Office of Research and Development

National Risk Management Research Laboratory, Ada, Oklahoma 74820

Natural Attenuation of the Lead 

Scavengers 1,2-Dibromoethane 

(EDB) and 1,2-Dichloroethane 

(1,2-DCA) at Motor Fuel Release 

Sites and Implications for Risk 

Management

EPA 600/R-08/107 | September 2008 | www.epa.gov/ada

Office of Research and Development

National Risk Management Research Laboratory, Ada, Oklahoma 74820

John T. Wilson 

National Risk Management Research Laboratory

Kenneth Banks

Dynamac Corporation

Robert C. Earle

Shaw Environmental and Infrastructure, Inc.

Yongtian He

Postdoctoral Associate, National Research Council

Tomasz Kuder

University of Oklahoma

Cherri Adair

National Risk Management Research Laboratory

Natural Attenuation of the Lead 

Scavengers 1,2-Dibromoethane 

(EDB) and 1,2-Dichloroethane 

(1,2-DCA) at Motor Fuel Release 

Sites and Implications for Risk 

Management
ii

Notice

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 5857. 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 or John Wilson.

Front Cover photos:

Courtesy of: Ron Falta and John Wilson

#1 Warning on a dispenser for leaded gasoline.

#2 Examining a water sample from a monitoring well to see if the sample has been 

collected properly.

#3 Vulnerability of shallow ground water wells to contamination with EDB.
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 threaten 

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.

Tetra-ethyl lead was widely used in leaded automobile gasoline from 1923 until 1987. To prevent lead 

deposits from fouling the engine, 1,2-dibromoethane (EDB) and 1,2-dichloroethane (1,2-DCA) were 

added to the gasoline to act as lead scavengers. These compounds reacted with lead in the engine to make 

volatile compounds that were discharged in the exhaust. If leaded gasoline is spilled to ground water from 

a leaking underground storage tank, there is a potential for EDB and 1,2-DCA to partition from the spill and 

contaminate ground water. The Maximum Contaminant Levels (MCLs) for EDB and 1,2-DCA are 0.05 and 

5.0 μg/L respectively. The concentrations of EDB and 1,2-DCA that would be expected in ground water in 

contact with unweathered leaded automobile gasoline are 1,900 and 3,700 μg/L respectively. 

Lead was effectively banned in gasoline in the USA before the underground storage tank program was fully 

implemented. As a result, only a portion of the state agencies that implement the federal UST program 

routinely monitor for EDB and 1,2-DCA at gasoline spill sites. In many states, little is known of the risk 

from EDB and 1,2-DCA at old leaded gasoline spill sites. Monitored Natural Attenuation (MNA) is widely 

used by State Agencies to manage the risk from other fuel components, such as benzene, in ground water. 

The appropriate application of MNA requires a solid understanding of the behavior of the contaminants in 

ground water. 

To provide a technical basis for application of MNA, this report reviews the current knowledge of the 

transport and fate of EDB and 1,2-DCA in ground water. This report also provides information on the 

distribution of EDB and 1,2-DCA at motor fuel release sites that was collected during a survey of sites 

coordinated by the U. S. EPA Office of Underground Storage Tanks and the Association of State and 

Territorial Solid Waste Management Officials (ASTSWMO) and evaluates the associated chance of 

contaminating ground water. 

Robert W. Puls, Acting Director

Ground Water and Ecosystems Restoration Division

National Risk Management Research Laboratory

v

Contents

Notice . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ii

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

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

Tables . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ix

Acknowledgments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . x

Executive Summary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . xi

1.0 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1

1.1 Use of EDB and 1,2-DCA in Leaded Motor Fuel . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1

1.2 Regulation of Motor Fuel Storage to Protect Ground Water and Drinking Water . . . . . . . . . . . . 2

1.3 Investigations of EDB and 1,2-DCA at Motor Fuel Release Sites . . . . . . . . . . . . . . . . . . . . . . . . 3

1.3.1 Evaluation of Data from South Carolina Performed at Clemson University . . . . . . . . . 3

1.3.2 EPA/ASTSWMO Lead Scavengers Team . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3

1.4 Purpose and Scope of This Report . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5

2.0 Transport and Fate of EDB and 1,2-DCA at Motor Fuel Release Sites . . . . . . . . . . . . . . . . . . . . . . . 6

2.1 Conceptual Model of a Motor Fuel Release in the Subsurface . . . . . . . . . . . . . . . . . . . . . . . . . . 6

2.1.1 Mathematical Description of Rates of Attenuation . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7

2.1.2 Relationship between a First Order Rate Constant and a Half Life . . . . . . . . . . . . . . . . 7

2.1.3 A Definition of a “Generally Useful” Rate of Attenuation . . . . . . . . . . . . . . . . . . . . . . . 8

2.2 Attenuation caused by physical processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8

2.2.1 Physical Weathering from Fuel Present at Residual Saturation . . . . . . . . . . . . . . . . . . . 8

2.2.2 Sorption on Native Organic Matter . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10

2.2.3 Rate Constants for Physical Weathering of EDB and 1,2‐DCA . . . . . . . . . . . . . . . . . . . 11

2.3 Attenuation Caused by Abiotic Transformation or Biodegradation . . . . . . . . . . . . . . . . . . . . . . . 12

2.3.1 Neutral Hydrolysis of EDB and 1,2‐DCA . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12

2.3.2 Abiotic Transformation of EDB and 1,2‐DCA by Iron(II) Sulfide . . . . . . . . . . . . . . . . . 12

2.3.3 Biodegradation of EDB and 1,2‐DCA . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15

2.3.3.1 Rate Constants for Biodegradation of EDB and 1,2-DCA . . . . . . . . . . . . . . . . . . . 16

2.3.3.2 Association of Geochemical Parameters with Removal of EDB and 1,2-DCA . . . 17

2.3.3.2.1 Dissolved Oxygen . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18

2.3.3.2.2 Nitrate........................................................ 18

2.3.3.2.3 Sulfate and Sulfide . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19

2.3.3.2.4 Methane . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20

2.3.3.2.5 pH . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 21

2.4 Applications of Compound Specific Isotope Analysis (CSIA) to Document Biodegradation 

and/or Abiotic Transformation of EDB and 1,2-DCA . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22

2.4.1 Theoretical Background for Using CSIA to Estimate the Extent of Biodegradation . . . 22

2.4.2 Isotopic Enrichment during Biodegradation of EDB and 1,2‐DCA . . . . . . . . . . . . . . . . 23

2.4.3 Isotopic Enrichment During Abiotic Transformation of EDB and 1,2‐DCA . . . . . . . . . 25

2.4.4 The Initial Value of δ13C of EDB or 1,2‐DCA Originally Released in Leaded 

Motor Fuel . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26

2.4.5 Measured Concentrations of EDB and δ13C of EDB in Ground Water . . . . . . . . . . . . . 26

2.4.6 Predictions of the Extent of Degradation or Transformation . . . . . . . . . . . . . . . . . . . . . 29

3.0 Distribution of EDB and 1,2-DCA at Motor Fuel Release Sites, and the Associated Chance of 

Contaminating Ground Water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 30
vi

3.1 Existing Distribution of EDB, 1,2-DCA, and Benzene in Ground Water at Selected Motor 

Fuel Release Sites . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 30

3.2 Relative Distribution of EDB, 1,2-DCA, and Benzene at Selected Motor Fuel Release Sites . . 31

3.3 The Chance that Conventional Monitoring Using EPA Method 8260 will fail to detect EDB . . 34

3.4 Co-Distribution of EDB with 1,2-DCA, Benzene, Xylenes, and Ethylbenzene at Selected 

Motor Fuel Release Sites . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 35

3.5 Local Vulnerability of Exposure to EDB based on Past Usage of Leaded Motor Fuel . . . . . . . . 36

4.0 References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 44

Appendix A. Partitioning of EDB and 1,2‐DCA Between Gasoline and Water . . . . . . . . . . . . . . . . 48

Appendix B. Materials and Methods for Laboratory Studies of Abiotic Degradationof EDB and 

1,2‐DCA . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 50

Appendix C. Method for Compound Specific Isotope Analysis to Determine the Ratio of Stable 

Carbon Isotopes in EDB and 1,2‐DCA . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 53

Appendix D. Analytical Methods and Quality Assurance . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 55

Appendix D.1 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 55

Appendix D.2 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 55

Appendix D.3 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 56

Appendix D.4 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 56

Appendix D.5 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 56

Appendix D.6 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 57

Appendix D.7 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 57

Appendix D.8 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 58
vii

Figure 1.1. Consumption of leaded motor fuel, EDB, and 1,2-DCA in the United States from 

1949 through 1989. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2

Figure 1.2. Distribution of EDB in ground water from monitoring wells at gasoline release sites 

in South Carolina, and in sites in the EPA/ASTSWMO Study. . . . . . . . . . . . . . . . . . . . . . . . 3

Figure 2.1. Relationship between a first order rate constant in units of per year and half-lives in 

units of days, weeks, and years. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8

Figure 2.2. Distribution of seepage velocity in MTBE plumes in South Carolina (2002). . . . . . . . . . . . 9

Figure 2.3. Thought experiment on the effect of the rate of ground water flow on the concentration 

of EDB and DCA in ground water in contact with residual leaded motor fuel. . . . . . . . . . . 9

Figure 2.4. Association of concentrations of EDB in the most contaminated wells at each of 

ten sites with estimates of the seepage velocity of ground water at the sites. . . . . . . . . . . . . 10

Figure 2.5. Rate of decline of concentrations of EDB over time in 55 monitoring wells in 

South Carolina (Bulsara, 2004). . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12

Figure 2.6. Effect of concentrations of sulfide at pH 7 on the rate of abiotic transformation of 

EDB and DCA. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13

Figure 2.7. Removal of EDB or DCA in the presence of biogenic FeS in batch experiments at 

pH near 7. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14

Figure 2.8. Expected rates of abiotic transformation of EDB or 1,2-DCA by FeS in aquifer 

sediment, assuming a water-filled porosity of 25%, and pH near 7. . . . . . . . . . . . . . . . . . . . 15

Figure 2.9. Co-occurrence of dissolved oxygen and reduced iron in water samples produced from 

monitoring wells at gasoline spill sites. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18

Figure 2.10. Association of concentrations of nitrate with concentrations of EDB in the most 

contaminated monitoring wells at sites. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18

Figure 2.11. Association of concentrations of sulfate with concentrations of EDB in the most 

contaminated monitoring wells at sites. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19

Figure 2.12. Panel A. Distribution of sulfate in the source area of the plume, and in background 

ground water, at gasoline spill sites in the Eastern United States. Panel B. Distribution 

of sulfide in the source area of the plumes, compared to the concentration of sulfide 

expected by the amount of sulfate remove. Data are from the survey of Kolhatkar et al., 

(2000).................................................................... 20

Figure 2.13. Association of concentrations of methane with concentrations of EDB in the most 

contaminated monitoring wells at sites. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 21

Figure 2.14. Association of pH with concentrations of EDB in the most contaminated monitoring 

wells at sites. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 21

Figure 2.15. Enrichment of the heavy isotope of carbon in EDB during anaerobic biodegradation 

of EDB in a microcosm study. (Redrawn from Henderson et al. (2008).) . . . . . . . . . . . . . . 23

Figure 2.16. Biological or Abiotic Transformations of EDB and 1,2-DCA. . . . . . . . . . . . . . . . . . . . . . . 25

Figure 2.17. A comparison of the enrichment of the heavy carbon isotope in EDB during anaerobic 

biodegradation of EDB in a microcosm study (Henderson et al., 2008) against enrichment 

during abiotic transformation of EDB by FeS. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 25

Figures
viii

Figure 2.18. Relationship between the concentrations of EDB in the most contaminated wells at 

each of five sites and the δ13C of EDB. (Note that the concentration on x-axis decreases 

from left to right). . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 28

Figure 2.19. A conservative estimate of the relationship between the measured value of δ13C of 

EDB in ground water at a Leaded Motor Fuel Release Site, and the extent of destruction of 

EDB by abiotic and biological processes. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 29

Figure 3.1. Distribution of states participating in the EPA/ASTSWMO Study. . . . . . . . . . . . . . . . . . . . 30

Figure 3.2. Distribution of the maximum concentrations of EDB, 1,2-DCA, and benzene in ground 

water from monitoring wells at motor fuel release sites in the EPA/ASTSWMO study, 

compared to the concentrations predicted for ground water in contact with unweathered 

leaded gasoline. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 32

Figure 3.3. Distribution of the relative chance that EDB, 1,2-DCA, and benzene in ground water at 

motor fuel release sites in the EPA/ASTSWMO Study will impact a water supply well at 

concentrations above the MCL, compared to the relative chance predicted for ground 

water in contact with unweathered leaded gasoline. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 33

Figure 3.4. Comparison of the relative chance that EDB will exceed the MCL in a water supply 

well to the relative chance that benzene will exceed the MCL in a water supply well for 

ground water from motor fuel release sites in the EPA/ASTSWMO Study. . . . . . . . . . . . . 33

Figure 3.5. Distribution of concentrations of EDB at the leaded gasoline spill sites in the 

EPA/ASTSWMO Study where the concentrations were above the MCL. . . . . . . . . . . . . . . 34

Figure 3.6. Association of concentrations of EDB with concentrations of ethylbenzene (Panel A) 

or total xylenes (Panel B). . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 35

Figure 3.7. Association of concentrations of EDB with concentrations of 1,2-DCA (Panel A) and 

benzene (Panel B). . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 36

Figure 3.8. Consumption of EDB in leaded motor fuel in five small states in the USA in the 

forty-five years before lead was banned in automotive motor fuel. . . . . . . . . . . . . . . . . . . . 37

Figure 3.9. Consumption of EDB in leaded gasoline in five large states in the USA in the forty-five 

years before lead was banned in motor gasoline. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 37

Figure 3.10. Relationship between heavy use of leaded motor fuel and the use of shallow ground 

water for public water supplies in the contiguous USA. . . . . . . . . . . . . . . . . . . . . . . . . . . . . 40

Figure 3.11. Relationship between heavy use of leaded motor fuel and the use of shallow ground 

water for public water supplies in New England, the Mid-Atlantic States, and the Great 

Lakes region of the USA. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 40

Figure 3.12. Relationship between the population of a state in the USA (2000 census) and the sales 

of gasoline (1995 data). . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 43

Figure A.1. Predicted maximum groundwater concentrations of EDB and DCA for a range of 

possible concentrations of un-weathered residual gasoline in aquifer sediment. . . . . . . . . . 48

Figure A.2. Expected relationship between the fraction of EDB or DCA that is dissolved in ground 

water, and can be flushed away by ground water flow, and the concentration of residual

leaded motor fuel. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 49

 Figure B.1. Consumption of sulfate during incubation of microcosms. . . . . . . . . . . . . . . . . . . . . . . . . . 52
ix

Tables

Table 1.1. Occurrence of EDB in ground water provided by the state agencies that participated in the 

EPA/ASTSWMO State Study. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4

Table 2.1. Retardation in ground water due to sorption that is expected from the organic matter 

content of the aquifer solids. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11

Table 2.2. Comparison of the rate of transformation of EDB and DCA by iron(II) sulfide 

to the rate or removal of TCE. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14

Table 2.3. Comparison of first order rate constants for biodegradation of EDB and 1,2-DCA in 

anaerobic aquifer sediment to rate constants for overall removal with ground water flow 

in anaerobic aquifers. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17

Table 2.4. Comparison of enrichment factors for EDB and DCA to the range of isotopic 

enrichment factors (ε) for carbon isotopes during reductive dehalogenation of halogenated 

organic compounds.......................................................... 24

Table 2.5. Range of δ13C in samples of commercial chlorinated solvents. . . . . . . . . . . . . . . . . . . . . . . 27

Table 2.6. Relationship between the concentrations of EDB in the most contaminated wells at five 

motor fuel release sites, enrichment of stable carbon isotopes in EDB, and a conservative 

estimate of the fraction of EDB destroyed. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 28

Table 3.1. The distribution of EDB in the sites included in the EPA/ASTSWMO State Study 

compared to distribution in South . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 31

Table 3.2. Occurrence of EDB, benzene, and DCA in sites sampled during the EPA/ASTSWMO 

State Study. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 31

Table 3.3. Exposure to Drinking Ground Water Contaminated with EDB from Leaded Motor Fuel. . . 38

Table 3.4. Vulnerability to Drinking Ground Water Contaminated with EDB from Leaded 

Motor Fuel. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 41

Table B.1: Distribution of iron, AVS, and CRS along the Column with Mulch and Hematite as 

described in Shen and Wilson (2007). . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 50

Table B.2: Distribution of pore water, AVS, and CRS in microcosms. . . . . . . . . . . . . . . . . . . . . . . . . . . 51

Table C.1. Reproducibility of δ13C values for EDB and 1,2-DCA prepared by a purge and trap 

sampler from ground water containing aqueous solutions of EDB and 1,2-DCA and 

gasoline. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 54
x

Acknowledgment

Peer reviews of this document were provided by Hal White (U.S. EPA Office of Underground Storage 

Tanks), Ronald W. Falta (Department of Geological Sciences, Clemson University), Read S. Miner (South 

Carolina Dept. of Health & Environmental Control), Harley Hopkins (American Petroleum Institute), George 

DeVaull (Shell Global Solutions), Ileana Rhodes (Shell Global Solutions) and an anonymous reviewer. 

Significant technical support was provided by Mark Blankenship, Chuck Hoover, John Cox, Tracy Pardue, 

Lisa Hudson, Kelly Bates, Vanessa Scroggins, and Lynda Callaway (Shaw Environmental), Daniel Pope 

(Dynamac Corporation), and Kevin Smith (Student Contractor). Jim Weaver (U.S. EPA/ORD/NERL) 

provided samples of leaded motor fuel for analysis.
xi

Executive Summary

The lead scavengers Ethylene Dibromide (EDB) and 1,2-Dichloroethane (1,2-DCA) were added to leaded 

motor gasoline to prevent the buildup of deposits of lead oxide inside internal combustion engines. Recent 

studies demonstrate that lead scavengers may persist for long periods of time in certain ground water 

environments. Although lead and lead scavengers were phased out in conventional motor gasoline by the 

end of the 1980s, the lead scavengers from old releases may continue to contaminate ground water at many 

gasoline service station sites. In addition, aviation gasoline (Avgas) contains lead scavengers, and gasoline 

containing lead scavengers is still used for certain off-road applications such as automobile racing. There is 

a significant possibility that lead scavengers from releases of leaded gasoline pose an ongoing risk to ground 

water quality. 

Domestic ground water wells and certain small public water supply wells that are in close proximity to sites 

where leaded gasoline may have been released should be of particular concern. These wells often produce 

ground water from shallow aquifers, which makes them more vulnerable to contamination than larger 

municipal water supply wells which usually produce water from deeper aquifers.

EPA has formed a team with the Association of State and Territorial Waste Management Officials to 

determine the scope and magnitude of the occurrence of lead scavengers at leaking UST sites. The team 

developed a three-phased approach to this problem: (1) developing an understanding of the magnitude 

of the potential problem by compiling existing background information, (2) assessing gaps in current 

knowledge, based on the findings of Phase 1, and implementing appropriate measures to fill the gaps, and 

(3) determining an appropriate response based on evaluation of the results of Phases 1 and 2. 

Phase 1 culminated in development of a document entitled Lead Scavengers Compendium: Overview of 

Properties, Occurrence, and Remedial Technologies (U.S. EPA, 2006). Phase 2 consisted of collecting 

and analyzing ground water samples from 102 old gasoline release sites spread across the 19 states 

that chose to participate in the investigation. This report Natural Attenuation of the Lead Scavengers 

1,2-Dibromoethane (EDB) and 1,2-Dichloroethane (1,2-DCA) at Motor Fuel Release Sites and Implications 

for Risk Management represents the culmination of Phase 2. It fills some of the data gaps on the expected 

distribution of lead scavengers at gasoline release sites, it discusses mechanisms for abiotic transformation 

and biodegradation of EDB and 1,2-DCA, and it provides new tools to recognize and use natural 

transformation and degradation of EDB and 1,2-DCA as part of a risk management strategy.

The survey found that significant concentrations of EDB continue to persist at many old leaded gasoline 

spill sites. Both EDB and 1,2-DCA were present at concentrations above their respective Maximum 

Concentration Level (MCL) at a significant number of sites; EDB was detected above its MCL of 0.05 μg/L 

at 42% of the sites sampled, and 1,2-DCA was detected above its MCL of 5.0 μg/L at 15% of the sites 

sampled. Benzene (with an MCL of 5.0 μg/L) was present at 100% of the sites sampled and was the primary 

risk driver at 75% of the sites where both benzene and EDB were present in ground water; EDB was the 

primary risk driver in the remaining 25% of sites. 

The persistence of EDB at UST spill sites is consistent with its expected behavior in ground water. Simple 

physical weathering of EDB and 1,2-DCA from residual gasoline is a slow process that may require decades 

to centuries to reduce high concentrations of EDB or 1,2-DCA to their MCLs. At some sites, anaerobic 

biodegradation can provide substantial reductions in the concentrations of EDB and 1,2-DCA. At some sites, 

abiotic degradation caused by reaction with Iron(II) sulfide minerals in aquifer material can also produce 

substantial reduction in the concentration of EDB, particularly in ground water at neutral pH.

Although it is theoretically possible that anaerobic biodegradation or abiotic degradation will remove EDB 

at a particular site, it is frequently difficult to prove that degradation is occurring based on conventional 

monitoring data. Compound Specific Isotope Analysis (CSIA) can be useful to recognize biodegradation and 
xii

abiotic transformation of EDB ground water. Degradation is recognized and documented by a change in the 

ratio of stable isotopes of carbon in the molecules of EDB that remain in the ground water after degradation. 

The change in the ratios can put a conservative boundary on the extent of degradation that has occurred in 

the ground water sampled by a particular well. This makes CSIA a useful tool to prove that degradation has 

happened at field scale at a particular site. 

If the concentrations of EDB and 1,2-DCA in ground water in the source area of plumes do not attenuate, 

the hazard associated with these contaminants will persist indefinitely. MNA is most cost effective as 

a remedy when the concentrations of contaminants attenuate to their MCLs in a reasonable period of 

time. The concentrations of EDB and 1,2-DCA that would be expected in ground water in contact with 

unweathered leaded automobile gasoline are 1,900 and 3,700 μg/L respectively. To bring these initial 

concentrations to their MCL within 20 years, the first order rate of attenuation in concentration in the 

most contaminated well at a site should be 0.5 per year or greater for EDB and 0.33 per year or greater 

for 1,2-DCA. At certain sites, and under some circumstances, rates in excess of 0.5 per year for EDB or 

0.33 per year for 1,2-DCA can be attained through anaerobic biodegradation or by abiotic reactions. To 

apply MNA at a specific site, rate constants for attenuation over time should be extracted from site-specific 

data and should be verified and validated by continued long-term monitoring.

Monitoring for concentrations of EDB in ground water can be a major cost of risk management at gasoline 

spill sites. The MCL for EDB is one hundred fold lower than the MCLs for Benzene or 1,2-DCA. Because 

the MCL for EDB is so low, not all analytical methods can detect EDB when it is present at its MCL. The 

EPA Method that is most commonly used to analyze for gasoline constituents in ground water (Method 

8260B) has a detection limit for EDB of approximately 3.0 μg/L, which is sixty fold higher than the MCL. 

As a result, Method 8260B cannot be used to document that ground water is free of contamination from 

EDB. In contrast, EPA Method 8011 has a method detection limit for EDB of approximately 0.01 μg/L, 

which is sufficiently sensitive to measure EDB at its MCL. 

