Report to Congress on Public Health,
Air Quality, and Water Resource Impacts
of Fuel Additive Substitutes for MTBE
United States
Environmental Protection
Agency
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Report to Congress on Public Health,
Air Quality, and Water Resource
Impacts of Fuel Additive Substitutes for
MTBE
Office of Research and Development
Office of Transportation and Air Quality
U.S. Environmental Protection Agency
SER&
United States
Environmental Protection
Agency
EPA-420-R-09-001
February 2009
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Report to Congress on Public Health, Air Quality, and Water Resource Impacts
of Fuel Additive Substitutes for MTBE
Table of Contents
Executive Summary
I. Introduction
II. Potential Public Health Effects of MTBE, Its Substitutes and Related Emissions Products
III. Commercial Feasibility of Ethanol and Other Additives
IV. Potential Air Quality Effects
V. Potential Water Quality Effects
VI. Potential Approach for Further Analyses: Comprehensive Environmental Assessment of
Fuel Additive Substitute Options
VII. Catalog of Scientific Activities Related to MTBE and Its Substitutes and Information
Gaps
Appendix A: Summary of the Impacts of Gasoline Regulation on Fuel Composition
Appendix B: Phase Partitioning of Compounds from Gasoline
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Executive Summary
In the Energy Policy Act of 2005 (EPAct), Congress directed the Administrator of the
Environmental Protection Agency to conduct a study on the effects on public health, air quality
and water resources of the increased use of replacements for the fuel additive methyl tertiary
butyl ether (MTBE), as well as the feasibility of such replacements. Specifically, Congress
directed attention to ethyl tertiary butyl ether (ETBE), tertiary amyl methyl ether (TAME), di-
isopropyl ether (DIPE), tertiary butanol (TEA), other ethers and heavy alcohols, ethanol, iso-
octane, and alkylates. EPAct also required other actions that significantly altered the commercial
fuels market and the feasibility for broad commercial use of many of the potential replacement
additives, including removal of the oxygenate standard in the federal reformulated gasoline
program and establishment of the national renewable fuels standard program. These actions
ultimately affected the priority focus of this report toward ethanol, which currently dominates the
gasoline market as a replacement for the use of MTBE. This report summarizes scientific
information regarding the health effects, air quality impacts, and water quality impacts of ethanol
and the other additives but does not attempt to weigh their relative risks and benefits. The focus
of the report is limited to the feasibility and impacts of MTBE replacements once the additive
has been blended with gasoline and distributed to service stations.
Consistent with Congressional intent, the report is not meant to address the larger issues
surrounding the increased use of alternative or renewable fuels. However, due to legislative
mandates for renewable fuels and the primacy of ethanol in the marketplace, particular attention
is given to the potential impacts of its use as a fuel additive.
Health Effects
This report highlights the adverse health effects that have been linked to MTBE and its
alternatives. All of these compounds have the potential to cause adverse health effects,
depending on the nature and degree of exposure. In general, humans can be exposed to fuel
additives by way of evaporative or exhaust emissions (leading to potential inhalation exposure)
or by spills or leaking tanks (leading to potential dermal and ingestion exposure). However,
characterization of exposure levels was beyond the scope of this report. For this reason and
because MTBE and several other fuel additive compounds are currently undergoing health
effects assessment under EPA's Integrated Risk Information Service (IRIS) program, this report
does not attempt to estimate quantitative risks or hazards of these chemicals. Section 1505
specifies that health effects include effects on "children, pregnant women, minority or low-
income communities, and other sensitive populations." Where possible, such information is
included in this report, but for many fuel additive compounds detailed information for sensitive
human populations is lacking, although extrapolations from animal data in IRIS assessments may
take sensitive populations into account.
In addition to the potential direct effects of the fuel additives themselves, use of these fuel
additives can affect (either positively or negatively) environmental levels of other pollutants such
as particulates, ozone, benzene, formaldehyde, and acetaldehyde. This report presents brief
summaries of the adverse health effects associated with such secondary pollutants as well.
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EPA is currently involved in numerous scientific activities related to fuels and fuel
additives. A health effects testing program for gasoline and mixtures of gasoline with six
respective oxygenates (MTBE, ethanol, ETBE, DIPE, TAME and TEA), required pursuant to
Section 21 l(b) of the Clean Air Act, is currently being conducted by a consortium of fuel and
fuel additive manufacturers. This testing program is administered by the American Petroleum
Institute with oversight by EPA. EPA will use the results of this testing program to help make
better informed risk assessment and risk management decisions regarding the future use of fuel
oxygenate compounds.
Air Quality
For criteria air pollutants and their precursors, increased use of ethanol at 10% volume
(E10) can lead to: an increase in VOC emissions due to increased fuel evaporation and
permeation (even though exhaust VOC emissions decrease slightly); a decrease in CO emissions,
and an increase in NOx emissions (primarily from nonroad engines). Further, the decrease in the
aromatic content of E10 fuel is likely to lead to a decrease in secondary PM formation. Ambient
ozone and PM levels are of particular concern for achieving and maintaining attainment with
National Ambient Air Quality Standards (NAAQS). For air toxics, present analyses suggest that
the increase in E10 usage will decrease emissions of benzene, 1,3-butadiene, and possibly
formaldehyde and increase emissions of others such as acetaldehyde. Other air toxics have not
been examined in this report. Potential public health impacts resulting from the increase in
acetaldehyde depend upon exposure levels, which are not known at this time, and a fuller
understanding of acetaldehyde toxicity, which is currently being reviewed under the IRIS
program. Given that both formaldehyde and acetaldehyde are also formed through atmospheric
reactions, a more complete analysis of potential exposures to these pollutants would require an
assessment of those atmospheric reactions and is beyond the scope of this report.
Further analyses of emission and evaporative impacts due to increased ethanol usage and
resulting impacts on air quality are being conducted to accompany the second renewable fuels
standard (RFS2) required by EISA. These standards mandate the use of 36 billion gallons of
ethanol by 2022 and are anticipated to be proposed early in 2009. EPA is also currently
conducting tests on the effect of ethanol blends on exhaust and evaporative emissions from
newer model vehicles and nonroad engines.
Water Quality
The potential for exposure to fuel components and/or additives can occur when
underground fuel storage tanks leak fuel into ground water that is used for drinking water
supplies or when spills occur that contaminate surface drinking water supplies. MTBE was
banned or phased out in several states due to its appearance in drinking water supplies and
exceedance of odor/taste thresholds. Unlike MTBE, which itself is a contaminant in water
supplies, ethanol biodegrades quickly and is not necessarily the pollutant of greatest concern.
Instead, its high biodegradability can cause the plume of BTEX (benzene, toluene, ethylbenzene
and xylenes) compounds in fuel to extend farther (by as much as 70%) and persist longer in
ground water, thereby increasing potential exposures to these compounds.
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EPA is beginning to assess the impacts of ethanol on storage tanks and associated
infrastructure. First, material incompatibility of existing underground storage tanks systems with
ethanol-fuel blends can result in equipment failure and releases to ground water. Second, storage
of denatured ethanol in aboveground storage tanks and steel piping can lead to stress corrosion
cracking. Further, ethanol-blended fuels may have an impact on the existing leak detection
systems that are designed for petroleum hydrocarbons. Although some of these issues are
beyond the scope of this report, these data gaps need to be addressed to prevent impacts to
ground water resources.
EPA is presently exploring the possibility of collaborative field studies with states to
examine the fate and transport of ethanol releases from underground storage tanks and their
potential impact on water supply wells. This work will assess the influence of water supply
pumping rates on the movement of the contaminated plume and allow for the development of a
tool for remediation site managers and water utilities that allows for integration of contaminant
source modeling and water supply pumping rates to determine the rate and extent of remediation
needed to protect existing and future water supplies.
An additional potential hazard from spills from fuels containing ethanol is currently being
evaluated. Laboratory and field studies have found biodegradation of ethanol can produce
concentrations of methane in excess of the water solubility of methane (i.e., more methane was
produced than could be dissolved by the available water). This methane could bubble out of the
ground water and enter the soil gas (gases that occur in the small spaces between particles of
soil) at potentially explosive concentrations, although it is not possible to quantify the risk at this
time. EPA is developing modeling software for the assessment of fuels of varying composition
on ground water, with simulation of methane production being one component of this work.
Further Analyses
In order to truly weigh the benefits and trade-offs between various fuels and fuel
additives, a comprehensive multimedia analysis of the environmental impacts is needed. Due to
the extensive qualitative and/or quantitative analyses this would require, which are beyond the
scope and timeframe of this report, we suggest other potential methods for obtaining this
information. Section 204 of EISA 2007 requires that EPA produce a report every three years on
the multimedia impacts of increased renewable fuel production and use, which will enhance our
understanding of the impacts of ethanol and other renewable fuels and fuel additives. In this
regard, Comprehensive Environmental Assessment (CEA) can be used to compare the
advantages and disadvantages of various fuels and fuel additives. CEA combines the risk
assessment paradigm with a product life cycle framework to evaluate multimedia environmental
impacts, including both ecological and human health effects, and uses collective expert judgment
methods to deal with limitations in empirical data, thus providing a qualitative comparison
between options. Although not necessarily quantitative in nature, this methodology can
nevertheless inform decision makers as to future policy direction.
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I. INTRODUCTION
This Report to Congress was prepared in response to Section 1505 of the 2005 Energy
Policy Act (EPAct), which directs the Administrator to:
(i) conduct a study on the effects on public health (including the effects on children, pregnant
women, minority or low-income communities, and other sensitive populations), air quality, and
water resources of increased use of, and the feasibility of using as substitutes for methyl tertiary
butyl ether in gasoline
(I) ethyl tertiary butyl ether;
(11) tertiary amyl methyl ether;
(III) di-isopropyl ether;
(IV) tertiary butyl alcohol;
(V) other ethers and heavy alcohols, as determined by then Administrator;
(VI) ethanol;
(VII) iso-octane; and
(VIII) alkylates....
Methyl tertiary butyl ether (MTBE) serves as an impetus as well as a point of reference
for this report. MTBE is an ether-based gasoline additive that has been used in the U.S. since the
late 1970s. It improves octane, provides other positive blending properties and also contains
oxygen, which has been shown to reduce certain harmful emissions from vehicles. MTBE use
expanded greatly in the mid-1990s as a result of the Clean Air Act Amendments mandate that
clean burning reformulated gasoline (RFG) be used in cities with the worst air pollution. (See
Appendix A for more details on how the RFG requirement affected fuel composition.) RFG
accounts for about 30% of the total US gasoline pool.
Under the 1990 amendments, every gallon of RFG was required to contain a minimum of
2 weight percent oxygen. While there are a number of oxygenates that could be used to satisfy
the requirement, a significant portion of the RFG pool contained MTBE. At the height of its use,
MTBE was used in about 60 percent of the RFG gasoline pool. Ethanol, another oxygenate, was
used in the remaining RFG gasoline pool. With passage of the Energy Policy Act of 2005
(EPAct), Congress removed the statutory oxygenate requirement from the RFG program. EPA
made corresponding changes to its regulations in the spring of 2006. Since that time, MTBE has
basically been removed from use in the U.S. gasoline market.
While MTBE helps to reduce some emissions from vehicles, its chemical properties
cause it to persist in ground water if it is spilled or leaked into the environment. As a result of
numerous fuel leaks, primarily from underground gasoline storage tanks, MTBE contaminated
drinking water supplies in several areas of the country. The threat it posed to water resources
was found to be unacceptable and led to a call to ban or limit its use (Blue Ribbon Panel, 1999)
at both state and national levels and, ultimately, Congressional removal in EPAct 2005 of the
statutory oxygenate requirement for RFG.
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This report takes a first step in examining some of the possible public health, air quality,
and water resource impacts related to other gasoline fuel replacement additives. It is important
to emphasize that this document does not present a set of final conclusions. Although work
investigating MTBE and other fuel additives has been underway at EPA and in other
organizations over the past several years, a comprehensive evaluation of the comparative
advantages and disadvantages of substitutes for MTBE remains to be undertaken. Therefore,
definitive judgments regarding the relative risks and/or benefits of various fuel additives cannot
be offered at present. Nevertheless, a substantial body of information, including work carried
out by EPA, will be summarized here, along with a suggested approach for conducting a
comprehensive environmental assessment of fuel additive substitute options.
In looking at the air and water quality effects associated with the use of MTBE and its
substitutes, this report is focused strictly on direct effects from usage as an additive (i.e., vehicle
emissions and associated air quality impacts and leaks from underground storage tanks), not the
effects associated with additive production (e.g., air quality impacts from refineries and water
quality impacts from increased corn production) or other upstream effects
This document is organized in seven chapters comprising (I) Introduction, (II) Potential
Public Health Effects of MTBE, Its Substitutes and Related Emissions Products, (III)
Commercial Feasibility of Ethanol and Other Additives, (IV) Potential Air Quality Effects, (V)
Potential Water Quality Effects, (VI) Potential Approach for Further Analyses: Comprehensive
Environmental Assessment of Fuel Additive Substitute Options, and (VII) Catalog of Scientific
Activities Related to MTBE and Its Substitutes and Information Gaps.
Chapter II, Potential Public Health Effects, first summarizes available information on the
health effects of MTBE as a point of reference for other fuel additives and blending components.
In addition to a section on MTBE, Chapter II includes sections on ethanol, ethyl tertiary butyl
ether (ETBE), tertiary amyl methyl ether (TAME), di-isopropyl ether (DIPE), tertiary butanol
(also referred to as tertiary butyl alcohol) (TEA), and alkylates/iso-octane, as well as some
mention of other alcohols and ethers for which information is quite limited. Moreover, Chapter
II provides brief summaries on the health effects of various emission products associated with
gasoline and the primary additives listed here. These products include aldehydes, peroxyacetyl
nitrate (PAN), particulate matter, ozone, and other substances that could be produced through
chemical transformation processes in air or water.
Chapter III, Commercial Feasibility of Ethanol and Other Additives, discusses current
indications on which additives and blending components are thought most likely to be present in
the U.S. marketplace. In particular, with the removal of the oxygen requirement and passage of
the Renewable Fuels Standard, ethanol currently dominates the market as a major component of
the gasoline pool. We provide information on where it is currently made, and the current
prevalence of use across the U.S. Chapter III also discusses other additives such as ETBE,
TAME, DIPE, and TEA. Since these historically constitute a small portion of the gasoline
additive market and given their historical use as well as currently passed renewable fuel volume
requirements in the Energy Security and Independence Act of 2007 (EISA), it does not currently
appear that these substances will attain commercially significant volumes. As the renewable
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fuels market continues to expand and second generation biofuels develop, additional analysis of
the fuels and additive market will be necessary.
Chapter IV, Potential Air Quality Effects, summarizes results of analyses on emissions
and associated air quality effects related to fuel additives and blending components identified in
Chapter II. Information is provided on projected changes in air pollutant emissions as a result of
increased use of ethanol and the results of an analysis that project the effect of such emissions on
air quality.
Chapter V, Potential Water Quality Effects, covers subsurface water effects identified
with various fuel additives and blending components. This section focuses on potential ground
water impacts as a result of spills from leaking underground storage tanks. The potential impact
of spills into surface water during shipment of fuel grade ethanol (typically E95) is beyond the
scope of this report, which focuses on impacts once ethanol has been blended and stored in
underground storage tanks at service stations. A qualitative discussion of surface water impacts
from spills of ethanol may be included in the Notice of Proposed Rulemaking for the upcoming
Renewable Fuels Standard that is required by EISA.
Chapter VI, Potential Approach for Further Analyses: Comprehensive Environmental
Assessment of Fuel Additive Substitute Options, describes an approach for evaluating
environmental trade-offs among fuel additive substitute options in a systematic manner that
incorporates a product life cycle framework with the risk assessment paradigm and uses expert
judgment methods to take advantage of the collective insights of an array of technical experts
and stakeholders.
Chapter VII, Catalog of Scientific Activities Related to MTBE and Its Substitutes and
Information Gaps, lists recent and current projects that EPA and other organizations have been
carrying out and highlights some of the more salient questions about fuel additives and
blendstocks that remain to be addressed if more definitive judgments are to be reached about the
comparative risks and benefits of specific fuel additives and blendstocks.
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References
Blue Ribbon Panel. (1999) Achieving clean air and clean water: the report of the Blue Ribbon
Panel on oxygenates in gasoline. Washington, DC: U.S. Environmental Protection
Agency, Office of Transportation and Air Quality; report no. EPA420-R-99-021.
Available: http://www.epa.gov/OMS/consumer/fuels/oxypanel/blueribb.htm [2000,
December 5].
U. S. Environmental Protection Agency. (2007) Regulation of fuels and fuel additives:
renewable fuel standard program. Final rule. F. R. 72 (May 1): 23899-24014.
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II. POTENTIAL PUBLIC HEALTH EFFECTS OF MTBE, ITS SUBSTITUTES AND
RELATED EMISSION PRODUCTS
This chapter highlights information on the health effects of MTBE as a point of reference
for other fuel additives. In addition, this chapter includes sections on ethanol, ethyl tertiary butyl
ether (ETBE), tertiary amyl methyl ether (TAME), di-isopropyl ether (DIPE), tertiary butanol
(TEA), and alkylates/iso-octane, as well as BTEX compounds present in gasoline. Furthermore,
this chapter provides brief summaries on the health effects of several secondary pollutants
associated with the primary additives listed here. These secondary pollutants include ozone,
particulate matter, formaldehyde, acetaldehyde, and peroxyacetyl nitrate (PAN). Where
possible, health effects information has been taken from existing EPA documents, including the
Integrated Risk Information System (IRIS) files for MTBE, iso-octane (2,2,4-trimethylpentane),
benzene, toluene, ethylbenzene, xylenes, formaldehyde, and acetaldehyde, as well as Air Quality
Criteria Documents for ozone and particulate matter. Information on the status of ongoing IRIS
assessments is available at http://cfpub.epa.gov/ncea/iristrac/index.cfm. The Regulatory Impact
Analysis for the Mobile Source Air Toxics rulemaking also presents brief summaries of these
pollutants as well as other mobile source pollutants that may be affected to a lesser degree by the
replacement of MTBE from vehicle fuels (U.S. EPA, 2007a).
Note that this chapter does not provide an exhaustive review of the literature but focuses
on assessments by governmental and other agencies that have undergone peer review and public
comment. Readers should look to existing EPA assessments (such as those supporting the IRIS
database and Integrated Science Assessments for criteria air pollutants) for more extensive
reviews of the scientific literature. Moreover, not all of the health effects noted in this chapter
are necessarily associated with exposure levels commonly seen (or anticipated) under
environmental conditions. Because this report does not attempt to estimate ambient
concentrations or exposures, the reader should not assume that all the effects described here are
actually occurring in the U.S. population; however, some effects may be. Such an evaluation is
beyond the scope of this report. The Regulatory Impact Analysis for the Mobile Source Air
Toxics rulemaking contains a summary of this type of information (U.S. EPA, 2007b).
A program of inhalation toxicity studies of several oxygenate-gasoline mixtures was
established under the Fuel/Fuel Additive (F/FA) Rule (CFR, 1994) under the authority of Section
21 l(b) of the Clean Air Act. The fuels under investigation are baseline gasoline, i.e., gasoline
without added oxygenates, MTBE-gasoline, ETBE-gasoline, TAME-gasoline, DIPE-gasoline,
ethanol-gasoline, and TBA-gasoline. A summary of the requirements and results of these studies
to date may be found in Chapter VII.
Figure II-1 shows the chemical structure for several chemicals that are or could be of
relevance as additives or gasoline blending components. The chemicals containing oxygen are
ethers or alcohols and have been or could be used as oxygenates in gasoline. The one chemical
shown in Figure II-1 that is not an oxygenate is iso-octane, which is an example of an alkylate,
and is an important blending component of gasoline. Alkylates, along with other compounds
such as aromatics and other high-octane hydrocarbon components including other branched
chained paraffins, may be added to gasoline to offset changes in fuel volume that accompany
changes in the percentage volume of different oxygenates. To illustrate, if the percentage
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volume of MTBE in gasoline is reduced from 11% to 0%, the loss of volume can be made up by
adding more iso-octane. Thus, iso-octane or other alkylate usage may be coupled with changes
in oxygenate usage.
FIGURE II-1. Chemical structure of several chemicals of interest as oxygenates or blendstock
components
a. tertiary
CH3
CH3~O~C~CH3
CH3
c.. Ethyl tertiary (ETBE)
CH3
CH3
e, Diisopropyl ether (DIPE)
CH3 CH3
CH3
CH3
g. Iso-octane or 2,2,4-Trimethylpentane
CH3 CH3
CH3-C-CH2-CH^
CH3 %CH3
i. Butanol
CH3-CH2-CH2-CH2-OH
k, tert/aryAmyl ethyl ether
(TAEE)
CH3
^
O
C
CH3
b. Ethanol
CH3-CH2-OH
d. tert/aryAmyl methyl
(TAME)
CH3
O
i
CH3-CH2-C-CH3
CH3
f, tertiary Butyl alcohol (TBA)
CH3
CH3~C-CH3
OH
h. Propanol
CH3-CH2-CH2-OH
j, tertiary Amy\ (TAA)
CH3
CH3-CH2-C-OH
CH3
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A. MTBE
In general, the health effects database for MTBE is more complete than for most other
fuel additives, but it has limitations, particularly the lack of epidemiological studies of the
long-term effects of MTBE in humans. Also, the majority of animal studies have focused on the
inhalation route of exposure, although extrapolation from inhalation to oral exposure may be
feasible for at least some endpoints by using pharmacokinetic modeling. Route-to-route
extrapolation is being evaluated as part of the current update of the IRIS file for MTBE.
When MTBE use significantly expanded in the United States with the initiation of the
1992 winter oxygenated fuels program and the 1995 reformulated gasoline program, public
health attention was primarily focused on symptom reports (e.g., eye and nose irritation,
headache) associated with acute exposures such as those that occur during refueling or while
near exhaust emissions (U.S. EPA, 1993a, 1994). In response to these concerns, several
assessments were undertaken by EPA and other organizations (e.g., U.S. EPA, 1993a, 1994;
Interagency Oxygenated Fuels Assessment Steering Committee, 1996, 1997; Health Effects
Institute, 1996; National Research Council, 1996; Blue Ribbon Panel, 1999). In essence, these
assessments, which generally were conducted by panels of independent experts, concluded that
no imminent public health threat was posed by inhalation of MTBE, but the possibility of effects
in sensitive individuals or from longer term exposure could not be ruled out based on the
available evidence. Subsequent efforts to identify and evaluate self-described MTBE-sensitive
individuals (e.g., Fiedler et al., 2000; Hong et al., 2001) have provided some support for the
possibility of biological differences among the general population in sensitivity to MTBE, but
results are equivocal.
During the mid- to late-1990s attention was drawn to the contamination of water
resources by MTBE in gasoline leaking from underground storage tanks (U.S. EPA, 1998a).
EPA's Drinking Water Advisory on MTBE (U.S. EPA 1997) analyzed the cancer and non-cancer
data that was available on MBTE at the time, but the primary outcome of this action was to
provide a consumer advisory level for MTBE based on taste and odor. Studies under controlled
conditions indicated that some individuals can detect the taste of MTBE in water at
concentrations as low as around 2 ug/L (e.g., Dale et al., 1997, as cited in EPA, 1997), although
group mean average detection levels are of course higher. Based on data from several studies,
the 1997 EPA Advisory recommended that levels of contamination should not be higher than 20
to 40 ug/L to protect consumer acceptance of the water resource and indicated that such levels
would also provide a large "margin of exposure (safety)" from toxic effects1.
1 The EPA Advisory used a risk characterization method called "Margin of Exposure (or safety)" which is different
from traditional slope factors and reference doses (RfDs) as estimates of response to defined exposures. The
"margin" is how far the environmental exposure of interest is from the lower end of the exposures at which animals
or humans have shown some toxicity effect. The use of the margin of exposure approach is helpful in the following
ways: (1) It allows for comparison of exposures associated with carcinogenic potential to those associated with non-
cancer health effects; (2) It provides the risk manager with a quick check to decide if the margin of exposure (safety)
appears to be adequate even when mathematical extrapolation of data from high to low dose cannot be done; and (3)
It gives a better understanding of the degree of risk associated with extrapolation of exposure data from animal
studies to humans. For example, given the limited number of animals that usually can be used in experiments, they,
at best, would detect a one in ten response (1 x 10"1). A common procedure for carcinogens is to mathematically
extrapolate from the exposure levels of animal tests to estimate risk at lower, environmental exposure levels. If the
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Since then, the State of California established a secondary maximum contaminant level (MCL)
of 5 ug/L (to address taste and odor concerns) in 1999 and a primary MCL of 13 ug/L (for
protection of health) in 2000 (CDHS, 2006). Note that the California health-based MCL is
higher (less stringent) than the secondary MCL. While EPA has not developed a Reference Dose
for MTBE, in 1996 the Agency for Toxic Substances and Disease Registry (ATSDR) developed
ingestion Maximal Risk Levels for MTBE (0.4 mg/kg/day for acute exposures, 0.3 mg/kg/day
for intermediate-duration exposures).
EPA has an ongoing effort to update the IRIS file for MTBE, but given that the IRIS
assessment is not yet complete, the present summary draws from past assessments and reviews.
As noted above, most of the toxicology studies of MTBE have been conducted via inhalation.
The existing inhalation reference concentration (RfC)2 for MTBE was last updated in 1993 and
was based on a chronic (24-month) inhalation study in rats (Chun et al., 1992) that showed
increased liver and kidney weights along with increased severity of spontaneous lesions of the
kidney in females, increased prostration in females, and swelling around the eyes in both males
and females (U.S. EPA, 1993b). These effects were seen at MTBE exposure levels of-10,900
mg/m3 and higher, with a no-observed-adverse-effect level (NOAEL) of-1450 mg/m3. An
uncertainty factor of 100 (comprising a factor of 10 for sensitive subpopulations, a factor of 3 for
extrapolation across species, and a factor of 3 for deficiencies in data reporting), along with
adjustment of the NOAEL to a human equivalent concentration, was applied to yield an RfC of 3
mg/m3.
Cancer risks of MTBE have been examined in various reports (e.g., U.S. EPA 1994,
1997; Interagency Oxygenated Fuels Assessment Steering Committee, 1997; National Research
Council, 1996; WHO, 1998; California EPA, 1999; IARC, 1999; U.S. DHHS, 2000a). In some
cases these assessments have characterized MTBE as a potential human carcinogen (e.g., U.S.
