1 11.5 11.8 18.7

Sample Concentration 10.5 0.53 <0.331 12.1

Spike Recovery (Percent) 103% 90.0% 94.0% 101%
39

EPA/600/R-07/100

October 2007

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 Research Laboratory

Cincinnati, OH 45268

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