Method 8260B would have only discovered 40% of the survey sites with concentrations of EDB above its 

MCL. At sites where benzene is the primary risk driver, Method 8260B would be appropriate to monitor 

the quality of ground water during active remediation. However, to determine if the site has reached the 

MCL for EDB, it is necessary to use Method 8011 or its equivalent.

Keywords: EDB, 1,2-dibromoethane, DCA, 1,2-DCA, 1,2-dichloroethane, ground water, UST, 

underground storage tank, MNA, Monitored Natural Attenuation
1

1.0

Introduction

This section reviews the use of the lead scavengers 

1,2-dibromoethane (also called ethylene bromide or 

EDB) and 1,2-dichloroethane (1,2-DCA) in leaded 

motor fuel , and briefly describes the regulatory 

framework developed to protect ground water 

resources from releases of leaded motor fuel1 stored 

in underground storage tanks. Also, this section 

describes two investigations of EDB and 1,2-DCA 

at motor fuel release sites. Finally, this section 

describes the scope and intended purpose of this 

report.

1.1 Use of EDB and 1,2-DCA in Leaded 

Motor Fuel

Internal combustion engines burn a mixture of 

fuel and air to create mechanical energy that turns 

a crankshaft. The most common automotive 

engine operates on a four-stroke cycle: intake, 

compression, combustion, and exhaust. During the 

compression cycle, a mixture of air and fuel vapor 

is compressed by a piston moving upward in its 

cylinder. Ideally, at the height of the compression 

cycle, the mixture is ignited by a spark from the 

spark plug, thus initiating the “combustion” stroke, 

whereby the piston is pushed downward in the 

cylinder producing the mechanical energy that 

turns a crankshaft. During the next upstroke of the 

piston, exhaust gases are expelled from the cylinder. 

Sometime during the combustion stroke, pockets 

of unburned fuel outside the advancing flame front 

within the cylinder are heated and pressurized 

leading to sudden ignition (“detonation”) resulting 

in engine “knock”. Engine knock is damaging to 

the mechanical parts of the engine and it wastes 

fuel. 

To reduce the tendency to knock, various additives 

have been used to increase the octane of the motor 

fuel. These additives have included tetra-ethyl lead 

(TEL) since the 1920s, and since the 1960s tetramethyl lead (TML), tri-methyl-ethyl lead (TMEL), 

di-methyl-di-ethyl lead (DMDEL), and methyltri-ethyl lead (MTEL). The additives to increase 

1 “Leaded motor fuel” is a more inclusive term that 

includes leaded gasoline for automobiles plus aviation 

gasoline, which still contains lead and a lead scavenger 

package, and some grades of racing fuel. Where this 

report refers more specifically to “gasoline” it is because 

the data and information pertain to leaded gasoline for 

automotive purposes.

octane also included methyl tertiary-butyl ether 

(MTBE) and ethanol. 

Tetra-ethyl lead was widely used in motor gasoline 

from 1923 to 1987 (Falta, 2004). Lead oxide 

deposits produced during the combustion of leaded 

motor fuel can accumulate and damage the engine. 

To make the lead volatile and thus reduce the 

accumulation of lead deposits, the lead scavengers 

EDB and 1,2-DCA were added to gasoline along 

with the TEL. With these additives, the lead 

forms lead dihalides which are volatile and can be 

expelled from the engine. 

Starting in 1975, automobiles in the U.S. were fitted 

with catalytic converters to treat the exhaust gas 

and allow the vehicles to meet U.S. EPA standards 

for emissions to control air pollution. Because 

lead in motor fuel can poison the catalyst and ruin 

the catalytic converter, in 1973 EPA (a) required 

that one grade of unleaded gasoline be available 

to protect catalytic converters that were to appear 

on new cars in 1975, and (b) re-proposed annual 

reductions in lead content of all other grades of 

gasoline to protect public health2. Figure 1.1 

presents estimates of gasoline consumption in 

the U.S. that were collected and collated by 

Falta (2004). It also estimates the consumption 

of EDB and 1,2‐DCA in gasoline, based on the 

estimates of Falta (2004) for lead consumed in 

gasoline, and his observation “Since the early 

1940s, leaded automotive gasoline has contained 

EDB and 1,2‐DCA in proportion to the amount 

of tetraalkyllead with a molar ratio of Pb:Cl:Br of 

1:2:1...” 

The proportion of EDB and 1,2-DCA consumed 

each year to the total gasoline consumed each 

year changed little from 1949 to 1972. The peak 

years for use of EDB and 1,2‐DCA were 1969 

through 1972. After 1972, the total amount of EDB 

and 1,2‐DCA consumed in automobile gasoline 

declined as the content of lead declined in gasoline. 

After 1988, much less EDB and 1,2‐DCA were 

added to conventional automobile gasoline in the 

2 Even though leaded gasoline has not been used 

for on-road automobiles for nearly two decades, leaded 

gasoline (which also contains lead scavengers) is still 

in use in aviation gasoline (avgas) and in some off-road 

applications such as racing fuel.
2

United States because leaded gasoline had largely 

been phased out. 

Figure 1.1. Consumption of leaded motor fuel, 

EDB, and 1,2‐DCA in the United 

States from 1949 through 1989.

Although lead and lead scavengers were phased 

out in conventional motor gasoline by the end of 

the 1980s, the lead scavengers from old releases 

may continue to contaminate ground water at many 

gasoline service station sites. In addition, aviation 

gasoline (Avgas) contains lead scavengers, and 

gasoline containing lead scavengers is still used for 

certain off-road applications such as automobile 

racing (Aronson and Howard, 2008).

A portion of the EDB produced in the US was used 

as pesticide and fumigant (Aronson and Howard, 

2008), and ground water contamination continues 

in certain agricultural areas from the past use of 

EDB as an agricultural chemical. EDB was used 

on citrus crops, on vegetables, on grain crops, and 

to protect golf courses (U.S. EPA, 2006). In 1977 

approximately 136 million kilograms of EDB was 

produced in the USA; 127 million kilograms was 

used in fuel, approximately 8 million kilograms 

was used as a soil fumigant, and approximately 

0.9 million kilograms was used to fumigate stored 

grain and grain milling machinery, and quarantined 

citrus and other tropical fruits (U.S. EPA, 2006). 

1.2 Regulation of Motor Fuel Storage to 

Protect Ground Water and Drinking 

Water

In 1974, Congress passed the Safe Drinking Water 

Act, which required U.S. EPA to determine safe 

levels of hazardous chemicals in drinking water. 

These safe levels are called Maximum Contaminant 

Level Goals or MCLGs. Because of the difficulty 

in achieving MCLGs, MCLs (Maximum 

Contaminant Levels) have been established for 

most contaminants; MCLs are a compromise based 

on best available treatment technology, limitations 

of analytical methods, and cost. In 1989, U.S. EPA 

promulgated MCLs for benzene and for 1,2‐DCA 

of 5 μg/L. In 1992, U.S. EPA promulgated an MCL 

of 0.05 μg/L for EDB.

In 1984, Congress added Subtitle I to the Resource 

Conservation and Recovery Act (RCRA), which 

required U.S. EPA to develop a regulatory program 

for underground storage tank systems (USTs) 

that contained petroleum or certain hazardous 

substances (collectively referred to as “regulated 

substances”). The federal underground storage 

tank program is administered by the Office of 

Underground Storage Tanks (OUST), within the 

Office of Solid Waste and Emergency Response 

(OSWER). Subtitle I of the Resource Conservation 

and Recovery Act (RCRA) allows state UST 

programs approved by EPA to operate in lieu of 

the federal program. The U.S. EPA has granted 

State Program Approval to most of the states and 

the others implement their own program under 

cooperative agreements with EPA. 

Most USTs are used for the storage of motor 

fuel (gasoline and diesel fuel) and the regulated 

substance that escaped from most leaking 

USTs was gasoline. The primary petroleumderived contaminants of concern in gasoline are 

the aromatic hydrocarbons benzene, toluene, 

ethylbenzene, and xylenes (collectively referred 

to as “BTEX”). Most state programs treat the 

individual MCLs for the BTEX compounds as the 

concentration below which the compounds are 

not a concern at gasoline release sites. Even today 

ground water monitoring at gasoline release sites is 

focused on BTEX. 

Most state agencies have not routinely monitored 

for EDB or 1,2‐DCA in ground water. This may 

have been due to the fact that lead in gasoline, and 

therefore EDB and 1,2‐DCA, was being phased 

out, or was altogether banned, at the time the state 

agencies put their monitoring programs in place. 

The South Carolina Department of Health & 

Environmental Control (SDHEC) was an exception. 

Beginning in 2001, SDHEC started collecting data 

on the concentrations of EDB in monitoring wells 

at gasoline service stations that were in existence 

at a time when leaded gasoline was still available 

in the USA. EPA Methods 8260 or 8021, which 

are conventionally used for analysis of BTEX 

compounds and fuel oxygenates such as MTBE, 

do not have adequate sensitivity to determine 

concentrations of EDB at its MCL (0.05 μg/L). 
3

The SDHEC required that analyses for EDB be 

performed by EPA Method 8011, which has a 

method detection limit that is near 0.01 μg/L. 

1.3 Investigations of EDB and 1,2-DCA 

at Motor Fuel Release Sites

1.3.1 Evaluation of Data from South 

Carolina Performed at Clemson 

University

Professor R.W. Falta and his students at Clemson 

University evaluated the monitoring data on the 

distribution of EDB in ground water in South 

Carolina, and found that many gasoline release 

sites had concentrations of EDB that far exceeded 

the MCL (Falta, 2004; Falta et al., 2005). They 

analyzed the data available as of December 2004, 

and found that 537 underground storage tank sites 

had ground water with concentrations in excess of 

the MCL (Falta et al., 2005). Figure 2 of Falta et 

al., (2005) presented a frequency distribution of the 

maximum concentration of EDB in any well at each 

individual site. Figure 1.2 plots the data from Falta 

et al., (2005). The concentrations of EDB followed 

a log-linear distribution with approximately half of 

the sites having concentrations of EDB that exceed 

the MCL. The median concentration of EDB in 

sites where EDB was detected was 4.3 μg/L and 

the maximum concentration was 6,550 μg/L. 

These results were unexpected and surprised many 

ground water scientists and engineers. Staff in 

EPA’s Region 4 office in Atlanta, GA, brought 

Dr. Falta’s findings to the attention of the Office 

of Underground Storage Tanks (OUST) at EPA’s 

headquarters in Washington, DC.

Figure 1.2. Distribution of EDB in ground water 

from monitoring wells at gasoline 

release sites in South Carolina, and in 

sites in the EPA/ASTSWMO Study.

1.3.2 EPA/ASTSWMO Lead Scavengers 

Team

The U.S. EPA’s OUST and Office of Research 

and Development’s National Risk Management 

Research Laboratory (NRMRL) in association 

with the Association of State and Territorial 

Solid Waste Management Officials (ATSWMO) 

formed a team to determine what problems, if any, 

these lead scavengers pose to public health and 

the environment. The team’s mission statement 

outlines an investigation comprised of the 

following three phases: 

1. Develop an understanding of the potential 

problem as it exists today by:

a. Compiling existing background information: 

toxicological data; historical usage 

information; and occurrence in drinking 

water supplies;

b. Evaluating selected state databases and 

case files for information on sampling, 

monitoring, and remediation at LUST sites; 

c. Conducting a study on the effectiveness 

and cost of treatment and remediation 

technology; and

d. Assess whether or not there are any gaps in 

our current knowledge, based on the results 

of Phase 1. If so, develop and implement 

appropriate measures to fill the gaps.

2. Identify next steps by evaluating the results of 

Phases 1 and 2. 

Phase 1 culminated in production of a compendium 

of information entitled Lead Scavengers 

Compendium: Overview of Properties, Occurrence, 

and Remedial Technologies (U.S. EPA, 2006). The 

compendium represents EPA’s state of knowledge 

on lead scavengers (through 2005) relating to 

historical usage, physical and chemical properties, 

analytical methods, environmental fate and 

transport, toxicology, occurrence in drinking water 

supplies, presence at leaking UST sites, and the 

effectiveness and cost of treatment technologies. 

In compiling information for the compendium, 

some gaps in knowledge were identified, including 

the lack of information on the occurrence of 

lead scavengers in domestic (private) wells, 

the effectiveness of remediation and treatment 

technologies, and the magnitude of the occurrence 

of lead scavengers in ground water at leaking UST 

sites. Filling in this last data gap became the focus 

of Phase 2 of the investigation.

To develop information on the distribution of EDB 

and 1,2‐DCA in ground water at leaking UST 

sites in states that did not routinely monitor for 
4

these contaminants, EPA offered to provide free 

analysis of samples collected by the states (or their 

contractors) from sites that met certain criteria: 

• sites that were used for storage and/or 

dispensing of leaded gasoline whether or 

not they were currently in use (i.e., sites 

where USTs were located in 1989 and 

earlier) or 

• sites where leaded aviation gasoline 

(AvGas) or leaded racing fuel was used or 

is still being used (i.e., airports, automobile 

race tracks) and

• sites that had existing monitoring wells 

on-site and were regularly scheduled for 

monitoring (this was done to minimize 

the burden on states and their contractors; 

however, no sites offered as candidates for 

sampling were turned down).

Sites meeting the criteria above that were also 

within close proximity to a private well or small 

community well were of particular interest because 

such wells are more vulnerable to contamination 

than larger municipal water supply wells.

States that volunteered to participate in the 

investigation were provided with a sampling kit for 

each candidate site. Typically the kit was shipped 

to the contractor that routinely sampled the wells at 

the site. Sample vials were filled by the contractor 

at the time the wells were sampled for normal 

compliance monitoring, and the samples were 

returned to NRMRL at the R.S. Kerr Environmental 

Research Center in Ada, Oklahoma, for analysis of 

EDB by Method 8011, and benzene and 1,2‐DCA 

by Method 8260. Details of sampling, preservation, 

shipment, storage, and analysis are presented in 

Appendix A. Samples were provided between 

October 2005 and July 2007. A total of 19 states 

provided samples from a total of 802 monitoring 

wells spread over 102 sites (Table 1.1).

Table 1.1. Occurrence of EDB in ground water provided by the state agencies that participated in the 

EPA/ASTSWMO State Study.

State Number Samples Number Sites Number Sites Number Sites

EDB detected EDB>MCL

Arizona 12 1 1 1

Colorado 107 9 4 2

Georgia 12 2 1 1

Maryland 27 3 2 2

Minnesota 29 4 3 3

Missouri 16 8 2 2

Mississippi 28 2 2 2

Montana 31 11 5 3

North Carolina 25 6 6 6

North Dakota 34 6 2 0

New Hampshire 63 4 3 3

New Mexico 15 1 1 1

Oklahoma 57 10 6 5

South Carolina 50 5 5 5

South Dakota 67 5 2 0

Tennessee 90 7 3 1

Utah 37 5 3 3

Vermont 19 3 1 1

Wisconsin 83 10 2 2

Total 802 102 54 43

 

 
5

1.4 Purpose and Scope of This Report

This report represents the culmination of Phase 2 

of the EPA/ASTSWMO study. It fills some of 

the data gaps on the expected distribution of EDB 

at gasoline release sites, it identifies a previously 

unsuspected role of iron(II) sulfides in the abiotic 

transformation of EDB in ground water, and 

provides new tools to recognize and use natural 

transformation and degradation of EDB and 

1,2‐DCA as part of a risk management strategy.
6

This section is intended for technical personnel who 

will conduct a risk evaluation for EDB or 1,2‐DCA 

at specific motor fuel release sites, and for the 

technical staff of regulatory agencies that review 

the risk evaluations and make decisions concerning 

risk management and cleanup of the contamination.

The implementation of Monitored Natural 

Attenuation (MNA) as part of a program of riskbased corrective action for contaminants in ground 

water requires a robust understanding of the 

exposure to the contaminant, which in turn requires 

a robust understanding of the transport and fate 

of the contaminant from the point of release to 

the receptor. This section presents a conceptual 

model for the behavior of a release of EDB and 

1,2‐DCA in leaded gasoline into the subsurface 

environment. It discusses the available information 

on the attenuation of EDB and 1,2‐DCA caused 

by weathering of the original mass of gasoline 

released, and the relationship between weathering 

and the persistence of EDB and 1,2‐DCA in source 

areas at motor fuel release sites. This section also 

discusses the effect of sorption on the attenuation 

of concentrations of EDB and 1,2‐DCA along a 

flow path in ground water, and it also presents 

new information on the rate of abiotic degradation 

caused by sulfide in solution in ground water, or 

by FeS mineral phases precipitated in the aquifer 

matrix as a result of sulfate reducing activity. This 

section also discusses prospects for biodegradation 

of EDB and 1,2‐DCA, and provides data on the rate 

of anaerobic biodegradation of EDB and 1,2‐DCA 

in ground water at gasoline spill sites. It also 

describes the application of Compound Specific 

Isotope Analysis (CSIA) for documenting the 

degradation of EDB and 1,2‐DCA. 

2.1 Conceptual Model of a Motor Fuel 

Release in the Subsurface

Gasoline released from an underground storage 

tank seeks the water table. If it is released above 

the water table it drains downward by gravity. 

If it is released below the water table it rises 

by buoyancy. Gasoline can act as a wetting 

phase for particles in the unsaturated zone. As a 

consequence, it tends to move into the capillary 

fringe just above the water table where it is held by 

capillary attraction. Over time, capillary attraction 

will re-distribute the gasoline in a roughly circular 

or oval shape just above the water table. As the 

re-distribution continues, the area contaminated 

with liquid phase gasoline gets larger and the 

concentration of liquid phase gasoline declines until 

the gasoline can no longer maintain a continuous 

wetting phase on the aquifer solids. At this point 

the gasoline breaks into discrete droplets that are no 

longer in contact with each other, the redistribution 

of gasoline stops or slows, and the gasoline is said 

to be at residual saturation. 

There are several ways that soluble contaminants 

from gasoline can enter the ground water. Soluble 

contaminants can diffuse from the gasoline in 

the capillary fringe down into the ground water. 

Recharge water draining down through the 

unsaturated zone can perfuse the gasoline and carry 

soluble contaminants to the ground water. Finally, 

variations in the elevation of the water table can 

inundate the residual gasoline, allowing soluble 

contaminants to partition directly into the ground 

water. At most sites, the elevation of the water 

table will vary a few inches to a few feet in a year’s 

time. As the water table moves up and down, the 

gasoline in the capillary fringe moves up and down, 

creating a “smear zone” that contains gasoline 

at residual saturation. Under typical conditions, 

the lower portion of the smear zone is below the 

average elevation of the water table. 

The most contaminated wells at a site are usually 

screened in material that has gasoline at residual 

saturation. Over time, residual gasoline tends to 

accumulate in the geological material with the 

finest texture: silt and clays rather than sands or 

gravels. This is probably because the material with 

2.0 

Transport and Fate of EDB and 1,2-DCA at 

Motor Fuel Release Sites
7

finer texture has a higher capillary attraction for the 

residual gasoline.

The concentration of soluble contaminants in 

ground water produced by the wells is controlled 

by equilibrium partitioning of the contaminants 

between the ground water and residual gasoline. 

As a consequence, the rate of attenuation of 

concentrations of EDB over time in monitoring 

wells in the source area of releases is controlled by 

the rate at which EDB dissolves from the residual 

gasoline into the flowing ground water and is either 

flushed away by ground water flow, or destroyed by 

biodegradation or abiotic transformation.

The rate of attenuation over time in sediment 

contaminated with residual gasoline determines 

how long a release of gasoline can continue to 

contaminate ground water. Once a contaminant 

enters the flowing ground water and moves away 

from the residual fuel in the source area, the 

concentration of the contaminate can be attenuated 

by processes such as dilution and dispersion, 

sorption to native organic matter, biodegradation 

by aerobic or anaerobic microorganisms, by neutral 

hydrolysis, or by non biological reactions with 

either sulfide in solution, or with sulfide minerals. 

The rate of removal in the flowing ground water 

will determine how far the plume of contamination 

can reach. 

2.1.1 Mathematical Description of Rates 

of Attenuation

The exposure assessment that is conducted as 

part the RBCA processes usually incorporates a 

mathematical description of the behavior of the 

contaminants at a site. This can be as simple as 

a few calculations with equations that describe 

the transport and fate of the contaminants, or 

it can require the calibration of a computer 

model to describe the behavior of the plume of 

contaminated ground water in time and space. In 

any case, the RBCA process requires a quantitative 

understanding of the behavior of the contaminant. 

The following section reviews the mathematics 

typically used to describe the behavior of organic 

contaminants in ground water, identifies rates of 

removal that are needed for MNA to be a practical 

alternative for EDB and 1,2‐DCA at many sites, 

and reviews the literature on the rates of attenuation 

of EDB and 1,2‐DCA over time in the source area 

of plumes in ground water at gasoline spill sites. 

2.1.2 Relationship between a First Order 

Rate Constant and a Half Life

Attenuation processes that follow a first order rate 

law can be described with either a half life or a 

pseudo first order rate constant. Readers that are 

familiar with these relationships can skip to the 

next section. 

When the rate of reaction is proportional to the 

concentration of the contaminant, the progress of 

the reaction can be described by equation 2.1;

t kt

o

C

e

C = (2.1)

where t is the time elapsed during the reaction, 

Ct is the concentration after time t, Co is the 

original concentration, and k is the first order rate 

constant for the instantaneous rate of change in 

concentration over time. When k has a negative 

value, concentrations are declining over time. The 

rate constant for the rate of attenuation, the rate of 

abiotic transformation, or the rate of biodegradation 

is the negative of the rate constant for the rate 

of change (k) in Equation 2.1. The solution of 

Equation 2.1 for k yields Equation 2.2.

k CC t = ln ( t o ) (2.2)

The half life corresponds to the value of t where 

one half of the contaminant has been destroyed, as 

described in Equation 2.3.

1 ( ) 2

t kk = =− ln 0.5 0.693

(2.3)

A half life can be converted to a first order rate 

constant or vice versa by dividing one parameter 

into -0.693 to calculate the other parameter. The 

unit for a half life is time (e.g. years), and the unit 

for a first order rate constant is reciprocal time (per 

year). 

Most readers have an intuitive grasp of a half 

life, and as a consequence, microbiologists and 

other life scientists commonly describe first 

order processes with a half life. Engineers tend 

to describe first order processes with a first order 

rate constant for two important reasons. The 

rate of change is directly related to the constant. 

In addition, if several first order processes are 

operating simultaneously, such as hydrolysis and 

biodegradation, the rate constant for the combined 

effect is simply the sum of the individual rate 

constants. This property is particularly convenient 

when calibrating transport and fate models 

because several processes can be combined into 

one calibration parameter. The remainder of this 

report will describe first order processes using first 

order rate constants. For readers who are more 

comfortable with half lives, and do not have access 

to a calculator, Figure 2.1 can be used to convert 

first order rate constants in units of per year to half 

lives in days, weeks or years.
8

Figure 2.1. Relationship between a first order rate 

constant in units of per year and halflives in units of days, weeks, and years.

2.1.3 A Definition of a “Generally 

Useful” Rate of Attenuation 

To put a rate of attenuation into context for natural 

attenuation, it is necessary to define a rate that 

might be useful for MNA. Any determination 

of “useful” is site specific depending on the 

hydrological context of a particular release, and the 

proximity of receptors. On a site-specific basis, 

the useful rate of attenuation is the rate that will 

achieve the remedial objectives within a reasonable 

time frame (U.S. EPA, 1999). If the degradation 

follows first order kinetics, then the first order 

rate constant (k) for a “useful” rate of removal is 

defined by the relationship in Equation 2.4.