EPA, 1994, 1997; Interagency Oxygenated Fuels Assessment Steering Committee, 1997),
whereas other assessments concluded that the currently available data did not allow that
conclusion (e.g., NRC, 1996; WHO, 1998; IARC, 1999; U.S. DHHS, 2000a). The database for
MTBE cancer effects consists of two chronic inhalation studies, one in rats and one in mice (both
reported in Bird et al., 1997), and a chronic oral (gavage) study in rats (Belpoggi et al., 1995,
1998). The inhalation studies reported tumors in the kidneys and testes of male rats and in the
liver of female mice, while the oral study reported lymphomas and leukemias in female rats and
testicular tumors in male rats. The oral study was conducted using olive oil as the medium for
administration of MTBE; no study to date has investigated the chronic effects of MTBE in
drinking water.
One of the issues that has complicated evaluations of MTBE cancer risks for humans is
the relevance of kidney tumors found in male rats. Some evidence implicates alpha-2u-globulin,
a protein found only in male rats, as a significant factor in the formation of kidney tumors in
extrapolation is done as a straight line, a risk estimate of 1 x 10~6 generally corresponds to a margin of exposure of
100,000. If the true, but unknown, relationship is downward sloping, not a straight line, the risk at a 100,000 margin
of exposure would be less than 1 x 10"6 and might be zero.
2 The inhalation reference concentration (RfC) is an estimate (with uncertainty spanning perhaps an order of
magnitude) of a daily inhalation exposure of the human population (including sensitive subgroups) that is likely to
be without an appreciable risk of deleterious effects during a lifetime.
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male rats exposed to MTBE (e.g., Prescott-Matthews et al., 1997). If male rat kidney tumors
were to result from exposure to a chemical such as MTBE solely because of the mechanistic role
of alpha-2u-globulin, such tumors would not be considered predictive of cancer in humans and
would not contribute to the qualitative weight of evidence regarding the human carcinogen!city
potential of the chemical (U.S. EPA, 199 la). EPA's position on this issue will be examined in
conjunction with the updating of the IRIS file on MTBE. Questions also remain about other
(e.g., testicular, hepatic) tumor effects in laboratory rodent studies that have been the subject of
post-hoc analysis and differing conclusions (e.g., Goodman et al., 2007; Kissling et al., 2007).
These and other issues will be addressed in the updated IRIS file for MTBE.
Attempts to analyze the quantitative carcinogenic potency of MTBE have been limited,
perhaps reflecting the mixed qualitative picture. However, the few analyses reported to date do
not suggest that MTBE is a potent carcinogen (U.S. EPA, 1994; Interagency Oxygenated Fuels
Assessment Steering Committee, 1997).
B. Ethanol
The chemical structure of ethanol (Figure II. 1 .b) comprises an alcohol group (hydroxyl)
attached to an ethyl group, making the molecule small and weakly polar (Wallgren, 1967). It
blends well with gasoline, although it is freely miscible with water and insoluble in fats and oils,
and it can pass freely through biological membranes because of its small size (Riveros-Rosas,
1997; Wallgren, 1967). Odor and taste thresholds for ethanol in water are >100,000 and -50,000
|ig/L, respectively (Fazzalari, 1978; Amoore and Hautala, 1983), which implies that ethanol is
less readily detected than many other chemicals, notably ethers such as MTBE, ETBE, and
TAME.
There are a wide variety of health effects that have been linked to ethanol exposures,
which are summarized below. Note that in general, these data are derived from ingestion
exposures and that route-to-route extrapolations of these findings generally are either not
possible at present or have not been undertaken.
1. Noncancer Effects
The health risks of drinking alcohol have been extensively studied and evaluated (WHO,
2004; U.S. DHHS, 2000b, 2005; IARC, 1988). (Note that although alcoholic beverages may
predominantly consist of ethanol and water, other constituents may also be present, and thus
alcoholic beverages and ethanol are not necessarily equivalent chemically or lexicologically.
This point is primarily relevant to epidemiological studies; animal toxicity studies typically
administer ethanol in water to avoid potential confounding by other substances in alcoholic
beverages.) High levels of alcoholic beverage consumption have been shown to result in, among
other things, immune system effects, reproductive dysfunction, cardiovascular disease (e.g.,
hypertension and possibly stroke), and liver and pancreas toxicity. However, as is the case with
many high-level exposure studies, the relevance of such effects to lower level exposures likely
associated with ethanol-gasoline mixture use is unclear.
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Unfortunately, the toxicological effects of low-level ethanol exposure, especially by the
inhalation route, have not been well studied. As part of EPA's ongoing IRIS assessment,
pharmacokinetic models are being evaluated as a means of extrapolating across routes and
exposure levels to better characterize the health hazards and dose-response relationships for low-
level ethanol exposure by both the oral and inhalation routes. In some people, genetic factors
influencing the metabolism of ethanol can lead to differences in internal levels of ethanol and
may render some subpopulations more susceptible to risks from the effects of ethanol.
The neurological consequences of prenatal ethanol exposure on the developing fetus
(both animal and human) are one of the most prominent hazards of ethanol and have been
examined in numerous reviews (Goodlett et al., 2005; Riley and McGee, 2005; Zhang et al.,
2005; Riley et al., 2004; Gunzerath et al., 2004; WHO, 2004; Chen et al., 2003; U.S. DHHS,
2000b; Driscoll et al., 1990). Neurobehavioral problems encountered in children prenatally
exposed to alcohol include hyperactivity and attention deficits, impaired motor coordination, a
lack of regulation of social behavior or poor psychosocial functioning, and deficits in cognition,
mathematical ability, verbal fluency, and spatial memory. Facial abnormalities, growth
retardation, and spontaneous abortion may occur at higher levels of alcohol exposure; however,
less obvious but potentially serious effects on the developing brain have been reported across a
continuum of exposure levels with no clear threshold yet determined (Sampson et al., 2000).
Neurodevelopmental effects of ethanol have been investigated in relation to oral ethanol
exposure in a large number of laboratory animal and epidemiology studies, with results in animal
models corresponding closely with the effects observed in humans. Neuroanatomical studies in
laboratory animals indicate that prenatal exposure to ethanol may lead to microencephaly (small
brain relative to body size), neuronal loss, fewer dendritic spines, modification of neural
circuitry, disrupted mitochondrial membranes, and changes in cell adhesion molecules (Chen et
al., 2003). Ethanol exposure during the brain growth spurt (the early postnatal period in rats)
decreased the weight of the forebrain, brainstem, and cerebellum. Neuronal cell loss was
observed in the Purkinje cells of the cerebellum, cells of the olfactory bulb, and the pyramidal
cells in the CA1 region of the hippocampus. The structural integrity of dendrites and the number
of dendritic spines were altered in neurons located in the substantia nigra (a brain region
involved in movement), cortex, and hippocampus (involved in spatial learning and memory)
(Berman and Hannigan, 2000).
In human studies, primary evidence of neurodevelopmental toxicity stemming from
prenatal exposure includes neuroanatomical data from autopsy and neuroimaging studies
demonstrating changes in brain structure and data from epidemiological studies documenting
neuropsychological and behavioral effects in children over time. Those studies have suggested
that the corpus callosum, cerebellum, and basal ganglia are areas of the developing brain most
susceptible to prenatal alcohol exposure (Riley and McGee, 2005; U.S. DHHS, 2000b). Prenatal
exposure to alcohol in humans has also been associated with a wide range of neuropsychological
deficits and behavioral changes including impairments in cognitive function (e.g., decreased IQ),
learning, memory, attention, language development, reaction time, visual-spatial abilities,
executive functioning, fine and gross motor skills, and social functioning (Coles et al., 1992,
1991, 1987a, b, 1985; Sood et al., 2001; Jacobson et al., 1998, 1994, 1993a, b; Sampson et al.,
2000; Streissguth et al., 1989, 1986a, b). In addition to neuroanatomical parallels,
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neurobehavioral outcomes observed in laboratory animals are similar to the cognitive and
behavioral outcomes observed in humans (Driscoll et al., 1990, Slawecki et al., 2004, Berman
and Hannigan, 2000).
The above information pertains to the qualitative effects of ethanol on health but does not
provide a dose-response perspective. Most epidemiology studies report alcohol intake as drinks
per day or per week. However, studies vary in their definitions of low, moderate, and high
intake. Laboratory animal studies generally express the oral intake of ethanol in units of grams
per kilogram body weight per day (g/kg bw-day) accompanied by an estimate of the blood
ethanol concentration (BEC). The BEC is generally presented in units of milligrams of ethanol
per deciliter (100 milliliters) of blood (mg/dL). The BEC can be a more reliable measure of
exposure than the amount of alcohol consumed, due to the large degree of variability in ethanol
metabolism and tolerance. However, epidemiological studies typically do not measure intake or
BEC in such a precise manner, and thus dose-response information may be uncertain. Note that
ethanol is also produced endogenously in the human body as a metabolic by-product of intestinal
flora, and is therefore always present in human blood at a low background concentration.
Longitudinal prospective studies on prenatal alcohol exposure and developmental toxicity
were performed in five areas of the United States (Atlanta, Cleveland, Detroit, Pittsburgh, and
Seattle). Study populations were followed for extended time periods ranging from 5 to 14 years
and were evaluated for measures of growth and physical development, cognitive ability,
academic achievement, attention, learning, memory, and adaptive behavior. These studies have
generally shown that maternal age (Jacobson et al., 1998) and trimester of pregnancy
(Streissguth et al., 1989, 1986a,b, Coles et al., 1992, 1991, 1987a,b, 1985) are important
considerations, with early pregnancy exposure of greatest concern. Sampson et al. (2000)
characterized the dose-response relationship for the Seattle cohort and reported that their analysis
suggests a linear dose-response for some of the alcohol-associated neurobehavioral effects, that
is, without any clear threshold down to the lowest nonzero levels of exposure. However, they
also noted that there was considerable variability in low-dose exposure data and that the slope of
the regression line appeared to be heavily influenced by higher-level exposures.
Only three prenatal inhalation exposure animal studies are known to have used multiple
air concentrations for ethanol (Nelson et al., 1985a,b, 1988). Exposure levels were relatively
high at 10,000, 16,000, and 20,000 ppm (7 hours/day throughout gestation) and resulted in peak
BECs of 3, 50, and 180 mg/dL, respectively. Although several reproductive and developmental
endpoints were examined (e.g., fertility, number of pups per litter, fetal body weight, incidence
of soft tissue or skeletal anomalies, tests of neuromotor coordination, activity level, and learning
ability in offspring), no clear treatment-related effects were seen. However, significant
biochemical changes were found in offspring of rats (paternal as well as maternal) that had been
exposed at air concentrations of 10,000 and 16,000 ppm (Nelson et al., 1988).
2. Cancer
Epidemiological data supporting the role of alcohol consumption in cancer development
in humans have been extensively reviewed (e.g., Boffetta and Hashibe, 2006; Gunzerath et al.,
2004; WHO, 2004; Poschl and Seitz, 2004; IARC, 1988) and support a clear association between
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low-to-moderate consumption of alcohol (2 drinks/day) and increased risk of cancers of the oral
cavity and pharynx, larynx, esophagus, and female breast. Several other studies suggest that low
to moderate consumption may also be associated with an increased risk of liver and colorectal
cancers (Boffetta and Hashibe, 2006; Corrao et al., 2004; Cho et al., 2004), although the data
supporting these findings are not as strong (as reviewed in Gunzerath et al., 2004; WHO, 2004).
Although an increase in risk of breast cancer associated with alcohol consumption is
modest, the increase has been demonstrated in a number of large, well-conducted prospective
cohort studies (e.g., Horn-Ross et al., 2002; Willett et al., 1987) as well as large case-control
studies (e.g., Longnecker et al., 1995).
Despite epidemiologic evidence relating breast cancer with alcohol consumption, animal
studies on the carcinogenic effect of ethanol have largely been inconclusive. However, most
animal studies suffer from a variety of limitations, most importantly the failure to use controls on
equivalent caloric intake diets. Of the few studies conducted using isocalorically fed controls,
Hackney et al. (1992) reported no effect of ethanol on mammary tumorigenesis, but this study
had a number of other methodological limitations. In a well-conducted chronic bioassay,
Holmberg and Ekstrom (1995) reported a significant increase in pituitary tumors in high-dose
females as well as a significant increase in benign tumors of the mammary glands in low-dose
females. Several animal studies using initiation-promotion or co-carcinogenicity protocols
report an increased incidence or multiplicity of mammary tumors with ethanol administration,
but also without a clear dose-response relationship (Singletary and Gapstur, 2001). However,
Hilakivi-Clarke et al. (2004) reported a dose-related increase in mammary tumors in rats exposed
to ethanol in utero and treated with an initiator within a few weeks of maturation. Finally, a
study by Soffritti et al. (2002) provides indications that early life exposure could be associated
with increased risk of oral cancers compared with exposure during adulthood. Rats treated with
10% ethanol in drinking water beginning in utero and continuing into adulthood had a
significantly increased incidence of tumors of the oral cavity, tongue, and lips, among other
carcinogenic effects relative to a similarly treated group of animals treated from age 39 weeks.
Genotoxicity data for ethanol have been reviewed in Phillips and Jenkinson (2001) and
IARC (1988). Ethanol was not mutagenic in bacterial or mammalian cell systems, but mixed
findings were reported in other test systems (S. cerevisiae, A. nidulans, and D. melanogaster).
Acetaldehyde, a metabolite of ethanol, has been observed to produce DNA adducts, point
mutations, SCE, and chromosome aberrations; and to interfere with DNA repair (reviewed in
Brooks and Theruvathu, 2005; Poschl and Seitz, 2004; Homann, 2001; Phillips and Jenkinson,
2001).
Finally, genetic factors may predispose some persons to greater susceptibility to cancer
from alcohol exposure. Polymorphisms have been reported to modify the risk of alcohol-
induced cancers, including both breast and oral cancers (Dumitrescu and Shields, 2005; Peters et
al., 2005).
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C. ETBE
Ethyl tertiary butyl ether (ETBE) has not been investigated to the extent that MTBE has,
but the limited information available on ETBE suggests that it is comparable to MTBE in various
respects. The chemical structures of ETBE (Figure II.l.c) and MTBE (Figure II. 1.a) are similar
except for ETBE having an ethoxy group (-O-CH2-CH3) instead of the methoxy group (-O-CH3)
in MTBE.
Like MTBE, the odor of ETBE can be detected at relatively low concentrations, with
thresholds of <1 ppm for ETBE odor and taste in both air and water (Vetrano, 1993). Limited
studies of odor thresholds suggest that ETBE may be detected at even lower concentrations than
MTBE (Vetrano, 1993) and that a 15% blend of ETBE in gasoline lowers the odor threshold well
below that for gasoline alone (TRC Env. Corp., 1993). At relatively high air concentrations (50
ppm), volunteer subjects reported throat and airway irritation for ETBE but not MTBE (Nihlen et
al., 1998a,b).
ETBE is currently under assessment as part of EPA's IRIS program. Although final
conclusions cannot be stated at this time regarding the qualitative or quantitative health hazard
potential of ETBE, some information on the health effects of ETBE can be summarized here
based on the available literature.
Although no chronic inhalation studies of ETBE have been conducted, inhalation studies
have been carried out with mice and rats for shorter periods of time. In a 4-week study by White
et al. (1995), the only treatment-associated responses in rats exposed to concentrations of 0 -
4000 ppm ETBE in air were increased liver weight at 2000 and 4000 ppm in females and
increased liver, kidney, and adrenal weight at 4000 ppm in males. Similar results were found by
Medinsky et al. (1999), who exposed rats and mice for 13 weeks to ETBE by inhalation and
found increased liver weight (with associated centrilobular hypertrophy) at 5000 ppm in female
rats, at 1750 and 5000 ppm in male rats, and at 1750 and 5000 ppm in male and female mice,
albeit in the absence of hepatic lesions, elevated serum enzymes characteristic of hepatotoxicity,
or other evidence of irregular pathology. Likewise, an increase in adrenal weight was seen in
both male and female rats at 5000 ppm but was not supported by histopathological findings.
In the same study, Medinsky et al. (1999) observed increased kidney weight at 1750 and
5000 ppm in male rats and at 5000 ppm in female rats; no change in kidney weight occurred in
mice. Female rats showed increased kidney weight unaccompanied by significant
histopathology and a small but unsustained increase in cell proliferation. Male rats showed
increased kidney weight associated with the presence of regenerative foci, hyaline droplet
accumulation, and alpha-2u-globulin immunoreactivity. This, along with other findings, raises
the question that was discussed in relation to MTBE, namely the role of alpha-2u-globulin in
accounting for nephropathy in male rats. This issue is undergoing evaluation by EPA in
connection with the IRIS assessment of ETBE.
In their 13-week inhalation study, Medinsky et al. (1999) also observed testicular effects,
namely a slight increase in the percentage of seminiferous tubules with degenerated
spermatocytes, at ETBE concentrations of 1750 and 5000 ppm. Neurotoxicological studies have
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found transient ataxia (staggering or unsteady gait) at air concentrations of 4000 to 5000 ppm
ETBE but no indications of long-term neurotoxicity (Dorman et al., 1997; White et al., 1995).
Few studies have examined the toxicity of ETBE by ingestion. Maltoni et al. (1999)
focused on the carcinogenicity of chronic oral-route exposure to ETBE administered in olive oil
at 0, 250, and 1000 mg/kg body weight. The authors reported increased incidence of malignant
uterine tumors in the low-dose group and increased incidence of pathologies of oncological
interest of the mouth in high-dose male rats. Uncertainties in the Maltoni et al. (1999) study
including the low-dose only effect of the uterine tumors and the combined inclusion of
precancers and tumors in the pathologies in the mouth contribute to uncertainty in the evaluation
of the results. Berger and Horner (2003) reported no effects on fertilizability of oocytes in
female rats exposed to ETBE or 2M2P (a metabolite of ETBE) in drinking water for 2 weeks.
Unpublished studies of ETBE administered to rats via the oral route at doses of 250, 500,
or 1000 mg/kg-day in corn oil evaluated reproductive, developmental, and other effects over two
generations (CIT, 2003, 2004a,b). At the two highest doses, significant decreases in body weight
gain in pregnant dams and some parental males were recorded but the effects were not observed
consistently in other rats exposed over a longer time period or in the subsequent generation.
Increased kidney weights and liver weights were observed in male and female rats, with more
pronounced effects in males, but the findings are difficult to interpret due to limited histological
evaluation. No effects were reported for fertility, gonadal function, reproductive performance,
parturition, lactation in the parental generations, and development of the offspring to weaning or
sexual maturity. The lack of testicular effects in particular contrasts with male reproductive
effects in the 13-week inhalation study of Medinsky et al. (1999) noted above. A transient
increase in salivation was also noted.
D. TAME
In terms of chemical structure (Figure Il.l.d), tertiary amyl methyl ether (TAME) and
MTBE (Figure II. 1 .a) are both ethers with a basic structure of an oxygen atom bonded to two
carbon atoms. TAME contains a methyl group on one side of the ether oxygen atom and a
tertiary amyl group on the other. The taste of TAME in water has been described as "highly
objectionable" by a panel of volunteer testers (Vetrano, 1993). Odor threshold studies suggest
that, as with ETBE and MTBE, mixing TAME with gasoline results in a much lower (by about
half) threshold compared to gasoline alone (TRC Environmental Corp., 1993). Also, compared
with odor thresholds for MTBE and ETBE, TAME appears to be intermediate.
TAME is currently under assessment as part of EPA's IRIS program; therefore, final
conclusions cannot be stated regarding its qualitative or quantitative health hazard potential.
Information on the health effects of TAME is limited, particularly for ingestion exposure. In a
subchronic oral study, rats were dosed with TAME for 7 days/week for 29 days (Daughtrey and
Bird, 1995). Treatment-related effects occurred at 1000 mg/kg-day with the deaths of 2 out of 10
animals and a decreased body weight gain in female rats. Absolute and relative adrenal weights
increased in females, and relative kidney weight increased in males at 500 and 1000 mg/kg-day.
A preliminary report by Belpoggi et al. (2002) evaluated the carcinogenic effects of chronic oral
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dosing rats 4 days/week for 78 weeks. The authors noted a statistically significant increase in the
number of total lymphomas and leukemias in TAME-treated females at 1000 mg/kg-day.
Inhalation studies of TAME include a subchronic study by White et al. (1995), who
investigated neurotoxicologic and other effects in rats inhaling TAME for 6 h/day, 5 days/week
for 4 weeks. The authors reported transient clinical signs (e.g., sedation, coma, ataxia) and other
indications of neurotoxicity (e.g., reduced tail pinch response, righting reflex, body temperature)
at 2000 ppm; however, none of the effects remained by 18 h after the last exposure. Other
transient effects noted at 2000 ppm included increased serum cholesterol in males and an
increased relative liver weight in both sexes unaccompanied by clear treatment-related
histopathological changes in the liver or in other tissues.
The American Petroleum Institute (API) funded a study conducted by Huntingdon Life
Sciences (1997) to evaluate effects of subchronic TAME inhalation exposures in rats and mice.
Rats inhaled concentrations of TAME ranging from 0 to 3500 ppm for 6 h/day, 5 days/week for
13-14 weeks. Clinical signs of neurotoxicity and hematological and clinical chemistry changes
occurred at 1500 ppm. Transient signs of neurotoxicity (lethargy and prostration) occurred
during and immediately after daily exposure for the first 4 weeks. Increased platelet counts and
serum protein (total protein, albumin, and globulin) levels persisted to the end of the 13- to 14-
week exposure period. Transient increases in liver and adrenal weights occurred at exposure
levels as low as 250 ppm but were not accompanied by histopathologic changes or increased
liver enzyme levels. Pathological changes related to alpha-2u-globulin accumulation in the
kidney were noted at and above 250-ppm exposures in male rats (see discussion regarding
alpha-2u-globulin in Section A on MTBE). In another portion of the study, mice inhaled
concentrations of TAME on the same regimen as the rats. High mortality rates (-56% male,
-30% female) occurred within three exposures at the 3500-ppm level. At 1500 ppm, clinical
signs of acute neurotoxicity such as lethargy and prostration occurred during and immediately
after daily exposure periods for the first 3 weeks. An effect on the liver (increased serum alanine
transaminase) occurred at 2500 ppm.
Reproductive and developmental effects of TAME in rodents were also investigated
under API sponsorship (Tyl et al., 2003; Welsch et al., 2003). Adult systemic toxicity was
evident at 1500 and 3000 ppm in rats, along with offspring toxicity at the same levels. However,
the only effects on reproductive function were at 3000 ppm in male rats (Tyl et al., 2003).
Developmental effects were more pronounced in mice than in rats and included increased
incidences of cleft palate at 1500 and 3500 ppm, along with increased late fetal deaths, reduced
fetal body weights, and enlarged lateral ventricles of the cerebrum at 3500 ppm (Welsch et al.,
2003).
E. DIPE
The chemical structure of di-isopropyl ether (DIPE) is shown in Figure II.I.e. DIPE
contains a secondary propyl group on each side of the oxygen atom. Odor and taste thresholds
for DIPE have not been investigated. Little toxicity information has been collected for DIPE,
and no chronic exposure studies have been conducted. DIPE is currently under assessment as
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part of EPA's IRIS program; therefore, final conclusions cannot be stated regarding its
qualitative or quantitative health hazard potential.
The primary study of DIPE toxicity was a subchronic inhalation study conducted for
Mobil and published in reports by Dalbey and Feuston (1996) and Rodriguez and Dalbey (1997).
Rats inhaled DIPE at concentrations ranging from 0 to 7100 ppm for 6 h/day, 5 days/week for 14
weeks (Dalbey and Feuston, 1996). At 3300 ppm, liver and kidney weights were slightly
increased, but no discernable morphological changes in liver and kidney were noted. Increased
liver weight with liver cell hypertrophy and increased kidney weight with increased incidences
of hyaline droplets in the proximal tubules occurred at 7100 ppm in males. This study also
evaluated developmental toxicity and found a concentration-related increase in the incidence of
rudimentary 14th ribs in the fetuses, although this effect was not supported by any other evidence
of teratogenicity. Dams showed a treatment-related, slightly decreased body weight gain and
decreased food intake at 6700 ppm.
Neurotoxic effects of DIPE were also evaluated (Rodriguez and Dalbey, 1997). At 450
ppm, males showed a decreased reflexive response to touch after two weeks of exposure, but this
response was not seen in other groups at any time point. A biphasic response was seen for
general activity levels in female rats, with decreased activity at week 4 and increased activity at
week 8. No treatment-related morphological changes were detected in the central or peripheral
nervous system structures examined.
Brooks et al. (1988) reported no positive responses in evaluations of DIPE in bacterial
mutation assays, a yeast mitotic gene conversion assay, and mammalian cell cultures for
structural chromosome damage.
F. TEA
Tertiary butyl alcohol (TEA) is a four-carbon alcohol with a methyl group and an alcohol
group (hydroxyl) attached to the second carbon of its three-carbon skeleton (Fig. Il.l.f), making
it miscible in water. In addition to being a fuel oxygenate additive and constituent of various
consumer products, TEA is a metabolite and a degradation by-product of MTBE. Odor detection
thresholds for TEA have been reported at roughly half the levels for ethanol (Amoore and
Hautala, 1983). TEA is currently under assessment as part of EPA's IRIS program; therefore,
final conclusions cannot be stated regarding its qualitative or quantitative health hazard potential.
Most of the information on the health effects of TEA comes from animal studies. Oral
and inhalation toxicity studies in rodents were conducted by the National Toxicology Program
(1995, 1997), including a chronic drinking water bioassay (Cirvello et al., 1995). Among the
more significant effects of long-term exposure to TEA in drinking water were kidney tumors in
male rats and thyroid tumors in female mice, along with dose-related nephropathy in male and
female rats and chronic inflammation and hyperplasia of the urinary bladder in male and female
mice (Cirvello et al., 1995). Subchronic TEA exposure also yielded signs of male rat
nephropathy and urinary tract inflammation and lesions in rats and mice of both sexes
(Lindamood et al., 1992). The issue of alpha-2u-globulin in male rat kidney effects arises here
as well (see discussion in Section A on MTBE).
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A few reproduction and/or developmental studies of TEA have been performed. In a
study with mice gavaged at a high dose (>1500 mg/kg-day) on gestational days 6-18, Faulkner et
al. (1989) found treatment-related increases in resorptions and reductions in fetal viability. In
another oral exposure study (Huntingdon Life Sciences, 2004), reductions in litter size, pup
survival, and neonatal body weight were seen in the offspring treated at 1000 mg/kg-day. Of two
inhalation studies of TEA developmental toxicity in rats (Nelson et al., 1989, 1991), one showed
reduced fetal body weight at 2000 ppm, albeit accompanied by some evidence of maternal
toxicity.