(MCL Current Concentration)

Time Available

Ln k = −

(2.4)

Any release of conventional leaded gasoline is at 

least 20 years old. For the purposes of discussion 

and comparison, we will arbitrarily define a 

“generally useful” rate as the rate that will bring 

the concentration of EDB or 1,2‐DCA that would 

be expected in ground water in contact with 

unweathered leaded gasoline to the MCL for EDB 

or 1,2‐DCA within an additional 20 years. 

Falta (2004) used the average composition of 

gasoline and partitioning theory to estimate the 

concentration of EDB and 1,2‐DCA, that would 

be expected in ground water in contact with 

unweathered leaded gasoline. The expected 

concentrations for EDB and 1,2‐DCA, were 

1,900 μg/L and 3,700 μg/L respectively. The first 

order rate of removal that would bring EDB from 

the maximum concentration expected (1,900 μg/L) 

to the MCL (0.05 μg/L) in 20 years is 0.5 per year, 

and the corresponding rate that would bring the 

maximum concentration of 1,2‐DCA of 3,700 μg/L 

to 5 μg/L in 20 years would be 0.33 per year. 

Therefore the “generally useful” rate of degradation 

of EDB would be 0.5 per year or greater and the 

“generally useful” rate of degradation of 1,2‐DCA 

would be 0.33 per year or greater. 

2.2 Attenuation caused by physical 

processes 

The flow of ground water through the residual 

gasoline should weather (“leach”) EDB and 

1,2‐DCA from the residual source material. With 

each exchange of the pore water in contact with 

residual gasoline, some fraction of the total 

amount of EDB and 1,2‐DCA would be flushed 

away from the source area. Typical values for the 

concentration of gasoline at residual saturation 

vary between 2,000 and 10,000 mg/kg of Total 

Petroleum Hydrocarbon (TPH). Appendix A 

provides equations that can be used to predict the 

distribution of EDB and 1,2‐DCA between residual 

gasoline and ground water. At these typical values 

for TPH, the fraction of the total amount of EDB 

that is dissolved in ground water would vary 

between 30% and 7%, and the fraction of 1,2‐DCA 

in ground water would vary from 40% and 12%. 

Because a relatively small proportion of EDB or 

1,2‐DCA is in the ground water, these contaminants 

would be expected to weather slowly from residual 

gasoline. This leads to two questions: How fast 

is EDB and 1,2‐DCA weathered from gasoline? 

What are the expected concentrations of EDB and 

1,2‐DCA in the plume? 

2.2.1 Physical Weathering from Fuel 

Present at Residual Saturation

The rate of exchange of pore water in contact 

with residual gasoline will depend on the seepage 

velocity of ground water and on the distance 

ground water must move to traverse the region 

with residual gasoline. The faster the ground water 

moves, the faster the EDB and 1,2‐DCA should 

be removed, therefore, higher seepage velocities 

should be associated with lower concentrations 

of EDB. The average seepage velocity of ground 

water is usually calculated from an estimate of 

hydraulic conductivity that is acquired from a 

slug test on a monitoring well. Because most 

monitoring wells are screened across materials with 

different textures, the average seepage velocity may 

underestimate the local seepage velocity through 

the material contaminated with residual gasoline. 
9

The South Carolina Dept. of Health & 

Environmental Control collected data on the 

hydrological characteristics of 150 MTBE plumes 

in South Carolina (Personal Communication, 

Read Miner, South Carolina Dept. of Health & 

Environmental Control). The seepage velocity 

at each plume was estimated by multiplying the 

hydraulic gradient by the hydraulic conductivity, 

then dividing by 0.3 as an estimate of the effective 

porosity. Figure 2.2 presents the frequency 

distribution of seepage velocity in the plumes. The 

median seepage velocity was 4 m/yr, and only 

10% of plumes exceeded a velocity of 50 m/yr. 

Measurements in the same data set indicated that 

the median width of the source area for MTBE 

plumes was 100 feet or 30 meters.

Figure 2.2. Distribution of seepage velocity in 

MTBE plumes in South Carolina 

(2002).

The following thought experiment is offered to 

put the rate of weathering of EDB and 1,2‐DCA in 

context for risk management. Assume, following 

Falta (2004), that the initial concentrations of EDB 

and 1,2‐DCA are 1,900 μg/L and 3,700 μg/L.

The distribution of EDB and 1,2‐DCA between 

water and non-aqueous phase gasoline is controlled 

by the partitioning coefficient between water and 

gasoline, and by the relative proportions of pore 

water and gasoline in the aquifer. Appendix A 

derives equations that can be used to predict the 

distribution of EDB and 1,2‐DCA based on the 

parameters. Assume that the concentration of 

residual gasoline is near 5,000 mg/kg, and that the 

total porosity is 30%. The equations in Appendix 

A would predict that 12% of EDB and 20% of 

1,2‐DCA is removed whenever one pore volume of 

ground water in contact with the residual gasoline 

is exchanged. The rate of exchange is the length 

of the region with residual gasoline divided by the 

seepage velocity of ground water. 

Assume that the footprint of residual gasoline from 

releases of leaded gasoline had the same geometry 

as the releases of gasoline with MTBE, and that the 

length of the region with residual gasoline in the 

direction of ground water flow is equal to the width 

perpendicular to the flow. The assumed length is 

30 meters. 

Figure 2.3 projects the time course of 

concentrations of EDB and 1,2‐DCA in ground 

water in contact with residual gasoline if the 

ground water seepage velocity is 4 m/yr (median 

velocity in South Carolina) or 50 m/yr (faster than 

90% of sites). The year 1987 was taken as the last 

year for a plausible release of EDB from leaded 

gasoline in an underground storage tank. The 

EPA/ASTSWMO study was conducted twenty 

years later in 2006 and 2007.

Figure 2.3. Thought experiment on the effect of the 

rate of ground water flow on the concentration of EDB and DCA in ground 

water in contact with residual leaded 

motor fuel. (Consult the text for assumptions in the comparisons.)

If the seepage velocity of ground water was 

50 m/yr, the concentration of EDB in 2007 would 

be near 27 μg/L. While this represents a 70-fold 

reduction in concentration from the original 

concentration of EDB, the concentration is still 

540-fold greater than the MCL. The concentration 

of EDB would not be expected to reach the MCL 

until 2037. If the seepage velocity were 4 m/year, 

the concentration of EDB would be reduced by 

less than 30% by 2007. It is reasonable to expect 

significant concentrations of EDB to remain in 

ground water at many releases of leaded gasoline.

In the case of 1,2‐DCA, when the seepage velocity 

is 50 m/yr, concentrations of 1,2‐DCA would be 
10

expected to meet the MCL before samples were 

collected for the study in 2006 and 2007. If the 

seepage velocity was 4 m/yr, concentrations of 

1,2‐DCA would only be reduced to 50% of the 

original concentration. Because 1,2‐DCA partitions 

to ground water more readily than EDB, it weathers 

more rapidly, but there is little practical difference 

in their behavior as it will take decades to centuries 

to reach their respective MCLs. 

The survey conducted by the EPA/ASTSWMO 

Lead Scavengers Team provides a small data set 

that can be used to validate the thought experiment. 

Case workers in the state agencies were asked 

to review files on the sites, and provide either an 

estimate of the seepage velocity of ground water 

from the file, or data on the hydraulic gradient and 

hydraulic conductivity that could be used with a 

reasonable estimate of effective porosity to estimate 

the seepage velocity of ground water. Data are 

available from ten sites. 

Figure 2.4 compares the average seepage velocity 

of ground water at the sites to the highest 

concentrations of EDB at each site in 2006. In 

general, the concentrations were lower than would 

be expected from the thought experiment. In eight 

of the ten sites, the maximum concentration of 

EDB was 1 μg/L or less, regardless of the seepage 

velocity of the ground water. This would indicate 

that some process other than leaching controlled the 

concentrations of EDB at these eight locations. The 

sites with the highest concentrations of EDB had 

seepage velocities of 11.5 and 71 meters per year. 

These velocities are relatively high, exceeding 72% 

and 92% of sites in South Carolina respectively 

(compare Figure 2.2). Despite these high seepage 

velocities, the concentrations of EDB were above 

400 μg/L, consistent with the slow rates of physical 

weathering predicted by the thought experiment. 

Figure 2.4. Association of concentrations of EDB 

in the most contaminated wells at 

each of ten sites with estimates of the 

seepage velocity of ground water at the 

sites.

2.2.2 Sorption on Native Organic Matter

As long as the ground water is in contact with 

residual gasoline, the concentration of EDB or 

1,2‐DCA is controlled by the concentration of EDB 

and 1,2‐DCA in the gasoline. Once contaminated 

ground water moves away from the area with 

gasoline at residual saturation, the concentrations 

of EDB or 1,2‐DCA are controlled by sorption 

to solids in the aquifer matrix, by biological 

degradation and abiotic transformation, and by 

dilution and dispersion in ground water. 

Retardation due to sorption is not an important 

process contributing to natural attenuation of EDB 

or 1,2‐DCA in ground water. Table 2.1 compares 

the retardation ratio of EDB, 1,2‐DCA, MTBE, 

benzene, toluene, and xylenes that is expected 

from sorption of the contaminant to the native 

organic carbon in the aquifer. The retardation ratio 

is simply the rate of movement of water divided 

by the rate of movement of the contaminants. 

As discussed in Wiedemeier et al., (1999), the 

estimates of the retardation ratio in Table 2.1 are 

based on Equation 2.5; 

1 R Kf = +ρ ∗ ∗ θ b oc oc (2.5)

where R is the retardation ratio, ρb is the bulk 

density, Koc is the partition coefficient of the 

organic contaminant between ground water and 

native organic matter, foc is the fraction organic 

carbon in the aquifer matrix material, and θ is 

the water filled porosity. In Table 2.1, values for 

Koc are taken from Wiedemeier et al., (1999), 

except for the value of Koc for EDB. Aronson and 

Howard (2008) concluded that “Soil-water partition 

coefficients (Koc values) for EDB in the solution 
11

phase range from 12 to 160 L/kg ... but average 

about 50 to 65 L/kg.” The calculation in Table 2.1 

assumes a value of 65 L/kg for the Koc of EDB.

The extent of sorption of EDB and 1,2‐DCA 

is intermediate between the extent of sorption 

of benzene and toluene. At concentrations of 

organic carbon in the aquifer solids that are near 

0.01%, there should be little retardation of EDB or 

1,2‐DCA. At concentrations of organic carbon in 

the aquifer solids near 0.1%, retardation of EDB 

and 1,2‐DCA is of no practical consequence. These 

concentrations are typical of aquifers with low or 

average concentrations of organic matter in the 

aquifer solids. At relatively high concentrations of 

organic carbon in the aquifer solids near 1%, the 

expected retardation ratio for EDB and 1,2‐DCA is 

near four to one. 

Sorption can reduce concentrations in plumes that 

are expanding, but when a plume reaches steady 

state, sorption does not influence the concentration 

of the contaminant in ground water. As discussed 

in Section 1, the use of leaded gasoline for 

conventional motor gasoline was minimal after 

1988. At the time of this writing, almost every 

plume of EDB or 1,2‐DCA from a release of 

leaded gasoline from an underground storage tank 

at a conventional gasoline service station is at 

least twenty years old. In this time interval, it is 

reasonable to presume that most plumes of EDB or 

1,2‐DCA have reached (or are approaching) steady 

state. Sorption can not be an important mechanism 

that will bring high concentrations of EDB or 

1,2‐DCA to their MCLs. 

2.2.3 Rate Constants for Physical 

Weathering of EDB and 1,2‐DCA

There is very little data in the literature on trends of 

EDB concentrations in monitoring wells at gasoline 

release sites. Bulsara (2004) fitted first order rate 

constants for attenuation of EDB to monitoring 

data from 55 wells in South Carolina. The data are 

replotted in Figure 2.5. The rates varied widely, 

from approximately - 4 per year to + 4 per year. In 

32 of the 55 wells, the rate constants were negative, 

indicating that the concentrations of EDB increased 

over time. 

As a comparison, Figure 2.5 plots data for the 

attenuation of MTBE in monitoring wells at 

gasoline release sites (Wilson et al., 2005). At 

6 out of 20 release sites, the rate constants were 

negative, indicating that concentrations of MTBE 

were increasing. However, the variation in rates of 

attenuation was less, from -0.52 to +0.75 per year. 

The range in rates of attenuation of MTBE was 

much less than the range in rates of attenuation of 

EDB. There is no obvious reason why this should 

be the case. The compounds have very similar 

physical properties.

Table 2.1. Retardation in ground water due to sorption that is expected from the organic matter content of 

the aquifer solids. 

 Fraction of Organic Carbon in sediment

Compound Koc (L/kg) 0.01% 0. 1% 1% 10%

Low for Aquifers Median for 

Aquifers

High for 

Aquifers

Typical of soils

MTBE 12 1.0 1.1 1.6 7

Benzene 38 1.0 1.2 2.9 20

DCA 58 1.0 1.3 3.8 29

EDB 65 1.0 1.3 4.1 33

Toluene 135 1.1 1.7 7.6 68

Xylenes 240 1.1 2.2 13 120
12

Figure 2.5. Rate of decline of concentrations of 

EDB over time in 55 monitoring wells 

in South Carolina (Bulsara, 2004). 

Rate of decline of concentrations of 

MTBE at 20 sites in the USA (Wilson 

et al., 2005) provided for reference.

Example data provided in Falta (2004) indicated 

that the rates for attenuation of EDB were extracted 

from three or four measurements extending over a 

monitoring period of one or two years. In contrast, 

the rates of MTBE attenuation were extracted from 

6 to 34 measurements (mean of 14 measurements) 

extending over a minimum of two years. The 

comparison of the EDB rate constants to the 

MTBE rate constants suggests that rate constants 

extracted from small data sets collected over short 

time periods may have large uncertainty. If a 

rate constant extracted from monitoring data at a 

particular release site is used to estimate the time 

required for concentrations of EDB to decline 

below the MCL, the rate should be verified and 

validated by continued long term monitoring.

2.3 Attenuation Caused by Abiotic 

Transformation or Biodegradation

For many processes that destroy contaminants, 

the rate of the reaction is proportional to the 

concentration of the contaminant at any moment 

in time. This is particularly true for abiotic 

reactions such as hydrolysis or abiotic reactions 

with minerals. For biodegradation reactions at low 

concentrations, the diffusion of the contaminant to 

the organism, and association of the contaminant 

with the enzymatic machinery of the organism, is 

rate limiting. At these lower concentrations, the 

rate of reaction is proportional to the concentration 

of the contaminant. At higher concentrations, the 

processing of the contaminant by the enzymatic 

machinery becomes rate limiting, and the rate of 

reaction is proportional to the density of cells, 

not the concentration of the contaminant. This 

transition in the rate limiting step for anaerobic 

biodegradation of EDB should be near 1,000 μg/L. 

At typical concentrations at releases of leaded 

gasoline, the rate of EDB biodegradation should be 

proportional to the concentration of EDB, and the 

process can be described with first order kinetics.

2.3.1 Neutral Hydrolysis of EDB and 

1,2‐DCA

Both EDB and 1,2‐DCA are slowly hydrolyzed in 

water. Barbash and Reinhard (1989) reported a 

half life for neutral hydrolysis of EDB of 22 years 

at 15 °C and 4.6 years at 25 °C. These half lives 

correspond to first order rate constants of 0.073 

per year at 15 °C and 0.15 per year at 25 °C. The 

rate of neutral hydrolysis of 1,2‐DCA was slower; 

the half life was 300 years at 15 °C and 64 years at 

25 °C, corresponding to first order rate constants 

of 0.0023 per year at 15 °C and 0.0108 per year at 

25 °C.

Hydrolysis of EDB can produce either 

2-bromoethanol or vinyl bromide (Pignatello and 

Cohen, 1990; Aronson and Howard, 2008), with 

2-bromoethanol being the dominant product. 

Similarly, 1,2‐DCA hydrolyzes to 2-chloroethanol 

(Jeffers et al., 1989) and vinyl chloride with vinyl 

chloride being a minor product (Barbash and 

Reinhard, 1989). 

Both EDB and 1,2‐DCA can undergo base 

catalyzed hydrolysis; however, the rate of the base 

catalyzed hydrolysis is not important at the pH of 

natural ground waters. At pH 9, the base catalyzed 

reaction represents only 10% of the total hydrolysis 

reaction (Jeffers and Wolfe, 1996).

2.3.2 Abiotic Transformation of EDB and 

1,2‐DCA by Iron(II) Sulfide

Both EDB and 1,2‐DCA can also react with sulfide 

as H2S and HS- to produce various thiols and 

thioethers (Schwarzenbach et al., 1985). Barbash 

and Reinhard (1989) compared the expected rates 

of hydrolysis and reaction with sulfide at 15 °C and 

25 °C. In Figure 2.6, the combined rate constants 

for hydrolysis and reaction with sulfide are used 

to project the effects of the concentration of total 

sulfide at pH 7 on the rate of removal of EDB and 

1,2‐DCA. In general, the rate of reaction of EDB 

is approximately ten times the rate of reaction of 

1,2‐DCA (compare the scales of the vertical axes 

in Figure 2.6). The reaction of both EDB and 

1,2‐DCA is sensitive to temperature; rates at 25 °C 

are approximately five fold faster than at 15 °C. 

Reactions with sulfide only become important at 

concentrations above 0.2 mg/L for EDB and 1 mg/L 

for 1,2‐DCA.
13

Figure 2.6. Effect of concentrations of sulfide at 

pH 7 on the rate of abiotic transformation of EDB and DCA.

Depending on temperature, the “generally 

useful” rate of EDB degradation of 0.5 per year 

is attained at concentrations between 2 and 10 

mg/L total sulfide at pH 7. The “generally useful” 

rate for 1,2‐DCA of 0.33 per year is attained at 

concentrations between 90 and 400 mg/L of sulfide. 

Sulfide produced as the end product of sulfate 

reduction can react with iron(III) minerals in the 

aquifer matrix to produce various mineral phases of 

iron(II) sulfide according to the following reactions 

(Shen and Wilson, 2007).

4H2 + SO4

-2 → S-2 + 4H2O

2Fe+3 + 3S-2 → 2FeS + S0

FeS + S0 → FeS2

The overall reaction consumes three moles of 

sulfate to produce one mole of FeS and one mole of 

FeS2.

12H2 + 3SO4

-2 + 2Fe+3 → FeS + FeS2 + 24H2O

The transformation of chlorinated hydrocarbons 

such as TCE and cis-DCE by FeS and FeS2 is well 

documented in the literature (Butler and Hayes, 

1999; Butler and Hayes, 2001; Shen and Wilson, 

2007; Liang et al., 2007). Shen and Wilson (2007) 

reported abiotic transformation of TCE in four 

laboratory columns constructed with river sand and 

shredded plant mulch. Sulfate reduction supported 

by the plant mulch produced sulfide, which reacted 

with Fe(III) minerals in the river sand to produce 

non crystalline or poorly crystalline FeS. Two of 

the columns were supplemented with hematite to 

encourage precipitation of FeS in the columns. In 

these two columns, the first order rate constant for 

reaction of TCE with the FeS varied from 0.53 to 

2.3 per day per mole of FeS in contact with a liter 

of pore water.

The EPA/ASTSWMO Lead Scavengers Team 

conducted an experiment to determine whether 

EDB or 1,2‐DCA could be degraded by FeS. To 

evaluate the capacity of biogenic FeS to degrade 

EDB and 1,2‐DCA, the column described by Shen 

and Wilson (2007) as the Column with Mulch and 

Hematite was frozen, cut into sections while frozen, 

and then the sections were allowed to thaw under 

an oxygen free atmosphere in a glove box. To 

remove the confounding effects of sorption to the 

plant mulch on concentrations of contaminants, the 

sections were sieved to remove the plant mulch. 

The sediment and pore water were transferred to 

20 ml serum vials, dosed with solutions of EDB 

or 1,2‐DCA, and incubated at room temperature. 

Experimental details are provided in Appendix B.

The removal of EDB and 1,2‐DCA followed 

first order kinetics without a lag, indicating that 

removal was an abiotic process that did not 

require acclimation of an active biological process 

(Figure 2.7). The removal of EDB was more 

rapid than the removal of 1,2‐DCA. The first 

order rate constants for degradation of EDB or 

1,2‐DCA were extracted from the data as the slope 

of a linear regression of the natural logarithm of 

concentration on time of incubation. The rate 

constants are presented in Table 2.2. At the end of 

the incubation, the concentration of Acid Volatile 

Sulfide or AVS was measured in each of the vials 

(mg/kg dry sediment). The AVS was taken as an 

estimate of FeS in the sediment. The water content 

of the wet sediment was determined by drying the 

sediment. The concentration of AVS in each vial 

was expressed as the millimoles AVS exposed to 

each liter of pore water. Finally, the rate of removal 

was normalized to the millimoles AVS exposed to 

the pore water.
14

Figure 2.7. Removal of EDB or DCA in the presence of biogenic FeS in batch experiments at pH near 7.

The normalized rate constants for transformation 

of EDB were consistent between the experimental 

vials. The average normalized rate for EDB 

degradation was 0.285 ± 0.091 per year per 

millimole AVS exposed to a liter of water at 95% 

confidence. The rate of transformation of EDB 

was in the same range as the rates of abiotic 

transformation of TCE in the original column. 

The rate of 1,2‐DCA degradation was an order of 

magnitude lower at 0.0263 per year per millimole 

AVS exposed to a liter of water. 

Analysis of acid volatile sulfide is simple 

to perform, and is commercially available. 

Figure 2.8 presents the rate of abiotic degradation 

of EDB and 1,2‐DCA that would be expected 

if the kinetics of degradation follow the rates 

of degradation presented in Figure 2.7. The 

detection limit for the analysis is near 2 to 3 mg/kg. 

The rates of abiotic degradation of EDB are 

“generally useful” at concentrations of AVS near 

10 mg/kg. Concentrations of AVS must approach 

100 mg/kg to provide “generally useful” rates of 

transformation of 1,2‐DCA.

Table 2.2. Comparison of the rate of transformation of EDB and DCA by iron(II) sulfide to the rate or 

removal of TCE.

Compound Experimental unit Rate of Removal Concentration FeS 

as AVS*

Rate of Removal

yr-1 mM** yr-1mM-1

EDB 2-10 76.4 198 0.386

EDB 2-11 65.1 272 0.239

EDB 2-12 64.9 278 0.233

EDB 6-15 62.6 381 0.164

EDB 6-s4 94.8 236 0.402

DCA 6-16 10.2 293 0.0348

DCA 6-17 6.7 140 0.0479

Data from Table 1 of Shen and Wilson (2007)

TCE Entire Column

Day 383

80 138 0.584

TCE Entire Column

Day 578

193 230 0.840

TCE Entire Column

Day 793

55 283 0.193

*Acid Volatile Sulfide

** millimoles AVS in contact with 1.0 liter pore water.
15

Figure 2.8. Expected rates of abiotic transformation of EDB or 1,2‐DCA by FeS in 

aquifer sediment, assuming a waterfilled porosity of 25%, and pH near 7.

The abiotic reaction of EDB or 1,2‐DCA occurs on 

the surface of the solid iron(II) sulfide. As a result, 

the rate of the reaction should be proportional to 

the surface area and not to the mass of the iron(II) 

sulfide, and the reaction rate will be faster if the 

iron(II) sulfide is finely divided. The rate of abiotic 

transformation of EDB and 1,2‐DCA has only been 

determined for this one preparation of biogenic 

iron(II) sulfide. It is not known at present if the 

properties of biogenic iron(II) sulfide in aquifer 

sediments vary from site to site. The projections 

in Figure 2.8 are provisional until more data can be 

obtained from field sites.

Butler and Hayes (2001) found that the rate of 

transformation of TCE on iron(II) sulfide decreased 

as the pH decreased. There is no information on 

the effect of pH on the rate of transformation of 

EDB and 1,2‐DCA. However, it is possible that 

the rates will be substantially lower at pH less than 

6.5. Based on sustained concentrations of sulfide in 

ground water and the temperature, reactions with 

sulfide can make a major contribution to removal 

of EDB and 1,2‐DCA in ground water, once the 

water moves away from the area of the plume with 

residual gasoline.

2.3.3 Biodegradation of EDB and 

1,2‐DCA

Both EDB and 1,2‐DCA can be rapidly degraded if 

oxygen is available (Aronson and Howard, 2008). 

Oxygen may be available in the far down gradient 

portion of the plume where metabolism of the 

petroleum hydrocarbons is complete, and diffusion 

and dispersion have mixed oxygen from unimpacted ground water into the plume. However, 

the oxygen demand of the petroleum hydrocarbons 

make it unlikely that oxygen will be available in 

the source area of a plume from a release of leaded 

gasoline, or in the mid gradient portions of the 

plume. 

Methane producing bacteria can co-metabolize 

EDB and 1,2‐DCA to ethylene when the cells are 

grown on molecular hydrogen and carbon dioxide 

(Belay and Daniels, 1987). The reaction does 

not support growth of cells, and the halogenated 

compounds can be harmful to the methane 

producing bacteria. Concentrations of EDB 

near 1,300 μg/L inhibit growth of the methane 

producing bacteria by 90%. In contrast, 1,2‐DCA 

has less effect; higher concentrations of 1,2‐DCA 

near 11,000 μg/L are required to inhibit growth.

Bacteria in the Dehalococcoides group can use 

EDB or 1,2‐DCA as a substrate. These bacteria 

grow with molecular hydrogen as a source of 

reducing power, and use EDB or 1,2‐DCA as 

the electron acceptor, basically as something to 

breathe. Strain Dehalococcoides BAV1 (He et 

al., 2003a) can grow while metabolizing EDB to 

ethylene, and Dehalococcoides ethanogenes 195 

can grow on 1,2‐DCA (Maymó-Gatell et al., 1999) 

and metabolize either EDB or 1,2‐DCA to ethylene 

(Maymó-Gatell et al., 1997).

No information is available in the literature on the 

effect of concentrations of EDB on the rate of EDB 

degradation by Dehalococcoides strains. The effect 

is usually described by the half saturation constant, 

the concentration at which the rate is one-half of the 

maximum possible rate. This is the concentration 

where the rate of degradation becomes dependent 

on the density of active organisms instead of the 

concentration of the substrate. He et al., (2003b) 

published half saturation constants for degradation 

of vinyl chloride, cis-DCE, and trans-DCE by 

the mixed culture from which Dehalococcoides 

BAV1 was isolated. The values were 5.8, 8.9, and 

8.5 μM for vinyl chloride, cis-DCE, and trans-DCE 

respectively. Cupples et al., (2004) determined half 

saturation constants for vinyl chloride and cis-DCE 

in a mixed culture containing the bacterium VS 

strain of Dehalococcoides. The values of the half 

saturation constant were 2.6 and 3.3 μM for vinyl 

chloride and cis-DCE respectively. 

The value of the half saturation constant is 

controlled in part by the affinity of the substrate 

for microbial tissue, and for the active site of 

the enzyme. The affinity is inversely related to 

the water solubility of the substrate. The water 

solubility of EDB, vinyl chloride, and cis-DCE 

are 4,300 mg/L, 2,700 mg/L and 800 mg/L 

respectively.
16

The closest match to EDB is vinyl chloride. If 

the half saturation constant of vinyl chloride is 

considered the best estimate of the half saturation 

constant for EDB, the constant would vary between 

5.8 millimolar and 2.6 millimolar, or 1,100 to 

490 μg/L. 

Considering the effect of EDB on the growth of 

methane producing bacteria along with the half 

saturation constant for Dehalococcoides strains, 

a value of 1,000 μg/L can be considered an upper 

boundary where biodegradation of EDB can be 

described by a first order rate law. 

2.3.3.1 Rate Constants for 

Biodegradation of EDB and 

1,2‐DCA

Table 2.3 compares the first order rate constants for 

removal of EDB, 1,2‐DCA, and benzene that were 

extracted from laboratory studies using anaerobic 

aquifer sediment. The table is restricted to data 

from sites that were actively methanogenic and 

would reflect the geochemical environment in 

the area of a release of leaded gasoline that had 

residual gasoline, as well as, the area in the plume 

that is immediately down gradient of the area 

with residual gasoline where BTEX compounds 

persisted in the ground water. The rates of 

degradation of EDB, 1,2‐DCA, and benzene in 

the laboratory microcosm studies all had a fairly 

narrow range, extending from 17 per year to 0.3 per 

year. 

In three different studies, the rate of anaerobic 

biodegradation of EDB was equivalent to the rate 

of biodegradation of benzene. The rate constants 

attained in the laboratory studies meet the arbitrary 

criteria established earlier for “generally useful” 

rate constants for applications to MNA (0.5 per 

year for EDB and 0.33 per year for 1,2‐DCA). 

Anaerobic biodegradation can make an important 

contribution to MNA of EDB and 1,2‐DCA in 

contaminated ground water, particularly under 

methane producing conditions.

Table 2.3 also compares rate constants that were 

extracted from the distribution of contaminants 

along the flow path in the aquifer. In Table 2.3, the 

rate constant was calculated from data provided 

in Mayer (2006) by multiplying the first order rate 

constant for attenuation of concentrations with 

distance along the flow path by the upper value of 

the range of the ground water seepage velocity at 

the site. 

Washington and Cameron (2001) warn that field 

scale rates of attenuation include the effects of 

sorption and dispersion and dilution. Data in 

Table 2.3 are from old releases that have reached 

their maximum extent and are stable. They have 

likely reached sorptive equilibrium. The length of 

plumes reported by Henderson et al., (2008) and 

Mayer et al., (2006) is relatively short compared 

to the likely width of the source area. The rate 

constants reported by Ravi et al., (1998) are 

corrected for dispersion and dilution. In the data 

from Table 2.3, it is most likely that attenuation at 

field scale is dominated by biodegradation, which 

would explain the reasonable correspondence 

between the field scale rates and the laboratory 

studies. In general, the rate constants in the field 

are smaller than those extracted from the laboratory 

studies, corresponding to the lower end of the range 

of laboratory rates.

There is one important exception to the reasonable 

correspondence in rate constants. Falta (2004) 

extracted rates of attenuation of EDB and 

benzene from the distribution of their plumes 

of contamination that resulted from a release of 

aviation gasoline at the Massachusetts Military 

Reservation on Cape Cod, Massachusetts. The 

rate constant for EDB biodegradation was ten fold 

to one hundred fold slower than the rate constants 

seen at other sites (Table 2.3). The release was into 

an aquifer comprised of poorly sorted fine to coarse 

sands. The seepage velocity of ground water was 

in the range of 0.3 to 0.6 m per day. There may 

have been no opportunity to develop the strongly 

reducing conditions that would lead to BTEX 

fermentation to produce the H2 that is needed by the 

organisms that biodegrade EDB.

 The study of Henderson et al., (2008) is careful 

and comprehensive and is published in a respected 

peer reviewed journal. The study of Mayer 

(2006) uses conventional and generally accepted 

approaches to analyze data that was collected at 

the site for other purposes. These two studies are 

the only useful case studies in the literature on 

natural biodegradation of EDB and 1,2‐DCA at 

gasoline release sites. It would be inappropriate 

to uncritically extrapolate the behavior of EDB 

and 1,2‐DCA at these two sites to hundreds of 

thousands of sites across the United States. At 

present, the general contribution of biodegradation 

at sites across the United States is unknown. 

The appropriate application of MNA, or risk 

management, requires a site-specific knowledge of 

the behavior of the contaminants of concern. It is 

important to remember that the data in Table 2.3 

illustrate what might be possible at other sites. 

Data from Table 2.3 should not be used in the place 

of site specific data to conduct a risk evaluation at 

other gasoline release sites.
17

2.3.3.2 Association of Geochemical 

Parameters with Removal of 

EDB and 1,2‐DCA

Both EDB and 1,2‐DCA are biodegradable under 

aerobic and anaerobic conditions. Based on the 

approach taken in the Technical Protocol for 

Evaluating Natural Attenuation of Chlorinated 

Solvents in Ground Water (Wiedemeier et al., 

1998), it may be possible to use geochemical 

parameters to identify environments where 

processes that destroy EDB are favored. This subsection examines the association of concentrations 

of EDB with concentrations of the biogeochemical 

parameters nitrate-nitrogen, sulfate and sulfide, and 

methane. The association is evaluated to determine 

if there is information in these parameters that can 

be used to predict concentrations of EDB.

The sub-section is intended for staff of state 

regulatory agencies that determine which 

parameters are monitored in ground water 

contaminated by releases of leaded gasoline. The 

cost of sampling ground water and analyzing 

the samples for EDB could be reduced if it were 

possible to predict beforehand whether high 

concentrations of EDB should be expected at a 

particular gasoline release site.

To determine whether geochemical or hydrological 

parameters can be used to predict concentrations 

of EDB, certain sites from the EPA/ASTSWMO 

study that were sampled in 2006 were selected to be 

resampled in 2007 for concentrations of EDB and 

for concentrations of nitrate, sulfate, and methane. 

In the data set there was one site each from 

Colorado, Missouri, Montana, Tennessee, Utah, 

Table 2.3. Comparison of first order rate constants for biodegradation of EDB and 1,2-DCA in anaerobic 

aquifer sediment to rate constants for overall removal with ground water flow in anaerobic aquifers.

Material First Order Rate Constant for 

Attenuation (per year)

Reference

EDB DCA Benzene

Microcosm studies in laboratory, all are conducted with methanogenic material.

sediment from source zone of a spill of 

leaded gasoline, South Carolina 1.5±1.0 1.3±0.3 1.4±0.2 Henderson et al., 2008, SI

sediment from mid gradient zone of a spill 

of leaded gasoline, South Carolina 5.4±0.3 0.3±0.1 3.5±0.8 Henderson et al., 2008, SI

sediment impacted by leachate from a solid 

municipal waste landfill, Norman, Oklahoma 17 2.6 Wilson et al., 1986

sediment impacted by leachate from a solid 

municipal waste landfill, Norman, Oklahoma 1.7 Klećka et al., 1998

sediment from manufacturing site 

contaminated with DCA in Louisiana 4.4 Klećka et al., 1998

sediment from manufacturing site 

contaminated with DCA in Texas 1.2 Klećka et al., 1998

Field studies, flow path in aquifer

spill of leaded gasoline, South Carolina 1.3 0.9 1.0

Henderson et al., 

2008, Supporting 

Information

spill of leaded gasoline, North Carolina 

(1995 data) 0.63 0.71 0.9 Mayer, 2006

spill of leaded gasoline, North Carolina 

(2004 data) 0.22 0.26 Mayer, 2006

leachate from municipal solid waste landfill, 

Michigan 0.22±0.19 0.42±0.32 Ravi et al., 1998

fs-12 spill of aviation gasoline on Cape Cod, 

Massachusetts 0.03 0.14 Falta, 2004
18

and Vermont. Two sites each were selected from 

Oklahoma, New Hampshire, and South Carolina, 

and three sites were selected from Minnesota. 

2.3.3.2.1 Dissolved Oxygen

Oxygen is particularly difficult to determine in 

monitoring wells. Collecting water samples 

with a bailer tends to introduce oxygen from the 

atmosphere into the sample so that the measured 

dissolved oxygen content of the sample is higher 

than that of the ambient ground water. This 

concern is illustrated in data that were collected 

by Kolhatkar et al., (2000) to investigate the role 

of various geochemical parameters in the natural 

biodegradation of MTBE in ground water. In 1999 

and 2000, ground water samples were collected 

from wells at 74 sites in Pennsylvania (41 sites), 

New Jersey (8), New York (5), Florida (5), Indiana 

(7), Maryland (3), Washington DC (2), and Ohio 

(3). Analyses of dissolved oxygen, iron(III), and 

hydrogen sulfide were performed on site by staff 

of the contractors or consultants that normally 

sampled the monitoring wells using procedures 

for purging the wells that were prescribed by the 

appropriate state agencies. Because iron(II) is 

rapidly oxidized to iron(III) in the presence of 

molecular oxygen, detectable concentrations of 

iron(II) and dissolved oxygen should not occur 

together in ground water. Yet, in water produced 

from the monitoring wells (Figure 2.9), the 

sampling technicians reported high concentrations 

of oxygen in samples with high concentrations of 

iron(II). Because of the difficulty with measuring 

dissolved oxygen in water from monitoring wells, 

data on dissolved oxygen was not collected during 

the survey conducted by the EPA/ASTSWMO 

Lead Scavengers Team. 

Figure 2.9. Co-occurrence of dissolved oxygen 

and reduced iron in water samples 

produced from monitoring wells at 

gasoline spill sites.

2.3.3.2.2 Nitrate

Concentrations of nitrate were used as a surrogate 

for the availability of dissolved oxygen in ground 

water. Higher concentrations of nitrate (and 

by inference of oxygen) would support aerobic 

biodegradation of EDB, and should be correlated 

with lower concentrations of EDB. 

Figure 2.10 compares concentrations of nitrate to 

concentrations of EDB in the most contaminated 

well at each release site in the EPA/ASTSWMO 

study. Notice that the same data are plotted on an 

arithmetic and logarithmic scale for concentrations 

of nitrate nitrogen. The sites were selected to 

include both sites with high concentrations of EDB 

as well as sites with low concentrations of EDB. 

At five of the selected sites, the concentration 

of EDB was less than the reporting limit of 

0.01 μg/L. These sites had low but detectable 

concentrations of EDB in 2006.

Figure 2.10.Association of concentrations of 

nitrate with concentrations of EDB 

in the most contaminated monitoring 

wells at sites.

There was no obvious relationship between 

concentrations of EDB and concentrations of 

nitrate nitrogen. The concentration of nitrate 

nitrogen in three of fifteen sites was less than the 

detection limit of 0.01 mg/L, and the concentration 
19

in eleven of fifteen sites was less than 1.0 mg/L. 

At these sites, the concentration of EDB was 10 

μg/L or less. Contrary to what was expected, in the 

two sites with the highest concentration of EDB 

(711 and 147 μg/L), the concentration of nitrate 

nitrogen was above 1 mg/L. There was no evidence 

that high nitrate concentrations (and by inference 

high oxygen concentrations) supported aerobic 

biodegradation of EDB. 

The presence of detectable concentrations of nitrate 

nitrogen in the wells with high concentrations 

of EDB is problematic. Presumably the high 

concentrations of EDB were associated with wells 

screened near the source of residual gasoline. It is 

possible that the screens of the monitoring wells 

intercepted uncontaminated ground water over 

some depth interval, and contaminated water over 

another depth interval. This resulted in a sample of 

mixed water being collected and analyzed. When 

waters from different geochemical environments 

are mixed in a monitoring well, it will be difficult 

or impossible to use geochemical parameters to 

interpret both the geochemical nature of the ground 

water environment and the contaminant plume.

2.3.3.2.3 Sulfate and Sulfide

The consumption of sulfate during sulfate reduction 

can produce reactive HS- that can destroy EDB or 

react with iron(III) minerals to produce FeS which 

also reacts to degrade EDB. Lower concentrations 

of sulfate would indicate more extensive sulfate 

reduction and should be correlated with lower 

concentrations of EDB. Figure 2.11 compares the 

distribution of sulfate and EDB at the sites in the 

EPA/ASTSWAMO survey. 

In general, the relationship seems to be as expected; 

the two highest concentrations of EDB were 

associated with concentrations of sulfate below 

41 mg/L, and the very high concentrations of 

sulfate were associated with EDB concentrations 

below 0.1 μg/L. However, for the sites where 

the concentrations of both EDB and sulfate were 

above the detection limit, the correlation coefficient 

between the logarithm of EDB concentration 

and the logarithm of sulfate concentration is low 

(-0.15) and the sign of the correlation coefficient 

is in the wrong direction (concentrations of EDB 

go down as concentrations of sulfate go up). 

The concentration of sulfate in the wells that are 

contaminated with EDB can not be used to predict 

concentrations of EDB. 

Figure 2.11. Association of concentrations of sulfate with concentrations of EDB in the 

most contaminated monitoring wells at 

sites.

In the survey conducted by EPA/ASTSWMO 

Lead Scavengers Team, the comparison was made 

between the concentrations of EDB and sulfate in 

the contaminated ground water. A better estimate 

of the extent of sulfate reduction might have been 

the reduction in concentration of sulfate between 

the contaminated ground water and ambient ground 

water up gradient or side gradient of the release 

of gasoline. To illustrate the extent of sulfate 

reduction that might be expected at gasoline spill 

sites, Figure 2.12 compares the distribution of 

sulfate and sulfide in the sites sampled by Kolhatkar 

et al. (2000). The maximum concentration of 

sulfate in any of the wells at a site was considered 

the background concentration. The median of 

the background concentrations was 107 mg/L, 

and 75% of sites had background concentrations 

above 57 mg/L. At most of the sites, at least one 

contaminated well had concentrations of sulfate 

that were below the detection limit of 1 mg/L. In 

every site, there was a substantial reduction in the 

concentration of sulfate between the contaminated 

wells and the background wells.
20

Figure 2.12. Panel A. Distribution of sulfate in 

the source area of the plume, and in 

background ground water, at gasoline 

spill sites in the Eastern United States. 

Panel B. Distribution of sulfide in the 

source area of the plumes, compared to 

the concentration of sulfide expected 

by the amount of sulfate remove. Data 

are from the survey of Kolhatkar et al., 

(2000).

The concentration of sulfide expected from sulfate 

reduction was calculated by subtracting the 

lowest concentration of sulfate in any well from 

the greatest concentration of sulfate in any well 

at the site, then multiplying by the ratio of the 

molecular weight of sulfide and sulfate. Sulfate 

was generally depleted in the contaminated 

wells in the sites (Figure 2.12, Panel A, and 

significant concentrations of sulfide were expected 

(Figure 2.12 Panel B).

The concentrations of sulfide actually measured 

were much lower than expected (Figure 2.12 

Panel B). In half the wells, sulfide was not present 

above the detection limit of 0.1 mg/L. Only 10% 

of sites had a sulfide concentration above 3 mg/L 

and the maximum concentration of sulfide was 

10 mg/L. The concentration of iron(II) was high in 

ground water at these sites (Figure 2.9). It is likely 

that the sulfide precipitated as iron(II) sulfides. 

If these minerals persist as Acid Volatile Sulfide 

(AVS), they may have a substantial contribution to 

natural attenuation of EDB and DCA at gasoline 

spill sites.

2.3.3.2.4 Methane

The source area of gasoline release sites usually 

contains high concentrations of methane that is 

produced by anaerobic metabolism of the BTEX 

compounds in gasoline. After the soluble electron 

acceptors such as oxygen, nitrate, and sulfate are 

depleted, the metabolism of the BTEX compounds 

shifts to a fermentation reaction that produces H2

and fatty acids such as acetate. Methane producing 

bacteria require H2 or acetate as substrates to 

produce methane. The H2 and acetate produced 

during the fermentation of the BTEX compounds 

create a geochemical environment where EDB or 

1,2‐DCA can be metabolized by methane producing 

bacteria, or by Dehalococcoides strains, or by 

both. Higher concentrations of methane should be 

associated with lower concentrations of EDB.

Henderson et al. (2008) constructed microcosms 

from the source area of a release of leaded gasoline 

and from a mid gradient location. They noted 

the greatest removal of EDB in microcosms with 

the greatest production of methane. Anaerobic 

biodegradation of EDB in the microcosms of 

Henderson et al. (2008) produced significant 

fractions of 2-bromoethanol instead of ethylene. 

Bromoethanol was not identified as a product in 

the pure culture studies of methanogenic bacteria 

conducted by Belay and Daniels (1987), and has 

not been reported as a product in mixed culture 

studies of Dehalococcoides strains (Tandol et al., 

1994). However, Bouwer and McCarty (1985) 

conducted 14C label studies of the fate of EDB in 

a mixed methanogenic culture that was provided 

with acetate as the primary substrate. A significant 

fraction of label (59±6 %) was recovered in 

material that had the properties of bromoethanol. 

It is possible that the bromoethanol was produced 

by methane producing bacteria because certain 

aerobic organisms can metabolize EDB and 

produce bromoethanol as an intermediate, using 

an enzyme that does not require molecular oxygen 

(Poelarends et al., 1999; Scholtz et al., 1987). 

It is also possible that EDB was metabolized 

to bromoethanol by other anaerobic organisms 

that are yet to be characterized. In any case, the 

activity of methane producing bacteria and strains 

of Dehalococcoides can account for the observed 

anaerobic biodegradation of EDB and 1,2‐DCA. 
21

Figure 2.13 compares the association of methane 

and EDB in sites in the EPA/ASTSWMO study. 

Methane was present above its detection limit 

of 0.001 mg/l at fourteen of sixteen sites. Five 

sites met the criterion of greater than 0.5 mg/L 

of methane. In general there were lower 

concentrations of methane in the EPA/ASTSWMO 

study than was found in the study of Kolhatkar 

et al. (2000). This may reflect a depletion of 

the supply of BTEX compounds to produce the 

H2 and acetate needed as substrates for methane 

production.

Figure 2.13.Association of concentrations of methane with concentrations of EDB in the 

most contaminated monitoring wells at 

sites.

On first examination, there appears to be a 

relationship between the concentrations of methane 

and the concentrations of EDB in the wells in the 

EPA/ASTSWMO study. The two sites with the 

highest concentrations of EDB (711 and 147 μg/L) 

had concentrations of methane less than 0.01 mg/L, 

and the five sites with concentrations of methane 

greater than 0.5 mg/L had EDB concentrations 

of 1 μg/L or less. However, there are only eight 

sites where the concentrations of both EDB and 

methane were above the detection limit. For these 

eight sites, the correlation coefficient between 

the logarithm of the EDB concentration and the 

logarithm of the methane concentration is -0.13. 

The sign of the correlation coefficient is in the 

right direction (concentrations of EDB go down 

as concentrations of methane go up) but the 

association is too weak to have any predictive value 

at other sites.

2.3.3.2.5 pH

The pH of ground water has a strong effect on 

the two processes that can destroy EDB and 

1,2‐DCA in ground water. The rate of abiotic 

transformation of TCE by iron(II) sulfide increases 

as the pH increases (Butler and Hayes, 2001). 

Transformation of EDB and 1,2‐DCA by iron(II) 

sulfide should proceed by the same mechanism and 

pH should have the same effect on the rate of their 

transformation. The rates at pH of 5 or less should 

be one hundred fold lower than rates at pH 7. 

Similarly, the primary organism that is responsible 

for biological reductive dechlorination of EDB and 

1,2‐DCA is sensitive to acid conditions (Maymo ́-

Gatell et al., 1997). The EPA/ASTSWMO study 

compared concentrations of EDB to pH in the most 

contaminated well at eleven sites (Figure 2.14). 

The concentration of EDB was high at one site that 

was strongly acid (pH of 4), and the concentration 

of EDB was lower in wells with pH near 7. 

However, there were also several sites where the 

concentration of EDB was high and the pH was 

near 7; indicating that pH by itself cannot be used 

to predict the concentration of EDB in old leaded 

motor fuel release sites. 