The pharmacokinetics of TEA have been extensively investigated, both as a parent
compound and as a metabolite of both MTBE and ETBE (e.g., Borghoff et al., 1996; Nihlen et
al., 1998a; Amberg et al., 2000; Prah et al., 2004; Blancato et al., 2007). In the latter regard, it
has not been determined to what extent, if any, the observed effects of MTBE and ETBE are
mediated by the metabolite they share. Unlike ethanol, TEA is not primarily metabolized by
alcohol dehydrogenase and is metabolized more slowly than ethanol (Thurman et al., 1980) or
the ether oxygenates (Prah et al., 1994; Nihlen et al., 1998a).
Some animal studies have suggested that TEA is more potent than ethanol in terms of
neurobehavioral effects such as psychomotor performance and learned behavior (Daniel and
Evans, 1982; Witkin andLeander, 1982; Dudek and Phillips, 1983; Witkin, 1984)
G. Alkylates/iso-octane
Alkylates are a class of organic chemicals, of which iso-octane is a leading example. The
chemical structure of the most prominent isomer of iso-octane (2,2,4-trimethylpentane) contains
no oxygen (Figure X.g), unlike the oxygenates in this discussion. Instead, this alkane comprises
a five-carbon chain (pentane) with three methyl groups attached, two to the second carbon and
one to the fourth carbon. Iso-octane is pertinent to this document because it (along with similar
branched-chain paraffins) could be used in increased amounts in some formulations of gasoline
using alcohols in place of ethers. As such it would help increase the volume of gasoline but
would not itself add oxygen to the fuel.
A ToxicologicalReview of2,2,4-Trimethylpentane (a synonym for iso-octane) was
recently posted on the EPA IRIS database (http://www.epa.gov/IRIS/toxreviews/0614-tr.pdf).
The Review indicates that insufficient information is available to allow quantification of
carcinogenic or non-carcinogenic risks for iso-octane. The following qualitative information is
summarized from the Toxicological Review of2,2,4-Trimethylpentane (U.S. EPA, 2007c).
Epidemiological or poisoning case studies of iso-octane in humans are not available, but
animal studies (mice, rats, guinea pigs) indicate that high levels (>8300 ppm) of iso-octane can
be lethal. Limited data in rats suggest that the chemical is readily absorbed via ingestion and
distributed to the kidneys, fat, and liver, with higher concentrations detected in the kidneys of
males compared with females (Kloss et al., 1986). Absorption via the respiratory tract appears to
be around 7-12% of the inhaled concentration (Dahl, 1989).
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Although no subchronic or chronic oral animal studies are available, a number of acute
and short-term oral studies point to altered renal function, an increase in alpha-2u-globulin
protein and hyaline droplet accumulation in the proximal tubules, necrosis of the tubule
epithelium, increased cell turnover, and foci of regenerative epithelium in male rats (Blumbach
et al., 2000; Saito et al., 1992; Borghoff et al., 1992; Burnett et al., 1989; Lock et al., 1987a,b;
Short et al., 1986; Stonard et al., 1986; API, 1983). However, no increases in alpha-2u-globulin
protein and hyaline droplet accumulation and no necrosis of the tubule epithelium were noted to
occur in female rats (Blumbach et al., 2000; Lock et al., 1987a,b) or in mice, guinea pigs, dogs,
or monkeys (Alden, 1986). In the only known subchronic study of iso-octane via the inhalation
route (Short et al., 1989) rats were exposed for 3-50 weeks to 50 ppm iso-octane. As with oral
studies, alpha-2u-globulin protein levels increased and hyaline droplets accumulated in males but
not females. Based on criteria described by U.S. EPA (1991a), it appears that the renal effects of
iso-octane are associated with alpha-2u-globulin accumulation and therefore are not relevant to
humans (U.S. EPA, 2007c). Note that this determination is made on a case-by-case basis and is
therefore specific to iso-octane; other chemicals (e.g., MTBE, ETBE, TEA) must be
independently evaluated regarding the role of alpha-2u-globulin.
One acute study also reported effects on the liver in a small number of male rats given
iso-octane by gavage for 2-3 days (Fowlie et al., 1987). Liver weights were significantly
increased, with centrilobular necrosis and hydrophobic degeneration of hepatocytes. No studies
are available on the reproductive or developmental effects of iso-octane. Limited genotoxicity
data are negative (Richardson et al., 1986; McLaren et al., 1994; Loury et al., 1986).
H. Associated chemicals and pollutants
This section summarizes health effects information for chemicals and pollutants that are
either constituents of fuels or are produced as a result of combustion and/or evaporation. BTEX
compounds (benzene, toluene, ethylbenzene and the xylenes) are components of fuels and enter
the atmosphere as evaporative emissions. The BTEX compounds can also contaminate ground
water from spills or when leaking occurs from underground fuel storage tanks, and their
concentrations are affected by presence of fuel additives (see Chapter V for more detail). Ozone,
particulate matter, acetaldehyde, peroxyacetyl nitrate (PAN), and formaldehyde are created in the
atmosphere as secondary products of fuel combustion or evaporation; particulate matter,
formaldehyde and acetaldehyde are also directly emitted as combustion products.
Concentrations are dependent on the fuel additive mixture (see Chapter IV for more detail).
1. Ozone
Tropospheric or ground-level ozone (Os) has been the subject of a great deal of scientific
study and evaluation. The following information is taken from Air Quality Criteria for Ozone
and Photochemical Oxidants (U.S. EPA, 2006), which provides a scientific basis for the
National Ambient Air Quality Standards for Os.
When found in the earth's troposphere (as opposed to the stratosphere, where it is
beneficial in shielding the Earth from harmful solar ultraviolet radiation), ozone generally
originates from photochemical reactions that involve the interaction of sunlight with precursor
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compounds, especially nitrogen oxides (NOx), carbon monoxide (CO), and volatile organic
compounds (VOCs) such as hydrocarbons emitted by various sources, including evaporative and
combustion emissions from motor vehicle fuels. Tropospheric Os is the primary constituent of
smog and can have adverse effects on humans, nonhuman animal species, and vegetation.
Inhalation of O?, can trigger a variety of health problems including chest pain, coughing,
throat irritation, and congestion. It can worsen bronchitis, emphysema, and asthma and can also
reduce lung function, inflame the linings of the lungs, and increase susceptibility to respiratory
infection. Repeated exposure to 63 may permanently scar lung tissue. In some people, these
effects can lead to increased medication use (among asthmatics), more frequent doctor visits,
school absences, and increased emergency room visits and hospital admissions. Moreover,
breathing ozone may contribute to premature death in people with heart and lung disease.
Ozone can also have harmful effects on plants and ecosystems. When sufficient Os
enters the leaves of a plant, it can interfere with the ability of sensitive plants to produce and
store food, making them more susceptible to certain diseases, insects, other pollutants,
competition and harsh weather. Further, 63 can visibly damage the leaves of trees and other
plants, harming the appearance of urban vegetation, national parks, and recreation areas.
Ultimately, the effects of Os may lead to reduced forest growth and crop yields.
Primary and secondary National Ambient Air Quality Standard (NAAQS) for ozone were
revised in 2008 to 0.075 parts per million (ppm) averaged over 8 hours for protection of both
public health and welfare effects, most notable those on vegetation.
2. Particulate Matter
The health effects of particulate matter (PM) have been intensively investigated and have
been characterized in depth in Air Quality Criteria for Particulate Matter (U.S. EPA, 2004a),
which provides a scientific basis for the PM National Ambient Air Quality Standards. Particles
in the ambient air come from a variety of sources and have a broad range of sizes, composition,
and other characteristics. Particles smaller than 2.5 //m in aerodynamic diameter (PM2.s) include
both directly emitted particles and particles that are products of chemical reactions of gases in
the atmosphere. Examples of directly emitted particles (both PM2.5, or "fine" particles, and
larger, or "coarse," particles with aerodynamic diameter of 2.5 - 10 //m) include those from
combustion sources such as vehicle tailpipe emissions, agricultural open burning, coal and oil
fired power plants and industries, as well as dust particles from roads and fields. Particles
formed in the atmosphere are referred to as "secondary" particles, and the majority of these
particles in many areas of the country are formed from gases from fuel combustion in
automobiles, trucks, and power plants such as sulfur dioxide (SO2) and nitrogen oxides (NOx)
released by anthropogenic and natural sources. In addition, aromatic compounds in fuels, such
as toluene and even benzene, have been shown to play an important role in secondary PM
formation. Other hydrocarbon components can also form secondary organic aerosol in
substantial amounts. These include long-chained alkanes as well as biogenic isoprene,
monoterpenes, and sesquiterpenes. These biogenic compounds can, especially in summer
conditions, form tremendous quantities of secondary organic aerosol.
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Particle pollution, especially fine particles, contains microscopic solids or liquid droplets
that are so small that they can get deep into the lungs and cause serious health problems.
Numerous scientific studies have linked particle pollution exposure to a variety of health
problems, including increased respiratory symptoms, such as irritation of the airways, coughing,
or difficulty breathing; decreased lung function; aggravated asthma; development of chronic
bronchitis; irregular heartbeat; nonfatal heart attacks; and premature death in people with heart or
lung disease. People with heart or lung diseases, children and older adults are the most likely to
be affected by particle pollution exposure. However, even healthy persons may experience
temporary symptoms from exposure to elevated levels of particle pollution. The PM Air Quality
Criteria Document also notes that the PM components of gasoline and diesel engine exhaust
represent one class of hypothesized likely important contributors to the observed ambient PM-
related increases in lung cancer incidence and mortality. Environmental effects of particulate
matter can be far-reaching. Fine particles (PM2.5) are the major cause of reduced visibility (haze)
in parts of the United States, including many national parks and wilderness areas. Particles can
also be carried over long distances by wind and then settle on ground or water. The effects of
this settling include acidification of lakes and streams, alteration of the nutrient balance in
coastal waters and large river basins, depletion of nutrients in soil, damage to sensitive forests
and farm crops, and reduction of the diversity of ecosystems. In addition, particle pollution can
stain and damage stone and other materials, including culturally important objects such as statues
and monuments.
The current National Ambient Air Quality Standards (NAAQS) for particulate matter, as
revised in 2006, are expressed for different averaging times and for different size fractions as
follows: PMio, 150 //g/m3 over 24 hours; PM2.5, 15.0 //g/m3 annually and 35 //g/m3 over 24
hours.
3. Acetaldehyde
While acetaldehyde is an emission product from combustion engines running on ethanol-
free fuels, it is emitted in increased amounts in the exhaust of vehicles burning gasoline/ethanol
blends. Acetaldehyde is also produced in substantial amounts in the atmosphere resulting from
photochemical reactions of ethanol and other hydrocarbons. Moreover, some atmospheric
acetaldehyde undergoes further photochemical reactions in the atmosphere to yield PAN (see
below). In addition, this compound is an important metabolite of ethanol in the body.
In humans, acute exposure to acetaldehyde vapor can cause eye irritation at a
concentration of 50 ppm for 15 minutes or as low as 25 ppm in sensitive individuals
(Verschueren, 1983). Dermal exposures in humans may result in contact dermatitis, erythema,
or burns, and in rabbits acute exposure has also been reported to cause mild skin and severe eye
irritation (ACGIH, 1991).
An inhalation reference concentration (RfC) of 0.009 mg/m3 was derived for
acetaldehyde in 1991 (U.S. EPA, 1991b). The RfC was based primarily on degeneration of
olfactory epithelium in rats that had been exposed to acetaldehyde for 4 weeks (Appleman et al.,
1982, 1986), consistent with effects seen at exposures lasting up to 52 weeks (Woutersen et al.,
1986; Woutersen and Feron, 1987). No information on reproductive or developmental effects
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was found at the time of the 1991 review. Some asthmatics have been shown to be a sensitive
subpopulation to decrements in forced expiratory volume and increased bronchoconstriction
upon acetaldehyde inhalation (Myou et al., 1993).
Acetaldehyde has been classified by EPA as a probable human carcinogen based on
animal studies (U.S. EPA, 1991b). Nasal epithelial cancers occurred in rats exposed by
inhalation for up to 28 months (U.S. EPA, 1991b), and hamsters exposed for 52 weeks had
significant increases in laryngeal tumors (IARC, 1985).
Note that acetaldehyde is currently under assessment as part of EPA's IRIS program;
therefore, conclusions are being revised regarding its qualitative or quantitative health hazard
potential.
4. PAN
Peroxyacetyl nitrate (PAN) is a photochemical oxidant associated with ground-level
ozone and smog. Information on the health effects of PAN is limited, but it appears to be less
toxic than ozone at comparable acute exposure levels (U.S. EPA, 2006 [Ozone Criteria
Document]). At environmental levels it can cause eye irritation, and at concentrations
approximately 100- to 1,000-fold higher than levels typically found in ambient air, it can cause
changes in lung morphology, behavioral modifications, weight loss, and susceptibility to
pulmonary infections. Cytogenetic studies indicate that PAN is not a potent mutagen, clastogen,
or DNA-damaging agent in mammalian cells in vivo or in vitro at concentrations several orders
of magnitude higher than those generally encountered in ambient air in most cities. Some studies
suggest that PAN may be a weak bacterial mutagen at concentrations much higher than exist in
present urban atmospheres.
5. Formaldehyde
Formaldehyde is the most prevalent aldehyde in engine exhaust. In addition, while it is
not a constituent of gasoline and thus is not a component of evaporative emissions, it is also
formed photochemically in the atmosphere from precursor hydrocarbons present in motor
vehicle exhaust. The primary route of human exposure is by inhalation.
Since 1987, EPA has classified formaldehyde as a probable human carcinogen based on
evidence in humans and in rats, mice, hamsters, and monkeys (U.S. EPA, 1987). EPA's IRIS
database provides an upper bound cancer unit risk estimate of 1.3xlO"5 per |ig/m3 (U.S. EPA,
1991d). In other words, there is an estimated risk of as many as thirteen excess leukemia cases
in one million people exposed to 1 |ig/m3 of formaldehyde over a lifetime.
In the past 15 years there has been substantial research on the inhalation dosimetry for
formaldehyde in rodents and primates by the CUT Centers for Health Research (formerly the
Chemical Industry Institute of Toxicology), with a focus on use of rodent data for refinement of
the quantitative cancer dose-response assessment (Conolly et al., 2003, 2004; CUT, 1999).
CIIT's risk assessment of formaldehyde incorporated mechanistic and dosimetric information on
formaldehyde. The risk assessment analyzed carcinogenic risk from inhaled formaldehyde using
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approaches that were consistent with the 1999 draft EPA guidelines for carcinogenic risk
assessment. In 2001, Environment Canada relied on this cancer dose-response assessment in
their assessment of formaldehyde (Health Canada, 2001). In 2004, EPA also relied on this
assessment during the development of the plywood and composite wood products national
emissions standards for hazardous air pollutants (NESHAPs) (U.S. EPA, 2004b). In these rules,
EPA concluded that the CUT work represented the best available application of the available
mechanistic and dosimetric science on the dose-response for portal of entry cancers due to
formaldehyde exposures. The value that EPA used for these assessments was 5.5 x 10"9/|ig3.
In 2004 the International Agency for Research on Cancer concluded that formaldehyde is
carcinogenic to humans (Group 1 classification), on the basis of sufficient evidence in humans
and sufficient evidence in experimental animalsa higher classification than previous IARC
evaluations. In addition, the National Institute of Environmental Health Sciences recently
nominated formaldehyde for reconsideration as a known human carcinogen under the National
Toxicology Program. Since 1981 it has been listed as a "reasonably anticipated human
carcinogen." More recently the German Federal Institute for Risk Assessment classified
formaldehyde as a known human carcinogen (BfR, 2006).
EPA is currently reviewing the health literature on formaldehyde, including several
recently published epidemiological studies. For instance, research conducted by the National
Cancer Institute (NCI) found an increased risk of nasopharyngeal cancer and
lymphohematopoietic malignancies such as leukemia among workers exposed to formaldehyde
(Hauptmann et al, 2003, 2004). NCI is currently performing an update of these studies. A recent
National Institute of Occupational Safety and Health (NIOSH) study of garment workers also
found increased risk of death due to leukemia among workers exposed to formaldehyde
(Pinkerton, 2004). Extended follow-up of a cohort of British chemical workers did not find
evidence of an increase in nasopharyngeal or lymphohematopoeitic cancers, but a continuing
statistically significant excess in lung cancers was reported (Coggon et al., 2003). EPA is
reviewing the recent work cited above from the NCI and NIOSH, as well as the analysis by the
CUT Centers for Health Research and other studies (e.g., Subramaniam et al., 2007), as part of a
reassessment of the human hazard and dose-response associated with formaldehyde.
Formaldehyde exposure also causes a range of noncancer health effects, including
irritation of the eyes, nose and throat. Decreased pulmonary function has been observed in
humans. Effects from repeated exposure in humans include respiratory tract irritation, chronic
bronchitis and nasal epithelial lesions (ATSDR, 1999b). Animal studies suggest that
formaldehyde may also cause airway inflammation - including eosinophil infiltration into the
airways (Coon et al., 1970; Fujimaki et al., 2004; Jung et al., 2007). There are several studies
suggesting that formaldehyde (up to about 3 ppm) may induce mild, reversible changes in human
pulmonary function (Green et al., 1987; Sauder et al., 1986; Horvath et al., 1988). Other studies
indicate that formaldehyde exposure may increase the risk of asthma - particularly in the young
(Franklin et al., 2000; Garret et al., 1999; Rumchev et al., 2002). However, EPA has not
developed an RfC for formaldehyde to date.
6. Benzene
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Benzene is an aromatic hydrocarbon that is a major constituent of gasoline, along with
toluene, ethylbenzene, and xylenes, collectively referred to as BTEX. As such, it is emitted into
the air both as exhaust and evaporative emissions. Besides being emitted in the exhaust as an
uncombusted gasoline component, benzene is also emitted in the exhaust as a combustion
product of other aromatic compounds. Also, it is emitted in evaporative emissions from the fuel
system. It can be emitted as evaporative emissions while the vehicle is running or even after it is
shut off. It is also emitted as refueling emissions and as fuel spillage. Inhalation is the most
common route of human exposure to benzene in the occupational and non-occupational setting;
however, leakage of gasoline from underground storage tanks into ground water has also posed a
source of exposure. Benzene has well-documented carcinogenic and other adverse health effects
in humans and laboratory animals.
The EPA's IRIS database lists benzene as a known human carcinogen (causing leukemia)
by all routes of exposure, based on convincing evidence in humans and supporting evidence
from animal studies (U.S. EPA, 2000). Epidemiologic studies and case studies provide clear
evidence of a causal association between exposure to benzene and acute nonlymphocytic
leukemia and also suggest evidence for chronic nonlymphocytic leukemia and chronic
lymphocytic leukemia. Other neoplastic conditions that are associated with an increased risk in
humans are hematologic neoplasms, blood disorders such as preleukemia and aplastic anemia,
Hodgkin's lymphoma, and myelodysplastic syndrome (MDS). These human data are supported
by animal studies. The experimental animal data add to the argument that exposure to benzene
increases the risk of cancer in multiple species at multiple organ sites (hematopoietic, oral and
nasal, liver, forestomach, preputial gland, lung, ovary, and mammary gland). It is likely that
these responses are due to interactions of the metabolites of benzene with DNA (Ross, 1996;
Latriano et al., 1986). Recent evidence supports the viewpoint that there are likely multiple
mechanistic pathways leading to cancer and, in particular, to leukemogenesis from exposure to
benzene (Smith, 1996). Exposure to benzene and/or its metabolites has also been linked with
chromosomal changes in humans and animals and increased proliferation of mouse bone marrow
cells (U.S. EPA, 1998b; Irons et al., 1992)
The excess risk of developing leukemia from inhalation exposure to benzene has been
estimated at 2.2 x 10"6 to 7.8 x 10"6 per |ig/m3. In other words, there is an estimated risk of about
two to eight excess leukemia cases in one million people exposed to 1 |ig/m3 of benzene over a
lifetime3. Based on the cancer assessment for benzene by inhalation, a cross-route extrapolation
yielded an oral unit risk estimate range of 4.4 x 10"7 to 1.6 x 10"6/|ig/L (U.S. EPA, 2000). These
estimates pertain to ingestion of drinking water but do not account for total exposure that could
include dermal uptake during bathing and inhalation of volatilized benzene from drinking water.
Data that were considered by EPA in its carcinogenic update suggested that the dose-response
relationship at doses below those examined in the studies reviewed in EPA's most recent
3 This range of unit risks reflects the maximum likelihood estimates (MLEs) calculated from different
exposure assumptions and dose-response models that are linear at low doses. At present, the true cancer risk from
exposure to benzene cannot be ascertained, even though dose-response data are used in the quantitative cancer risk
analysis, because of uncertainties in the low-dose exposure scenarios and lack of clear understanding of the mode of
action. A range of estimates of risk is recommended, each having equal scientific plausibility. There are confidence
intervals associated with the MLE range that reflect random variation of the observed data. For the upper end of the
MLE range, the 5th and 95th percentile values are about a factor of 5 lower and higher than the best fit value.
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benzene assessment may be supralinear4. This relationship could support the inference that
cancer risks are as high, or higher than the estimates provided in the existing EPA assessment.
However, since the mode of action for benzene carcinogenicity is unknown, the current cancer
unit risk estimate assumes linearity of the low-dose response.
In addition to carcinogenic effects, a number of adverse noncancer health effects,
particularly blood disorders and immunotoxicity, have been associated with long-term
occupational exposure to benzene (U.S. EPA, 2003a). People with long-term occupational
exposure to benzene have experienced harmful effects on the blood-forming tissues, especially in
the bone marrow. These effects can disrupt normal blood production and suppress the
production of important blood components, such as red and white blood cells and blood platelets,
leading to anemia (a reduction in the number of red blood cells), leukopenia (a reduction in the
number of white blood cells), or thrombocytopenia (a reduction in the number of blood platelets,
thus reducing the ability of blood to clot). Chronic inhalation exposure to benzene in humans
and animals results in pancytopenia,5 a condition characterized by decreased numbers of
circulating erythrocytes (red blood cells), leukocytes (white blood cells), and thrombocytes
(blood platelets). Individuals that develop pancytopenia and have continued exposure to benzene
may develop aplastic anemia, whereas others exhibit both pancytopenia and bone marrow
hyperplasia (excessive cell formation), a condition that may indicate a preleukemic state. The
most sensitive noncancer effect observed in humans, based on current data, is the depression of
the absolute lymphocyte count in blood.
Childhood may be a stage of life where individuals are at increased risk from benzene
exposure. Limited data suggest an increased risk to children whose parents were occupationally
exposed to benzene (Shu et al., 1988; McKinney et al., 1991), supported by animal studies
indicating that benzene exposures resulted in damage to the hematopoietic (blood cell formation)
system during development (Keller and Snyder, 1986, 1988; Corti and Snyder, 1996).
The EPA inhalation reference concentration (RfC) for benzene is 30 |ig/m3 (U.S. EPA,
2003a). A total uncertainty factor of 300 (10 for intraspecies variability, 3 for interspecies
4 Recent data on benzene adducts in humans, published after the most recent IRIS assessment, suggest that
the enzymes involved in benzene metabolism start to saturate at exposure levels as low as 1 ppm (Rappaport et al.,
2002; 2005; Lin et al., 2007). These data highlight the importance of ambient exposure levels and their contribution
to benzene-related adducts. Because there is a transition from linear to saturable metabolism below 1 ppm, the
assumption of low-dose linearity extrapolated from much higher exposures could lead to substantial underestimation
of leukemia risks. This is consistent with recent epidemiological data which also suggest a supralinear exposure-
response relationship and which "[extend] evidence for hematopoietic cancer risks to levels substantially lower than
had previously been established (Hayes et al., 1997; 2001; Lan et al., 2004). These data are from the largest cohort
study done to date with individual worker exposure estimates. However, these data have not yet been formally
evaluated by EPA as part of the IRIS review process, and it is not clear how they might influence low-dose risk
estimates. A better understanding of the biological mechanism of benzene-induced leukemia is needed.
5 Pancytopenia is the reduction in the number of all three major types of blood cells (erythrocytes, or red
blood cells, thrombocytes, or platelets, and leukocytes, or white blood cells). In adults, all three major types of
blood cells are produced in the bone marrow of the skeletal system. The bone marrow contains immature cells,
known as multipotent myeloid stem cells, that later differentiate into the various mature blood cells. Pancytopenia
results from a reduction in the ability of the red bone marrow to produce adequate numbers of these mature blood
cells.
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extrapolation, 3 for database deficiencies, 3 for subchronic-to-chronic extrapolation) was used in
deriving the RfC, with an overall confidence rating of medium. The RfC is based on suppressed
absolute lymphocyte counts seen in humans under occupational exposure conditions (Rothman et
al., 1996). The oral reference dose (RfD) for benzene was derived through route-to-route
extrapolation of the results of benchmark dose (BMD) modeling of the absolute lymphocyte
count data from the occupational epidemiologic study by Rothman et al. (1996), in which
workers were exposed to benzene by inhalation. Based on these and other analyses, an RfD of
4.0 x lo'mg/kg/day was derived, with a total uncertainty factor of 300 and overall confidence
rating of medium (U.S. EPA, 2003a). In addition, the Agency for Toxic Substances and Disease
Registry Minimal Risk Level for acute exposure to benzene is 160 |ig/m3 for 1-14 days exposure.
Since development of this RfC, there have appeared reports in the medical literature of
benzene's hematotoxic effects in humans that provide data suggesting a wide range of
hematological endpoints that are triggered at occupational exposures of less than 5 ppm (about
16 mg/m3) (Qu et al., 2002) and at air levels of 1 ppm (about 3 mg/m3) (Lan et al., 2004) or less
among genetically susceptible populations. These studies had large sample sizes and extensive
individual exposure monitoring. One recent study found benzene metabolites in mouse liver and
bone marrow at environmental doses, indicating that even concentrations in urban air may elicit
a biochemical response in rodents that indicates toxicity (Turtletaub and Mani, 2003). EPA has
not evaluated these recent studies as part of the IRIS review process to determine what role they
might play in any future revisions to the RfC.
7. Toluene
Toluene is found in evaporative as well as exhaust emissions from motor vehicles, and
can be a significant contributor to secondary PM. Other aromatic compounds contribute to
secondary organic PM as well (such as benzene), but the extent of their contribution has not been
quantified. As a major constituent of gasoline, toluene (in BTEX) may be found in ground water
due to leaking fuel storage tanks. Thus, exposure to toluene may occur by inhalation, ingestion,
and dermal contact.