Figure 2.14.Association of pH with concentrations of EDB in the most contaminated 

monitoring wells at sites. 
22

2.4 Applications of Compound Specific 

Isotope Analysis (CSIA) to 

Document Biodegradation and/or 

Abiotic Transformation of EDB and 

1,2-DCA

This sub-section reviews the application of 

Compound Specific Isotope Analysis (CSIA) 

to evaluate the biodegradation and abiotic 

transformation of EDB and 1,2‐DCA at motor 

fuel release sites. This sub-section is intended for 

technical staff that will interpret CSIA analyses to 

provide an estimate of the extent of biodegradation 

and/or abiotic transformation of EDB or 1,2‐DCA 

in ground water at gasoline release sites. 

Frequently, mathematical models of transport and 

fate are used to provide an exposure assessment 

as part of the risk management process. Because 

EDB and 1,2‐DCA do not sorb strongly to aquifer 

sediments, the transport and fate models are very 

sensitive to the rate of biodegradation and the rate 

of abiotic transformation of the contaminants. It 

is difficult to extract unequivocal estimates of the 

rate of biodegradation or abiotic transformation 

from conventional field monitoring data, and as a 

consequence, there is a large amount of uncertainty 

in the calibration of the transport and fate models. 

In many circumstances, a determination of 

the isotopic ratio of carbon can unequivocally 

document the extent of biodegradation or 

abiotic transformation of EDB or 1,2‐DCA that 

has occurred as the contaminant moves along 

a flow path in ground water. By providing an 

independent estimate of the extent of degradation 

or transformation, CSIA analysis can facilitate the 

calibration of transport and fate models by reducing 

the uncertainty in the estimates of exposure that are 

obtained from the models. 

2.4.1 Theoretical Background for Using 

CSIA to Estimate the Extent of 

Biodegradation

There are two naturally occurring stable isotopes 

of carbon: (1) Carbon-12, which has a weight of 

twelve Daltons, and (2) Carbon-13, which has a 

weight of thirteen Daltons. Carbon-12 (12C) is 

approximately one hundred times more abundant 

than Carbon-13 (13C). During biodegradation or 

abiotic transformation, EDB or 1,2‐DCA molecules 

with two 12C atoms are degraded more rapidly than 

molecules with one 13C atom and one 12C atom, 

which leads to an increase in the ratio of 13C to 12C 

in the residual EDB or 1,2‐DCA. The change in the 

ratio of 13C to 12C is described as enrichment of the 

carbon isotopes during biodegradation or abiotic 

transformation. In recent years, enrichment of 

stable carbon isotopes has been used to recognize 

and describe the natural biodegradation of MTBE 

in ground water (Hunkeler et al., 2001; Gray et al., 

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

Somsamak et al., 2006; Wilson et al., 2005; Zwank 

et al., 2005), and benzene in ground water (Mancini 

et al., 2003; Fischer et al., 2007). 

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 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, as 

described in Equation 2.6 (Meckenstock et al., 

2004; Schmidt et al., 2004).

( ) ( )

( )

13 12 13 12

13 sample standard

13 12

standard

1000

CC CC

C

C C

  −

δ =   ×

 

(2.6)

The units for δ13C are parts per thousand, often 

represented as ‰, or per mil, or per mill. The usual 

pronunciation of δ13C is “delta thirteen sea.” The 

substance used as the international standard for 

stable carbon isotopes has a ratio of 13C to 12C of 

0.0112372. During the course of biodegradation or 

abiotic transformation, the compound that is still 

remaining will have more of the heavy isotope 13C, 

and the value of δ13C will become more positive. 

The extent of isotopic enrichment is typically 

characterized by calculating a linear regression of 

the δ13C in the compound being degraded versus the 

natural logarithm of the fraction of the compound 

remaining (C/Co) after biodegradation or 

transformation (Meckenstock et al., 2004; Schmidt 

et al., 2004). The slope of the regression line is 

termed the isotopic enrichment factor (ε). When 

δ13C is expressed in units of ‰, the unit of ε is ‰.

When an appropriate value of ε is available, and 

when an estimate of the initial value of δ13C in the 

compound as originally released in the gasoline is 

available, the fraction (F) of material remaining 

in ground water after biodegradation or abiotic 

transformation can be calculated from the value of 

δ13C determined for the compound in ground water 

(Meckenstock et al., 2004; Schmidt et al., 2004; 

Wilson et al., 2005). 

( ) 13 13 C C field i 0

F C Co e

    δ −δ ε =   = = (2.7)
23

The fraction of material degraded is simply one 

minus F.

The value of δ13C in EDB or 1,2‐DCA in ground 

water is measured directly in a water sample from 

monitoring wells at the release site (δ13Cfield). 

Equation 2.7 also requires the isotopic enrichment 

factor (ε) and the initial value of δ13C (i.e., at 

t = 0). The following material will review available 

literature on the isotopic enrichment factor for 

biodegradation and abiotic transformation of EDB 

and 1,2‐DCA, followed by a discussion of the 

most plausible values for the initial δ13C of EDB or 

1,2‐DCA originally released in leaded gasoline.

2.4.2 Isotopic Enrichment during 

Biodegradation of EDB and 

1,2‐DCA

At gasoline release sites, the oxygen demand 

associated with petroleum hydrocarbons 

generally consumes all the available dissolved 

oxygen in ground water. As a consequence, 

biological degradation of EDB or 1,2‐DCA in 

close proximity to a spill of gasoline must occur 

through an anaerobic process. At this writing 

(spring 2008), there is only one value of ε for 

anaerobic biodegradation of EDB available in the 

literature (Henderson et al., 2008); the primary 

data are presented in Figure 2.15. The enrichment 

factor was extracted from microcosms that were 

constructed with sediment from a UST release site. 

The enrichment factor, as determined by the slope 

of the regression line, has a value of -5.7 ± 1‰. 

Figure 2.15.Enrichment of the heavy isotope of 

carbon in EDB during anaerobic biodegradation of EDB in a microcosm 

study. (Redrawn from Henderson et al. 

(2008).)

There are only two enrichment factors available 

in the literature for anaerobic biodegradation 

of 1,2‐DCA. Hunkeler et al. (2002) reported a 

value of -32.1 ± 1.1‰ and Hirschorn et al. (2007) 

reported a value of -25.8 ± 3.5 ‰. These values 

of ε for biodegradation of 1,2‐DCA are much more 

negative (show stronger enrichment) than is the 

case for EDB (ε = -5.7 ± 1‰). 

There are several ways to account for the stronger 

enrichment of 1,2‐DCA compared to EDB. 

Table 2.4 compares isotopic enrichment factors 

for biodegradation of a variety of halogenated 

compounds. In general, the value of ε becomes less 

negative as the molecular weight increases. The 

value of ε for anaerobic biodegradation of EDB is 

near the values for biological degradation of PCE 

and TCE, compounds that have a molecular weight 

similar to EDB. 

This relationship between molecular weight and the 

enrichment factor has been explained by Nijenhuis 

et al. (2005) as the effect of a mass transport 

limitation of the compound across the cell wall of 

the bacteria that degrade the contaminant. They 

showed that the enrichment factor for degradation 

of PCE was most negative for the pure PCE 

reductive dehalogenase enzyme, was less negative 

for cell extracts, and was even less negative for 

the living bacteria. The compounds with lower 

molecular weight are more soluble in water, and 

enter the bacterial cells less readily. 

Part of the difference in the value of the 

enrichment factor for anaerobic biodegradation 

of EDB compared to 1,2‐DCA can be attributed 

to differences in the mechanisms of degradation. 

Figure 2.16 compares three pathways that are 

available for anaerobic biodegradation of EDB 

or 1,2‐DCA. Biodegradation of EDB in the 

microcosm study of Henderson et al. (2008) 

was primarily through a mechanism of reductive 

dehalogenation, also called hydrogenolysis, with 

a minor amount of hydrolytic debromination. In 

contrast, in the study of Hunkeler et al. (2002) 

biodegradation of 1,2‐DCA was through a process 

of dichloroelimination. Elsner et al. (2005) showed 

that a dichloroelimination reaction should produce 

more negative values for the enrichment factor than 

a hydrogenolysis reaction. 
24

Table 2.4. Comparison of enrichment factors for EDB and DCA to the range of isotopic enrichment factors 

(ε) for carbon isotopes during reductive dehalogenation of halogenated organic compounds.

Halogenated 

hydrocarbon

Molecular 

Weight 

(Daltons)

Mechanism 13C/12C ε ‰ Reference

EDB 187.9

mostly biological 

hydrogenolysis, some 

hydrolytic debromination

-5.7 Henderson et al., 2008

EDB 187.9 -20.2 This Report

PCE 165.8 biological hydrogenolysis -5.2 Slater et al., 2001

PCE 165.8 biological hydrogenolysis -2 Hunkeler et al., 1999

TCE 131.4 biological hydrogenolysis -6.6

-2.5 Bloom et al., 2000

TCE 131.4 biological hydrogenolysis -7.1 Sherwood Lollar et al., 1999

TCE 131.4 biological hydrogenolysis -13.8 Slater et al., 2001

1,1,2-TCA 133.4 biological 

dichloroelimination -2.0 Hunkeler et al., 2002

cis-DCE 96.9 biological hydrogenolysis -14.1

-16.1 Bloom et al., 2000

cis-DCE 96.9 biological hydrogenolysis -19.9 Hunkeler et al., 2002

cis-DCE 96.9 biological hydrogenolysis -20.4 Slater et al., 2001

trans-DCE 96.9 biological hydrogenolysis -30.3 Hunkeler et al., 2002

1,1-DCE 96.9 biological hydrogenolysis -7.3 Hunkeler et al., 2002

1,2-DCA 98.9 biological 

dichloroelimination -32.1 Hunkeler et al., 2002

1,2-DCA 98.9 biological hydrolytic 

dechlorination -25.8 Hirschorn et al., 2007

VC 62.5 biological hydrogenolysis -21.5

-26.6 Bloom et al., 2000

VC 62.5 biological hydrogenolysis -22.4 Slater et al., 2001

VC 62.5 biological hydrogenolysis -31.1 Hunkeler et al., 2002
25

Figure 2.16.Biological or Abiotic Transformations 

of EDB and 1,2‐DCA.

Hirschorn et al. (2007) compared the enrichment 

factor for biodegradation under aerobic conditions 

to the enrichment factor under nitrate reducing 

conditions. The enrichment factor under aerobic 

conditions was strongly negative (-25.8 ± 0.4‰). 

The values of the enrichment factor under aerobic 

and nitrate reducing conditions were essentially 

identical. Hirschorn et al. (2007) used the 

correspondence in enrichment factors to argue that 

anaerobic biodegradation of 1,2‐DCA under nitrate 

reducing conditions proceeded through the same 

hydrolytic dechlorination reaction that was used 

under aerobic conditions.

A third possible explanation for the difference 

in the enrichment factors during anaerobic 

biodegradation of EDB and 1,2‐DCA would relate 

to the differences between the C-Cl and C-Br 

bond. Miller et al. (2001) compared the enrichment 

factors produced by aerobic biodegradation of 

methyl chloride, methyl bromide, and methyl 

iodide. The enrichment factors were essentially the 

same; the nature of the halogen in the compound 

does not seem to matter. 

2.4.3 Isotopic Enrichment During 

Abiotic Transformation of EDB and 

1,2‐DCA

As of this writing (spring 2008) there is no 

literature available on the isotopic enrichment of 

EDB during abiotic transformation. As described 

in Section 2.3.2, a laboratory study was conducted 

on the transformation of EDB in the presence of 

biogenic FeS. The biogenic FeS was formed in a 

laboratory column that simulated a passive reactive 

barrier constructed with plant mulch and sand. The 

plant mulch supported sulfate reduction to produce 

sulfide, and the sulfide reacted with hematite and 

other iron minerals in the column to produce FeS 

in the column. To determine the kinetics of EDB 

transformation by the biogenic FeS and carbon 

isotope enrichment during transformation, the 

column was sampled and FeS and sand were 

separated from the plant mulch under an oxygen 

free atmosphere and then incubated with a dose 

solution that contained approximately 50 mg/L 

EDB. Incubations were conducted in five separate 

preparations taken from two locations in the 

original column. The δ13C of EDB in the porewater 

was determined after 1 hour of incubation and after 

7, 15, 22, and 29 days of incubation.

The relationship between the extent of removal 

of EDB and the δ13C of EDB is presented in 

Figure 2.17. The isotopic enrichment factor, 

as determined by a linear regression of δ13C of 

EDB on the natural logarithm of the fraction of 

EDB remaining (where the data from all five 

incubations were pooled) was -20.2 ± 2.23‰ at 

95% confidence.

Figure 2.17.A comparison of the enrichment of the 

heavy carbon isotope in EDB during 

anaerobic biodegradation of EDB in 

a microcosm study (Henderson et al., 

2008) against enrichment during abiotic transformation of EDB by FeS. 
26

The enrichment of carbon isotopes during 

abiotic degradation of EDB during reaction 

with FeS (-20.2 ‰) is significantly greater than 

the enrichment during biodegradation (-5.7‰). 

This relationship has been observed for other 

halogenated organic compounds. Liang et al., 

(2007) compared the enrichment of stable carbon 

isotopes during biological and nonbiological 

degradation of PCE and TCE. Enrichment of PCE 

and TCE during transformation by FeS at pH 7 

was -30.2 ± 4.3‰ and -33.4 ± 1.5‰ respectively. 

Enrichment during biodegradation varied with the 

culture, ranging from -1.39‰ to -7.12‰ for PCE 

and from -4.07‰ to -15.27‰ for TCE. In each 

case, the enrichment during abiotic transformation 

was substantially greater than the enrichment 

during biodegradation. 

The contrast in behavior of stable carbon isotopes 

in EDB during biodegradation and abiotic 

transformation parallels the behavior of PCE and 

TCE. If the reaction EDB with FeS followed 

the dihaloelimination pathway (Figure 2.16) 

the enrichment factor should be close to the 

enrichment factor observed by Liang et al. (2007) 

for reaction of TCE and PCE with FeS. According 

to Equation 2.7, as the value δ13C of EDB in ground 

water becomes less negative, the predicted value 

of the fraction remaining (C/Co) is smaller, and the 

predicted extent of biodegradation is greater. For 

a given change in the value of δ13C, the predicted 

extent of biodegradation becomes greater as 

the value of the enrichment factor (ε) becomes 

less negative. Liang et al. (2007) warn that the 

approach used in Equation 2.7 will over predict the 

extent of biodegradation of PCE or TCE if a value 

of ε that is appropriate for biodegradation is used 

to interpret enrichment in ground water samples 

where abiotic processes are primarily responsible 

for the destruction of PCE or TCE. 

The same warning applies to EDB. If a range of 

values of ε are available for the same process, or 

if several processes can operate simultaneously 

and they vary widely in the value of ε, a more 

conservative estimate of the extent of EDB 

destruction is provided by choosing the most 

negative value of the isotopic enrichment factor. 

2.4.4 The Initial Value of δ13C of EDB or 

1,2‐DCA Originally Released in 

Leaded Motor Fuel

To use equation 2.7 to estimate the extent of 

degradation of EDB from the δ13C of EDB in 

ground water, it is necessary to have an initial 

value for δ13C of EDB in the leaded motor fuel 

that was spilled. The analytical techniques for 

CSIA were developed in the late 1990s, while 

EDB and 1,2‐DCA were banned in conventional 

motor gasoline after 1980. As a result, there is no 

direct information available on the δ13C of EDB 

or 1,2‐DCA originally released in leaded motor 

fuel. Five samples of modern leaded motor fuels 

were analyzed for δ13C of EDB. Four samples 

of aviation gasoline (Avgas LL 100 octane) 

returned values of -29.6‰, -30.1‰, -30.2‰, and 

-30.7‰. One sample of 110 octane automobile 

racing fuel returned a value of -30.2‰. In the 

absence of additional information, a value of -30‰ 

is recommended as the best estimate currently 

available of the original value of δ13C in EDB in 

leaded motor fuels. 

Similar processes are used to manufacture 

both EDB and 1,2‐DCA and PCE, TCE, and 

1,1,1‐TCA, and they often were produced at the 

same manufacturing plants. Table 2.5 summarizes 

published values for δ13C of PCE, TCE, and 

1,1,1‐TCA produced by a number of manufactures 

in the past. The most negative value reported was 

-37.2‰, the least negative value reported was 

-23.2‰. 

If the assumed initial value of δ13C in EDB is 

more negative than the true value, the extent 

of degradation predicted by Equation 2.7 will 

overestimate the true extent of degradation. To 

provide a conservative lower boundary on the 

fraction of EDB remaining after degradation, and 

a conservative upper boundary on the fraction 

degraded, it is most appropriate to use the least 

negative value within the plausible range of values 

of δ13C that would be expected in leaded motor 

fuel. In the absence of other direct information, 

a value of -23.2‰ will be taken as conservative 

upper boundary on the value of δ13C in EDB or 

1,2‐DCA in leaded gasoline. 

2.4.5 Measured Concentrations of EDB 

and δ13C of EDB in Ground Water 

If measured concentrations of EDB in ground water 

are related to the destruction of EDB in ground 

water, there should be a relationship between the 

concentration of EDB and the δ13C of EDB. To 

search for a relationship, water samples from 

the most contaminated wells at selected sites in 

the EPA/ASTSWMO study were analyzed for 

concentrations of EDB and for δ13C of EDB. The 

population of candidate sites was limited by two 

factors: (1) access to the sites so that they could be 

resampled, and (2) the relatively high concentration 

of EDB required for determination of δ13C in EDB 

in the sample. A concentration of at least 4 μg/L 

of EDB is required to attain the quality objective 
27

Table 2.5. Range of δ13C in samples of commercial chlorinated solvents. 

Compound Source δ13C (‰/PDB) Reference

PCE Manufacturer A -27.12±0.03 Jendrzejewski et al. (2001)

PCE Manufacturer B -35.27±0.12 Jendrzejewski et al. (2001)

PCE Manufacturer C -24.06±0.08 Jendrzejewski et al. (2001)

PCE Dow -23.19±0.10 van Warmerdam et al. (1995)

PCE ICI -37.20±0.03 van Warmerdam et al. (1995)

PCE PPG -33.84±0.03 van Warmerdam et al. (1995)

PCE Vulcan -24.1±0.04 van Warmerdam et al. (1995)

PCE Range -23.19 to -37.20

TCE Manufacturer A -31.53±0.01 Jendrzejewski et al. (2001)

TCE Manufacturer B -27.90±0.08 Jendrzejewski et al. (2001)

TCE Manufacturer C -29.93±0.18 Jendrzejewski et al. (2001)

TCE Aldrich -33.49±0.08 Jendrzejewski et al. (2001)

TCE Dow -31.90±0.05 van Warmerdam et al. (1995)

TCE ICI -31.32±0.03 van Warmerdam et al. (1995)

TCE PPG -27.80±0.01 van Warmerdam et al. (1995)

TCE Range -27.80 to -33.49

1,1,1-TCA Manufacturer A -31.64±0.09 Jendrzejewski et al. (2001)

1,1,1-TCA Dow -29.42±0.06 van Warmerdam et al. (1995)

1,1,1-TCA ICI -26.64±0.09 van Warmerdam et al. (1995)

1,1,1-TCA PPG -25.80±0.46 van Warmerdam et al. (1995)

1,1,1-TCA Vulcan -28.42±0.07 van Warmerdam et al. (1995)

Range -25.80 to -31.64

Summarized from Wilson (2008).
28

of a sample standard deviation of 0.5‰ or less. 

Five sites met both criteria, one each in Montana, 

Minnesota, Virginia, and two sites in New 

Hampshire. As was expected, the two sites with 

lower concentrations of EDB had higher values for 

δ13C of EDB (Figure 2.18 and Table 2.6).

Figure 2.18.Relationship between the concentrations of EDB in the most contaminated 

wells at each of five sites and the δ13C 

of EDB. (Note that the concentration 

on x-axis decreases from left to right).

Falta (2004) used the expected concentration 

of EDB in leaded gasoline, and Raoult’s Law 

to estimate the concentration of EDB in ground 

water that would be in equilibrium with typical 

leaded gasoline. The predicted concentration 

is 1,900 μg/L. Equation 2.7 was rearranged to 

produce Equation 2.8, and Equation 2.8 was in 

turn used to calculate the value of δ13C of EDB that 

would correspond to various concentrations of EDB 

(Table 2.6).

( ) ( ) 13 13

0 ln . 1900 C C Con field t= field δ = δ −ε∗

(2.8)

Where δ13Cfield is the expected value of δ13C of 

EDB in ground water in units of ‰, Conc.field is 

the expected concentration of EDB in ground water 

in units of μg/L, δ13Ct=0 is the initial value of δ13C 

of EDB in the gasoline (assumed to be -30‰) 

and ε is the appropriate value for the isotopic 

enrichment factors in units of ‰. Equation 2.8 

assumes that destruction of EDB is the only process 

that attenuates the concentration of EDB. There 

is no allowance for attenuation due to dilution in 

the monitoring wells. If there were dilution of 

the plume in the monitoring well, the expected 

concentration of EDB would shift to a lower value. 

In Figure 2.18, a value of -5.7‰ was used 

for ε to predict the relationship dominated by 

biodegradation and a value of -20.2‰ was used 

to predict the relationship dominated by abiotic 

transformation. The data best fit the projections 

of fractionation that would be expected from 

biodegradation of EDB. 

In general, the few data in Figure 2.18 are 

consistent with the assumption that reductions in 

concentration of EDB are caused by processes that 

Table 2.6. Relationship between the concentrations of EDB in the most contaminated wells at five motor 

fuel release sites, enrichment of stable carbon isotopes in EDB, and a conservative estimate of the fraction 

of EDB destroyed.

Location Concentration EDB δ13C Fraction 

Remaining*

Fraction 

Destroyed

μg/L ‰

Station A, New Hampshire 711 -22.8 0.70 0.30

Virginia 557 -13.9 0.45 0.55

Station B, New Hampshire 147 -18.9 0.58 0.42

Montana 8.0 6.1 0.17 0.83

Minnesota 6.8 3.9 0.19 0.81

* Fraction remaining calculated from Equation 2.7 assuming δ13Ct=0 = -30.‰ and ε = -20.2‰
29

destroy EDB, and can be recognized by analysis 

of δ13C in EDB. This suggests that CSIA will be a 

useful tool to recognize biodegradation or abiotic 

transformation of EDB in residual gasoline in the 

source area of plumes. 

2.4.6 Predictions of the Extent of 

Degradation or Transformation

At present there is no straightforward approach 

to attribute one portion of EDB destruction to 

biodegradation and the remaining portion to abiotic 

transformations. As a consequence, the most 

conservative approach is to attribute all of the 

destruction of EDB to abiotic transformation. 

Figure 2.19 presents a prediction of the extent of 

destruction of EDB based on various measured 

values of δ13C in EDB in ground water, and the 

assumption that abiotic processes are entirely 

responsible for destruction of EDB. Figure 

2.9 presents predictions of the fraction of EDB 

remaining that correspond to the best estimate of 

the initial value of δ13C in EDB in the leaded motor 

fuel that was originally spilled (-30‰) and to the 

most conservative estimate of the initial value 

of δ13C in EDB in the leaded motor fuel that was 

originally spilled (-23.2‰). As an example, at the 

best estimate of the initial value of δ13C in EDB 

in gasoline, a measured value of δ13C in EDB in 

ground water of +24‰ would predict that 7% of 

the original quantity of EDB remained. The most 

conservative estimate of the initial δ13C in EDB 

in ground water predicts that 10% of the original 

quantity of EDB remains.

Figure 2.19 was used to provide a conservative 

boundary on the extent of destruction of EDB at 

the five release sites listed in Table 2.6. Even with 

a small data set, CSIA could distinguish sites with 

minor amounts of EDB destruction from sites 

where a major fraction of EDB had been destroyed.

If the data are available, a better approach to 

evaluate EDB destruction at a particular site would 

use the most negative value of δ13C in EDB in any 

monitoring well at a site as the local best estimate 

of δ13C in EDB in gasoline, then use Equation 2.6 

to predict the extent of degradation in the other 

wells. This approach is particularly recommended 

if data are available from wells along a flow path in 

the aquifer that are down gradient of the area with 

known residual gasoline. Figure 2.19 is offered 

to provide a quick and conservative boundary on 

the extent of EDB destruction when the available 

data are few, or when the hydraulic relationships of 

wells are not readily apparent. 

Figure 2.19.A conservative estimate of the relationship between the measured value 

of δ13C of EDB in ground water at a 

Leaded Motor Fuel Release Site, and 

the extent of destruction of EDB by 

abiotic and biological processes. The 

heavy solid line projects destruction 

based on the best available estimate 

of the initial value of δ13C of EDB in 

leaded gasoline. The light solid line 

projects destruction based on the most 

conservative estimate for the initial 

value of δ13C of EDB in leaded gasoline. The diamonds represent estimates 

of the fraction of EDB remaining in 

monitoring wells in the spill sites 

described in Table 2.6. See text for 

further discussion.
30

3.0

Distribution of EDB and 1,2-DCA at Motor 

Fuel Release Sites, and the Associated 

Chance of Contaminating Ground Water

Figure 3.1. Distribution of states participating in the EPA/ASTSWMO Study.

This section describes the existing concentrations 

of EDB, benzene, and 1,2‐DCA in monitoring 

wells at gasoline service stations. It is intended 

for decision makers and other staff of regulatory 

agencies that must apportion resources for 

monitoring and risk management of the hazard 

associated with EDB, benzene, and 1,2‐DCA in 

ground water used for drinking water. It is also 

useful to case workers in regulatory agencies as a 

basis for comparison of the distribution of EDB and 

1,2‐DCA at their sites to concentrations that were 

determined at other sites. 

3.1 Existing Distribution of EDB, 

1,2-DCA, and Benzene in Ground 

Water at Selected Motor Fuel 

Release Sites

Figure 3.1 provides a map showing the states 

that participated in the study. Participation in the 

EPA/ASTSWMO study was entirely voluntary on 

the part of the state agencies. The coverage of the 

USA was reasonably representative; however, the 

states on the Pacific Coast chose not to participate. 

The number of stations that were sampled in each 

state was variable. Two of the states provided 

samples from only one station each, and two of 

the states provided samples from ten stations each 

(Table 1.1).
31

Because the sampling locations were not selected 

randomly, the samples can not be taken to represent 

the population of gasoline service station sites 

in the USA. At best, the samples represent a 

population of service station sites in the USA where 

the relevant case worker in the state agency was 

interested in the possibility that EDB was present 

in ground water at the site. Nevertheless, there is 

a close correspondence between the distribution of 

EDB concentrations in South Carolina, and in the 

sites that were sampled in the EPA/ASTSWMO 

study (Figure 1.2). The total fraction of sites 

that exceeded the MCL was very similar, and 

the distribution of sites with concentrations 

above 20 μg/L was almost identical. Table 3.1 

summarizes the data in Figure 1.2.

As discussed earlier, not all the EDB used in the US 

was used in motor fuel. Approximately 10% was 

used in agriculture. Background wells were not 

sampled at all of the sites that participated in the 

EPA/ASTSWMO study. As a consequence, it is not 

possible to exclude the possibility that a portion of 

the EDB detected in wells at the gasoline service 

stations resulted from agricultural application.

3.2 Relative Distribution of EDB, 

1,2-DCA, and Benzene at Selected 

Motor Fuel Release Sites

There were 39 sites in the EPA/ASTSWMO study 

with detectable concentrations of both EDB and 

benzene. In the 39 sites, there were seven sites 

from Colorado, two sites from Maryland, one site 

from Montana, three sites from North Carolina, 

four sites from North Dakota, two sites from New 

Hampshire, two sites from Oklahoma, five sites 

from South Carolina, one site from South Dakota, 

three sites from Tennessee, two sites from Utah, 

and seven sites from Wisconsin.

Data are summarized in Table 3.2. As was 

expected, benzene was detected above the MCL 

at every site that was sampled. In contrast, EDB 

was detected in 53% of sites, but detected above 

the MCL at only 42% of sites, and 1,2‐DCA was 

detected at 23% of sites, but detected above the 

MCL at only 15% of sites. Both EDB and 1,2‐DCA 

were present at concentrations above the MCL at a 

significant number of sites. 

Table 3.1. The distribution of EDB in the sites included in the EPA/ASTSWMO State Study compared to 

distribution in South Carolina.

Maximum Concentration at Site South Carolina

(Falta data)

EPA/ASTSWMO

State Study

Percent of Sites

Above 100 μg/L 7% 6%

Above 50 μg/L 11% 10%

Above 1 μg/L 35% 25%

Table 3.2. Occurrence of EDB, benzene, and DCA in sites sampled during the EPA/ASTSWMO State Study.

States Sites Samples

Method 8011 for EDB 19 102 802

EDB detected 19 (100%) 54 (53%) 151 (19%)

EDB above MCL 17 (89%) 43 (42%) 103 (13%)

Method 8260 for Benzene 12 39 335

Benzene detected 12 (100%) 39 (100%) 118 (40%)

Benzene above MCL 12 (100%) 39 (100%) 105 (36%)

Method 8260 for DCA 12 39 293

DCA detected 6 (50%) 10 (26%) 18 (6%)

DCA above MCL 5 (42%) 6 (15%) 8 (3%)
32

Figure 3.2 plots the frequency distribution of the 

concentrations of benzene, 1,2‐DCA, and EDB 

independent of each other at a particular site. For 

comparison, the figure also plots the concentration 

that would be predicted for benzene, 1,2‐DCA, 

and EDB in ground water that was in equilibrium 

with un-weathered leaded gasoline. The predicted 

values for concentrations were taken from Table 2 

of Falta (2004). He calculated the predicted values 

based on the average concentrations of benzene, 

1,2‐DCA, and EDB in leaded gasoline, and on the 

gasoline to water partition coefficient in the case 

of EDB and benzene, or based on Raoult’s Law for 

1,2‐DCA. 

Figure 3.2. Distribution of the maximum concentrations of EDB, 1,2‐DCA, and benzene 

in ground water from monitoring wells 

at motor fuel release sites in the EPA/

ASTSWMO study, compared to the 

concentrations predicted for ground 

water in contact with unweathered 

leaded gasoline.

The predicted concentration of benzene in ground 

water that is equilibrated with leaded gasoline is 

ten fold higher than the predicted concentration 

of 1,2‐DCA, and the predicted concentration of 

1,2‐DCA is two fold higher than the predicted 

concentration of EDB. The actual concentrations 

of benzene, 1,2‐DCA, and EDB are less than 

predicted. There are a number of possible 

explanations. Weathering of the residual gasoline 

over time may have reduced the concentrations 

of benzene, 1,2‐DCA, and EDB in the residual 

gasoline. Ground water that was exposed to the 

residual gasoline in the aquifer and contaminated 

with benzene, 1,2‐DCA, and EDB may have been 

diluted with uncontaminated ground water in 

the monitoring well. There may have been mass 

transfer limitations on dissolution of benzene, 

1,2‐DCA, and EDB from the residual gasoline into 

flowing ground water. In any case, the ground 

water samples that were most contaminated with 

EDB or benzene were within a factor of two of the 

predicted concentration. The highest measured 

concentration of 1,2‐DCA was more than 60 fold 

lower than the predicted concentration. 

With the exception of two lowest concentrations 

of benzene in the data set, the concentrations of 

benzene, EDB, and 1,2‐DCA all follow a log-linear 

frequency distribution. There were ten sites that 

reported concentrations of 1,2‐DCA above the 

method detection limit. As a basis of comparison, 

the geometric mean concentration was calculated 

for the ten sites with the highest concentrations 

of benzene, of EDB, and of 1,2‐DCA. On 

average the concentration of benzene was much 

higher than EDB or 1,2‐DCA. The geometric 

mean concentrations for benzene, EDB, and 

1,2‐DCA were 7,300 μg/L, 68 μg/L, and 5.0 μg/L 

respectively. 

The MCL for EDB (0.05 μg/L) is one hundred fold 

lower than the MCL for benzene and 1,2‐DCA 

(5 μg/L). As a result, it can be confusing and 

misleading to evaluate the chance for ground water 

contamination by comparing the concentrations 

in ground water. A better comparison that is 

relevant to beneficial use of an aquifer as a ground 

water supply is the chance that a source of EDB 

or benzene will contaminate a water supply well 

at any concentration above the MCL, making 

it useless for drinking water unless the water is 

treated.

It is possible to estimate a relative chance 

that a water supply well will be impacted at 

concentrations above the MCL if the following 

three assumptions are accepted. First, the chance 

that a plume of contamination will impact a water 

well is proportional to the surface area of the 

plume. Second, the surface area of a plume is 

proportional to the square of its length. Third, the 

length of the plume is proportional to the logarithm 

of the maximum concentration of the contaminant. 

Studies of the distribution of contaminants in 

ground water indicate that concentrations attenuate 

with distance following a pseudo first order rate 

law. As a consequence, the length of the plume 

is proportional to the logarithm of the ratio of the 

maximum concentration to the MCL. If the MCL 

is 0.05 μg/L and the maximum concentration is 

0.5 μg/L, the length would be some distance X. If 

the maximum concentration is 5 μg/L, the length 

would be 2X, if the maximum concentration is 

50 μg/L, the length would be 3X, if the maximum 
33

concentration is 500 μg/L, the length would be 4X 

and so on. 

Ignoring the particular hydrogeological context and 

cultural context of a release, the relative chance that 

a ground water supply well will be contaminated 

at concentrations above the MCL is described by 

Equation 3.1.

Relative Chance = (Ln[MaximumConc./MCL])2

(3.1)

Figure 3.3 plots the frequency distribution of the 

relative chance that EDB, benzene, and 1,2‐DCA 

from fuel spill sites in the EPA/ASTSWMO Study 

will contaminate a water supply well. The data are 

ranked independently of each other at a particular 

site. For comparison, the figure also plots the 

relative chance that would be predicted for EDB, 

benzene, and 1,2‐DCA in ground water that was 

in equilibrium with un-weathered leaded gasoline. 

The predicted chance that EDB in ground water 

that is equilibrated with leaded gasoline will exceed 

the MCL is greater than the predicted chance for 

benzene, which in turn is greater than the predicted 

chance for 1,2‐DCA. The greatest chance to 

exceed the MCL for a particular site and particular 

contaminant was associated with EDB. 

Figure 3.3. Distribution of the relative chance that 

EDB, 1,2‐DCA, and benzene in ground 

water at motor fuel release sites in the 

EPA/ASTSWMO Study will impact 

a water supply well at concentrations 

above the MCL, compared to the relative chance predicted for ground water 

in contact with unweathered leaded 

gasoline. The relative chance is proportional to the surface area of the plume 

of EDB or benzene, or 1,2‐DCA with 

concentrations above the MCL. See 

text for details.

Notice from Figure 3.3 that for approximately 

20% of the sites in the EPA/ASTSWMO study, the 

relative chance that EDB would impact a water 

supply well was roughly equivalent to the chance 

that benzene would impact a water supply well. 

However, for the remaining 80% of sites, the chance 

of EDB impacting a water supply well was much 

less than the chance that benzene would impact a 

well. Also notice that in the worst case, the chance 

that 1,2‐DCA would impact a water supply well was 

roughly 10% of the worst case chance for EDB or 

benzene, and that the relative chance for 1,2‐DCA 

fell off rapidly as a percent of the sites in the survey.

Although the relative chance that EDB or benzene 

would impact a water supply well is roughly the 

same in 20% of the sites in the EPA/ASTSWMO 

study, the 20% of sites are not the same sites 

for EDB and for benzene. Figure 3.4 compares 

the chance associated with the highest reported 

concentration of benzene in any well at a particular 

site to the chance associated with the highest 

reported concentration of EDB in any well at the 

same site. At 10 of 39 sites, the chance associated 

with EDB was greater, and the chance associated 

with benzene was greater at the remaining 29 sites. 

EDB was the risk driver for impact to water supply 

wells at approximately one fourth of the sites.

Figure 3.4. Comparison of the relative chance that 

EDB will exceed the MCL in a water supply well to the relative chance 

that benzene will exceed the MCL in 

a water supply well for ground water 

from motor fuel release sites in the 

EPA/ASTSWMO Study. 

The total chance that EDB, 1,2‐DCA, or benzene 

at the 39 sites in the EPA/ASTSWMO study might 
34

contaminate a water supply well above the MCL 

can be estimated by summing the individual 

relative chances for each site. The sum over 

the 39 sites for EDB is 1100, for benzene is 580 

and for 1,2‐DCA is 10. In aggregate, in the sites 

sampled in the EPA/ASTSWMO study, the chances 

of contaminating a water supply well with EDB 

or benzene at concentrations above the MCL are 

roughly equivalent, and the chance for 1,2‐DCA is 

orders of magnitude lower. These estimates can be 

compared to data released by the U.S. Geological 

Survey from their comprehensive survey of the 

occurrence of volatile organic compounds in 

ground water and drinking-water supply wells in 

the USA (Zogorski et al., 2006, see Appendix 7). 

They sampled ground water supplying domestic 

and public wells; EDB was not detected in 

462 samples of ground water supplying public 

water supply wells. However, EDB was detected 

in 3 of 2,085 samples of ground water supplying 

domestic water supply wells. All three detections 

were above the MCL for EDB of 0.05 μg/L. In 

comparison, benzene was detected in five of 

1,095 samples of ground water supplying public 

water supply wells, and in five of 2,401 samples 

of ground water supplying domestic water supply 

wells, but in each case the concentration was less 

than the MCL of 5 μg/L. In a sampling effort that 

was approximately equivalent, three samples were 

contaminated with EDB above the MCL and no 

sample was contaminated with benzene above the 

MCL. 

In summary, there were significant concentrations 

of EDB and 1,2‐DCA at the sites that were 

sampled during the EPA/ASTSWMO study. 

As a consequence, EDB was the risk driver at 

approximately one quarter of the sites in the study.

3.3 The Chance that Conventional 

Monitoring Using EPA Method 8260 

will fail to detect EDB

Method 8260B prepares a water sample for 

analysis by purge-and-trap, separates the volatile 

organic compounds by gas chromatography, 

and determines the chemical identity and the 

concentration of volatile organic compounds by 

mass spectrometry. It is widely applied for analysis 

of the petroleum components of gasoline in ground 

water. Method 8260B can also determine EDB or 

1,2‐DCA, however the effective method detection 

limit for EDB and 1,2‐DCA is near 3 μg/L. This 

detection limit is problematic for interpretation of 

concentrations of EDB because it is almost one 

hundred fold higher than the MCL. The detection 

limit is not a problem for determination of 

1,2‐DCA, which has an MCL of 5 μg/L. 

Method 8011 prepares the water sample by microextraction, separates volatile organic compounds 

by gas chromatography, and determines the 

concentration of the volatile organic compounds by 

the response of an electron capture detector. This 

detector is exquisitely sensitive to halogenated 

compounds, but responds hardly at all to 

conventional hydrocarbons. The detection limit for 

EDB is 0.01 μg/L, which is a reasonable margin 

lower than the MCL of 0.05 μg/L.

Analysis of EDB can be included in the routine 

analysis of BTEX compounds by Method 8260B 

at minimal extra cost. In contrast, monitoring 

for EDB by 8011 and monitoring for the BTEX 

compounds by Method 8260B can essentially 

double the total cost of analysis. The selection of 

one method over the other depends on the goals and 

priority in risk management. 

Figure 3.5 presents the practical consequence 

of the difference in detection limits using 

Method 8260B or Method 8011. It compares the 

distribution of EDB that was determined in the 

EPA/ASTSWMO survey to the distribution that 

would have been discovered if the survey had been 

conducted using Method 8260B. Method 8260B 

would have discovered only 40% of the sites with 

concentrations of EDB above the MCL. 

Figure 3.5. Distribution of concentrations of EDB 

at the leaded gasoline spill sites in the 

EPA/ASTSWMO Study where the concentrations were above the MCL. Note 

that the concentrations of EDB in only 

40% of the sites were above the method 

detection limit for EPA Method 8260B.
35

3.4 Co-Distribution of EDB with 

1,2-DCA, Benzene, Xylenes, and 

Ethylbenzene at Selected Motor 

Fuel Release Sites 

At many gasoline release sites no ground water 

samples have been analyzed for EDB or 1,2‐DCA, 

but analyses for BTEX compounds are common. 

Is there an association between concentrations 

of any of the BTEX compounds and EDB and 

1,2‐DCA that could be used to predict whether 

EDB and 1,2‐DCA might be present and at 

what concentrations? Using the idea that a cooccurring compound of gasoline might serve as 

a conservative tracer, Wiedemeier et al., (1996) 

compared the concentrations of benzene in a 

release of JP-4 jet fuel to concentrations of the 

xylenes. The xylenes and ethylbenzene are a 

major fraction of gasoline, they tend to partition 

from gasoline slowly, and they are resistant to 

biodegradation under anaerobic conditions. In 

a similar manner Falta (2004) used the average 

composition of gasoline and partitioning theory 

to estimate the concentration of EDB, 1,2‐DCA, 

benzene, ethylbenzene, and total xylenes that 

would be expected in ground water in contact 

with un-weathered leaded gasoline. The predicted 

concentrations for EDB, 1,2‐DCA, benzene, 

ethylbenzene, and xylenes were 1,900 mg/L, 

3,700 mg/L, 37,100 mg/L, 3,000 mg/L, and 

13,100 mg/L respectively. 

Figure 3.6 compares the predicted concentrations 

in contact with un-weathered gasoline, and the 

concentrations that would be predicted from 

dilution, to the measured concentrations of EDB, 

ethylbenzene, and total xylenes in wells sampled 

during the EPA/ASTSWMO study. The method 

detection limit for EDB, ethylbenzene, and 

individual xylenes was 0.005, 0.05, and 0.15 μg/L 

respectively, and values less than the method 

detection limit are plotted at the method detection 

limit. 

The dotted line in the panels of Figure 3.6 are the 

predicted concentrations of EDB and ethylbenzene 

(Panel A) or EDB and total xylenes (Panel B) 

that would be expected in water based on the 

assumptions and calculations of Falta (2004). If 

a datum falls above the line, there is more EDB in 

the ground water than would be predicted from the 

concentration of total xylenes or ethylbenzene. 

Figure 3.6. Association of concentrations of EDB 

with concentrations of ethylbenzene 

(Panel A) or total xylenes (Panel B).

The maximum concentrations of ethylbenzene 

and total xylenes were close to the concentrations 

that should be in contact in leaded gasoline. For 

any particular concentration of total xylenes or 

ethylbenzene, the highest measured concentrations 

of EDB were in good agreement with the 

concentrations that would be predicted from the 

calculations of Falta (2004). The only exceptions 

were for concentrations of ethylbenzene and total 

xylenes below 1 μg/L. The very low concentrations 

of ethylbenzene and total xylenes may have been 

influenced by biodegradation or by retardation due 

to differential sorption of ethylbenzene or total 

xylenes compared to EDB.

Most of the measured concentrations of EDB 

were many fold lower than the concentrations that 

would be predicted from measured concentrations 

of total xylenes or ethylbenzene. For a release of 

leaded gasoline, at concentrations in the range of 

1 μg/L or higher of ethylbenzene and 0.1 μg/L of 

total xylenes, the dotted lines in Figure 3.6 and the 

measured concentration of ethylbenzene or total 

xylenes can be used to estimate a conservative 

upper boundary on the concentration of EDB. 

These estimates could be used to conduct a 
36

preliminary RBCA assessment for EDB if data on 

the concentrations of EDB are not available. 

Figure 3.7 compares concentrations of EDB to 

concentrations of 1,2‐DCA and benzene wells 

sampled under the EPA/ASTSWMO study. The 

method detection limit for EDB, 1,2‐DCA, 

and benzene was 0.005, 0.13, and 0.07 μg/L 

respectively, and values less than the method 

detection limit are plotted at the method detection 

limit. 

Figure 3.7. Association of concentrations of EDB 

with concentrations of 1,2‐DCA (Panel 

A) and benzene (Panel B).

When EDB was measured in the ground water, 

the concentration of 1,2‐DCA was low or below 

the detection limit. When 1,2‐DCA was measured 

in the ground water, the concentration of EDB 

was low or below the detection limit. We have 

no explanation for the strong negative association 

of concentrations of EDB and 1,2‐DCA. The 

maximum concentration of EDB was near the 

maximum that would be predicted for ground water 

in contact with residual leaded gasoline (based 

on calculations of Falta, 2004). The maximum 

concentration of 1,2‐DCA was almost two orders 

of magnitude lower than would be predicted, 

suggesting extensive degradation of 1,2‐DCA 

compared to degradation of EDB. In any case, 

concentrations of 1,2‐DCA cannot be used to 

predict concentrations of EDB.

The maximum concentrations of benzene were 

near the concentrations that would be predicted 

for ground water in contact with leaded gasoline. 

There were several samples clustered at a benzene 

concentration near 400 μg/L where the EDB 

concentration was up to an order of magnitude 

higher than would be predicted from the 

concentration of benzene. This probably reflects 

preferential degradation of benzene compared 

to EDB. There was a second cluster of samples 

at benzene concentrations less than 1 μg/L that 

indicated a preferential degradation of benzene 

compared to EDB. As a consequence, benzene 

concentrations can not be used to calculate 

a conservative upper boundary on predicted 

concentrations of EDB.

3.5 Local Vulnerability of Exposure 

to EDB based on Past Usage of 

Leaded Motor Fuel

This section estimates the chance that drinking 

water will be contaminated with EDB from 

gasoline that was released from a leaking 

underground storage tank. In particular, it 

estimates on a state wide basis the relative chance 

that a resident of one of the forty eight contiguous 

states will be affected. This section presents maps 

relating the regions with high densities of residents 

that drink water from shallow wells to regions with 

high historical use of leaded gasoline. This section 

is intended for decision makers and other staff of 

regulator agencies that must apportion resources 

for monitoring and risk management of the hazard 

associated with EDB in ground water used for 

drinking water. 

In 2006, the U.S. Geological Survey released a 

comprehensive study of the occurrence of volatile 

organic compounds in ground water and drinkingwater supply wells in the USA (Zogorski et al., 

2006). They sampled 462 public water supply 

wells for EDB and EDB was not detected in any of 

them; however, EDB was detected in three of 2,085 

domestic water supply wells that were sampled. 

All three detections were above the MCL for EDB 

of 0.05 μg/L. In comparison, benzene was detected 

nine times in 997 public water supply wells, but in 

each case the concentration was less than the MCL 

of 5 μg/L. Based on the findings of Zogorski et al., 

(2006), the most likely scenario for contamination 

of drinking water by EDB from leaded motor fuel 

is contamination of private domestic water wells in 

a suburban residential landscape.
37

The chance that an individual will consume 

drinking water that is contaminated with EDB from 

a leaking underground storage tank is influenced 

by a number of site-specific factors including 

the source of drinking water, the proximity of 

underground storage tanks, and the amount of 

leaded gasoline that was stored in the tanks and 

might have leaked to contaminate ground water. 

It is impossible to calculate a precise estimate of 

the chances for exposure from data that are readily 

available. The following approach was taken to 

provide a broad estimate of the local vulnerability 

of exposure to EDB based on past usage of leaded 

gasoline in the state. 