The health effects of toluene were recently described in a Toxicological Review by EPA
(2005). The central nervous system (CNS) is the primary target for toluene toxicity in both
humans and animals for acute and chronic exposures. CNS dysfunction (which is often
reversible) and narcosis have been frequently observed in humans acutely exposed to low or
moderate levels of toluene by inhalation; symptoms include fatigue, sleepiness, headaches, and
nausea. Central nervous system depression has been reported to occur in chronic abusers
exposed to high levels of toluene. Symptoms include ataxia, tremors, cerebral atrophy,
nystagmus (involuntary eye movements), and impaired speech, hearing, and vision. Chronic
inhalation exposure of humans to toluene also causes irritation of the upper respiratory tract, eye
irritation, dizziness, headaches, and difficulty with sleep.
Human studies have also reported developmental effects, such as CNS dysfunction,
attention deficits, and minor craniofacial and limb anomalies, in the children of women who
abused toluene during pregnancy. A substantial database examining the effects of toluene in
subchronic and chronic occupationally exposed humans exists. The weight of evidence from
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these studies indicates neurological effects (i.e., impaired color vision, impaired hearing,
decreased performance in neurobehavioral analysis, changes in motor and sensory nerve
conduction velocity, headache, dizziness) as the most sensitive endpoints. The data from these
human studies were found to be sufficient for EPA to develop an RfC of 5 mg/m3 for toluene
exposure. The total uncertainty factor for the RfC is 10 (interspecies extrapolation) and the
overall confidence in this RfC is high.
Data on the effects of toluene in humans following oral exposure are limited to case
reports of accidental oral ingestions. A subchronic study of oral exposure to toluene in rodents
(rats and mice) showed significant (p<0.05) increases in absolute and relative weights of both the
liver and kidney in male rats at doses greater than or equal to 446 mg/kg-day (NTP, 1990).
Based on this study, EPA derived an RfD of 80 jig /kg-day for toluene oral exposure. The total
uncertainty factor for the RfD is 3000 (10 for interspecies extrapolation, 10 for intraspecies
variation, 10 for subchronic-to-chronic extrapolation and 3 for database insufficiencies and
contradictions in immunotoxicity data) and the overall confidence in the RfD is medium.
Under the 2005 Guidelines for Carcinogen Risk Assessment, there is inadequate
information to assess the carcinogenic potential of toluene (U.S. EPA, 2005). Toluene was not
found to be carcinogenic in inhalation cancer bioassays of rats and mice exposed for life, but
increased incidences of mammary cancer and leukemia were reported in a lifetime rat oral
bioassay. Toluene has generally not been found to be genotoxic in short-term testing.
8. Ethylbenzene
Ethylbenzene is one of the BTEX compounds that are the major constituents of gasoline.
It is found in both exhaust and evaporative emissions from gasoline-powered vehicles and may
also be found in ground water due to leaking fuel storage tanks. Thus, exposure to ethylbenzene
may occur by inhalation, ingestion, and dermal contact.
Limited information is available on the carcinogenic effects of ethylbenzene in humans
and animals. Under the 1987 Cancer Guidelines, EPA (1991c) classified ethylbenzene as a
Group D carcinogen, meaning it is not classifiable as a human carcinogen. This classification
was the result of the lack of animal bioassays or human data to evaluate its carcinogenic potential
at that time. In 1999, the National Toxicology Program (NTP) published the results of a 2-year
chronic bioassay that concluded that there was clear evidence of ethylbenzene carcinogenicity in
male rats (renal and testicular), and some evidence in female rats and both sexes of mice (NTP,
1999). These results, in the context of the full health science literature for ethylbenzene, are
being reviewed by EPA in its ongoing review of the IRIS file for ethylbenzene.
Chronic exposure to ethylbenzene by inhalation in humans may result in effects on the
hematological system, kidney and liver, and endocrine system (California EPA, 2007). No
information is available on the developmental or reproductive effects of ethylbenzene in humans,
although animal studies have reported developmental effects via inhalation. The data from these
studies were found to be sufficient for EPA to develop an RfC of 1 mg/m3 for ethylbenzene
exposure by inhalation (U.S. EPA, 1991c). The total uncertainty factor used was 300 (10 for
intraspecies variation, 3 for interspecies extrapolation, 10 database insufficiencies), and
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confidence in the RfC is considered low because no chronic studies or multi-generational
developmental studies were available at the time. An animal study has reported effects on the
blood, liver, and kidneys from oral exposure to ethylbenzene (Wolf et al., 1956). The data from
this study were found to be sufficient for EPA to develop an RfD for oral ethylbenzene exposure
of 100 |ig/kg-day (U.S. EPA, 1991c). The total uncertainty factor was 1000 (10 for intraspecies
variability, 10 for interspecies extrapolation, 10 for subchronic-to-chronic extrapolation), and
confidence in this RfD is considered low because rats (single species) of only one sex were
tested, no chronic studies were then available, and no other supporting oral toxicity data were
found. Since the 1991 review, the NTP (1999) completed a chronic animal bioassay, as noted
above; the relevance of the results from this study in reviewing the RfC will be evaluated in
EPA's ongoing IRIS assessment. In addition, more recent studies characterizing the toxicity of
ethylbenzene by the oral route are being considered in the ongoing review of the IRIS
assessment (e.g., Mallerat et al., 2007).
With respect to short-term inhalation exposures, ATSDR has derived a 10 ppm Minimal
Risk Level for exposure to ethylbenzene for 14 days or less, based on effects of the auditory
system in animals. (ATSDR, 1999a; Cappaert et al., 1999, 2000, 2001).
9. Xylenes
Mixed xylenes are blended into gasoline as part of the BTEX compounds. Xylenes are
emitted in the exhaust and evaporative emissions of gasoline-powered engines and may also be
found in ground water due to leaking fuel storage tanks. Thus, exposure to xylenes may occur
by inhalation, ingestion, and dermal contact.. Xylenes, along with other aromatics, may be
contributors to secondary organic PM as toluene and even benzene have been shown to be.
The health effects of xylenes were described in a Toxicological Review (U.S. EPA,
2003b). Chronic inhalation exposure in humans to mixed xylenes results primarily in central
nervous system effects, such as headache, nausea, fatigue and also included eye and nose
irritation and sore throat. Animal studies have reported developmental effects, such as an
increased incidence of skeletal variations in fetuses, and fetal resorptions via inhalation. EPA
(2003b) developed an RfC of 100 ug/m3 for xylenes based on impaired motor coordination in
rats. The total uncertainty factor is 300 (10 for interspecies extrapolation, 10 for intraspecies
variability, 10 for database deficiencies), and the confidence rating assigned to the RfC for
xylenes is medium. Data from animal oral exposure studies showing decreased body weight and
increased mortality were found to be sufficient for EPA (2003b) to develop an RfD of 200
|ig/kg-day for oral xylene exposure. The RfD included a total uncertainty factor of 1000 and was
assigned an overall confidence rating of medium.
Acute inhalation exposure to mixed xylenes in humans results in irritation of the nose and
throat, gastrointestinal effects such as nausea, vomiting, and gastric irritation, mild transient eye
irritation, and neurological effects (U.S. EPA, 2003b).
Inadequate information is available on the carcinogenic effects of mixed xylenes in
humans, and animal studies have been inconclusive. Thus, data are inadequate for an assessment
of the carcinogenic potential of xylenes (U.S. EPA, 2003b).
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III. COMMERCIAL FEASIBILITY OF ETHANOL AND OTHER ADDITIVES
This chapter focuses primarily on ethanol, a renewable fuel blending component that is
currently the predominant product blended into gasoline in the U.S. marketplace and is expected
to continue to be for the foreseeable future. Several factors have lead to the rapid expansion of
ethanol use. With passage of the Energy Policy Act of 2005 (EPAct), two important provisions
affecting ethanol use occurred. First, EPAct removed the oxygen requirement in the federal
reformulated gasoline (RFG) program. This allowed refiners and blenders to produce RFG with
or without an oxygenate. In areas where MTBE was still being blended into RFG, refiners
shifted away from the use of MTBE and began blending ethanol into RFG. Additionally,
passage of the Renewable Fuels Standard in EPAct 2005 required increased volumes of
renewable fuel be used in the transportation sector beginning in 2006, requiring a minimum of 4
billion gallons per year and increasing to 7.5 billion gallons per year in 2012. This, along with
positive economics for the use and blending of ethanol, saw the market exceed the mandated
volumes. Looking forward, with passage of the Energy Independence and Security Act of 2007
which established new market requirements for renewable fuels, ethanol use will continue to see
a rapid rise in use in the gasoline or alternative (E85) fuel pool.
Also in this chapter, information is provided on where ethanol is currently produced in
the U.S. Finally, other additives such as ETBE, TAME, DIPE, and TEA are discussed. These
currently constitute a very small portion of the additive market and given their historical use and
other policy and market realities, it does not appear that these products will attain commercially
significant volumes.
A. Ethanol
1. Where Ethanol is Produced
As of September 1, 2007, there were 133 fuel ethanol plants operating in the United
States with a combined production capacity of approximately 7 billion gallons per year.6 These
numbers have increased and will to do so in the coming years to support attaining the EISA
mandated renewable fuel volumes. Most ethanol currently produced in the US at commercial
levels comes from grain or starch-based feedstocks that can easily be broken down via traditional
fermentation processes. The majority of domestic ethanol is currently produced in the Midwest
within PADD (Petroleum Administration for Defense District) 2 - where most of the corn
feedstock is grown. In 2007, of the 133 ethanol production facilities, 119 were located in PADD
2. As a region, PADD 2 accounted for almost 96 percent (or about 6.7 billion gallons) of the
total production capacity in 2007, as shown in Table III.A-1. Figure III.A-1 shows the locations
of each of the five PADDs.
6 The September 1, 2007 ethanol production capacity baseline was generated based on a variety of data sources
including: Renewable Fuels Association (RFA), Ethanol Biorefinery Locations (updated August 22, 2007); Ethanol
Producer Magazine (EPM), Current plant list (last modified on August 13, 2007) and ethanol producer websites.
The baseline includes does not include ethanol plants whose primary business is industrial or food-grade ethanol
production. Where applicable, current ethanol plant production levels have been used to represent plant capacity, as
nameplate capacities are often underestimated. The baseline does not include idled plants located in Bartow, FL;
Hamburg, IA; or Plover, WI nor does it include plants that may be located in the Virgin Islands or U.S. territories.
49
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Table III.A-1. 2007 U.S. Ethanol Production Capacity by PADD
PADD
PADD1
PADD 2
PADD 3
PADD 4
PADD 5
Total
Capacity
MGY
0.4
6,679
30
108
170
6,987
%of
Capacity
0.0%
95.6%
0.4%
1 .5%
2.4%
100.0%
No. of
Plants
1
119
1
6
6
133
%of
Plants
0.8%
89.5%
0.8%
4.5%
4.5%
100.0%
Figure III.A-1. PADD Boundary and Regions
Petroleum Administration
for Defense Districts
5:
West
AK, Hf
4;
PA-DD 2:
MMwesf
1;
Coast
3:
Leading the Midwest in ethanol production are Iowa, Nebraska and Illinois followed by
South Dakota and Minnesota whose production levels are nearly equal. Together, as of
September 2007, these five states were home to 84 ethanol plants with a combined production
capacity of 5.2 billion gallons per year (contributing to 74 percent of the total domestic product).
Although the majority of ethanol production comes from the Midwest, there are a growing
number of plants located outside the traditional corn belt. In addition to the 15 states comprising
PADD 2, (all of which have operational ethanol plants besides Ohio7), in 2007, ethanol
production facilities in the contiguous states were located in Arizona, California, Colorado,
Georgia, Idaho, New Mexico, Oregon and Wyoming. Some of these facilities ship in feedstocks
(primarily corn) from the Midwest, others rely on locally grown/produced feedstocks, while
others rely on a combination of both.
7 Ohio does not currently have any operational ethanol plants; however, six plants are under construction.
50
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The U.S. ethanol industry is currently comprised of a mixture of company-owned and
locally-owned farmer cooperatives (co-ops). More than two-thirds (93) of ethanol plants are
owned by corporations, and, on average, these plants are larger in size than farmer-owned co-
ops. Accordingly, company-owned plants account for more than three-quarters of the total 7007
U.S. ethanol production capacity. Furthermore, 30 percent of the total domestic product came
from 27 plants owned by just two different companies - Archer Daniels Midland and POET
Biorefining (formerly Broin).
. A summary of ethanol production alphabetically by state is found in Table III. A-2.
Table III.A-2. Summary of Ethanol Production by State, 2007
State
Arizona
California
Colorado
Georgia
Iowa
Idaho
Illinois
Indiana
Kansas
Kentucky
Michigan
Minnesota
Missouri
North Dakota
Nebraska
New Mexico
Oklahoma
Oregon
South Dakota
Tennessee
Wisconsin
Wyoming
Total
Capacity
MMgy
50
80
96
0.4
2,031
3
887
300
289
38
214
634
160
136
997
30
2
40
636
67
289
9
6,987
%of
Capacity
0.7%
1.1%
1 .4%
0.0%
29.1%
0.0%
12.7%
4.3%
4.1%
0.5%
3.1%
9.1%
2.3%
1 .9%
14.3%
0.4%
0.0%
0.6%
9.1%
1 .0%
4.1%
0.1%
100.0%
No. of
Plants
1
4
4
1
29
1
8
4
9
2
4
16
4
4
18
1
1
1
13
1
6
1
133
%of
Plants
0.8%
3.0%
3.0%
0.8%
21 .8%
0.8%
6.0%
3.0%
6.8%
1 .5%
3.0%
12.0%
3.0%
3.0%
13.5%
0.8%
0.8%
0.8%
9.8%
0.8%
4.5%
0.8%
100.0%
In addition to the domestic ethanol production described above, the U.S. also receives a
small amount of ethanol from other countries.
51
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2. Projected Growth in Ethanol Production
Over the past 25 years, domestic ethanol production has steadily increased due to
environmental regulation, federal and state tax incentives, and market demand. More recently,
ethanol production has soared due to the phase-out of MTBE, an increasing number of state
ethanol mandates, and elevated crude oil prices. As shown in Figure III.A-2, over the past five
years, domestic ethanol production has almost tripled from 1.8 billion gallons in 2001 to 4.7
billion gallons in 2006.
Figure III.A-2. Historical Growth in U.S. Ethanol Production (Thru 2006)
5.0
4.5
4.0
3.5
ro
O)
c 2.5
O
1.0
0.5
0.0
n
nn
Year
Sources: RFA Ethanol Industry Outlook 2007 (1980-2003); EIA September 2007 Monthly Energy Review total
ethanol usage less EIA reported ethanol imports (2004-2006)
In addition, according to industry sources, 2007 domestic ethanol production reached
approximately 6.5 billion gallons. Not only was production at an all time high, industry growth
between 2006 and 2007 (-1.7 billion gallons) was also record-setting. By all indications, growth
in ethanol production is expected to continue.
As mentioned earlier, as of September of 2007, there was ~7 billion gallons of ethanol
plant capacity online. Even if no new plants/capacity were added, ethanol production would
grow from 2007 to 2008 (provided plants continue to operate at or above nameplate capacity).
As of September 1, 2007, there were 10 expansion and 67 new construction projects underway
52
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with a combined production capacity of an additional 5.9 billion gallons.8 These projects are at
various phases of construction from conducting land stabilization/foundation work, to
constructing tanks and installing associated structural, mechanical, and electrical components, to
completing start-up activities. Once these construction projects are complete, there will be 200
ethanol plants operating in the U.S. with a combined production capacity of nearly 13 billion
gallons per year, as shown in Table III.A-3. The growth from 2004 to that projected for "online
capacity" is shown in Figure III. A-2.
Table III.A-3. Summary of Under Construction Projects (2007)
PADD
PADD1
PADD 2
PADD 3
PADD 4
PADD 5
Total
Online
Capacity
MGY
0.4
6,679
30
108
170
6,987
No. of
Plants
1
119
1
6
6
133
Expansions
Capacity
MGY
0
286
0
2
0
288
No. of
Plants
0
9
0
1
0
10
New Construction
Capacity
MGY
264
4,686
385
120
168
5,623
No. of
Plants
3
54
5
3
2
67
Total
Capacity
MGY
264
11,651
415
230
338
12,898
No. of
Plants
4
173
6
9
8
200
8 Based on Renewable Fuels Association (RFA), Ethanol Biorefinery Locations (updated August 22, 2007); Ethanol
Producer Magazine (EPM), Under Construction plant list (last modified on August 13, 2007), ethanol producer
websites, and follow-up correspondence with ethanol producers.
53
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Figure III.A-2. Annualized Ethanol Production
with Online Capacities from 2004 to 2007 (projected)
8
-,
O 7
O
] Annualized Corn Ethanol Production (total ethanol - imports)
Estimated Corn Ethanol Capacity Online
SL
c
o
6 -
HI
Dec-04 Apr-05 Aug-05 Dec-05 Apr-06 Aug-06 Dec-06 Apr-07 Aug-07 Dec-07
While theoretically it only takes 12 to 18 months to build an ethanol plant, the rate at
which these facilities actually come online will likely be slowed as a result of many factors, one
of them being that the majority of plants are being built by a few design/build companies (e.g.,
ICM, Fagen, Delta-T). The schedules of these firms are limiting and require coordination. In
addition, construction rates are often slowed by environmental permitting, material/labor
constraints, and other unforeseen circumstances. Ultimately, the future growth of the corn
ethanol industry will be, to some extent, limited by feedstock pricing.
According to USD A, enough corn will be available in the U.S. to support 20 billion
gallons of ethanol production in 2017.9 With each new plant that is built, however, corn pricing
will affect the costs of ethanol production. Higher corn pricing translates to higher operating
costs, possibly lower profit margins, and higher risk for ethanol producers. Under such
circumstances, the industry may respond by slowing investments and/or delaying construction
plans until corn prices fall or until the market responds with more demand for ethanol (e.g.,
increase in mandated volumes, more Flexible Fuel Vehicles (FFVs) and more E85 refueling
stations) and a viable means for distributing the additional ethanola trend that has already
begun.
9 USD A, An Analysis of the Effects of an Expansion in Biofuel Demand on U.S. Agriculture, May 2007
54
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B. Other Additives
Other alcohols and ethers (e.g. ETBE, TAME, and T-butanol or TEA) have been less widely
used and less widely studied than MTBE and ethanol. To the extent that they have been studied,
they appear to have similar, but not identical, chemical and hydrogeologic characteristics.
Reporting forms that EPA requires gasoline producers to submit (RFG and anti-dumping forms)
contain information on ETBE, TAME and TEA usage as well as MTBE and ethanol. Based on
survey data completed for 2005, MTBE and ethanol represented 55 and 41% of oxygenate fuel
additive use, respectively, while ETBE, TAME and TEA were at 0.29, 3.3 and 0.1%,
respectively. Since that time, we know that MTBE use has decreased dramatically, while
ethanol use has increased dramatically. Survey data for 2007 are being processed. We also
know that usage of the other additives has remained at approximately the same levels. Based on
the history of the use of the other additives, it is not expected that they will have any appreciable
change in their levels of commercial use as gasoline additives for the foreseeable future. (See
Appendix A for more information on historical use of MTBE, ethanol and the other fuel
additives.)
U.S. refiners currently value ethanol for several reasons including its octane value and
other positive blending properties, particularly as a replacement or substitute for MTBE.
Favorable economics for ethanol blending also contribute to an increase in use over the past
several years. Production and use of another high value blending component, alkylate, has also
increased in the U.S., and trends also appear to be on the rise elsewhere. Alkylates are typically
high-octane branched chain hydrocarbons produced at refineries, typically in alkylation units.
Alkylates are one hydrocarbon refinery stream. Aromatics make up another common
hydrocarbon stream at refineries and consist of benzene and benzene-like structures. Like
alkylates, they are high in octane quality, a positive quality for gasoline production. Unlike
alkylate content, EPA has traditionally collected data on and controlled aromatic content, since
emissions tests have shown that aromatics have been found to contribute to criteria pollutants
and to toxic emissions like benzene. Although little specific work has been done on alkylates,
alkylate streams in past testing have not been shown to play a primary role in the production of
the most toxic materials nor contribute to vehicle emissions in the way aromatics do.
There may also be products developed in the future which could play an increased role in
the commercial gasoline market. These products will need to be identified and evaluated if a
commercial market develops for new products.
55
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IV. POTENTIAL AIR QUALITY EFFECTS
The bulk of the impact of ethanol use on emissions and air quality is expected to be
associated with emissions from spark-ignited vehicles and equipment using low level ethanol-
gasoline blends. We expect the use of high level ethanol-gasoline blends, like E85, to see an
increase in use and consumption over the next decade, however in the near- and mid-term, this
volume will be relatively small in comparison to that of gasoline ethanol blends, and the focus of
this Report is on the use of ethanol as a fuel additive, not as a fuel. This section therefore
focuses on emissions and air quality effects from the use of low level ethanol blends.
A. Emissions Associated with Ethanol
In April of 2007, EPA promulgated the final Renewable Fuels Standard (RFS1) rule
which requires specific volumes of ethanol to be in gasoline by year, with 7.5 billion gallons per
year required for 2012. This rule mandated an overall increase of 3.2 billion gallons of ethanol.
We projected emission changes and air quality effects associated with the increased use of
ethanol in the Regulatory Impact Analysis (RIA) conducted for the RFS1 rule (U.S. EPA,
2007b). EPA is now tasked with writing a new version of the RFS rule under the recently passed
Energy Information and Security Act (EISA) of 2007. The new version, referred to as RFS2,
mandates the use of 36 billion gallons of ethanol by 2022. Any new analyses of emissions and
air quality that we conduct for the RFS2 rule will be made publicly available as part of that
rulemaking.
For the current Report to Congress, we have relied on the RIA conducted for the RFS1
rule (U.S. EPA, 2007b). In that analysis, we considered three cases for the future use of ethanol-
blend gasoline: a Reference Case, a Renewable Fuels Standard (RFS1) Case, and an Energy
Information Administration (EIA) Case. Using the MOBILE model, we evaluated emissions for
each of these cases, assuming that exhaust emissions from Tier 1 and later vehicles do not
respond to changes in ethanol content of the gasoline. 10 We recognize that these analyses are
not strictly relevant to the use of ethanol as a oxygenate fuel additive, and as such will
significantly overstate the impacts of its use.
The main difference between the Reference, RFS1 and EIA Cases is the assumption
about how much ethanol will be used and where it will go. The Reference Case represents our
estimate of fuel quality by county in 2004 when approximately 3.5 billion gallons of ethanol
were consumed nationwide. The RFS1 case assumes 6.7 billion gallons of ethanol consumption
in 2012, in accordance with the requirements of the RFS1 (Renewable Fuel Standard) mandate.
The EIA case assumes 9.6 billion gallons of ethanol is used nationwide in 2012, based on
projections made in the Energy Information Agency's 2006 Annual Energy Outlook. For each
Case, fuel quality was predicted for each county in the U.S. in 2012. This 2012 fuel matrix was
then used for all inventory and air quality assessments.
10 The Tier 1 standards are emission standards applicable to federal light duty vehicles and trucks that were phased
in beginning in 1994 and were phased out beginning in 2004. An implementation schedule for these standards is
available at http://epa.gov/otaq/cert/veh-cert/b00001e.pdf. The Tier 0 standards were the emission standards
applicable prior to the Tier 1 standards. These are light duty vehicle standards required by the 1977 Clean Air Act
amendments and first implemented in 1981.
56
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Table IV. A-1 shows ethanol impacts on VOC inventories for each of the three cases of
renewable fuel use in years 2012, 2015, and 2020. In any given year, the data suggest that total
VOC emissions will increase as ethanol use increases. The largest increase is seen in the EIA
case, where the increase is about 1% of the Reference Case inventory.
Our analysis indicates that this increase is a result of VOC non-exhaust emissions, such
as those from evaporation or permeation. While VOC exhaust emissions decrease, they do not
decrease enough to counteract the increase from non-exhaust emissions.
Table IV.A-1. National VOC Emissions from Gasoline Vehicles and Equipment:
Reference Case Inventory and Change in Inventory for Control Cases (Tons/Year)
Primary Case
Total
Reference
RFS1 Case (Change)
EIA Case (Change)
On-Road
Reference
RFS1 Case (Change)
EIA Case (Change)
Non-Road
Reference
RFS1 Case (Change)
EIA Case (Change)
Tons/Year
2012
5,882,000
18,000
43,000
2012
3,417,000
10,000
32,000
2012
2,465,000
8,000
11,000
2015
5,569,000
25,000
49,000
2015
3,269,000
16,000
36,000
2015
2,300,000
9,000
13,000
2020
5,356,000
34,000
58,000
2020
3,244,000
23,000
42,000
2020
2,112,000
11,000
16,000
Table IV.A-2 shows ethanol impacts on CO inventories for each of the three cases of
renewable fuel use in years 2012, 2015, and 2020. In any given year, data suggest that total CO
emissions will decrease as ethanol use increases. The largest reduction is seen in the EIA case;
this decrease is still less than 3% of the Reference inventory.
57
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Table IV.A-2. National CO Emissions from Gasoline Vehicles and Equipment:
Reference Case Inventory and Change in Inventory for Control Cases (Tons/Year)
Primary Case
Total
Reference
RFS1 Case (Change)
EIA Case (Change)
On-Road
Reference
RFS1 Case (Change)
EIA Case (Change)
Non-Road
Reference
RFS1 Case (Change)
EIA Case (Change)
Tons/Year
2012
55,022,000
-483,000
-1,366,000
2012
37,656,000
-45,000
-359,000
2012
17,366,000
-438,000
-1,007,000
2015
53,702,000
-473,000
-1,329,000
2015
36,171,000
-39,000
-321,000
2015
17,531,000
-434,000
-1,008,000
2020
53,949,000
-460,000
-1,286,000
2020
35,723,000
-19,000
-252,000
2020
18,226,000
-441,000
-1,034,000
Table IV.A-3 shows ethanol impacts on NOx inventories for each of the three cases of
renewable fuel use in years 2012, 2015, and 2020. In any given year, the data suggest that total
NOx emissions will increase as ethanol use increases. The largest increase is seen in the EIA
case, which is around 2% of the Reference inventory.
Our analysis also indicates that nonroad NOx emissions increase much more than onroad
emissions. While onroad inventories increase less than one percent in control cases, nonroad
inventories increase up to 11% in the EIA case.
Table IV.A-3. National NOx Emissions from Gasoline Vehicles and Equipment:
Reference Case Inventory and Change in Inventory for Control Cases (Tons/Year)
Primary Case
Total
Reference
RFS1 Case (Change)
EIA Case (Change)
On-Road
Reference
RFS1 Case (Change)
EIA Case (Change)
Non-Road
Reference
RFS1 Case (Change)
EIA Case (Change)
Tons/Year
2012
2,487,000
23,000
40,000
2012
2,240,000
9,000
13,000
2012
247,000
14,000
27,000
2015
2,059,000
18,000
33,000
2015
1,797,000
3,000
4,000
2015
262,000
15,000
29,000
2020
1,695,000
17,000
32,000
2020
1,407,000
0
0
2020
288,000
17,000
32,000
Table IV.A-4 shows ethanol impacts on air toxic emissions for each of the three cases of
renewable fuel use in 2012. We emphasize that the toxics inventories are based on very limited
data, especially when it comes to emissions from nonroad equipment.
58
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For all air toxics shown, the most extreme changes occur in the EIA case. The data
suggest that, in 2012, total benzene emissions will decrease by about 4%. Total 1,3-butadiene
emissions decrease by less than 2% while total formaldehyde emissions decrease by up to 1.5%.
On the other hand acetaldehyde emissions increase by as much as 36%.
Generally, the emissions trends in 2015 and 2020 parallel those of 2012 shown in the
table below (U.S. EPA, 2007b). Benzene maintains a drop of up to about 6% with increased
ethanol use. Formaldehyde remains fairly flat, ranging from a 0.5% increase to a 1.2% decrease.
Acetaldehyde maintains an increase of as much as 37%. Finally, 1,3-butadiene remains fairly
flat, ranging from no change to a 0.5% increase.
Table IV.A-4. National Toxic Emissions from Gasoline Vehicles and Equipment in 2012:
Reference Case Inventory and Change in Inventory for Control (Tons/Year)
Primary Case
Benzene
1,3-Butadiene
Formaldehyde
Acetaldehyde
Total
Reference
RFS1 Case (Change)
EIA Case (Change)
178,000
-3,200
-7,200
18,900
-200
-300
40,400
-600
-200
19,900
3,400
7,100
Onroad
Reference
RFS1 Case (Change)
EIA Case (Change)
124,100
-2,300
-5,400
12,000
-200
-200
29,900
-600
-300
15,500
2,400
5,400
Nonroad
Reference
RFS1 Case (Change)
EIA Case (Change)
53,900
-900
-1,800
6,900
0
-100
10,500
0
100
4,400
1,000
1,700
1. Local and Regional CO, VOC and NOx Emissions (Summer 2015)
We also estimated the percentage change in VOC, NOx, and CO emissions from gasoline
fueled motor vehicles and equipment in those areas of the country where we expect the most
significant changes in ethanol use. Specifically, we focused on areas where the market share of
ethanol blends was projected to change by 50 percent or more and also focused on summertime
emissions, because these are most relevant to ozone formation. We modeled 2015 because the
ozone Response Surface Model (RSM) used for air quality modeling is based upon a 2015
emissions inventory, though we would expect similar results in 2012. Finally, we developed
separately estimates for: 1) RFG areas, including the state of California and the portions of
Arizona where their CBG fuel programs apply, 2) low Reid Vapor Pressure (RVP) areas (i.e.,
RVP standards less than 9.0 RVP), and 3) areas with a 9.0 RVP standard. This set of groupings
helps to highlight the emissions impact of increased ethanol use in those areas where emission
control is most important.
Table IV.A-5 presents our primary analysis estimates of the percentage change in VOC,
NOx, and CO emission inventories for these three types of areas when compared to the 2015
reference case.
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Table IV.A-5. Change in July 2015 Emissions from Gasoline Vehicles and Equipment in
Counties Where Ethanol Use Changed Significantly - Primary Analysis
Ethanol Use
RFS1 Case
EIA Case
RFC Areas
Ethanol Use
VOC
NOx
CO
Down
0.8%
-3.4%
6.1%
Up
2.3%
1.6%
-2.6%
Low RVP Areas
Ethanol Use
VOC
NOx
CO
Up
4.2%
6.2%
-12%
Up
4.6%
5.7%
-13%
Other Areas (9.0 RVP)
Ethanol Use
VOC
NOx
CO
Up
3.6%
7.3%
-6.4%
Up
4.6%
7.0%
-6.0%
As expected, increased ethanol use tends to increase NOx emissions. The increase in
NOx emissions in "Low RVP Areas" and "Other Areas" is greater than in "RFG Areas", since
the gasoline in the RFG Areas included in this analysis all contained MTBE. Also, increased
ethanol use tends to increase VOC emissions, indicating that the increase in non-exhaust VOC
emissions exceeds the reduction in exhaust VOC emissions. This effect is muted with RFG due
to the absence of an RVP waiver for ethanol blends.
Table IV. A-6 presents the percentage change in VOC, NOx, and CO emission inventories
under our sensitivity analysis (i.e., when we apply the emission effects of the EPA Predictive
Models to all motor vehicles).
Table IV.A-6. Change in July 2015 Emissions from Gasoline Vehicles and Equipment in
Counties Where Ethanol Use Changed Significantly - Sensitivity Analysis
Ethanol Use
RFS1 Case
EIA Case
RFG Areas
Ethanol Use
VOC
NOx
CO
Down
-1.0%
-0.9%
7.3%
Up
1.0%
5.6%
-3.0%
Low RVP Areas
Ethanol Use
VOC
NOx
CO
Up
3.4%
10.4%
-15.0%
Up
3.7%
10.8%
-16.4%
Other Areas (9.0 RVP)
Ethanol Use
VOC
NOx
CO
Up
3.0%
10.8%
-9.0%
Up
3.9%
11.0%
-8.9%
60
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Directionally, the changes in VOC and NOx emissions in the various areas are consistent
with those from our primary analysis. The main difference is that the increases in VOC
emissions are smaller, due to more vehicles experiencing a reduction in exhaust VOC emissions,
and the increases in NOx emissions are larger.
B. Potential air quality impacts
1. Ozone Impacts
We performed ozone air quality modeling simulations for the eastern United States using
the ozone Response Surface Model (RSM) to estimate the effects of the projected changes in
emissions from gasoline vehicles and equipment associated with the RFS1 rule. Details on how
the modeling was conducted are contained in Chapter 5 of the RIA (U.S. EPA, 2007b).
The ozone RSM used for assessing the air quality impacts of expanded ethanol use in fuel
was developed broadly to look at various control strategies with respect to attaining the 8-hour
ozone NAAQS. The experimental design for the ozone RSM covered three key areas: type of
precursor emission (NOx or VOC), emission source type (i.e., onroad vehicles, nonroad vehicles,
area sources, electrical generating utility (EGU) sources, and non-utility point sources), and
location in or out of a 2015 model-projected residual ozone nonattainment area. This resulted in
a set of 14 emissions factors.
The impact of the RFS1 rule's increased ethanol use on the 8-hour ozone design values in
2015 is presented in Table IV.B-1. The changes presented in Table IV.B-1 are for those counties
with 2001 modeled design values. 11
11 2001 design values were calculated as an average of the 1999-2001, 2000-2002 and 2001-2003 design values at
each monitoring site. Please see the Air Quality Modeling Technical Support Document for the final Clean Air
Interstate Rule for additional information.
61
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Table IV.B-1. Impact of Increased Ethanol Use (RFS1 Rule)
on 8-hour Ozone Design Values in 2015 (ppb)
RFS1 Case
El A Case
Primary Analysis
Minimum Change
Maximum Change
Average Change Across 37 States
Population-Weighted Change Across 37 States
Average Change Where Ethanol Use Changed
Significantly (greater than 50%) States
Population-Weighted Change in States Where Ethanol
Use Changed Significantly
-0.015
0.329
0.057
0.052
0.153
0.154
0.000
0.337
0.079
0.056
0.181
0.183
Sensitivity Analysis
Minimum Change
Maximum Change
Average Change Across 37 States
Population-Weighted Change Across 37 States
Average Change Where Ethanol Use Changed
Significantly States
Population-Weighted Change Where Ethanol Use
Changed Significantly States
-0.115
0.624
0.111
0.092
0.300
0.272
0.000
0.549
0.142
0.096
0.325
0.315
As can be seen, ozone levels generally increase with increased ethanol use. This is likely
due to the projected increases in both VOC and NOx emissions. Some areas do see a small
decrease in ozone levels. In our primary analysis, where exhaust emissions from Tier 1 and later
onroad vehicles are assumed to be unaffected by ethanol use, the population-weighted increase in
ambient ozone levels is 0.052-0.056 ppb. Since the 8-hour ambient ozone standard is 0.08 ppm
(85 ppb), this increase represents about 0.06 percent of the standard, a very small percentage!2.
While small, this figure includes essentially zero changes in ozone in areas where ethanol use did
not change. When we focus just on those areas where the market share of ethanol blends
changed by 50 percent or more, the population-weighted increase in ambient ozone levels rises
to 0.154-0.183 ppb. This increase represents about 0.2 percent of the standard.
In our sensitivity analysis, where exhaust emissions from Tier 1 and later onroad vehicles
are assumed to respond to ethanol like Tier 0 vehicles, the population-weighted increase in
ambient ozone levels across the entire 37 state area is slightly less than twice as high, or 0.092-
0.096 ppb. This increase represents about 0.11 percent of the standard. When we focus just on
those areas where the market share of ethanol blends changed by 50 percent or more, the
12 Appendix I of 40 CFR Part 50.
62
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population-weighted increase in ambient ozone levels rises to 0.272-0.315 ppb. This increase
represents about 0.35 percent of the standard.
It is important to note that the results of this ozone response surface metamodeling
exercise is meant for screening-level purposes only and does not represent the results that would
be obtained from full-scale photochemical ozone modeling. In addition, there are several
important caveats concerning our estimated ozone impacts using the ozone RSM. The ozone
RSM does not account for changes in CO emissions. As shown in Section IV.A, ethanol use
should reduce CO emissions, directionally reducing ambient ozone levels in areas where ozone
formation is VOC-limited. Accounting for the reduction in CO emissions in NOx-limited areas,
however, may have little impact on the ozone impact of ethanol use.
The ozone RSM does not account for changes in VOC reactivity. With additional
ethanol use, the ethanol content of VOC should increase. Ethanol is less reactive than the
average VOC. Therefore, this change should also reduce ambient ozone levels in a way not
addressed by the ozone RSM. Again, like the impact of reduced CO emissions, this effect
applies to those areas where ozone formation is VOC-limited. Another limitation is the RSM's
inability to simulate the spatial distribution of emission impacts associated with the proposed
standard. Instead, we are forced to make simplifying assumptions about the geographic
uniformity of RFS1 emissions impacts, explained above. The caveats and limitations associated
with the RSM highlight the fact that it should only be used as a screening-level tool to
characterize broad trends associated with changes in different source categories of ozone
precursors.
Keeping these limitations in mind, the expanded use of ethanol will impact the national
emissions inventory of precursors to ozone, such as VOCs and NOx. Exposure to ozone has
been linked to a variety of respiratory effects including premature mortality, hospital admissions
and illnesses resulting in school absences. Ozone can also adversely affect the agricultural and
forestry sectors by decreasing yields of crops and forests.
Overall, we estimate that the measurable changes in VOC and NOx that result from the
RFS1 rule's 3.2 billion gallon increase in ethanol use, will, on average, result in small increases
in ambient ozone formation. As we discussed above, the ozone modeling results indicate a net
increase in the average population weighted ozone design value metric measured within the
modeled domain (37 Eastern states and the District of Columbia). EISA of 2007 mandates the
use of 36 billion gallons of ethanol by 2022; the air quality of impacts of this new mandate will
be addressed by the RFS2 rule.
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2. Particulate Matter
a. Potential Impact of Changes in Direct PM Emissions
The amount of data evaluating the impact of ethanol and MTBE blending on
direct emissions of PM from gasoline-fueled vehicles is extremely limited. In the past,
most studies did not test PM emissions from vehicles fueled with unleaded gasoline
because the level of PM emissions from properly operating vehicles is usually very low,
less than 0.1 g/mi.
Two studies indicate that the addition of ethanol might reduce direct PM
emissions from gasoline vehicles (Ragazzi and Nelson, 1999; Mulawa, 2007). However,
both studies were performed under wintertime conditions, one at high altitude, and one of
the studies only consisted of three vehicles. The available data indicate that ethanol
blending might reduce exhaust PM emissions under very cold weather conditions (i.e., 0
F or less), particularly at high altitude. There is no indication of PM emission reductions
at higher temperatures or under conditions in which the engine is warmed up sufficiently.
Thus, the data are certainly too limited to support a quantitative estimate of the effect of
ethanol on direct PM emissions.
b. Potential Impact of Changes in Secondary PM Formation
In addition to being emitted directly from a combustion source, fine particles can
be formed through a series of chemical reactions in the atmosphere when SO2, NOx, and
VOC oxidize or otherwise react to form a wide variety of secondary PM. For example,
SC>2 oxidizes to SOs and sulfuric acid and NOx oxidizes to NOs and nitric acid which, in
turn, react with ammonia in the atmosphere to form ammonium sulfate and ammonium
nitrate. Particles generated through this gas-to-particle conversion are referred to as
secondary aerosols and represent a significant portion of ambient fine particulate matter.
Studies have shown that as much as 70% of the total organic carbon in urban particulate
matter can be attributed to secondary organic aerosol (SOA) formation although the
amount can also be less (Grosjean, 1992). Secondary PM tends to form more in the
summer with higher temperatures and more intense sunlight.
Source-receptor modeling studies conducted in the Los Angeles area in 1993 by
Schauer et al. (2002) indicated that as much as 67% of the fine particulate matter
collected could not be attributed to primary sources. The authors concluded that much of
this unidentifiable organic matter is SOA formed in the atmosphere. This is consistent
with previous studies conducted by Turpin and Huntzicker (1991) who concluded that
70% of the total organic carbon in urban PM measurements made in southern California
can be attributed to SOA.
Gas phase VOCs are oxidized by OH, NO2, peroxyacetylnitrate (PAN), and ozone
in the atmosphere, but their propensity to condense in the particle phase is a function of
two factors: volatility and reactivity. To accumulate as an aerosol, a reaction product
must first be formed in the gas phase at a concentration equal to its saturation
64
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concentration. This requirement will not be met if the relevant gas-phase reactions of the
VOC are too slow or if the vapor pressure of the reaction product is higher than the initial
concentration of its VOC precursor (Pun and Seigneur, 2006). Limited data for reaction
rate constants determined both experimentally and estimated by structural relationships
are available in the published literature. However, the atmospheric chemistry behind
SOA reaction rates and the estimated aerosol yield is highly complex and carries with it a
great deal of uncertainty. Research in this area is ongoing and thus the capacity to
quantitatively model SOA formation is not yet a straightforward process.
In general, all reactive VOC are oxidized by OH or other compounds.
Additionally, alkenes, cycloalkenes, and other olefinic compounds can react with ozone
and NO2 to form secondary aerosols. In fact, ozone is responsible for nearly all the SOA
formation from olefins, while OH plays little or no role at all (Grosjean and Seinfeld,
1989; Izumi and Fukuyama, 1990). Many VOC, however, will never form SOA under
atmospheric conditions regardless of their reactivity. This is because the reaction
products of these compounds have vapor pressures that are too high to form aerosols at
atmospheric temperatures and pressures. These include all alkanes and alkenes with up
to 6 carbon atoms and many low-molecular weight carbonyls, chlorinated compounds,
and oxygenated solvents (Grosjean, 1992).
The VOC that have the greatest propensity to form SOA include aromatic
hydrocarbons (such as toluene but even including benzene (Martin-Reviego and Wirtz,
2005)), higher molecular weight olefins and cyclic olefins, and higher molecular weight
paraffins. Kleindienst et al. (1999, 2007) suggest that a high fraction of SOA is due to
aromatic hydrocarbon precursors. Furthermore, "aromatic products having a single alkyl
group on the aromatic ring were found to represent a 'high-yield' family (e.g., toluene,
ethylbenzene); compounds having multiple methyl groups (e.g., m-xylene, 1,2,4-
trimethylbenzene) were found to represent a 'low-yield' family" (Kleindienst et al., 1999,
2007). All of the above mentioned VOC precursors are important either because there
are large amounts of these particular VOC emitted per day, or because a large fraction of
the VOC reacts, or a combination of the two. Based on VOC emissions inventory data
collected in the Los Angeles area, the most important aerosol precursors (in the LA area
using 1982 VOC emissions inventories) are listed in Table IV.B-4 below.
65
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Table IV.B-4. Predicted In Situ SOA Formation
During a Smog Episode in Los Angeles
VOC Functional Group Kg emitted daily*
Aromatics
Olefins
Paraffins
Alkenes
Cyclic Olefins
Terpenes
Alkanes
Cycloalkanes
223985
31163
3220
6000
140493
37996
Secondary PM %
Produced (kg)* yield
3061
608
144
626
368
96
1.37
1.95
4.47
10.43
0.26
0.25
'Source: Grosjean, 1992
These predictions are a function of input data collected in the Los Angeles area,
and assume ambient levels of [ozone] = 100 ppb, [OH] = l.OxlO6 molecules/cm3, and
[NOs] = 0 with 6 hours of reaction time. Aromatics are the largest functional group in
terms of the absolute quantities of VOC emitted daily, and thus they eventually form the
most SOA. Likewise, many high molecular weight paraffins (alkanes) form SOA on a
significant scale simply because their emissions are high. However, the relative fraction
of paraffins that react is less than that of aromatics in smog chamber experiments
simulating SOA formation in the atmosphere. For olefins, the alkenes exhibit a
combination of both relatively high emissions, and a high fraction of VOC reacted to
form SOA. Cyclic olefins, in contrast, are emitted in relatively low levels, but a high
fraction of these VOC react and the end result is a proportionally higher SOA yield than
with the alkenes. Lastly, there are several "miscellaneous" compounds and terpenes that
are emitted on a relatively small scale (in southern California), but that produce a
substantial amount of SOA.
Researchers at EPA recently completed a field study in the Raleigh/Durham area
of North Carolina that investigated the contribution of various sources to ambient PM 2.5
concentrations (Lewandowski et al., 2007; Kleindienst et al., 2007). In the study they
identified toluene as an SOA precursor. They estimate that mobile sources contribute
nearly 90% of the total toluene emissions in that region based on a chemical mass
balance approach. At the same time, however, SOA attributable to non-fuel-related VOC
(i.e., biogenic emissions) was found to be an even larger contributor to SOA (i.e., toluene
was not likely the dominant source of SOA in this area). This study is currently
undergoing peer review and will be published shortly. Qualitatively, however, this
information is still quite useful since the study identifies a contributing source of SOA
that is attributable almost entirely to mobile sources.
VOC reaction rates increase with increasing ambient temperature and sunlight
intensity, so the level of SOA formed is much higher in summer than in winter. Even in
the more temperate coastal climates of southern California, studies have found the
summertime concentration of SOA calculated through Chemical Mass Balance models to
be anywhere from 2-5 times higher in summer than winter. In a study conducted at both
66
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urban and rural locations in the southeastern United Sates, the concentration of SOA in
the summer and early fall was roughly 2-3 times that of colder months (Zheng et al.,
2002).
Due to its high octane, the addition of ethanol will reduce aromatics in gasoline,
which will in turn reduce the aromatics emitted in the exhaust. Fully quantifying the
emission reduction is not possible at this time due to a lack of speciated exhaust data for
newer vehicles (e.g., Tier 2 vehicles first introduced in the 2004 model year) running on
ethanol blends. These data are now being obtained (see Section VII.B). In addition,
increased NOx emissions, resulting from the increased use of ethanol, could increase the
formation of nitrate PM.
In contrast, gasoline-fueled vehicles and equipment comprise a significant
fraction of all national gaseous aromatic VOC emissionslS. The lower aromatics levels in
gasoline with gasoline/aromatic blends will result in less aromatic emissions in both
exhaust and evaporative emissions. We are presently doing testing and modeling to
calculate the magnitude of these reductions.
The research to facilitate this incorporation is currently underway. EPA ORD
scientists are currently carrying out a wide variety of laboratory studies to refine the SOA
chemistry mechanisms for use in the next version of the CMAQ model. This information
should be available in time for the comprehensive study of the Act's fuel requirements
which is due in 2009.14
13 Based on internal analyses of emissions inventories.
14 Subject to funding.
67
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References
Grosjean, D., Seinfeld, J. H. (1989) Parameterization of the formation potential of
secondary organic aerosols, Atmos. Environ. 23, 1733-1747.
Grosjean, D. (1992) In situ organic aerosol formation during a smog episode: estimated
production and chemical functionality. Atmos. Environ. 26A: 953-963.
Izumi, K., Fukuyama, T. (1990) Photochemical aerosol formation from aromatic
hydrocarbons in the presence of NOx. Atmos. Environ. 24A, 1433-1441.
Kleindienst, I.E.; Smith, D.F.; Li, W.; Edney, E.G.; Driscoll, D.J.; Speer, R.E.;
Weathers, W.S. (1999) Secondary organic aerosol formation from the oxidation
of aromatic hydrocarbons in the presence of dry submicron ammonium sulfate
aerosol. Atmos. Environ. 33(22): 3669-3681.
Kleindienst, T. E.; Jaoui, M.; Lewandowski, M.; Offenberg, J. H.; Lewis, C. W.; Bhave,
P.; Edney, E. O. (2007) Estimates of the contributions of biogenic and
anthropogenic hydrocarbons to secondary organic aerosol at a southeastern U.S.
location. Atmos. Environ.
Lewandowski, M.; Jaoui M.; Kleindienst T. E.; Offenberg, J. H.; Edney, E. O. (2007)
Composition of PM2.5 during the summer of 2003 in Research Triangle Park,
North Carolina. Atmos. Environ. 41(19): 4073-4083.
Martin-Reviego, M.; Wirtz, K. (2005) Is benzene a precursor for secondary organic
aerosol? Environ. Sci. Technol. 39: 1045-1054.
Mulawa, P. A. (1997) Effect of ambient temperature and E-10 on primary exhaust
particulate matter emissions from light-duty vehicles. Environ. Sci. Technol. 31:
1302-1307.
Pun, B. K.; Seigneur, C. (2006) Investigative modeling of new pathways for secondary
organic aerosol formation. Final Report. CRC Project A-59. Coordinating
Research Council.
Ragazzi, R.; Nelson, K. (1999) The impact of a 10% ethanol blended fuel on the exhaust
emissions of Tier 0 and Tier 1 light duty gasoline vehicles at 35F. Colorado
Department of Public Health and Environment. March 26, 1999. For additional
information or copies, contact Gerald L. Gallagher or Margie M. Perkins.
Schauer, J. J.; Fraser, M. P.; Cass, G. R.; Simoneit, B. R. T. (2002) Source reconciliation
of atmospheric gas-phase and particle-phase pollutants during a severe
photochemical smog episode." Environ. Sci. Technol. 36: 3806-3814.
68
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Turpin, B. J., Huntzicker, J. J., Larsen, S. M., and Cass, G. R. (1991) Secondary
formation of organic aerosol in the Los Angeles Basin: a descriptive analysis of
organic and elemental carbon concentration, Environ. Sci. Technol. 25: 1788-
1793.
Turpin, B. J.; Huntzicker, J. J. (1991) Secondary formation of organic aerosol in the Los
Angeles basin. A descriptive analysis of organic and elemental carbon
concentrations. Atmos. Environ. 25(22): 207-215.
Zheng, M.; Cass, G. R.; Schauer, J. J.; Edgerton, E. S. (2002) Source apportionment of
PM2.5 in the Southeast United States using solvent-extractable organic
compounds as tracers. Environ. Sci. Technol. 36: 2361-2371.
U. S. Environmental Protection Agency. (2007a) "National Emissions Inventory (NEI)
Air Pollutant Emissions Trends Data, 1970 - 2006 Average annual emissions."
U.S. Environmental Protection Agency. (2007b) Regulatory Impact Analysis:
Renewable Fuel Standard Program. EPA 420-R-07-004. April. Available from:
http://www.epa.gov/otaq/renewablefuels/420r07004.pdf
69
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V. POTENTIAL WATER QUALITY EFFECTS
This chapter focuses on the potential effects of fuel and fuel additives on
contamination of ground waters which serve as supplies for drinking water. There are
three mechanisms by which fuel additives can contaminate ground water: (1) the additive
can itself be a contaminant, (2) the additive can interfere with natural biodegradation of
hydrocarbons such as benzene, toluene, ethylbenzene and xylenes (BTEX) in ground
water allowing the plumes formed by these petroleum components to extend farther than
they otherwise would, and (3) the additive can encourage growth of anaerobic bacteria
that produce other toxic or undesirable substances such as hazardous biodegradation
products, sulfide, soluble forms of iron, manganese, arsenic, methane (at explosive
concentrations), or volatile fatty acids such as butyrate.
The probability of affecting ground water supply wells is related to the length of
the plume generated from the source of the contamination in ground water and the rate of
water supply withdrawal. Replacing MTBE with an alternative that is either more
hazardous than MTBE, has hazardous by-products or causes a significant increase in the
length of plumes of benzene and other BTEX compounds can be detrimental to the
overall quality of ground water and public health. Ethanol, when blended with gasoline,
has the potential to make the plume of benzene and other BTEX compounds persist
longer and travel farther than they would in the absence of ethanol.
Potential water quality and quantity impacts due to renewable fuel feedstock
production are beyond the scope of this report but will be included in a future report to
Congress as directed by Section 204 of the Energy Independence and Security Act of
2007.
A. Factors Influencing Plume Length and Ground Water Contamination
The length of plumes associated with MTBE and its replacements and the
possibility that they will adversely affect ground water used as drinking water is related
to (1) the hazard associated with the fuel additive and by-products and any resulting
concentration-based standard used to define the plume boundary, (2) the tendency of the
fuel additive to partition out of spilled gasoline, dissolve into ground water and establish
high concentrations in the area of the spill, and (3) the rate and extent of biodegradation
of the fuel additive in ground water.
The hazard associated with the individual fuel additives and by-products is
evaluated in Chapter II, and studies relating initial concentrations of fuel additive
contaminants to plume length will be discussed in the following two sections.
The second factor, the tendency of fuel additives to partition out of gasoline,
depends on a number of factors, including: gasoline composition, equilibrium phase
partitioning, mass transfer limitations and geologic heterogeneity. As the concentration
of an additive in gasoline increases, the concentration in ground water will increase.
Equilibrium phase partitioning describes the tendency of the additive to either remain
70
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behind in gasoline or to dissolve in ground water. Additives that are chemically similar
to petroleum, such as the alkylates, tend to remain in the gasoline. Additives that are
chemically more similar to water, such as the alcohols, tend to dissolve in water. Mass
transfer limitations refer to conditions that slow the rate that the additive can dissolve in
ground water. Geological heterogeneity refers to variations in the composition of the
aquifer that affects the flow of ground water. The flow of ground water carries an
additive away from the spill of gasoline to form a plume of contamination in the aquifer.
Ground water flows faster through sands and gravels than it does through silts and clays.