The first task was to access data on the content 

of lead in leaded gasoline. Beginning in the 

1930s, the National Institute for Petroleum and 

Energy Research (NIPER) has collected data on 

the composition of gasoline. The successor to 

NIPER (Northup-Grumman) continues to analyze 

approximately 1,000 gasoline samples taken 

twice per year. Data are available from 174 areas 

around the USA; however, the record is often 

incomplete. About 35 locations have been sampled 

on a continuous basis. This data base is the most 

extensive historical record (that is publically 

available) of the composition of conventional 

and reformulated gasoline in the USA. The data 

base provided concentrations of lead in gasoline 

from 1976 through 1995 (approximately 27,800 

samples), collected at locations within 101 cities 

representing 42 states. The weighted average lead 

concentrations (gram per liter) of all octane levels 

were calculated for five-year categories as follows: 

1976-1980, 1981-1985, 1986-1990, and 1991-

1995. The weighted averages for these categories 

are 0.43 g/L, 0.15 g/L, 0.038 g/L, and 0.015 g/L 

respectively. These averages are in good agreement 

with estimates provided in Figure 2 of Falta et al., 

(2005). Estimates provided in Figure 2 of Falta 

(2004) were applied as the estimate average lead 

concentrations for the 5-year categories 1951-1955, 

1956-1960, 1961-1965, 1966-1970, and 1971-1975. 

The second task was to estimate the amount 

of leaded gasoline sold in the USA. The U.S. 

Department of Transportation (2008) has published 

data on the volume of gasoline sold in each state in 

each year between 1949 and 1995. To estimate the 

amount of lead in gasoline sold in each state in each 

year category, the volume of gasoline sold in each 

state for all the individual years within each year 

category was summed, and the resulting sums were 

multiplied by the average lead concentration for 

each respective year category. 

The third task was to convert the amount of lead to 

the amount of EDB. In the interval 1951 to 1995, 

the molar ratio of lead and EDB in gasoline has 

been one to one (Falta, 2004). From stoichiometry, 

there was 0.435 gram of EDB for each gram of lead 

in leaded motor fuel sold in the USA in the period 

of interest. To estimate the amount of EDB in 

gasoline sold in each state in each year category, the 

amount of lead was multiplied by 0.453. Figure 3.8 

compares the total amount of EDB in gasoline sold 

in five states with small populations, and Figure 3.9 

compares EDB in gasoline sold in states with large 

populations. Notice the difference in the scale of 

the axis showing EDB sold in gasoline between 

the figures. In general, sales of EDB in gasoline 

increased more than two fold between 1951-1960 

and 1976-1980, and then declined substantially by 

1991-1995.

Figure 3.8. Consumption of EDB in leaded motor 

fuel in five small states in the USA in 

the forty-five years before lead was 

banned in automotive motor fuel.

Figure 3.9. Consumption of EDB in leaded gasoline in five large states in the USA in 

the forty-five years before lead was 

banned in motor gasoline.
38

Table 3.3. Exposure to Drinking Ground Water Contaminated with EDB from Leaded Motor Fuel.

State Total EDB Sold in 

Motor Fuel

EDB Sold in 

Motor Fuel

Population Drinking 

Water from Shallow 

Wells

Relative Exposure 

Index*

(kg per State) (kg per hectare) Percent of Total 

Population

Connecticut 27,528,745 21.35 8.4 179

New Jersey 66,527,161 34.21 3.8 130

Rhode Island 8,009,494 29.59 4.2 124

Delaware 5,856,848 11.00 8.7 96

Maryland 37,661,911 14.93 6.0 90

Massachusetts 50,397,289 9.06 3.2 77

Pennsylvania 106,478,053 23.80 8.5 76

Michigan 93,618,641 6.24 11.2 70

Ohio 107,154,665 10.04 6.7 67

Indiana 57,014,485 6.05 9.4 57

North Carolina 59,481,984 4.68 11.8 55

New Hampshire 7,865,305 3.28 15.9 52

New York 125,731,461 9.99 4.7 47

Vermont 5,011,479 2.01 21.9 44

Florida 125,947,001 8.71 5.0 44

Virginia 52,072,141 5.05 8.3 42

Wisconsin 45,375,403 3.12 12.6 39

West Virginia 18,960,163 3.02 12.0 36

South Carolina 31,454,595 3.93 8.1 32

Maine 11,364,807 1.36 21.3 29

The estimated total amount of EDB in leaded 

gasoline sold in each state in the period 1951-

1995 is presented in Table 3.3. Almost all of the 

leaded gasoline was stored in underground storage 

tanks prior to sale. The amount of EDB in leaded 

gasoline sold in a state will be taken as a predictor 

of the amount of EDB released from releases of 

leaded gasoline from underground storage tanks. 

The next task is to estimate the probability of 

encountering a release of leaded gasoline in an 

aquifer, expressed as the amount of leaded gasoline 

released per unit surface area. One obvious 

approach would be to normalize the amount 

of EDB sold by the surface area of each state. 

Table 3.3 lists, by state, the total amount of EDB 

sold in gasoline in that state divided by the surface 

area of the state. This is intended as an estimate 

of the chance that ground water under a suburban 

residential landscape has been contaminated by 

EDB from a release of leaded gasoline. The 

relative exposure of an individual in a particular 

state consuming ground water with EDB from 

leaded gasoline is estimated by dividing the total 

amount of EDB in gasoline sold in the state by the 

surface area of the state, and then by multiplying by 

the fraction of people in the state that drink ground 

water from shallow wells, as tabulated in the 

2000 census. The estimate is termed the Relative 

Exposure Index. 

In Table 3.3, the states are ranked by the value 

of the Relative Exposure Index. The older 

industrialized states on the Atlantic Coast and 

the Great Lakes area rank particularly high. The 

states in the mid-continent and in the arid western 

USA rank particularly low. Figure 3.10 compares 

the amount of EDB sold in gasoline divided by 

the surface area of the state to regions where 

shallow ground water is used for drinking water. 

Figure 3.11 provides a closer view of the New 

England and Mid-Atlantic states where the Relative 

Exposure Index is particularly high.
39

State Total EDB Sold in 

Motor Fuel

EDB Sold in 

Motor Fuel

Population Drinking 

Water from Shallow 

Wells

Relative Exposure 

Index*

(kg per State) (kg per hectare) Percent of Total 

Population

Illinois 104,486,891 7.16 3.5 25

Kentucky 37,360,830 3.58 6.9 25

Tennessee 48,344,985 4.43 5.0 22

Georgia 57,010,862 3.75 5.9 22

Missouri 55,806,135 3.08 6.3 19

Minnesota 41,275,045 1.89 10.0 19

Iowa 33,112,275 2.27 7.3 17

Alabama 41,607,861 3.11 4.9 15

Louisiana 40,237,026 3.39 4.2 14

Arkansas 24,123,265 1.76 6.9 12

Oklahoma 34,808,511 1.92 5.3 10

Washington 35,527,773 2.04 4.8 10

Mississippi 25,868,121 2.10 4.5 9.4

California 201,242,695 4.92 1.5 7.4

Oregon 24,419,926 0.97 6.3 6.1

Texas 149,485,144 2.18 2.8 6.1

Nebraska 17,431,282 0.87 6.4 5.6

Kansas 27,511,908 1.29 4.1 5.3

Idaho 9,045,453 0.42 9.3 3.9

North Dakota 7,622,837 0.42 9.0 3.7

Montana 9,297,566 0.24 13.6 3.3

South Dakota 8,589,734 0.43 7.2 3.1

Colorado 25,159,205 0.93 3.1 2.9

New Mexico 14,068,219 0.45 5.8 2.6

Wyoming 6,108,986 0.24 9.0 2.2

Arizona 18,415,405 0.63 1.7 1.1

Utah 11,781,586 0.54 1.1 0.59

Nevada 5,045,588 0.18 1.9 0.33

Alaska 5,265,059

District of 

Columbia

6,909,170 0

Hawaii 2,387,037

* The Relative Exposure Index is calculated as the product of total EDB sold in gasoline in the state (kg), divided by 

the surface area of the state (hectares), and then multiplied by the fraction of the people in the state that drink ground 

water from shallow wells.
40

Figure 3.10. Relationship between heavy use of leaded motor fuel and the use of shallow ground water for 

public water supplies in the contiguous USA.

Figure 3.11. Relationship between heavy use of leaded motor fuel and the use of shallow ground water for 

public water supplies in New England, the Mid-Atlantic States, and the Great Lakes region of 

the USA.
41

Table 3.4. Vulnerability to Drinking Ground Water Contaminated with EDB from Leaded Motor Fuel. 

State Total EDB Sold in 

Motor Fuel

EDB Sold in 

Motor Fuel

Population Drinking Water 

from Shallow Wells

Relative 

Vulnerability 

Index*

(kg per State) (kg per resident) Percent of Total 

Population

Vermont 5,011,479 8.23 21.9 181

Maine 11,364,807 8.95 21.3 191

New Hampshire 7,865,305 6.34 15.9 101

Montana 9,297,566 10.31 13.6 140

Wisconsin 45,375,403 8.47 12.6 107

West Virginia 18,960,163 10.48 12.0 126

North Carolina 59,481,984 7.39 11.8 87

Michigan 93,618,641 9.42 11.2 105

Minnesota 41,275,045 8.39 10.0 84

Indiana 57,014,485 9.38 9.4 88

Idaho 9,045,453 7.01 9.3 65

North Dakota 7,622,837 11.87 9.0 107

Wyoming 6,108,986 12.37 9.0 112

Delaware 5,856,848 7.47 8.7 65

Pennsylvania 106,478,053 8.66 8.5 73

Connecticut 27,528,745 8.07 8.4 68

The Relative Exposure Index takes no account of 

the fact that the states in the USA vary widely in 

the use and extent of development of the landscape. 

Within a state, some areas are urbanized, some 

areas are largely suburban residential land, some 

areas are farmland or pasture, and some areas are 

military reservations or parkland. As an alternative 

to normalizing the total amount of EDB sold 

in gasoline by the surface area of the state, the 

calculation presented in Table 3.4 normalizes the 

total amount sold to the population of the state in 

the 2000 census. 

The total amount of EDB sold in gasoline in a 

state divided by the population of the state in 

the 2000 census is intended as an estimate of 

the chance that ground water under a suburban 

residential landscape has been contaminated by 

EDB from a release of leaded gasoline. The 

relative vulnerability of an individual in a particular 

state consuming ground water with EDB from 

leaded gasoline is estimated by dividing the total 

amount of EDB in gasoline sold in the state by the 

population, and then by multiplying by the fraction 

of people in the state that drink ground water from 

shallow wells, as tabulated in the 2000 census. The 

estimate is termed the Relative Vulnerability Index 

in Table 3.4.

There is a reasonably close correlation between the 

population of a state and the volume of gasoline 

sold. Figure 3.12 compares the liters of gasoline 

sold in states (in 1995, the latest data available) to 

the population (in 2000, the latest data available). 

The ratio varies by less than a factor of two 

between the extremes. Gasoline service stations 

are clustered near major highways, shopping areas, 

and suburban residential areas, which is also where 

drinking water sources are located. The current 

population of a state should serve as a proxy for the 

surface area of the state that contains both gasoline 

service stations and suburban residential areas 

where the inhabitants might have shallow private 

wells.

The Vulnerability Index varies by a factor of 30 

across all the states in the USA. It is most sensitive 

to the fraction of the population that drinks ground 

water from shallow wells. The rankings change for 

some of the states between the Relative Exposure 

Index and the Relative Vulnerability Index. 

However, the general trends are the same. The 

older industrialized states on the Atlantic Coast and 

the Great Lakes area rank particularly high. The 

states in the mid-continent rank lower and the states 

in the arid western USA rank particularly low. 
42

State Total EDB Sold in 

Motor Fuel

EDB Sold in 

Motor Fuel

Population Drinking Water 

from Shallow Wells

Relative 

Vulnerability 

Index*

(kg per State) (kg per resident) Percent of Total 

Population

Virginia 52,072,141 7.35 8.3 61

South Carolina 31,454,595 7.84 8.1 64

Iowa 33,112,275 11.30 7.3 82

South Dakota 8,589,734 11.38 7.2 82

Kentucky 37,360,830 9.25 6.9 63

Arkansas 24,123,265 9.03 6.9 62

Ohio 107,154,665 9.40 6.7 63

Nebraska 17,431,282 10.19 6.4 65

Missouri 55,806,135 9.97 6.3 63

Oregon 24,419,926 7.14 6.3 45

Maryland 37,661,911 7.11 6.0 43

Georgia 57,010,862 6.96 5.9 41

New Mexico 14,068,219 7.73 5.8 45

Oklahoma 34,808,511 10.09 5.3 54

Tennessee 48,344,985 8.50 5.0 43

Florida 125,947,001 7.87 5.0 40

Alabama 41,607,861 9.35 4.9 46

Washington 35,527,773 6.03 4.8 29

New York 125,731,461 6.62 4.7 31

Mississippi 25,868,121 9.11 4.5 41

Louisiana 40,237,026 9.00 4.2 38

Rhode Island 8,009,494 7.63 4.2 32

Kansas 27,511,908 10.23 4.1 42

New Jersey 66,527,161 7.91 3.8 30

Illinois 104,486,891 8.43 3.5 30

Massachusetts 50,397,289 7.94 3.2 26

Colorado 25,159,205 5.85 3.1 18

Texas 149,485,144 7.15 2.8 20

Nevada 5,045,588 2.52 1.9 5

Arizona 18,415,405 3.59 1.7 6

California 201,242,695 5.94 1.5 9

Utah 11,781,586 5.28 1.1 6

Alaska 5,265,059 8.40

District of 

Columbia

6,909,170 12.08 0

Hawaii 2,387,037 1.97

*The Relative Vulnerability Index is calculated as the product of total EDB sold in gasoline in the state (kg) in the 

period 1951 through 1965, divided by the population of the state (2000 census data), and then multiplied by the 

fraction of the people in the state that drink ground water from shallow wells (1990 census data).
43

Figure 3.12. Relationship between the population 

of a state in the USA (2000 census) and 

the sales of gasoline (1995 data).
44

4.0

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Appendix A.

Partitioning of EDB and 1,2-DCA between 

Gasoline and Water

When leaded gasoline is spilled into ground water, 

the EDB and 1,2‐DCA in the residual gasoline 

diffuses out of the gasoline into the adjacent 

ground water until the concentrations of EDB 

and 1,2‐DCA come to equilibrium. As EDB and 

1,2‐DCA are transferred from gasoline to water, the 

concentrations remaining in gasoline decline. As a 

result, the final concentration of EDB or 1,2‐DCA 

in ground water depends on the ratio of gasoline 

to ground water in the aquifer, as described in 

Equation A.1.

,

,

original gasoline

water

water

gasoline water

gasoline

C

C

K

= θ

+

θ

 

(A.1)

Where Cwater is the final concentration in ground 

water, Coriginal, gasoline was the original concentration 

in the gasoline that was released, θgasoline is the 

porosity filled with residual gasoline (expressed 

as the volume of gasoline divided by the total 

volume of aquifer), θwater is the water-filled porosity 

(expressed as the volume of water divided by the 

total volume of aquifer), and Kgasoline,water is the 

distribution coefficient between gasoline and water 

(expressed as the concentration in gasoline divided 

by the concentration in water).

Equation A.1 was provided by personal 

communication from Dr. William Rixey at the 

University of Houston, see also Rixey and Joshi 

(2000). 

Falta (2004) provided estimates of K gasoline,water for 

EDB and 1,2‐DCA of 152 and 84 respectively. 

Typical original concentrations of EDB and 

1,2‐DCA in leaded gasoline were near 290 mg/L 

and 310 mg/L (Falta, 2004). At these estimated 

values, Equation 2.9 predicts maximum possible 

concentrations of EDB and 1,2‐DCA of 1,900 μg/L 

and 3,700 μg/L respectively. 

Figure A.1 extrapolates the predictions of Equation 

A.1 for a range of concentrations of un-weathered 

residual leaded gasoline. Calculations assume 

a total porosity of 0.3. The value of θgasoline was 

calculated from the value of Total Petroleum 

Hydrocarbons (mg/kg) by multiplying by the 

bulk density of aquifer sediment at a porosity of 

0.3 (1.855 g/cm3) and dividing by the density of 

gasoline (0.78 g/cm3). The value of θwater was 

calculated by subtracting θgasoline from 0.3.

Figure A.1. Predicted maximum groundwater 

concentrations of EDB and DCA for a 

range of possible concentrations of unweathered residual gasoline in aquifer 

sediment.

Typical values for the concentration of residual 

gasoline in unconsolidated sediments vary from 

2,000 to 10,000 mg/kg. The corresponding 

predicted concentrations of EDB under typical 

conditions vary from 1,100 μg/L to 1,400 μg/L and 

corresponding concentrations of 1, 2-DCA vary 

from 1,700 μg/L to 2,500 μg/L.

As EDB and 1,2‐DCA partition from residual 

gasoline to ground water, and as the moving ground 

water carries EDB and 1,2‐DCA away from the 
49

residual gasoline, the concentrations of EDB and 

1,2‐DCA in ground water will decline over time. 

The rate of this weathering process is controlled 

by the distribution of EDB and 1,2‐DCA between 

residual gasoline and ground water, and by the 

seepage velocity of the ground water.

Figure A.2 estimates the fraction of EDB and 

1,2‐DCA that is dissolved in ground water, and 

can be flushed away from the residual gasoline 

by the flow of ground water, as a function of the 

concentration of residual gasoline. For a given 

value of TPH (mg/kg), values of θgasoline and θwater

were calculated as described above, and Cwater was 

calculated using Equation A.1. An estimate of the 

concentration of the contaminant in the residual 

gasoline (Cgasoline) was calculated by multiplying 

Cwater by K gasoline,water . Then Equation A.2 was used 

to calculate the fraction of contaminant in ground 

water. 

( ) , water water

water water gasoline gasoline

C Fraction water

C C

∗θ = ∗θ + ∗θ

(A.2) 

At typical values of TPH between 2,000 and 10,000 

mg/kg, the fraction of EDB in ground water would 

vary between 30% and 7%, and the fraction of 

1,2‐DCA in ground water would vary from 40% 

and 12%. With each exchange of the pore water 

in contact with residual gasoline, roughly 12% of 

EDB and 20% of 1,2‐DCA would be flushed away 

from the source area. Because a relatively small 

proportion of the EDB or 1,2‐DCA is in the ground 

water, these contaminants would be expected to 

weather slowly from residual gasoline. 

 

Figure A.2. Expected relationship between the 

fraction of EDB or DCA that is dissolved in ground water, and can be 

flushed away by ground water flow, and 

the concentration of residual leaded 

motor fuel.
50

Appendix B

Materials and Methods for Laboratory Studies 

of Abiotic Degradation of EDB and 1,2-DCA 

Table B.1: Distribution of iron, AVS, and CRS along the Column with Mulch and Hematite 

as described in Shen and Wilson (2007). 

Section Distance 

from inlet H2

O Fe fractionation S fractionation

HCl* total Fe** AVS CRS

cm Volume % (mg/kg) (mg/kg) (mg/kg) (mg/kg)

s1 0-5 41.36 2216±183 3710±0 1020±100 2774±625

s2 5-10 40.12 4926±291 5095±1110 2642±371 104±219

s3 10-15 30.63 4303±936 5610±268 2087±101 530±254

s4 15-20 36.40 4612±1181 5110±282 2158±187 2120±370

s5 20-25 32.45 3645±925 6475±855 1630±74 987±275

s6 25-30 33.25 2967±123 4610±311 1449±153 1517±64

s7 30-35 30.79 2897±297 6755±7 1295±229 1793±463

s8 35-40 31.55 3181±510 5080±0 1190±231 955±336

s9 40-45 33.85 1702±125 5400±0 577±76 1583±395

Data from He et al., (2008).

*extracted for one hour in 0.5M HCl at pH <2.0, extracts amorphous Fe oxides, AVS

**extracted for 0.5 hr, microwave digestion in HNO3 

Batch experiments were conducted with material 

collected from the column described in Shen 

and Wilson (2007) as the Column with Mulch 

and Hematite (column B3). After 875 days of 

operation, Column B3 was put into a freezer for 

one week. The column was frozen solid, and the 

glass container broke apart. The glass was removed 

from the column of frozen pore water, shredded 

plant mulch, river sand, and hematite and then 

the frozen column was cut into nine sections with 

a cross cut saw. The original column was 45 cm 

long and 15 cm in diameter, each section was 5 cm 

long and 15 cm in diameter. The sections were 

numbered 1 through 9, with 1 being adjacent to 

the column influent and 9 adjacent to the column 

effluent. The sections were returned to the freezer 

before they could thaw. 

Each section was thawed in an anaerobic glovebox 

(N2:H2=92.5%:7.5%), homogenized and separated 

into 4 equal subsections, transferred to a plastic 

bag with a zip seal, sealed without headspace, 

and refrozen until used for chemical analysis or 

preparation of batch microcosms. 

For analysis of Fe and S partitioning, samples were 

thawed in the anaerobic glovebox in sealed plastic 

bags without headspace. Wet sediments were 

used directly for analysis. Analytical results were 

corrected for the water content; the water content 

was determined gravimetrically. The analysis of 

total iron and HCl extractable iron was modified 

after Kosta and Luther (1994). The analysis of acid 

volatile Sulfide (AVS) and Chromium Reducible 

Sulfide (CRS) followed procedures used by Wilkin 

and Bischoff (2006). Results are presented in 

Table B.1. Data are from He et al., (2008).

Acid Volatile Sulfide includes and is expected 

to be dominated by FeS, and CRS includes and 

is expected to be dominated by FeS2. Batch 

microcosms were constructed with column material 

from section 2 and section 6. The concentration 

of iron associated with AVS and CRS in section 2 

is 4706 mg/kg, and the concentration of iron that 

was associated with AVS and CRS in section 6 

is 3856 mg/kg. This represents 92% and 84% of 

the total iron remaining in the columns. Allowing 

for uncertainty in the determinations, the iron(III) 
51

minerals originally present in the river sand and 

the hematite added to the column were almost 

entirely converted to FeS and FeS2.

Material from section 2 and section 6 was selected 

to prepare batch microcosms. 

To remove the confounding effects of sorption to 

the plant mulch on concentrations of contaminants, 

the sections were sieved to remove the plant 

mulch. The frozen section (contained in a plastic 

bag without headspace) was allowed to thaw in a 

glovebag filled with oxygen free nitrogen, but no 

hydrogen. This was done to minimize enrichment 

of anaerobic bacteria in the microcosms. The 

atmosphere of the glovebag was exchanged three 

times before the bags were opened. 

Microcosms were prepared with material from 

section 2 and labeled as 2-10, 2-11, and 2-12. 

Material from section 6 was used to prepare 

microcosms labeled 6-15, 6-16, 6-17, and 6-19. 

The microcosms were prepared as follows. To 

separate the fine sand and iron sulfide precipitates 

from the plant mulch, each of the thawed sections 

was sonicated in a Branson 1510 sonicator for two 

cycles of ten minutes each. Then the sediment 

and pore water was sieved to remove the shredded 

plant mulch. The wet sediment that passed the 

sieve was blended thoroughly and then distributed 

to 20 ml serum vials. Each vial received 19 ml 

of wet sediment, and 1.0 ml of a solution that 

contained 14 gm per liter of Na2SO4. The serum 

vials were then sealed with a Teflon-faced butyl 

rubber septum and an aluminum crimp cap. 