These processes and interactions are described in more detail in Appendix B.
Compared to the natural components of petroleum in gasoline, MTBE has a
higher solubility in water. Benzene is one of the most soluble natural components of
gasoline; its solubility in water is 1.8 g/L. In contrast, the solubility of MTBE is near 50
g/L. The solubility of ETBE, TAME, and DIPE in water are 12, 11, and 7 g/L
respectively, and they produce concentrations in ground water contaminated by spilled
gasoline that are only slightly less than those produced by MTBE. The alcohols (ethanol
and TEA) produce concentrations in ground water that are even higher than those
produced by MTBE (see Table 3-1 in Wilson, 2003). Based on their higher solubility in
water compared to MTBE, the concentrations of 1-propanol, 2-propanol, and 1-butanol,
should also be higher than MTBE (Montgomery, 2000). In contrast, the branched chain
alkanes (alkylates) have little solubility in water (Montgomery, 2000). The expected
concentrations of alkylates is one thousand fold lower than the observed concentrations
of MTBE.
The third factor, the biodegradation of the fuel additive, can occur via a variety of
mechanisms. Organic compounds are metabolized through a series of
oxidation/reduction reactions, where one compound loses electrons and is oxidized, and
the other compound receives electrons and is reduced. If the organic compound is
entirely oxidized to carbon dioxide, the process is termed respiration. Oxygen, nitrate,
sulfate, and iron (Fe(III)) minerals, can serve as the terminal electron acceptors during
respiration. When molecular oxygen is reduced to water, the process is called aerobic
respiration; when sulfate is reduced to sulfide, the process is termed sulfate reduction:
when nitrate is reduced to ammonia or molecular nitrogen, the process is called nitrate
reduction or denitrification, and when insoluble Fe(III) minerals are reduced to form
soluble Fe(II) salts, the process is termed iron reduction. Organic compounds can also
serve as both electron donors and electron acceptors, and reactions where the electron
donor and electron acceptor are both organic compounds are called fermentations.
Information on the potential for biodegradation of ethanol and other alcohols, MTBE and
other ethers, and the BTEX compounds is discussed in detail later in the section and is
summarized in Table V.B-3.
Bacteria that carry out aerobic respiration are widely distributed in soil and
sediment and have a great metabolic diversity. However, oxygen has limited solubility in
water and is usually unavailable in contaminated aquifer sediments, and ambient
concentrations of nitrate are usually low in these sediments; as a result, the most
71
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important electron acceptors in contaminated ground water are sulfate and Fe(III)
minerals in the aquifer sediment (Wiedemeier et al., 1999; U.S. EPA, 2007).
The biodegradability for each additive and its affect on plume composition and
length will be discussed in the following two sections.
B. The Potential Effect of Fuel Additives on Plume Length and Ground Water
Contamination
A spill of gasoline may contaminate ground water with natural petroleum
hydrocarbons, including benzene and the BTEX compounds, as well as fuel additives
such as MTBE, TEA, ETBE, DIPE, heavy alcohols, iso-octane and alkylates, and
ethanol. Near the spill, the plume of contaminated ground water will contain all the
contaminants found in the particular fuel spilled. As the plume moves away from the
spill, individual contaminants will attenuate due to sorption, biodegradation and dilution
and dispersion. The concentration of individual ground water contaminants will drop
below the MCL, the action level, or other regulatory goal at various distances from the
spill, depending on the initial concentration in the source, and the particular rate of
attenuation of the individual contaminants. One contaminant will move the greatest
distance before it reaches an acceptable concentration, and this contaminant will define
the overall size of the plume of contamination from a particular spill of gasoline. The
presence of one contaminant, such as ethanol, can influence the size of the plume of other
contaminants, such as benzene. This section compares the sizes of plumes formed by
benzene, fuel additives such as MTBE, TEA, or ethanol, and the interaction between fuel
additives and the size of benzene plumes.
1. MTBE, TEA, ETBE, DIPE, iso-octane and alkylates
The U.S. Geological Survey recently summarized data collected under the
National Water Quality Assessment Program on the concentrations of selected volatile
organic compounds in untreated ground water (Zogorski et al., 2006). Two of the
important contaminants that might be derived from spills of gasoline from underground
storage tanks are benzene and MTBE. Benzene and MTBE were found to have a similar
distribution in the number of occurrences at unacceptable concentrations. In shallow
ground water in urban areas, benzene was detected 9 times in 847 samples at a
concentration above its Maximum Contaminant Level (MCL) of 5 |ig/L. The median
concentration of occurrences of benzene above the standard was 80 |ig/L, which is 16
fold above its MCL. In contrast, MTBE was detected 12 times in 847 samples at
concentration above 20 |ig/L (the lower limit for the U.S. EPA Drinking Water Advisory
for MTBE). The median concentration of occurrences of MTBE above the standard was
120 |ig/L, which is 6 fold above its lower Drinking Water Advisory. Therefore, as
measured by the National Water Quality Assessment Program, the extent of benzene and
MTBE contamination in shallow ground water in urban areas is very similar. It should be
noted that the MCL for benzene and the Drinking Water Advisory level for MTBE
address very different concerns. Benzene is a known human carcinogen, and the MCL is
health protective of this endpoint. The Drinking Water Advisory level for MTBE, on the
72
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other hand, is based on odor and taste thresholds. More health effects information on
benzene and MTBE can be found in Chapter II.
The published data on the length of plumes of benzene and MTBE produced from
leaking underground storage tanks are also consistent with one another. Table V.B-1
compares published data on the lengths of benzene and MTBE plumes in Texas and
California. In each of the studies, the plumes were associated with the entire population
of gasoline spills in the relevant geographic area that was subject to monitoring by the
regulatory authority that collected the data. As a consequence, the population in each
study included old and new spills, spills of conventional gasoline, spills of gasoline with
MTBE and other ethers added to enhance octane, spills of RFG gasoline, and spills of
E10. On average, the MTBE plumes are longer, but not many fold longer, than the
plumes of benzene from the same spills of gasoline. Data are also available on the length
of plumes of TEA in Greater Los Angeles. This compound is used to manufacture
MTBE, has been used in gasoline as an alternative to MTBE, and is produced by natural
biodegradation of MTBE in ground water. On average, the length of the TEA plumes is
on the same scale as the length of benzene and MTBE plumes. In greater Los Angeles,
the average lengths of plumes of ETBE, DIPE and TAME are also on the same scale as
the length of benzene and MTBE plumes (Shih et al., 2004). There have been no
observational studies of iso-octane or other alkylates.
Table V.B-1. Comparison of plume lengths of volatile organic contaminants in
ground water originating from gasoline spills from underground storage tanks.
Mean Plume Length In Texas
(to 10 |ig/L) (feet)a
Mean Plume Length In
California (feet)b
Geometric mean of measured
Plume Length in Greater Los
Angeles (feet)0
Median estimated plume
length in 90 plumes in Greater
Los Angeles (feet)d
MTBE
174
170
272
242
Benzene
144
113
167
148
TEA
207
aMace and Choi, 1998
kHappel et al., 1998
cShih et al., 2004
^ong and Rong, 2002
As discussed previously, the extent of a contaminant plume depends upon the
rates of release of mass from the source, transport in the aquifer, and biodegradation.
The effect of each transport parameter on the extent and concentration of contaminant
plumes can be shown through a sensitivity analysis. Maier and Grathwohl (2006)
developed a relationship for plume length from empirical results generated from one- and
73
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two-dimensional model runs. The models were developed for dissolved contamination
with reaction between the contaminant and an electron acceptor. 15
For the conditions represented in an example provided by Maier and Grathwohl,
assumed source concentrations of MTBE and ethanol and for stoichiometric constants are
shown in Table V.B-2. The results show that in every case the result of lowering source
concentration is to reduce the plume length. The reduction in plume length, however, is
not proportional to the reduction in source concentration. For example, for the
stoichiometric conversion of MTBE to TEA a change in source concentration from 1000
mg/L to 100 mg/L resulted in a reduction in plume length from 1560 m to 490 m. In
other words, reduction of the source concentration by a factor of 10 resulted in reduction
of plume length by a factor of 3.2.
Table V.B-2. Results from Maier and Grathwohl (2006) plume length analysis for
an aquifer of 4.5 m contaminated thickness, vertical dispersivity of 0.032, and
electron acceptor concentration of 8 mg/L. The concentration and plume length
ratios are relative to the first result for each reaction.
Reaction
MTBE
to TEA
Ethanol
toCO2
Ethanol
toCO2
Electron
Acceptor
Oxygen
Oxygen
Sulfate
Source
Concentration
(mg/L)
1000
100
2500
1400
250
25
2500
250
P
0.54
2.08
3.13
Plume
Length
(m)
1560
490
4840
3620
1530
484
5940
1880
Concentration
Ratio
1
10
1
1.8
10
100
1
10
Plume
Length
Ratio
1
3.2
1
1.3
3.2
10
1
3.2
Although Maier and Grathwohl's results are instructive, the analysis does not
include the effect of dissolution from a non-aqueous phase liquid or fuel. The
Hydrocarbon Spill Screening Model (HSSM), developed by the EPA (U.S. EPA, 1994,
1995) includes flow of the fuel as a separate phase and dissolution of its components into
flowing ground water. In a modeling exercise done by the EPA (U.S. EPA, 2005b), the
composition of the released gasoline was varied to represent three scenarios: a state with
an MTBE-ban (New York) where MTBE was present at 0.28% volume, which is below
15 The resulting relationship has the form
M2fQ/?Y
L = a
a. ( Cea )
where L is the plume length, a and b are empirical constants determined by Maier and Grathwohl (2006) to
be approximately 0.3 and 0.5, respectively, M is the thickness of the contaminant plume, Ot is the vertical
transverse dispersivity, Cs is the source concentration, Cea is a representative electron acceptor
concentration and (3 is the stoichiometric constant representing the use of the electron acceptor for
degradation of the source material.
74
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the maximum allowable MTBE content of 0.5% volume; a state where oxygenated
gasoline was not required (Georgia) and MTBE's presence in gasoline at 4.89% volume
was likely for use as an octane enhancer, and a state where reformulated gasoline was
required (parts of Virginia) with a concentration of MTBE at 13.13% volume (U.S. EPA,
2005b). The results of these simulations show that the MTBE concentrations in the
aquifer are proportional to the MTBE content in the gasoline (see Figure V.B-1).
However, the extent of contamination, defined by the plume boundaries where the
concentration of MTBE is 0.02 mg/L or higher, does not appear to be particularly
sensitive to source strength. The maximum extent of the plumes increased from about
675 feet to 780 feet as the MTBE content increased from 0.28% volume to 13.13%
volume.
Figure V.B-1. HSSM simulation results for three gasoline scenarios. At top, New
York premium containing 0.28% volume MTBE; at center, Georgia premium
containing 4.89% volume MTBE, and at bottom, Virginia premium containing
13.13% volume MTBE. The concentrations are given in units of mg/L, and the 0.02
mg/L contour is highlighted as it corresponds to the lower limit of the EPA drinking
water advisory for MTBE.
d
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75
-------
long and Rong (2002) conducted a statistical analysis of 90 plumes in Greater
Los Angeles that had adequate monitoring to establish the behavior and distribution of
the plume. They calibrated the analytical model of Domenico (1987) to best fit the
monitoring data. The distribution of MTBE plume lengths was strongly asymmetric
(Figure V.B-2). Most plumes were short, but a few plumes were much longer than the
average. They found that the length of the MTBE plumes did not correlate strongly with
the velocity of ground water flow (correlation coefficient of 0.17) or with the maximum
concentration at the source (correlation coefficient of 0.31). Instead, the strongest
correlation was with the calibrated first order rate constant for biodegradation (correlation
coefficient of-0.65). The correlation is negative because the shortest plumes were
associated with the highest rates of biodegradation.
Figure V.B-2. Distribution of the lengths of MTBE plumes in ground water in 90
plumes in Greater Los Angeles (Tong and Rong, 2002).
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Length MTBE Plume (feet)
Extrapolating from the analysis of Tong and Rong (2002) for MTBE, the length
of a plume of an alternative fuel additive in ground water will be largely controlled by the
rate of natural biodegradation of the fuel additive. Table V.B-3 provides a qualitative
summary of the potential for biodegradation of MTBE and of the substitutes for MTBE
identified by Congress for evaluation. Because ground water contaminated with gasoline
is generally devoid of oxygen, the evaluation is limited to the potential for natural
anaerobic biodegradation.
As summarized in Table V.B-3, there is no evidence from the literature that the
other ether oxygenates (ETBE, TAME, or DIPE) are any more degradable than MTBE.
76
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Anaerobic biodegradation of MTBE produces TEA as the first degradation product
(Wilson et al., 2005b; Somsamak et al., 2005; Mackay et al., 2007). As far as is known,
the organisms that degrade MTBE do not further metabolize the TEA that they produce
(Wilson et al., 2005a; Somsamak et al., 2005). TEA is degraded by other organisms,
including sulfate reducing bacteria, iron-reducing bacteria, or bacteria that ferment the
TEA to methane (Finneran and Lovley, 2001; U.S. EPA, 2007). For MTBE to be
entirely degraded, the TEA produced must be entirely degraded. With respect to iso-
octane and the other alkylates!6, there are no reports in the literature for anaerobic
biodegradation of branched chain alkanes (alkylates) in the molecular weight range of
gasoline components.
Organisms that are capable of anaerobic MTBE biodegradation are not
universally distributed in contaminated ground water. Many plumes fail to acclimate to
anaerobic biodegradation, and many plumes that do acclimate are heterogeneous, with
rapid and extensive degradation in some portions of the plume and limited degradation in
others (U.S. EPA, 2005a). Anaerobic biodegradation of TEA in ground water follows
the same pattern (U.S. EPA, 2007).
There are a few publications available in the literature that allow for a quantitative
comparison of the rates of anaerobic biodegradation of benzene, MTBE and the
alternative fuel additives in ground water. Table V.B-4 summarizes what data are
available from the literature for biodegradation of benzene, MTBE, and TEA in ground
water, and compares the average of the rates of anaerobic biodegradation in the literature
to the average rate of biodegradation estimated by EPA for a large population of plumes
in Greater Los Angeles.
Shih et al. (2004) provided data on the distribution of plumes of benzene, MTBE,
and TEA in Greater Los Angeles. Data were reported on the average length of 95
benzene plumes, 96 MTBE plumes and 86 TEA plumes, and the maximum concentration
of organic compounds in the source area of 716 benzene plumes, 718 MTBE plumes and
530 TEA plumes. Using this information, combined with hydrological information from
a subset of the plumes observed in the same area by Tong and Rong (2002), we extracted
an estimate of the average rate of biodegradation at these release sites.
Tong and Rong (2002) estimated that the mean ground water velocity for 90 of
the sites in Greater Los Angeles was 55 feet per year. This estimate is similar in
magnitude to the median velocity of 100 feet per year extracted by McNab et al. (2000)
from a population of chlorinated solvent plumes, most of which were in the Western
United States.
16 Iso-octane is a component of alkylates.
77
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Table V.B-3, Part 1 of 3. Potential for natural anaerobic biodegradation of
additives, the impact of additives on potential for natural anaerobic biodegradation
of BTEX, and secondary pollutants associated with additive.
Additive or
Category of
Additive
Methyl tertiary
butyl ether (MTBE)
Ethyl tertiary butyl
ether (ETBE)
Tertiary amyl
methyl ether
(TAME)
Di-isopropyl Ether
(DIPE)
Tertiary butyl
alcohol (TEA)
Potential for Natural
Anaerobic Biodegradation
of Additives
Degradation rapid and
extensive at some sites but
only after microbial
acclimation.
Does not degradea'b'c'd'e
Does degradef'g'h'1J'Um'n
Does not degrade21' ''
Does degradek
Degradation rapid and
extensive at some sites but
only after microbial
acclimation.
Does not degrade21^
Does degrade1'-1
No report of anaerobic
biodegradation in the
literature.
Does not degrade21
Degradation rapid and
extensive at some sites but
only after microbial
acclimation.
Does not degrade21'11''1'
Does degradeg'k'n'p'q'r's
Potential for
Impact on
Natural
Anaerobic
Biodegradation
of BTEX
Yes, metabolism
produces TEA
and probably
acetate.
Metabolism of
TEA and acetate
may consume
sulfate, and ferric
iron that
otherwise could
be used to
degrade benzene.
None
demonstrated.
Yes, metabolism
produces TAA
and acetate.
Metabolism of
acetate may
consume sulfate
and ferric iron
that otherwise
could be used to
degrade benzene.
No known
potential
Yes, metabolism
may consume
sulfate and ferric
iron that
otherwise could
be used to
degrade benzene
Secondary
Pollutants
Associated with
Degradation of
Additive
TEA0
TBAk
Tertiary amyl
alcohol
(TAA)iJ
None
demonstrated
Acetone1, sulfide,
ferrous iron
78
-------
Table V.B-3, Part 2 of 3. Potential for natural anaerobic biodegradation of
additives, the impact of additives on potential for natural anaerobic biodegradation
of BTEX, and secondary pollutants associated with additive.
Additive or Category
of Additive
Potential for Natural
Anaerobic
Biodegradation of
Additives
Potential for
Impact on
Natural
Anaerobic
Biodegradation
of BTEX
Secondary
Pollutants
Associated
with
Degradation
of Additive
Other ethers and heavy alcohols as determined by Administrator
Other
Ethers
Alcohols
Other
Ethers
Alcohols
Other
Ethers
Alcohols
Other
Ethers
Alcohols
Other
Ethers
Alcohols
Tertiary
amyl alcohol
(TAA)
1-propanol
2-propanol
(isopropyl
alcohol)
1-butanol
Methanol
No known potential
Does not degrade3^
Degradation rapid and
extensive at almost all
sites after microbial
acclimation.
Does degradeu'v
Degradation rapid and
extensive at almost all
sites after microbial
acclimation.
Does degradea'f'u'v
Degradation rapid and
extensive at almost all
sites after microbial
acclimation.
Does degradeu'v
Degradation rapid and
extensive at almost all
sites after microbial
acclimation.
Does degradea'f'r's'v'w'x
No known
potential
Metabolism can
be expected to
consume sulfate
and ferric iron
that otherwise
could be used to
degrade benzene
Metabolism can
be expected to
consume sulfate
and ferric iron
that otherwise
could be used to
degrade benzene
Metabolism can
be expected to
consume sulfate
and ferric iron
that otherwise
could be used to
degrade benzene
Yesw'x,
metabolism may
consume sulfate
and ferric iron
that otherwise
could be used to
degrade benzene
None
demonstrated
Methane,
sulfide, ferrous
iron
Methane,
sulfide, ferrous
iron
Methane,
sulfide, ferrous
iron
Methane,
sulfide, ferrous
iron
79
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Table V.B-3, Part 3 of 3. Potential for natural anaerobic biodegradation of
additives, the impact of additives on potential for natural anaerobic biodegradation
of BTEX, and secondary pollutants associated with additive.
Additive or Category
of Additive
Ethanol
Iso-octane (an alkylate)
2,2,4-trimethylpentane
Also other alkylates
2,3-dimethylpentane,
2,4-dimethylpentane
Potential for Natural
Anaerobic
Biodegradation of
Additives
Degradation rapid and
extensive at almost all
sites after microbial
acclimation.
Does degradea'b'k'y'z'aa'bb
Unclear. Degradation of
unbranched alkanes
rapid and extensive after
microbial
acclimationff'gg'hh.
However, branched
chain alkanes (alkyates)
expected in the range of
gasoline are not shown
to degradeff.
Potential for
Impact on
Natural
Anaerobic
Biodegradation
of BTEX
Yes and by
inferencew'x.
Metabolism may
consume sulfate
and ferric iron
that otherwise
could be used to
degrade benzened
None
demonstrated
Secondary
Pollutants
Associated
with
Degradation
of Additive
Methaneb'cc'dd,
Acetate,
propionate,
butyratecc'ee,
sulfide,
ferrous iron
None
demonstrated
"Suflita and Mormile, 1993
bDa Silva and Alvarez, 2002
cRuiz-Aguilar et al., 2002
dLandmeyeretal., 1998
eAmerson and Johnson, 2002
fMormile et al., 1994
gUS EPA, 2005a
hWilsonetal.,2005a
'Somsamaketal., 2001
JSomsamaketal.,2005
kYeh and Novak, 1994
'Kuder et al., 2005
mKolhatkaretal., 2002
"Finneran and Lovley, 2001
"Wilson etal., 2005b
PUS EPA, 2007
qFinneran and Lovley, 2003
'Novak etal., 1985
"Hickman and Novak, 1989
'Wilson, 2003
"Hovious et al., 1973
vChouetal., 1978
"Barker etal., 1990
"Barker etal., 1992
yAlvarez and Hunt, 1999
zMraviketal., 2003
aaZhang et al., 2006
bbMcDowell etal., 2003
ccMackay et al., 2006
ddBuschecketal., 2001
eePowers etal., 2001
ffRueteretal., 2002
ggCaldwell et al., 1998
"Widdel and Rabus, 2001
To estimate an average rate of biodegradation from the data reported by Shih et
al. (2004), the reported mean plume length was first divided by the average ground water
velocity of 55 feet per year. This provided an estimate of the average travel time in
ground water from the source to the down gradient margin of the plume. The length of
the plume was the distance to the point where the concentrations dropped below
California's concentration based goals (5 ug/L for MTBE and benzene and 10 ug/L for
TEA). Degradation was assumed to follow a pseudo first order rate law; in other words,
80
-------
concentrations declined over time following a half life, much like radioactive decay. 17
These rates are reported in Table V.B-4.
Tong and Rong (2002) also reported the average rate of biodegradation necessary
to calibrate the analytical equations of Domenico (1987) to the field data from the 90
plumes selected from the larger data set reported in Shih et al. (2004). Their average rate
of biodegradation is also presented in Table V.B-4. For MTBE there was good
agreement between Tong and Rong's mean and median of the rate constants calibrated to
the 90 individual plumes, and our calculated estimates made for the entire set of plumes
(described above). There was also good agreement between the average of the rates of
biodegradation available in the literature and our estimated rates of biodegradation for the
population of plumes in Greater Los Angeles. For benzene, there was also good
agreement between the average rate of benzene biodegradation published in the literature
for anaerobic biodegradation of benzene at field scale, and our estimated rate of
attenuation in the Greater Los Angeles plumes.
The close agreement does not prove that the rates of biodegradation of MTBE and
benzene are correct, but they do suggest that the rates published in the literature are
plausible. They also suggest that biodegradation is a primary control on the length of
benzene and MTBE plumes at field scale. In the absence of a field biodegradation rate
constant of at least 1 per year (for substances with a first-order degradation rate), the
plumes formed by substitute additives may be longer than those of MTBE and benzene.
For TEA, the agreement between the calculated literature values for
biodegradation and our estimated average rate of biodegradation in the Greater Los
Angeles plumes are not as close, with the literature rates about an order of magnitude
higher. This may be due to the fact that the rates reported in U.S. EPA (2007) are
dominated by laboratory studies, and rates extracted from laboratory studies are often
faster than rates in the field (Suarez and Rifai, 1999), particularly for degradation limited
by the supply of a critical requisite for microbial metabolism of TEA, such as sulfate.
The published rates of TEA biodegradation are probably less representative of the
achieved rates of TEA biodegradation in contaminated aquifers.
2. Ethanol
Just as with MTBE and the substitutes discussed in the previous section, the
impact of ethanol will depend on the concentration of ethanol that is attained in ground
water in contact with the spill of gasoline and the extent of the plume of ethanol. The
extent of the plume will in turn be controlled by the rate of biodegradation of ethanol.
Anaerobic MTBE biodegradation is dependent upon the presence of a particular
variety of acetogenic anaerobic bacteria. In contrast, methanol, ethanol, and the heavy
17 The half life is equivalent to 0.693 divided by the first order rate constant. The first order rate constant
was estimated as follows:
rate of decay = -natural logarithm [Go%ighest Concentration]'residence time
81
-------
alcohols other than TEA (1-propanol, 2-propanol, and 1-butanol) are readily biodegraded
by anaerobic bacteria that are common and widely distributed, including sulfate-reducing
bacteria and organisms that ferment the alcohols to fatty acids (Eichler and Schink, 1984)
and ultimately to methane (Chou et al., 1978). See Powers et al. (2001) for a review of
anaerobic ethanol biodegradation.
Biodegradation of benzene in ground water is supported by aerobic respiration,
nitrate reduction, sulfate reduction and iron reduction (Suarez and Rifai, 1999;
Wiedemeier et al, 1999). Many of the substitutes for MTBE are metabolized more
readily than benzene; this is particularly true for the alcohols. If the metabolism of the
compounds that substitute for MTBE consumes the available supplies of the electron
acceptors, then the rate and extent of biodegradation of benzene should be restricted, and
the plumes of benzene should be longer than they would have been in the absence of the
substitute (Corseuil et al., 1998; Alvarez and Hunt, 1999; Powers et al., 2001, Da Silva
and Alvarez, 2002; Ruiz-Aguilar et al., 2002).
Ulrich (1999) used parameters from the literature to model the effect of 10,000
mg/L of ethanol on the migration of 10 mg/L of benzene in an aquifer. In a simulation of
conditions relevant to the Borden Aquifer in Ontario, Canada, the presence of ethanol
would make the benzene plume longer by a factor of 1.8. McNab et al. (1999) developed
a screening model for a release of E10 gasoline. Compared to gasoline without ethanol,
the benzene plume was longer by a factor of 1.2 at the 10 |ig/L boundary and by a factor
of 2.0 at the 1 |ig/L boundary. Following up on McNab et al. (1999), McNab (2001) used
a finite-difference-based reactive transport model to evaluate the role of depletion of
electron acceptors on the behavior of ethanol and benzene. He modeled the release of
E10 and conventional gasoline, both containing 1.5% benzene, into an aquifer with 9
mg/L dissolved oxygen, 10 mg/L nitrate and 20 mg/L sulfate. The benzene plume
simulated with E10 was longer than the plume simulated for gasoline without ethanol by
a factor of 2.5.
Deeb et al. (2002) showed in laboratory cultures that ethanol inhibited the aerobic
biodegradation of benzene. They constructed a two-dimensional mathematical model
comparing the aerobic and anaerobic rates of biodegradation of benzene in the presence
and in the absence of ethanol. Their simulations indicated that the presence of ethanol
could make the benzene plumes increase in length by a factor 1.16 to 1.34. Molson et al.
(2002) simulated the effect of the consumption of oxygen by ethanol from an E10 spill on
the length of the benzene plume. Under some circumstances, the simulated benzene
plume was longer than the plume simulated in the absence of ethanol by a factor of 1.5.
82
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Table V.B-4. Comparison of first order rates of anaerobic biodegradation of TEA,
MTBE, and benzene and the zero order rate of biodegradation of ethanol, and the
consequences for time of travel along a contaminated flow path in ground water.
MTBE
Benzene
TEA
Methanol
Ethanol
Summary data from literature on the rate of biodegradation under anaerobic conditions
Mean
Median
Number of rates
Mean plume
length
Geometric mean
of highest
concentration at a
site
Plume boundary
definition
Estimated travel
time along flow
path in plume
Estimated rate
constant for
overall
attenuation
1.0
per yeara
0.41
per year a
10a
1.1
per yearb
45b
9.2
per
year0
7.3
per
year0
14°
Greater Los Angeles*
272
feet
900
Hg/L
5
Hg/L
4.9
years
1.1
per year
167
feet
700
Hg/L
5
Hg/L
3.0
years
1.6
per year
207
feet
1730
Hg/L
10
Hg/L
3.8
years
1.4
per year
Ontario,
Canada8
220
feet
7,000
mg/L
1 mg/L
(Detection
Limit)
1.3
years
15
mg/L per
day
17
mg/L per dayd'e
(Table V.B-6)
13
mg/L per dayd'e
(Table V.B-6)
1Qd,e
(Table V.B-6)
Northwestern
USA11
260
feet
16,000
mg/L
50 mg/L
(Detection
Limit)
0.8
years
55
mg/L per day
Summary data on rate of biodegradation in 90 plumes in Greater Los Angeles Area1
Mean
Median
Number of rates
1.6
per year
1.3
per year
90
aU.S. EPA,2005a
bSuarez and Rifai, 1999
CU.S. EPA, 2007
dSuflita and Mormile, 1993
"Somsamaketal., 2001
'Shihetal., 2004
gBarkeretal., 1990
hBuscheck et al., 2001
'Tong and Rong, 2002
The dominant fuel additives detected by Shih et al. (2004) in ground water in
Greater Los Angeles were MTBE and TEA. Consistent with the estimations in Table
V.B-4 on rates of biodegradation of MTBE and TEA, Shih et al. (2004) recognized that
biodegradation of MTBE and TEA should be expected to consume electron acceptors
83
-------
such as sulfate and Fe(III) in ground water and have an effect on the length of the
benzene plume. In an attempt to evaluate the possible effect of readily degradable
compounds on the length of benzene plumes, Shih et al. (2004) compared the lengths of
benzene plumes that also contained fuel oxygenates to the lengths of plumes that did not
contain fuel oxygenates. The median length of plumes without oxygenates was 130 feet,
while the median length of plumes with oxygenates was 186 feet. Shih et al. (2004)
recognized and acknowledged that the plumes without oxygenates might be shorter
because these plumes would be older and further along in their lifecycle and should
therefore have fewer plumes that were growing and more plumes that were stabilized and
shrinking. Nevertheless, their observation can put a useful upper boundary on the
expected increase in size of benzene plumes due to the influence of a biodegradable fuel
oxygenate at something on the order of 40% longer.
This observation for TEA and the ether oxygenates can be extrapolated to ethanol
by comparing the amount of sulfate required for complete metabolism of MTBE or its
substitutes in gasoline. Table V.B-5 compares the amount of sulfate that would be
required for complete metabolism of various fuel oxygenates in gasoline. In general, the
requirement for sulfate to metabolize the higher alcohols is less than that for MTBE, and
the requirement for ethanol at 2% oxygen (i.e., 5.8% volume) is even less. The
requirement for sulfate to metabolize the ethanol in gasoline at 10% volume is
approximately two-thirds the requirements for metabolism of MTBE at 2% oxygen (i.e.,
11% volume).
Table V.B-5. Comparison of the amount of electron acceptor required to
metabolize MTBE present in gasoline as a fuel oxygenate to the amount required for
TBA and the other butanols, propanols and ethanol.
Fuel Additive
MTBE
at 2% oxygen in
gasoline (11% volume)
TBA and other butanols
at 2% oxygen in
gasoline
Propanols
at 2% oxygen in
gasoline
Ethanol at 2% oxygen in
gasoline (5.8% volume)
Ethanol
at 10% volume in
gasoline
Mass
Fraction
0.11
0.093
0.075
0.058
0.10
Sulfate Requirement
(mass sulfate per mass
oxygenate)
4.08
3.90
3.59
3.13
3.13
Sulfate Requirement
(kilogram per liter
gasoline)
0.33
0.27
0.20
0.13
0.23
On average, the requirement for soluble electron acceptors will be slightly less for
gasoline with 10% ethanol than for gasoline with 11% MTBE. However, the requirement
for soluble electron acceptors for MTBE and ethanol is likely to be satisfied in different
84
-------
ways. Much of the ethanol will stay in the unsaturated zone or the capillary fringe (the
interface between the unsaturated zone and the ground water aquifer) and never enter
ground water. On the other hand, if the release is large and sudden, the initial
concentration of ethanol in ground water can be very high. In general, the plume formed
from the initial release of ethanol will be more concentrated and may travel further in the
aquifer but will last for a short period. MTBE is not as soluble as ethanol and therefore
tends to stay with the residual gasoline. Continued dissolution of MTBE from the
residual gasoline will sustain concentrations of MTBE in ground water in contact with
the residual gasoline. The plume of MTBE and the requirement for electron acceptors to
metabolize MTBE will not extend as far into the aquifer, but the effect on the plume of
benzene will last much longer.
Ruiz-Aguilar et al. (2003) collected statistics on the length of benzene plumes
from 29 sites in Kansas that were contaminated with gasoline that contained ethanol
(most likely E10) and 217 sites in Iowa that were contaminated with gasoline without
ethanol. The median benzene plume length for sites with ethanol use was 263 feet and
for sites without ethanol use was 156 feet. The presence of ethanol in the gasoline
mixture may have increased the length of the benzene plume by approximately 70%.
This postulated effect has now been documented at field scale (Mackay et al.,
2006). A plume with known concentrations of benzene, toluene and o-xylene (BToX)
was created by pumping ground water, amending the ground water with 1 to 3 mg/L of
each of the organic compounds, and injecting the ground water back into the source area
of an old plume emanating from a spill of gasoline at Vandenberg Air Force Base,
California. A second plume was created adjacent to the first that had the same
concentrations of BToX, as well as approximately 50 mg/L ethanol. In the side by side
comparison, the pseudo first order rate of biodegradation of benzene was twenty-fold
slower, and the benzene plume extended up to ten times further in the presence of
ethanol.
There is relatively little direct information available on the concentration of
ethanol, methanol, 1-propanol, 2-propanol, or 1-butanol in ground water in contact with
gasoline spills. Most states do not routinely monitor for these constituents, and when
they are monitored, concentrations are usually below the analytical detection limit. The
limited information that is available is either from releases of fuel grade ethanol
(typically E95) from tank farms, or from controlled-release experiments at pilot scale or
field scale where ethanol or methanol is intentionally introduced at a predetermined
concentration.
There are two controlled release experiments where gasoline with ethanol was
inserted below the water table into aquifer sediment. These experiments were designed
to simulate a spill of gasoline below the water table from an underground storage tank.
Brazilian gasoline with 24% ethanol was injected at a site in Florianopolis, Brazil
(Corseuil et al., 2000), and gasoline with 10% ethanol was injected into an aquifer in
Ontario, Canada (Mocanu et al., 2006). The maximum recorded concentration of ethanol
in ground water down gradient of each of these releases was approximately 2,500 and
85
-------
1,400 mg/L respectively. These concentrations are approximately one thousand-fold
higher than the highest concentrations of benzene, MTBE or TEA typically seen in
ground water at gasoline spill sites (compare Table V.B-4). The high concentrations of
ethanol are caused by the fact that ethanol is miscible in water, and almost all the ethanol
originally in the gasoline dissolves in the ground water. In contrast, benzene and MTBE
have much lower solubility in water relative to ethanol, and therefore most of the benzene
and MTBE remains in the gasoline.
Because the concentrations of ethanol that would be expected from a spill of E10
gasoline are so much higher than concentrations of benzene and MTBE, the rate of
biodegradation of ethanol tends to follow a different pattern. Benzene and MTBE decay
following a pseudo first order rate law (decay described by a half-life), where the rate of
decay depends on the concentration at that particular time. In contrast, the concentration
of ethanol declines at a steady rate that is independent of the concentration at any time.
When the rate of decay is steady and independent of the concentration of the compound,
it is termed a zero order rate. As a consequence, the rate of decay of ethanol should not
show the tailing effect associated with a first order rate constant. Table V.B-6
summarizes rate constants extracted from seven reports of field scale releases of ethanol
and ethanol in fuel (all at varying concentrations).
With one exception, the ethanol decay rates fell into a range between 61 mg/L per
day and 1.4 mg/L per day. The exceptionally high rate of 500 mg/L per day was
observed at a site where the ethanol was continuously introduced as a solution in ground
water released from an injection well (Mackay et al. 2006). All the other rates were
extracted from sites where the ethanol was released as a slug. As bacteria degrade
ethanol, they increase in number. It is likely that the continued source of ethanol allowed
the accumulation of higher and higher densities of ethanol-degrading bacteria in the
aquifer just down gradient of the injection well. This effect has also been noted in the
laboratory. The rate of ethanol degradation in a laboratory column study with a
continuous supply of ethanol reached 13,000 mg/L per day (Da Silva and Alvarez, 2002).
Table V.B-6 compares the average decay rate of the six field studies where
ethanol was released as a slug, plus rates extracted from laboratory studies (Corseuil et
al., 1998; Suflita and Mormile, 1993). The mean and median rates derived from these
studies were 17 and 13 mg/L per day, respectively. As mentioned earlier, the maximum
concentrations attained from spills of gasoline containing ethanol were approximately
2,500 and 1,400 mg/L. The rates above can degrade these ethanol concentrations in 3 to
6 months.
As pointed out by Mocanu et al. (2006), the maximum concentration of ethanol in
ground water (1,400 mg/L) obtained in their experiment, which injected gasoline plus
10% ethanol directly into an aquifer, is likely higher that would be attained in a real spill
of gasoline containing ethanol. If gasoline with 10% ethanol is released above or near the
water table, the ethanol tends to leave the gasoline and accumulate in the capillary fringe
just above the ground water in the aquifer (McDowell and Powers, 2003; Capiro et al.,
86
-------
2007). Dakhel et al. (2003) showed that ethanol from a gasoline release into the
unsaturated zone was readily biodegraded.
Table V.B-4 also compares a large plume of ethanol from a release of fuel grade
ethanol from a tank farm at a gasoline distribution terminal (Buscheck et al., 2001;
McDowell et al., 2003). Although the concentration of ethanol was about ten-fold higher
than would be expected from a gasoline spill from a leaking underground storage tank,
the maximum plume length was only 260 feet, which is roughly comparable to the
average length of benzene, MTBE and TEA plumes in Greater Los Angeles (Shih et al.,
2004). This is the longest ethanol plume documented in the literature (see Table V.B-6).
Table V.B-4 also presents data from a long plume of methanol created in a sand aquifer
at the Canadian Forces Base Borden in Ontario, Canada (Barker et al., 1990). This is a
low-nutrient, oligotrophic aquifer, where we should expect less biodegradation. The field
scale rate of biodegradation of methanol was comparable to the mean of the rates of
biodegradation of ethanol (see Table V.B-6). Again, the length of the plume was
comparable to the average length of benzene, MTBE and TEA plumes in Southern
California. In general, the impact of plumes of alcohol in ground water will not be from
the alcohol as a contaminant but from the effect on the alcohol on the behavior of
benzene plumes. Ethanol can also affect MTBE plumes by promoting natural
biodegradation of MTBE to TEA (Mackay et al., 2007).
As discussed above, most of the available data concerning the impact of ethanol
on benzene plumes are either from spills of technical grade ethanol at storage facilities,
releases of ethanol in pilot scale experiments, or projections of computer models. In
contrast, the survey data summarized in Ruiz-Aguilar et al. (2003) were collected from
spills of gasoline from actual underground storage tanks at automobile service stations.
Therefore, EPA believes that the observations of Ruiz-Aguilar et al. (2003) are the best
available description of the current impact of ethanol in gasoline on the size of benzene
plumes in ground water. The ethanol in an average release of gasoline with 10% ethanol
can be expected to increase the length of the benzene plume by approximately 70%.
However, it is likely that the distribution of plume lengths will be strongly skewed.
Particularly large plumes of benzene can be expected when there is a large area of the
aquifer that is contaminated with liquid phase gasoline [area with floating product more
than 200 feet wide], when the background concentration of sulfate in the ground water is
low [less than 20 mg/L], when the rate of ethanol biodegradation is low [less than 1 mg/L
per day], and when the flow velocity of the ground water is high [more than 1 foot per
day].
C. Secondary Pollutants Associated with Releases of Gasoline with Additives
The first biotransformation product of many of the ether fuel oxygenates (MTBE,
ETBE, TAME, DIPE) is a compound that is often as problematic (Table V.B-3). TEA
accumulates from the initial anaerobic biodegradation of MTBE (Wilson et al., 2005b);
or ETBE (Yeh and Novack, 1994). Tertiary amyl alcohol (TAA) accumulates during
biodegradation of TAME (Somsamak et al., 2001, 2005). Under these conditions, initial
87
-------
biodegradation of the ether oxygenate does not restore the environmental quality of the
ground water.
Because the alcohols are readily degradable and they can enter ground water at
high concentrations, their metabolism can provide significant concentrations of a variety
of metabolic end products. If sulfate is available in the ground water, or biologically
available Fe(III) minerals are present in the aquifer solids, then hydrogen sulfide and
ferrous iron can be produced. Fermentation reactions of the alcohols can also produce
methane and fatty acids. These same materials are produced from the normal anaerobic
metabolism of the BTEX compounds in gasoline (Wiedemeier et al. 1999). However,
adding alcohols at concentrations in the hundreds to thousands of mg/L increases the
production of these materials many fold.
There are other potential hazards than those associated with chemical toxicity;
some spills of gasoline with ethanol may produce methane concentrations in the soil that
pose a risk for explosion (Da Silva and Alvarez, 2002; Powers et al., 2001). Ethanol fed
to a laboratory column at a concentration of 1,000 mg/L was degraded to produce
concentrations of methane that were in excess of the water solubility of methane (i.e.,
more methane was produced than could be dissolved by the available water) (Da Silva
and Alvarez, 2002). At field scale, the methane could bubble out of the ground water and
enter the soil gas (gases that occur in the small spaces between particles of soil) at
explosive concentrations. At the controlled release study at Vandenberg AFB, California,
a release of 500 mg/L ethanol was biologically degraded to produce concentrations of
methane in excess of the water solubility of methane (Mackay et al., 2006). At the spill
of fuel grade ethanol at the Northwest Terminal Site, the concentration of methane in
many wells exceeded the water solubility of methane, and the concentration of methane
in the soil gas above the plume exceeded the upper explosive limit (Buscheck et al.,
2001).
Methane is not routinely monitored in ground water at gasoline spill sites or in the
soil gas above the spills, and there is not enough information currently available to
quantitatively evaluate the risk from explosions associated with methane produced from
alcohols in gasoline. This is an area where more research is needed as the use of ethanol
expands (see Chapter VTI).
88
-------
Table V.B-6. Comparison of rates of anaerobic biodegradation of ethanol in ground water and the consequences for time of
travel along a contaminated flow path in ground water.
Nature of
Release
Site
Initial
concentration
Travel
distance, final
concentration
Estimated
travel time
along flow path
Estimated rate
of removal
Ethanol
dissolved in
water3
Florida
10,000
mg/L after
pumping
entire plume
3.5 years
9 mg/L per
day**
Ethanol
dissolved in
waterb
Platte Valley,
Nebraska
190
mg/L
20 feet to
4.4 mg/L
13 days
14 mg/L per
day*
6 1 mg/L per
day**
Ethanol
dissolved in
water0
Vandenberg
AFB,
California
-500
mg/L
1.5 feet to
1 mg/L
1 day
500 mg/L per
day*
72,000 L
Denatured
Ethanold'e
Pacific
Northwest
Terminal
Tigard, Oregon
16,000
mg/L
260 feet to
<50 mg/L
0.8 years
55 mg/L per
day*
100 L Brazilian
Gasoline
24% Ethanolf
Florianopolis,
Brazil
2503
mg/L
after 16
months
28 feet to
non-detect
3 years
2.3 mg/L per
day*
2.9 mg/L per
day **
50 L Gasoline
with 10%
Ethanol,
followed by
2200 L water8
Borden Field
Site, Ontario,
Canada
1390
mg/L
5 1 feet to
560 mg/L
1.6 years
1 .4 mg/L per
day*
50 L of Ethanol
Denatured with
5% Gasoline,
followed by
2200 L water8
Borden Field
Site, Ontario,
Canada
15700
mg/L
51 feet to
5030 mg/L
1.6 years
1 8 mg/L per
day*
* rate calculated by dividing reduction in concentration by residence time in ground water along the path in the aquifer
** rate estimated by multiplying published pseudo first order rate constant by the initial concentration
aMraviketal.,2003
bZhangetal.,2006
TVIackay et al., 2006
dMcDowelletal.,2003
"Buschek et al., 2001
fCorseuiletal.,2000
gMocanuetal., 2006
89
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VI. POTENTIAL APPROACH FOR FURTHER ANALYSES: COMPREHENSIVE
ENVIRONMENTAL ASSESSMENT OF FUEL ADDITIVE SUBSTITUTE OPTIONS
Any choice among fuel options is accompanied by inherent trade-offs. The
experience with the environmental trade-offs of MTBE over the past two decades amply
illustrates why fuel additive substitute options require careful evaluation in a systematic,
holistic manner. This report presents information on the potential health, water and air
pollution effects of fuel additive usage, but it does not constitute a full, comparative
evaluation of the environmental and health implications of the fuel additives in question,
nor does it examine in depth the secondary and indirect impacts potentially associated
with different fuel additive substitute options. To evaluate the multimedia environmental
benefits and risks of fuels and fuel additives in a scientifically rigorous manner would
require a more extensive effort. For example, EPA anticipates that future reports
required under Section 204 of 2007's Energy Independence and Security Act will go
much further in investigating the full environmental impacts in the context of the
production and use of renewable fuels. One approach for such an analysis is
comprehensive environmental assessment (CEA), which grew out of the MTBE
experience (Davis and Thomas, 2006).
The CEA approach to examining the comparative advantages and disadvantages
of various fuel additives and formulations would help avert substituting a different
problem or set of problems for those associated with MTBE. The present chapter
discusses this CEA approach as a paradigm for future analyses of these issues.
In essence, CEA combines the risk assessment paradigm with a product life cycle
framework and uses collective expert judgment methods to deal with limitations in
empirical data. Some general features of CEA are illustrated in Figure VI-1. Typical
stages of a product life cycle are listed in the first column, including feedstock
production, manufacturing processes, distribution, storage, use, and disposal of the
product and/or waste by-products. At any of these stages, pollutants (whether the
primary chemicals or waste by-products) may be released into one or more environmental
media or pathways, namely air, water, soil, and the food web (second column). Transport
and transformation processes may then come into play, translocating materials from the
immediate area of release and possibly transforming the original pollutants into
secondary pollutants (third column). Exposure can occur when humans or other
organisms (biota) come into contact with these primary and secondary contaminants in
the environment (fourth column). Ultimately, the effects on human health and
ecosystems associated with exposures in the aggregate (across routes) and cumulative
(across pollutants) need to be considered as part of a CEA (fifth column).
Note that CEA should not be equated with life cycle analysis or life cycle impact
assessment. Although CEA can incorporate and build upon such analyses, it involves a
greater range of expertise and is more qualitative in nature than typical life cycle
assessments. Therefore, while CEA can inform decisions, it does not necessarily provide
a definitive quantitative assessment of all impacts associated with different policy
options. However, as illustrated by statements more than 15 years ago from the
97
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Alternative Fuels Research Strategy (U.S. EPA, 1992), a precursor to the CEA approach,
it is possible to identify potential problems in a qualitative sense even before quantitative
assessment is feasible. For example, the Strategy noted:
"Compared to gasoline, the ethers MTBE and ETBE have relatively large aqueous
solubilities and would likely leach more rapidly through soil and ground water.
Also, limited data suggest that ethers may be persistent in subsurface
environments."
and
"Very little is known about emissions and releases from MTBE and ETBE storage
and distribution, making this area an appropriate target for research. Effects on
existing equipment and controls...need to be evaluated."
Even though these statements were not conclusive and offered no quantitative estimation
of risks, they turned out to be prescient forewarnings of problems that ultimately
contributed to the de-selection of MTBE as a fuel additive. In a sense, they offer a proof
of the CEA concept and its applicability to a contemporary comparative assessment of
fuel additive substitute options.
Figure VI-1. Schematic representation of Comprehensive Environmental
Assessment (Source: adapted from Davis and Thomas, 2006)
v>EPA Comprehensive Environmental Assessment
$usa, ^EAj
Life Cycle Environmental Fate &
Stages Pathways Transport
Exposure
Effects
Feedstocks
Manufacture
Distribution
Storage
Use
Disposal
Air ~\
Water
Soil
Food chain
J
Primary
contaminants
Secondary
contaminants
J
Biota
Human
populations
Eco-systems
Human Health
I Office of Research and Development
CEA provides a means to examine both environmental and public health impacts
potentially associated with different fuel formulation options. A more extensive
discussion of the CEA approach is presented in a paper by Davis and Thomas (2006).
That discussion focuses on a comparison of oxygenate options for reformulated gasoline,
98
-------
namely MTBE, ethanol, and a no oxygenate formulation; however, the CEA approach is
easily extended to other fuel formulations. Table VI-1 highlights some examples of
issues to be considered in a CEA of fuel additive substitute options. Beginning with
feedstocks, for example, one might consider impacts associated with each of the three
choices. In North America, methanol is primarily produced from methane, which is an
important greenhouse gas, and can be produced from coal. Is an increase, decrease, or no
change in methane emissions likely to occur if the amount of MTBE derived from
methanol is reduced? In the case of ethanol, the primary feedstock in North America is
currently corn, the production of which typically involves the application of fertilizers
and pesticides and subsequent run-off with precipitation. The herbicide atrazine, for
example, is one of the most prevalent water pollutants in the United States (U.S. EPA,
2006: http://www.epa.gov/safewater/contaminants/dw_contamfs/atrazine.html), which
raises the question of whether increased production of ethanol from corn would result in
greater use of atrazine and, hence, more or more severe contamination of water resources.
Issues about increased nitrous oxide from additional fertilizer utilization have also been
raised (Crutzen et al., 2008). Comparable issues surround increased production of iso-
octane, which could imply a greater potential for transport of hydrofluoric acid. With
transport (e.g., by tank cars) and transfers (e.g., loading and unloading) of hazardous
materials comes some risk of ecological, health and safety impacts associated with
releases.
In addition to the examples highlighted in Table VI-1, Chapter VII of this report
also identifies questions that merit examination as part of a CEA of fuel issues.
A key feature of CEA is the use of the collective intelligence of an array of
technical experts and stakeholders in addressing questions for which information is
lacking. In this respect, diversity of perspectives is critical for reaching sound collective
judgments. Such diversity typically requires consideration of a spectrum of expertise and
viewpoints encompassing several scientific and technical disciplines as well as
stakeholder groups. Although subjective conjecture cannot substitute for empirical data,
the tapping of collective wisdom to deal with scientific uncertainties is a critical
component of the CEA approach. Moreover, the assembled group needs to interact in a
structured manner (Shatkin and Qian, 2004). Simply having a group engage in free-form
discussion is unlikely to achieve a coherent or useful outcome, but structured discussion
affords greater potential for deriving sound collective judgments.
99
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Table VI-1. Illustrative issues to be considered in a CEA of fuel additive/
blending options for reformulated gasoline.
Reformulated Gasoline with:
Source
Characterization
Environmental
Quality
Exposure
Human Health
Effects
Ecosystem
Effects
Global Climate
Feedstocks
Production
Distribution
Storage
Use (evaporative
& combustion
emissions)
Air
Subsurface
Surface Water
Human
Biota
Acute
Chronic
Terrestrial
Aquatic
(marine &
freshwater)
CO2
methane
N2O
CO
NO2
VOCs
MTBE
Methanol:
methane
VOCs
Small/chronic
releases
Air toxics, NO,
CO, ozone, etc.
Formaldehyde,
tert butyl formate
MTBE,
tert butanol
MTBE
Ethanol
Corn: pesticides,
fertilizers
Air toxics, odors
Large/acute releases
Materials
compatibility, product
containment
Acetaldehyde, PAN,
alkylates, etc.
Acetaldehyde, ozone
Increased BTEX,
reduced water supplies
Oxygen depletion,
increased nitrates,
increased dead zones
No Oxygenate
Iso-octane:
hydrofluoric
acid
?
?
Alkylates,
toluene, etc.
VOCs,
secondary
organic aerosols
Acute/chronic; personal and population exposures;
cumulative and aggregate; single chemical and mixtures
Acute/chronic, terrestrial/aquatic
Neurobehavioral, respiratory,
organoleptic (sensory), etc.
Cancer potency,
inhalation reference concentration (RfC),
oral reference dose (RfD)
Organism, population, community, ecosystem
Increase? decrease? no net change?
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The outcome of a CEA of fuel additives would be, at a minimum, a cataloging of
potential environmental and health risks and benefits associated with each option. For
aspects of the problem that are data-rich, a more quantitative approach may be taken.
Although it is possible that a given additive might have a clear preponderance of
advantages or disadvantages, such that little question would exist about its suitability or
lack thereof, more often a mix of pros and cons would need to be weighed against a
different set of characteristics for an alternative choice. Thus, the CEA outcome may not
point to a single option, but it will nevertheless provide decision-makers more complete
information to help understand the potential implications of policy options. Most
importantly, to the extent that potential problems can be anticipated through the CEA
process, it should enable risk managers to take mitigative actions to reduce the
occurrence of adverse outcomes. For example, targeted environmental monitoring might
be indicated for detecting contaminant releases as soon as possible.
The CEA approach is iterative. As information and experience accumulate with
the use of a chosen fuel additive substitute option, it is imperative to reexamine
judgments that were based on earlier, less complete data. In this respect, CEA is
consistent with the idea of adaptive management, which refers to a continuing evaluation
of the implications of a policy decision and adjustment of such decisions as may be
warranted (Linkov et al., 2006; Shatkin and Barry, 2006). Adaptive management
strategies have typically been applied to major natural resource interventions or
conservation actions taken in the face of significant scientific uncertainties. However, the
concept seems equally applicable to fuel formulation options and other complex
technological issues (Davis, 2007). The key point is that periodic reevaluations and
appropriate adjustments need to accompany policy decisions on fuel options.