An additional microcosm was prepared using 

material from section 6 and labeled as 6-s4. The 

microcosm was prepared as follows. The mulch 

from section 6 was placed in boiled reverse 

osmosis water and sonicated again for 10 minutes 

(500 gm wet mulch, 150 ml water). Then the 

mulch was removed from the water and sediment 

by sieving. The material that passed the sieve 

was allowed to stand to separate the sediment 

from the water. The sediment was collected, 

while the water was added back to the mulch and 

sonicated again for 10 minutes. After the water 

and sediment were removed from the mulch be 

sieving, the water and sediment were combined 

with the previous sediment, and the mixture 

was allowed to settle overnight. The water was 

decanted, the settled sediment was mixed well, and 

then 19 ml of wet sediment was added to a 20 ml 

serum vial. The vial received 1.0 ml of a solution 

that contained 14 gm per liter of Na2SO4. Then the 

serum vial was sealed with a Teflon-faced butyl 

rubber septum and an aluminum crimp cap.

The water content of the microcosms and the 

concentrations of AVS and CRS are presented in 

Table B.2. 

Dose solutions of EDB and 1,2‐DCA were 

prepared as follows. A 165-ml serum bottle was 

filled with 3.42 grams of Na2SO4, a magnetic spin 

bar, and boiled reverse osmosis water to make a 

positive meniscus. Then 53 microliters of EDB 

or 92 microliters of 1,2‐DCA were added with a 

syringe, and the bottle was sealed with a Teflon 

faced septum and an aluminum crimp cap. The 

dose solution was stirred overnight to dissolve the 

EDB or 1,2‐DCA. 

The dose solution contained Na2SO4 to ensure 

that adequate sulfate was available to sustain 

sulfate reducing conditions over the course 

of the incubations. Figure B.1 presents data 

for the microcosms dosed with EDB. Sulfate 

concentrations above 100 mg/L were sustained on 

any date where EDB was detected.

Table B.2: Distribution of pore water, AVS, and CRS in microcosms.

Microcosm Water content CRS AVS AVS

% volume basis mg/kg mg/kg mM/L*

2-10 43.5 2222 1907 198

2-11 43.1 2417 2577 272

2-12 43.8 3349 2700 278

6-15 46.0 6319 4054 381

6-s4 40.2 3454 1983 236

6-16 48.7 3309 2537 279

6-17 57.4 2309 1488 121

*millimole in contact with a liter of pore water.
52

The microcosms and the dose solutions were taken 

into a glovebag with an atmosphere of oxygen free 

nitrogen. The microcosms were opened; 1.0 ml 

of standing pore water was removed and replaced 

with 1.0 ml of the appropriate dose solution. The 

microcosms were sealed with new septa and crimp 

caps.

The dosed microcosms were removed from the 

glove bag and incubated at room temperature in the 

dark on a roller that slowly stirred the contents of 

the microcosms. The microcosms were sampled 

after one hour and then after one, two, three, and 

four weeks of incubation. At each sampling period, 

the microcosms were removed from the roller and 

allowed to settle. The microcosms were placed 

inside the anaerobic glove bag and the septa were 

removed. For each microcosm, 0.5 ml of standing 

water was collected and transferred to a 1.5 ml 

micro centrifuge tube for analysis of sulfate. An 

additional 0.5 ml aliquot of standing water was 

collected and transferred to a 20 ml serum vial 

containing 14.5 ml of reverse osmosis water for 

analysis of EDB or 1,2‐DCA. Three drops of 1:1 

HCl were added to the 20 ml serum vials to adjust 

the pH to less than 2. The vials were then sealed 

with a septum and crimp cap and stored at 4 °C 

until analyzed.

Concentrations of EDB and 1,2‐DCA were 

determined as described earlier by head space gas 

chromatography/mass spectrometry (GC/MS). 

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. The 

lowest calibration standard was 1.0 mg/L.

 Figure B.1. Consumption of sulfate during incubation of microcosms.
53

Appendix C

Method for Compound Specific Isotope 

Analysis to Determine the Ratio of Stable 

Carbon Isotopes in EDB and 1,2-DCA 

This method applies compound specific isotope 

ratio mass spectrometry to determine the ratio of 

the stable isotopes of carbon in 1,2-dibromethane 

(EDB) and in 1,2-dichloroethane (1,2‐DCA) 

dissolved in ground water that is contaminated with 

gasoline.

Because there was no available protocol for 

analysis of EDB and 1,2‐DCA in ground water, 

U.S. EPA contracted with the University of 

Oklahoma (Dr. Paul Philp, pphilp@ou.edu) to 

develop a protocol. Because the method for 

analysis of stable isotopes requires baseline 

separation of peaks in the gas chromatogram, 

it was necessary to separate EDB from the 

other fuel components by two-dimensional gas 

chromatography. To our knowledge, at this 

writing the University of Oklahoma is the only 

commercially available source for analysis of 

stable isotopes of carbon in EDB in ground water 

contaminated with gasoline. It is possible that other 

vendors will choose to provide this service in the 

future. U.S. EPA makes no endorsement of the 

services provided by the University of Oklahoma. 

The analytes were extracted by a purge and trap 

(P&T model OI 4660) interfaced to a GC‐IRMS 

instrument (Finnigan MAT 252 IRMS). Due to 

the chromatographic complexity of the samples, 

satisfactory resolution of EDB and 1,2‐DCA 

required a 2-dimensional chromatographic 

approach (separation on polar GC phase followed 

by separation on non-polar GC phase). The 

cryogenic focuser at the P&T-GCIRMS interface 

described in Kuder et al., (2005b) was programmed 

for collecting 2 minute heart-cuts from the retention 

window of EDB or DCA of the sample eluting from 

the polar pre-column. The heart-cuts were directed 

onto a non-polar phase GC column for final 

separation followed by on-line combustion and 

analysis of the isotope composition. The analyses 

were otherwise performed as described in Kuder et 

al., (2005a) and Kuder et al., (2005b). 

For determination of the d13C values for EDB 

and 1,2‐DCA, the analytes were purged from 

a 25 ml water sample on to a Vocarb 3000 trap 

for 12 minutes. The sample temperature was 

50 °C. The sample was desorbed over a 3 

minute period. The trap temperature was 25 °C 

at purge, 240 °C at desorption, and the trap was 

baked for 15 minutes at 260 °C before the next 

cycle. The initial GC separation was achieved 

on a DB-Wax column, 30 m, 250 μm i.d., film 

thickness 0.5 μm with Helium as the carrier gas 

at an initial flow of 6 ml/min (constant flow). 

The GC temperature program was isothermal at 

40 °C during 5 minutes, then 4 °C/min up to the 

time of elution of 1,2‐DCA or EDB. The second 

stage of separation was achieved on a DB-MTBE 

column, 60 m, 320 μm i.d, film thickness 1.8 μm, 

with Helium as the carrier gas at an initial flow of 

1.5 ml/min (constant flow). The GC temperature 

program was isothermal at 40 °C during 5 minutes, 

then 2.5 °C/min up to 120 °C and 25 °C/min up 

to 220 °C (hold time 15 min.). The combustion 

reactor for combustion of the components to CO2

and water was maintained at 980 °C. 

The isotopic composition of the samples was 

measured relative to a CO2 standard directly 

introduced as a reference into the ion source. A 

standard solution of EDB and 1,2‐DCA was 

run randomly between the tests to check the 

reproducibility of the PT-GC-IRMS method. 

To eliminate problems with method linearity 

(relationship between signal strength and the 

obtained δ13C), the concentrations of diluted 

samples and standards were kept within a narrow 

range of concentrations, approximately 12 μg/L of 

EDB. For samples that were lower in concentration, 

standards were run at a corresponding 

concentration. It was observed that a portion of 

EDB was degraded on active surfaces of the P&T 

and/or the GC-IRMS interface. This resulted in 

isotope fractionation, and as a result, the overall 

method bias for EDB was larger than the typical 
54

range for volatile organic compounds. The samples 

discussed in this report were analyzed in several 

batches, with varying degrees of analytical bias 

(from +1 to +4 ‰). The bias affected the samples 

and standards in the same extent. The bias remained 

steady over the duration of analytical work (over 

several days) and could be corrected based on the 

EDB standards. 

To evaluate the method, standards of EDB 

and 1,2‐DCA were purchased, and standard 

solutions were prepared in water at concentrations 

of 4.2 μg/L, 8.4 μg/L, and 12 μg/L. The 

reproducibility was good, the sample standard 

deviation of duplicate analyses of δ13C in EDB 

standards averaged 0.5‰. There were several 

outliers with standard deviations of 1‰.

To determine whether the presence of gasoline 

hydrocarbons would bias the samples, a “standard” 

of contaminated ground water was prepared by 

dissolving gasoline (4 μl) into methanol (43 ml) 

and then by spiking 15 μl of this diluted gasoline 

into 25 ml of distilled water. The standards of 

EDB and DCA were spiked into the “standard” 

contaminated ground water, and measured values 

were compared to each other. Results are in 

Table C.1. Concentrations of EDB and 1,2‐DCA 

varied from 3.6 to 48 μg/L. Concentrations at 

4.2 μg/L, 8.4 μg/L, and 36 μg/L were determined 

in replicate. The values of δ13C in EDB in 19 

replicate samples varied from -27.0‰ to -28.5‰, 

and values for 1,2‐DCA varied form -27.8‰ to 

-28.9‰. The sample standard deviation for δ13C 

in EDB and 1,2‐DCA was the highest at the lowest 

concentration (4.2 μg/L). A sample standard 

deviation of 0.5‰ was be taken as the data quality 

objective for determination of δ13C in EDB and 

1,2‐DCA, as a consequence, 4 μg/L is the lowest 

concentration of EDB and 1,2‐DCA that could be 

analyzed for δ13C with acceptable precision.

Table C.1. Reproducibility of δ13C values for EDB and 1,2‐DCA prepared by a purge and trap sampler 

from ground water containing aqueous solutions of EDB and 1,2‐DCA and gasoline. 

Run # Conc EDB DCA

μg/L δ13C δ13C

Mean Stan dev 

n=2 Mean Stan dev n=2

1975 3.6 -27.7 -28.4

1976 4.2 -27.9

-27.475 0.5

-28.4

-28.05 0.4

1979 4.2 -27.5 -27.8

1981 4.2 -27.2 -27.9

1982 4.2 -27.3 -28.1

1973 4.8 -27.8 -28.6

1984 8.4 -27.3

-27.26 0.4

-28.2

-28.46 0.4

1986 8.4 -27.6 -28.6

1988 8.4 -27.3 -28.3

1990 8.4 -27.1 -28.8

1991 8.4 -27 -28.4

1971 12 -27.6 -28.5

1972 21.6 -28.2 -28.4

1974 36 -28.3

-28.42 0. 1

-28.7

-28.76 0.1

1978 36 -28.3 -28.8

1980 36 -28.5 -28.7

1985 36 -28.5 -28.9

1989 36 -28.5 -28.7

1977 48 -28.0 -28.5
55

Appendix D.

Analytical Methods and Quality Assurance 

Six compounds were considered critical parameters 

in the Quality Assurance Project Plan; EDB, 

1,2‐DCA, benzene, nitrate, sulfate, and methane.

Appendix D.1

1,2-Dibromoethane (EDB) in Water 

Samples from Field Sites

When EDB was determined in samples of ground 

water from field sites, EDB was determined by 

EPA Method 8011. In some analyses, the method 

detection limit was 0.010 μg/L, and the lowest 

calibration standard was 0.03 μg/L. In other 

analyses the method detection limit was 0.005 and 

the limit of quantitation was 0.020 μg/L. 

The acceptance value for the method blank was a 

value less than the method detection limit for the 

particular report for that sample set. A total of 

507 method blanks were analyzed in sample sets 

analyzed on 137 different dates, and all 507 method 

blanks were below the method detection limit 

reported with the sample set. 

A total of 414 continuing calibration checks were 

performed for EDB in sample sets analyzed on 130 

different dates. The acceptance value was 60% to 

140% of the nominal value of the calibration check 

standard. The range of reported values was 62.6% 

to 255% of the nominal values. Two continuing 

calibration checks were out of the acceptable range, 

a value of 206% and a value of 255% in samples 

reported 05-07-2007.

A total of 112 performance standards (secondary 

source standards) were evaluated for EDB in 

sample sets analyzed on 84 different dates. The 

acceptable range of recoveries was 60% to 140% of 

the nominal value of the secondary source standard. 

None of the secondary source standards were out of 

the acceptable range. 

A total of 74 matrix spikes were performed for 

EDB on samples analyzed on 58 different dates. 

The acceptable range of recoveries was 60% to 

140% of the spiked concentration. The actual range 

of recoveries was 48% to 309%; five samples were 

out of the acceptable range.

A total of 11 laboratory control spikes were 

analyzed for EDB on samples analyzed on nine 

different dates. The acceptable range of recoveries 

was 60% to 140% of the spiked concentration. The 

actual range of recoveries was 72.4% to 134%. No 

samples were out of the acceptable range.

Appendix D.2

1,2-Dibromoethane (EDB) in Water 

Samples from Microcosm Studies

In samples from laboratory microcosms, EDB was 

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 head space sampler. Analytes 

were determined by gas chromatography/mass 

spectrometry using an Ion Trap Detector. The 

lowest calibration standard was 0.5 μg/L; the 

method detection limit was 0.1 μg/L.

The acceptance value for EDB in the method blank 

was less than the method detection limit. A total 

of 29 method blanks were analyzed in sample sets 

analyzed on five different dates, and all 29 method 

blanks were below the method detection limit. 

A total of 52 continuing calibration checks were 

performed for EDB in sample sets analyzed on five 

different dates. The acceptance value was 60% 

to 140% of the nominal value of the calibration 

check standard. The range of reported values 

was 88% to 112% of the nominal values. No 

continuing calibration check samples were out of 

the acceptable range.

A total of 15 matrix spikes were performed for 

EDB on samples analyzed on five different dates. 

The acceptable range of recoveries was 60% to 

140% of the spiked concentration. The actual range 

of recoveries was 96% to 107%. No spike samples 

were out of the acceptable range. A total of 46 

secondary source standards were analyzed for EDB 

on samples analyzed on five different dates. The 

acceptable range of recoveries was 60% to 140% 

of the spiked concentration. The actual range of 

recoveries was 90% to 113%. No samples were out 

of the acceptable range. A total of seven laboratory 

duplicates were analyzed on five different dates. 

The acceptable range of agreement was for the 

laboratory duplicates to agree with each other 

with a relative percent difference of ± 25%. Two 

of the seven laboratory duplicates had detectable 
56

concentrations of EDB; the relative percent 

difference was 5% and 11%. 

Appendix D.3

1,2-Dichloroethane (DCA) in Water 

Samples from Field Sites

Concentrations of 1,2‐DCA 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 head space 

sampler. Analytes were determined by gas 

chromatography/mass spectrometry using an Ion 

Trap Detector. The lowest calibration standard was 

0.5 μg/L; the method detection limit was 0.13 μg/L.

A total of 79 method blanks were analyzed for 

1,2‐DCA in sample sets analyzed on 44 different 

dates. All method blanks were less than the method 

detection limit. 

A total of 86 continuing calibration checks were 

analyzed in sample sets on 50 different dates. The 

range of acceptable values was 80% to 120% of the 

nominal value of the check samples. The values 

of the continuing calibration check samples ranged 

from 81% to 120% of the nominal value. All 

samples were within the acceptable range.

A total of 39 performance evaluation standards, 

or secondary source standards, were analyzed on 

29 different dates. The range of acceptable values 

was 80% to 120% of the nominal value of the 

check samples. The values of the secondary source 

standards varied from 90% to 117% of the nominal 

value. All samples were within the acceptable 

range.

A total of 25 matrix spikes were performed in 

water samples analyzed on 19 separate dates. The 

acceptable range of recoveries was 70% to 130% 

of the spiked concentration. The actual range was 

83% to 119% of the spiked concentration. All 

samples were within the acceptable range.

Appendix D.4

1,2-Dichloroethane (DCA) in Water 

Samples from Microcosm Studies

Samples from microcosms were analyzed using 

the same method as described above for samples 

from field sites. A total of 29 method blanks were 

analyzed for 1,2‐DCA in sample sets analyzed on 

five different dates. All method blanks were less 

than the method detection limit. 

A total of 52 continuing calibration checks were 

analyzed in sample sets on five different dates. The 

range of acceptable values was 80% to 120% of the 

nominal value of the check samples. The values 

of the continuing calibration check samples ranged 

from 91% to 112% of the nominal value. All 

samples were within the acceptable range.

A total of 46 performance evaluation standards, or 

secondary source standards, were analyzed on five 

different dates. The range of acceptable values 

was 80% to 120% of the nominal value of the 

check samples. The values of the secondary source 

standards varied from 91% to 120% of the nominal 

value. All samples were within the acceptable 

range.

A total of 15 matrix spikes were performed in 

water samples analyzed on five separate dates. The 

acceptable range of recoveries was 70% to 130% 

of the spiked concentration. The actual range was 

91% to 115% of the spiked concentration. All 

samples were within the acceptable range.

A total of seven laboratory duplicates were 

analyzed on five different dates. The acceptable 

range of agreement was for the laboratory 

duplicates to agree with each other with a relative 

percent difference of ± 25%. Five of the seven 

laboratory duplicates had detectable concentrations 

of 1,2‐DCA; the relative percent differences were 

6%, 0.8%, 6%, 0.6% and 4.4%. 

Appendix D.5

Benzene in Water Samples from 

Field Sites

Concentrations of benzene were determined by 

headspace 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 head space 

sampler. Analytes were determined by gas 

chromatography/mass spectrometry using an Ion 

Trap Detector. The lowest calibration standard was 

0.5 μg/L; the method detection limit was 0.07 μg/L.

A total of 78 method blanks were analyzed for 

benzene in sample sets analyzed on 45 different 

dates. All method blanks were less than the method 

detection limit. 

A total of 88 continuing calibration checks were 

analyzed in sample sets on 50 different dates. The 

range of acceptable values was 80% to 120% of the 

nominal value of the check samples. The values 

of the continuing calibration check samples ranged 
57

from 89% to 115% of the nominal value. All 

samples were within the acceptable range.

A total of 38 performance evaluation standards, 

or secondary source standards, were analyzed on 

29 different dates. The range of acceptable values 

was 80% to 120% of the nominal value of the 

check samples. The values of the secondary source 

standards varied from 87% to 107% of the nominal 

value. All samples were within the acceptable 

range.

A total of 25 matrix spikes were performed in 

water samples analyzed on 19 separate dates. The 

acceptable range of recoveries was 70% to 130% 

of the spiked concentration. The actual range was 

89% to 109% of the spiked concentration. All 

samples were within the acceptable range.

Appendix D.6

Nitrate

Samples were analyzed for nitrate plus nitrite 

nitrogen using an automated hydrazine reduction 

method based on EPA Method 353.2. The method 

detection limit was 0.01 mg/L and the lowest 

calibration standard was 1.0 mg/L

A total of 14 method blanks were analyzed for 

nitrate plus nitrite nitrogen in sample sets analyzed 

on 6 different dates. There were four method 

blanks below the method detection limit, and 

concentrations in the remainder of the samples were 

below 0.047 mg/L. All data were considered of 

adequate quality for the purpose of this report. 

A total of 19 continuing calibration checks were 

analyzed in sample sets on six different dates. The 

range of acceptable values was 90% to 110% of the 

nominal value of the check samples. The values 

of the continuing calibration check samples ranged 

from 94% to 105% of the nominal value. All 

samples were within the acceptable range.

A total of seven performance evaluation standards, 

or secondary source standards, were analyzed on 

six different dates. The values of the secondary 

source standards varied from 93% to 101% of the 

nominal value. No data were flagged as being out 

of the acceptable range.

A total of 10 matrix spikes were performed in 

water samples analyzed on six separate dates. The 

acceptable range of recoveries was 80% to 120% 

of the spiked concentration. The actual range was 

88% to 115% of the spiked concentration. All 

matrix spikes were less than the method detection 

limit. 

A total of 13 laboratory duplicates were performed 

on six separate days. The acceptable range was a 

percent relative difference of ± 10%. The relative 

percent difference ranged from 0.05% to 1.05%. 

Appendix D.7

Sulfate

Concentrations of sulfate were determined by two 

separate methods. In some samples, 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. The 

lowest calibration standard was 1.0 mg/L.

A total of 14 method blanks were analyzed for 

sulfate in sample sets analyzed on seven different 

dates. All method blanks were less than the method 

detection limit. Two field blanks were analyzed on 

one day. Both were less than the method detection 

limit. 

A total of 14 continuing calibration checks were 

analyzed in sample sets on seven different dates. 

The range of acceptable values was 90% to 110% 

of the nominal value of the check samples. The 

values of the continuing calibration check samples 

ranged from 94% to 102% of the nominal value. 

All samples were within the acceptable range.

A total of seven performance evaluation standards, 

or secondary source standards, were analyzed on 

six different dates. The values of the secondary 

source standards varied from 86% to 97% of the 

nominal value. No data were flagged as being out 

of the acceptable range.

A total of six matrix spikes were performed in 

water samples analyzed on six separate dates. The 

acceptable range of recoveries was 80% to 120% 

of the spiked concentration. The actual range was 

89% to 97% of the spiked concentration. One 

matrix spike was out of range by 1%. Despite this 

one excursion out of the acceptable range in the 

SOP, all data were considered of adequate quality 

for the purpose of this report.

A total of six laboratory duplicates were performed 

on five separate days. The acceptable range was a 

percent relative difference of ± 10%. The relative 

percent difference ranged from 0% to 2.74%. 

Water samples that had been preserved with HCl 

were analyzed by an automated turbidimetric 

method based on EPA Method 375.4. The method 

detection limit was 1.14 mg/L and the lowest 

calibration standard was 2.0 mg/L. 
58

A total of eight method blanks were analyzed for 

sulfate in sample sets analyzed on three different 

dates. All method blanks were less than the method 

detection limit. 

A total of 13 continuing calibration checks were 

analyzed in sample sets on three different dates. 

The range of acceptable values was 90% to 110% 

of the nominal value of the check samples. The 

values of the continuing calibration check samples 

ranged from 93% to 104% of the nominal value. 

All samples were within the acceptable range.

A total of four performance evaluation standards, or 

secondary source standards, were analyzed on three 

different dates. The values of the secondary source 

standards varied from 86% to 97% of the nominal 

value. No data were flagged as being out of the 

acceptable range.

A total of four matrix spikes were performed in 

water samples analyzed on two separate dates. The 

acceptable range of recoveries was 80% to 120% 

of the spiked concentration. The actual range was 

85% to 114% of the spiked concentration. All 

matrix spikes were less than the method detection 

limit. 

A total of four laboratory duplicates were 

performed on four separate days. The acceptable 

range was a percent relative difference of ± 10%. 

The relative percent difference ranged from 0.70% 

to 7.75%. 

 Appendix D.8

Methane

Concentrations of methane were analyzed using 

a headspace equilibration technique based on 

Kampbell and Vandegrift (1998). The method 

detection limit was 0.0001 mg/L and the lowest 

calibration standard was 0.0010 mg/L.

A total of 18 method blanks were analyzed for 

methane in sample sets analyzed on five different 

dates. All method blanks were below the method 

detection limit. 

A total of 27 continuing calibration checks were 

analyzed in sample sets on seven different dates. 

The range of acceptable values was 85% to 115% 

of the nominal value of the check samples. The 

values of the continuing calibration check samples 

ranged from 93% to 105% of the nominal value. 

All samples were within the acceptable range.

A total of six performance evaluation standards, 

or secondary source standards, were analyzed on 

five different dates. The expected values of the 

secondary source standards varied from 85% to 

115% of the nominal value. The measured values 

of the secondary source standards varied from 99% 

to 101% of the nominal value. 

A total of eight laboratory duplicates were 

performed on six separate days. The acceptable 

range was a percent relative difference of ± 20%. 

The relative percent difference ranged from 0.95% 

to 3.17%. Two field duplicates were analyzed; 

however, they both were below the detection limit. 

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