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References
Crutzen, P. I; Mosier, A. R.; Smith, K. A.; Winiwarter, W. (2008) N2O release from
agro-biofuel production negates global warming reduction by replacing fossil
fuels. Atmos. Chem. Phys. 8: 389-395.
Davis, J. M.; Thomas, V. M. (2006) Systematic approach to evaluating trade-offs among
fuel options: the lessons of MTBE. Ann. N.Y. Acad. Sci. 1076: 498-515.
Davis, J. M. (2007) How to assess the risks of nanotechnology: learning from past
experience. J. Nanosci. Nanotechnol. 7: 402-409.
Linkov, I; Satterstrom, F. K.; Kiker, G. A.; Bridges, T. S.; Benjamin, S. L.; Belluck, D.
A. (2006) From optimization to adaptation: shifting paradigms in environmental
management and their application to remedial decisions. Integr. Environ. Assess.
Manage. 2: 92-98.
Shatkin, J. A.; Qian, S. (2004) Classification schemes for priority setting and decision
making: a selected review of expert judgment, rule-based, and prototype methods.
In: Linkov, I; Ramadan, A. B., eds. Comparative risk assessment and
environmental decision making: proceedings of the NATO advanced research
workshop; October, 2002; Rome (Anzio), Italy. Dordrecht, The Netherlands:
Kluwer; pp. 213-243. (NATO science series IV, earth and environmental
sciences: v. 136).
Shatkin, J. A.; Barry, B. E. (2006) Approaching risk assessment of nanoscale materials.
In: Nanotech 2006: technical proceedings of the 2006 NSTI nanotechnology
conference and trade show, volume 1. Cambridge, MA: Nano Science and
Technology Institute (NSTI); pp. 553-556.
U.S. Environmental Protection Agency. (1992) Alternative fuels research strategy
[external review draft]. Washington, DC: Office of Research and Development;
report no. EPA/600/AP-92/002. Available:
http://www.epa.gov/ncea/pdfs/mtbe/altfuel.pdff28 January, 2008].
U.S. Environmental Protection Agency. (2006) Consumer factsheet on: ATRAZINE.
Washington, DC: Office of Ground Water and Drinking Water. Available:
http://www.epa.gov/safewater/contaminants/dw_contamfs/atrazine.html [29
January, 2008].
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VII. CATALOG OF ONGOING SCIENTIFIC ACTIVITIES RELATED TO MTBE
AND ITS SUBSTITUTES AND INFORMATION GAPS
A. Health Effects Research
1. Ongoing Scientific Activities
a. Alternative Tier 2 Health Effects Testing Program for Gasoline-Oxygenate
Blends
A number of studies were done by the American Petroleum Institute (API) in the
1980s that set the stage for current gasoline-oxygenate blends health effects testing.
These studies were conducted on rats and mice with vaporized unleaded gasoline
(MacFarland et al., 1984; Haider et al., 1984; HEI, 1988). EPA also prepared a major
report on the health effects of unleaded gasoline during the same time period (U.S. EPA,
1987). These studies informed the development of the Alternative Tier 2 health effects
testing program.
In the mid-1990's, EPA initiated a program to evaluate the potential toxicity of
exposure to gasoline vapors alone and containing several potential oxygenate fuel
additives including MTBE, ETBE, ethanol (EtOH), TAME, DIPE and TEA. This
program was authorized under the Alternative Tier 2 provision of the fuel and fuel
additives health effects testing requirements as described in sections 21 l(b)(2) and 21 l(e)
of the Clean Air Act. The American Petroleum Institute (API) was notified of EPA's
intention to develop this testing program on August 20, 1997. Subsequent Federal
Register notices provided opportunities for public comments on the proposal, and a letter
dated November 2, 1998, from Margo Oge, Director of EPA's Office of Transportation
and Air Quality to Carol Henry of API formally established the requirements of the
testing program. The API administered a consortium of fuel and fuel additive
manufacturers, referred to as the Section 21 l(b) Research Group, to share compliance
burdens and costs of fulfilling the testing requirements.
The 21 l(b) health effects testing program required that health effects toxicity
testing be conducted in laboratory animals exposed to the evaporative emissions of
baseline gasoline alone and gasoline plus each of the specified fuel oxygenates added in
proportions reflecting their anticipated typical usage. The program focused on
evaporative rather than combustive emissions because it was thought that the toxicity
from the carbon monoxide concentrations present in gasoline exhaust would likely
present a practical limit to the hydrocarbon concentrations achievable in exhaust.
Therefore, a testing program based on exposure to exhaust from combustion gasoline
plus fuel oxygenates would not yield meaningful data on the relative toxicity of
hydrocarbon combustion products arising from presence of the different fuel oxygenates
in gasoline. The intention of the testing program, then, was to evaluate and compare the
relative toxicity of evaporative emissions of gasoline alone and gasoline with the
specified fuel additives present. A separate program of studies was required to evaluate
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personal exposure levels of combustion emissions in cities using gasoline with or without
oxygenate additives, viz. MTBE and ethanol.
An array of health effects testing was specified in the 21 l(b) program that was
based on specific research needs identified in reports of several expert panels, including
the National Science and Technology Council in 1996/1997, the National Research
Council in 1996, and the Health Effects Institute Oxygenates Evaluation Committee in
1997. In addition, several data gaps were identified in risk assessments and other
analyses performed by EPA (1993, 1994, 1996, 1998). Consistent with these various
expert analyses, the testing program sought to gain information on acute health effects,
carcinogenicity neurotoxicity, reproductive/developmental effects, pharmacokinetics and
potential exposures via air and drinking water. The toxicity testing portion of the
program specified more extensive testing for baseline gasoline and MTBE-gasoline than
the other oxygenate-gasoline mixtures. Testing required for MTBE included the standard
Tier 2 regimen of subchronic toxicity, carcinogenicity, mutagenicity, teratogenicity and
neurotoxicity, and also an Alternative Tier 2 requirement of additional neurotoxicity
testing, a two-generation reproductive study, developmental toxicity testing in two
species, a two-year carcinogenicity study and screening for immunological effects. The
Alternative Tier 2 testing required for oxyfuels other than MTBE-gasoline included the
standard Tier 2 tests plus additional screening for immunological and histopathological
effects. In addition, studies of the pharmacokinetic behavior, consisting of uptake,
distribution metabolism and elimination, were required for the neat oxygenates: ethanol,
ETBE, TAME, DIPE and TEA. For all the fuel oxygenates, EPA would determine if
additional Tier 3 testing would be required after evaluating the results of Tier 2.
The testing rule was designed to incorporate scientific peer review and evaluation
of the study protocols and reports (except for protocols where previously reviewed
standard guidelines were employed). In addition, quality assurance programs were
required to be in accordance with federal guidelines for Good Laboratory Practices and
Vehicle Emissions Inhalation Exposures. The scientific peer review required assessment
of draft protocols and reports by an independent panel of qualified experts prior to
submitting the reports to EPA. EPA then would review the draft reports, the comments
of the panel members, and API's response and revisions made in reply to the peer review
comments. If EPA's assessment indicated that the response to the peer reviewer was
insufficient, or if additional issues of scientific or technical quality were identified, then
those comments would be returned to API, who had an opportunity to address them in a
revised report. Once the issues were resolved, EPA would determine whether the studies
were conducted in a manner sufficient to satisfy the testing requirements. Accepting that
the study reports satisfy the testing requirements, however, would not imply necessarily
that EPA concurred with the interpretation of the study results or the conclusions stated
by the authors.
As of June 2008, some of the study reports required under the Alternative Tier 2
health effects testing program have been received in final form by EPA, while others are
in the process of peer review or revision. A summary of the current status of the
individual study reports is provided in Table VII. A-1.
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Under Section 8e of the Toxic Substances Control Act (TSCA), companies that
observe significant new toxicological outcomes are required to report those findings to
EPA. There have been several TSCA 8e reports filed to EPA as a result of the 21 l(b)
Alternative Tier 2 testing program. A summary of those reports is provided in Table
VII.A-2.
EPA will make an independent evaluation of the results of the studies conducted
under the 21 l(b) program after all of the final reports are received from the Alternative
Tier 2 testing of the substances evaluated. EPA will conduct these evaluations as a
component of a comprehensive assessment of all the data available for each of the
substances involved including, where available, data from other sources and the peer-
reviewed scientific literature. In addition, EPA will determine whether substantial data
gaps remain regarding potential human health effects that necessitate additional testing.
In combination with data on anticipated usage of the fuel oxygenates and estimated levels
of human exposures, the toxicity data from the 21 l(b) testing program, when available,
should improve the ability to make informed risk assessment and risk management
decisions regarding the use of fuel oxygenate compounds.
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Table VII.A-1. Status of reports from health effects portion of the Section 211(b) Alternative Tier 2 testing program
Test
Material
Baseline
gasoline
Gasoline +
MTBE
Gasoline +
EtOH
Gasoline +
TAME
Gasoline +
ETBE
Gasoline +
DIPE
Gasoline +
TEA
Subchronic#
Final Report
Received
Final Report
Received
2nd Draft
Reviewed
1st Draft
Reviewed
1st Draft
Reviewed
No Report
Received
No Report
Received
Developmental
Rat
Final Report
Received
2nd Draft
Reviewed
2nd Draft
Reviewed
2nd Draft
Reviewed
2nd Draft
Reviewed
No Report
Received
1st Draft
Reviewed
Developmental
Mouse
2nd Draft
Reviewed
2nd Draft
Reviewed
Reproductive
2nd Draft
Reviewed
1st Draft
Reviewed
2nd Draft
Reviewed
2nd Draft
Reviewed
1st Draft
Reviewed
1st Draft
Reviewed
1st Draft
Reviewed
Carcinogenicity
1st Draft
Reviewed
1st Draft
Reviewed
ADME/PK*
Sufficient PK data
submitted. No
additional testing.
Sufficient PK data
submitted. No
additional testing.
Reviewed final study
protocols
Reviewed pilot study
protocols
Reviewing draft pilot
study report
# Includes genotoxicity, neurotoxicity and immunotoxicity outcomes
* Absorption, Distribution, Metabolism and Excretion (ADME) or PK testing is done on "neat" oxygenates only (not mixture)
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Table VII.A-2. TSCA 8e reports filed to EPA as a result of the 211(b) Alternative Tier 2 testing program
Substance
Baseline
Gasoline
Vapors
Gasoline +
EtOH
Gasoline +
MTBE
Gasoline +
ETBE
Gasoline +
TAME
Gasoline +
DIPE
Test
Genotoxicity
Developmental
toxicity
Immunotoxicity
Neurotoxicity
Developmental
toxicity
Developmental
toxicity
Immunotoxicity
Reproductive
Toxicity
Immunotoxicity
Findings
Statistically significant increases in sister chromatid exchange in peripheral blood
lymphocytes cultured from rats exposed for 4 weeks to vapor concentrations of 10,000 or
20,000 mg/m3.
Statistically significant reductions of body weight and body weight gain of pregnant mice
during gestational exposure to 20,000 mg/m3. Statistically significant reductions in weight
of gravid uterus and litter size, and an increase in fetal resorption ratio at 20,000 mg/m3.
Statistically significant reduction in the number of antibody forming cells in spleens of rats
exposed to vapor concentrations of 20,000 mg/m3 for 4 weeks.
Statistically significant increases in the concentration of glial fibrillary acidic protein
(GFAP) in dissected brain regions of male rats exposed for 13 weeks, increases over control
were observed at all dose levels in the cerebellum, but were not dose related in other brain
regions, and were not observed in females.
Incidents of fetal malformations in one of 25 litters of mice exposed to low dose (2,000
mg/m3) and one of 25 litters exposed to mid dose (10,000 mg/m3), but not observed in
control or high dose (20,000 mg/m3).
Follow-up to the previous study reports no fetal malformations following exposure to 0,
10,000 or 20,000 mg/m3, but one incident of a similar malformation in 33 liters of mice
exposed to 30,000 mg/m3.
Statistically significant reduction in the number of antibody forming cells in spleens of rats
exposed to vapor concentrations of 10,000 or 20,000 mg/m3 for 4 weeks.
Statistically significant decrease in pup weight gain on postnatal days (PND) 7-14 in litters
of rats exposed to 10,000 and 20,000 mg/m3 during development. Significant reduction in
male pup weight at PND-21 at 2,000 mg/m3, but not at higher concentrations.
Statistically significant reduction in the number of antibody forming cells in spleens of rats
exposed to vapor concentrations of 20,000 mg/m3 for 4 weeks.
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b. National Health and Environmental Effects Research Laboratory
Within EPA's Office of Research and Development, the National Health and
Environmental Effects Research Laboratory (NHEERL) conducts an intramural research
program on the health effects of exposure to air pollution, including those effects caused
by fuels and fuel additives. The major focus of the NHEERL air research program is to
evaluate the health effects of paniculate matter in support of establishing National
Ambient Air Quality Standards. A component of this work focuses on gasoline and
diesel engine emissions, which are major sources of particulate matter in urban air. In
addition, a portion of this research program focuses on near roadway environments, for
which fuel emissions are a major contributor. Finally, a portion of the research program
is evaluating specific fuel components, and is currently assessing pharmacokinetic and
neurotoxic properties of acute inhalation exposure to iso-octane (2,2,4-trimethylpentane).
The first results of the research regarding iso-octane were presented at the annual meeting
of the Society of Toxicology in Seattle, Washington, March 16-20, 2008.
c. IRIS Assessments
The IRIS files for a number of the additives or compounds covered in this report
are either under revision or being developed at this time. A summary of the available
health effects information for these chemicals is provided in Chapter II of this report.
The status of these ongoing assessments can be accessed at
http ://cfpub. epa. gov/ncea/iristrac/index. cfm.
2. Information Gaps
Although scientific uncertainties and questions accompany even the most well
studied chemicals, clearly some fuel additives have larger information gaps than others.
In particular, more toxicity and pharmacokinetic data are much needed for DIPE and iso-
octane. Information gaps also exist for the health effects of TAME, ETBE, and TEA,
albeit not to the degree they exist for DIPE and iso-octane. However, these compounds
are not presently projected to be used as much as those in renewable fuels, which are
mandated by the Energy Independence and Security Act of 2007.
B. Emissions and Air Quality Research
1. Ongoing Activities
EPA's Office of Transportation and Air Quality has initiated a Light Duty Vehicle
Fuels Effects Testing Program for exhaust emissions along with a separate program for
evaporative emissions. These programs are in response to Section 1506 of the 2005
Energy Policy Act (EPAct) which contained a provision requiring the EPA to produce an
updated fuel effects model representing the 2007 light duty gasoline fleet. Such a model,
which will include factors to determine the emissions impacts of ethanol use, will be used
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to support future fuel programs and numerous other regulatory activities in the coming
years, such as:
MOVES or Motor Vehicle Emission Simulator model (SIP, inventory and air
quality analyses)
EPAct and EISA studies (anti-backsliding, fuel harmonization, others)
Future regulatory programs, legislative and policy discussions
The EPAct Program is designed to establish the effects of RVP, T50, T90,
aromatic content and ethanol content on exhaust emissions from Tier 2 vehicles. Its
scope includes exhaust emissions testing of 19 vehicles on 29 fuels over the California
Unified Cycle (LA92). Regulated emissions will be measured along with CC>2, NC>2,
VOCs (volatile organic compounds), alcohols, carbonyl compounds, N2O (nitrous oxide),
NHs (ammonia) and HCN (hydrogen cyanide). Limited testing of high emitter/mileage
vehicles will be done. Detailed speciation of over 100 individual VOC compounds,
including those for the carbonaceous portion of PM and SVOC (semi-volatile organic
compounds), is also being done and will be used for speciation profiles for ozone (and
PM) modeling with CMAQ.
The following levels of investigated fuel parameters have been included in the
fuel matrix with the 29 fuels:
RVP: 7, 9, and 10 psi
T50: 150, 160, 190, 220 and 240°F
T90: 300, 325, and 340°F
Aromatic content: 15, 23, 25, 30, and 40%
Ethanol content: 0, 10, 15 and 20%
Among these 29 fuels are three fuels designed to represent typical gasoline with no
ethanol and ethanol blends of 10% (and also 15%).
The test program consists of three phases:
1. Phase 1: 19 Tier 2 vehicles and 3 high emitter/mileage vehicles tested at 75°F
on three US "average" EO, E10 and El 5 fuels
2. Phase 2: 19 Tier 2 vehicles and 3 high emitter/mileage vehicles tested at 50°F
on three US "average" EO, E10 and El 5 fuels
3. Phase 3: 19 Tier 2 vehicles, including 4 FFVs (flexible-fueled vehicles),
tested on 26 fuels, including an E85 blend.
The EPAct Program is a collaboration between the EPA and DOE and is being
conducted at the Southwest Research Institute in 2008 - 2009.
The above testing is for exhaust emissions. Evaporative emissions testing is also
being done largely through the Coordinating Research Council for newer vehicles over
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the next several years to address similar issues related to the effect of fuels on VOC
emissions, including some information on detailed VOC speciation.
A more limited test program is being done on 12 nonroad engines with 18 fuels
representing a variety of gasoline/alcohol blends.
2. Information Gaps
Although these test programs are very comprehensive, there are still some data
gaps, including the following:
Emission impacts on older vehicles. Older vehicles (i.e., pre-Tier 2 vehicles)
are not being tested in this program. While such vehicles have been tested in
prior programs (such as the cooperative Auto/Oil program done in the early
1990s), the earlier testing did not, of course, specifically address current fuels
(i.e., Tier 2 low sulfur, RFG) with ethanol.
Emissions under cold temperature. Emission testing is being done at 75 and
50°F. This program will not have results for colder temperatures, such as 20°F,
where we know emissions are different.
Emissions under extreme driving conditions. The exhaust testing is being done
using the California Unified Cycle (LA 92), which is similar to the Federal Test
Procedure (FTP) used in vehicle certification to determine compliance with EPA
emission standards. There are other driving conditions, such as extremely high
load (pulling a trailer up a hill) or constant low speed driving, that are not
included here.
Long term emission impacts. The long term emission impacts of using these
blends for, say, 100,000-150,000 miles are not being determined.
C. Water Quality Research
1. Ongoing Activities
Within EPA's Office of Research and Development, the National Risk
Management Research Laboratory (NRMRL) and National Exposure Research
Laboratory (NERL) have examined the fate and transport of fuel and fuel additives in
laboratory and field studies (Mackay, 2006; USEPA, 2005). NRMRL and NERL have
also developed conceptual and predictive models to understand the behavior of these
plumes in the environment (see the OnSite tool for modeling subsurface petroleum
hydrocarbon content http://www.epa.gov/athens/onsite and the Center for Subsurface
Modeling Support http://www epa gov/ada/csmos.html). This work has included
conducting research on the composition, fate and transport, and active and passive
treatment of fuel spills from leaking underground storage tank (LUST) sites. Specific to
ethanol, EPA has conducted laboratory and field studies (including at Vandenberg Air
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Force Base [Mackay, 2006]). These studies assessed the fate and transport of ethanol in
the environment and its impact on the behavior of the conventional constituents in
gasoline, such as benzene, toluene and xylene.
2. Information Gaps
In the area of impacts of fuels leaking from storage tanks and associated
infrastructure, research is needed in the following areas:
Ethanol may contribute to the deterioration and corrosion of materials used in the
storage of fuels, including tanks, pipes, fittings, and seals. These failures could
lead to releases of fuel to ground water. EPA is undertaking an assessment of
available data in the public and private sectors to determine the vulnerabilities in
the system as a function of the fuel type, e.g. E20, E85. This evaluation will
highlight the key areas where further research is needed to reduce these
vulnerabilities.
Further data is needed to determine the impact of ethanol on the performance of
leak detection systems for gas stations. There are two aspects to this problem.
First, ethanol may affect the integrity of these leak detection systems that have
been designed for petroleum hydrocarbons. Performance of these systems needs
to be ascertained to ensure their integrity with various fuel blends. Second, as
ethanol can extend BTEX plume lengths, the sensitivity of these systems needs to
be assessed to minimize ground water impacts.
A tool is needed for remediation site managers and water utilities that allows them
to integrate the contaminant source modeling and the water supply pumping rates
to determine the rate and extent of remediation needed to protect existing and
future water supplies. EPA is presently exploring the possibility of collaborative
field studies with states to examine the fate and transport of ethanol releases from
underground storage tanks and their potential impact on water supply wells. This
work will assess the influence of water supply pumping rates on the movement of
the contaminated plume.
As documented in the literature, some spills of ethanol may produce methane
concentrations at levels that potentially pose an explosion risk. A laboratory
column fed with 1,000 mg/L ethanol produced methane concentrations in excess
of the solubility of methane (Da Silva and Alvarez, 2002). At the controlled
release study at Vandenberg AFB, California, a release of 500 mg/L ethanol was
biologically degraded to produce concentrations of methane in excess of the water
solubility of methane (Mackay et al., 2006). Once the solubility is exceeded,
there is the potential for gas-phase methane to occur at explosive concentrations.
At a spill of fuel grade ethanol at the Northwest Terminal Site, the concentration
of methane in many wells exceeded the water solubility of methane, and the
concentration of methane in the soil gas above the plume exceeded the explosive
limit (Buscheck et al., 2001). As part of EPA's research field study assessing the
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impact of ethanol on hydrocarbons in fuel, the formation of methane was
examined. EPA is developing of modeling software for the assessment of fuels of
varying composition on ground water, with simulation of methane production
being one component of this work.
In the area of impacts of increased biofuel feedstock production on water quality,
which is beyond the scope of this report, more research is needed in the following areas:
The large-scale production of ethanol from corn may have significant water
quality impacts. From a recent NAS report on the water implications of biofuels,
expansion of corn production on marginal lands or soils that do not adequately
retain nutrients can result in increased loading of nutrients as well as sediment in
waterways. This may further accentuate the hypoxia in the Gulf.
An analysis is needed, at the local and regional scales, to better identify the
available water supplies in areas impacted by corn production. Corn is being
grown and ethanol produced in areas already managing water shortages. This
analysis will be important for the U.S. in managing our existing and future water
supply challenges.
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References
Buscheck, T. E.; O'Reilly, K.; Koschal, G.; O'Regan, G. (2001) Ethanol in groundwater
at a Pacific Northwest terminal. In: Ground water: prevention, detection, and
remediation: proceedings of the petroleum hydrocarbons and organic chemicals
conference; November, 2001, Houston, TX; National Ground Water Association.
pp. 55-66.
Da Silva, M. L. B.; Alvarez, P. J. J. (2002) Effects of ethanol versus MTBE on benzene,
toluene, ethylbenzene, and xylene natural attenuation in aquifer columns. J.
Environ. Eng. 128: 862-867.
Haider, C. A.; Warne, T. M.; Hatoum, N. S. (1984) Renal toxicity of gasoline and related
petroleum naphthas in male rats. In: Mehlman, M. A.; Hemstreet, G. P.Ill;
Thorpe, J. J.; Weaver, N. K., eds. Advances in modern environmental toxicology.
Volume VII: Renal effects of petroleum hydrocarbons. Princeton, N.J.: Princeton
Scientific Publishers, 73-88.
MacFarland, H. N.; Ulrich, C. E.; Holdsworth, C. E.; Kitchen, D. N.; Halliwell, W. H.;
Blum, S. C. (1984) A chronic inhalation study with unleaded gasoline vapor. Int.
J. Toxicol. 3:231-248.
Mackay, D. M.; de Sieyes, N. R.; Einarson, M. D.; Feris, K. P.; Pappas, A. A.; Wood, I.
A.; Jacobson, L.; Justice, L. G.; Noske, M. N.; Scow, K. M.; Wilson, J. T. (2006)
Impact of ethanol on the natural attenuation of benzene, toluene, and o-xylene in a
normally sulfate-reducing aquifer. Environ. Sci . Technol. 40: 6123-6130.
U.S. Environmental Protection Agency (1987) Evaluation of the carcinogenicity of
unleaded gasoline. Washington, DC: Office of Research and Development; report
no. EPA/600/6-87/001.
U.S. Environmental Protection Agency (2005) Predicted ground water, soil and soil gas
impacts from U.S. gasolines, 2004: First analysis of the autumnal data. Athens,
GA: Office of Research and Development; report no. EPA/600/R-05/032.
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APPENDIX A. SUMMARY OF THE IMPACTS OF GASOLINE REGULATION ON
FUEL COMPOSITION
The Clean Air Act Amendments of 1990 called on EPA to regulate gasoline
composition throughout the United States. These amendments created the reformulated
and winter oxygenated gasoline programs and contained "anti-dumping" provisions that
specified requirements for conventional gasoline. Effectively, all areas of the United
States were covered by one or more of these programs. This appendix reviews aspects of
these regulations and resulting changes in gasoline composition. Characteristics and
trends in reformulated gasoline (RFG) composition can be found in the Fuel Trends
Report: Gasoline 1995 - 2005, released by the EPA on January 2, 2008 (U.S. EPA,
2008a). Aggregated gasoline parameter data from the RFG property survey data may be
found at the "Gasoline Properties Data" link at http://www.epa.gov/otaq/rfg.htm (U.S.
EPA, 2008b). Information from the Fuel Trends Report and RFG property survey data
were not included in this report due to time constraints.
A. Gasoline formulations
1. Reformulated Gasoline
Among other requirements of the Clean Air Act, reformulated gasoline (RFG)
used in certain parts of the country (Figure A-l) was nominally 18 required to contain at
least 2% oxygen by weight until 2006. This requirement was removed effective May,
2006 (71 FR26691, May 8, 2006) for most of the federal program areas in the U.S. and
earlier in California (April 24, 2006 in California, see U.S. EPA, 2006). Because
different chemicals could be used to meet the RFG oxygen requirement, the
concentrations used of specific oxygenates were variable (see Table A-l). In addition the
benzene content of RFG could not exceed a nominal level of 1% by volume. When the
oxygenate mandate was in force, for example, because of the ability to choose a per
gallon or average basis for meeting the requirements, compliant RFG could contain
between 8.4% and 11.7% MTBE, or as much as 1.3 % benzene.
2. Conventional Gasoline
Conventional gasoline (CG) is gasoline that does not meet the requirements of
RFG. The benzene content of conventional gasoline is determined from either a default
producer/importer baseline or a baseline based on 1990 production Ievelsl9. Since the
baselines are associated with the producer/importer and not specific locations, the
benzene levels in conventional gasoline at any given location might change, due to
supply considerations.
18 The word nominal is used because the standards could be met either on a per gallon basis or averaged
basis.
19 Note that this refers to gasoline produced before promulgation of the 2007 Mobile Source Air Toxics
rule. The impacts of this regulation are not discussed in this report.
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Table A-l. Oxygen and approximate amounts of oxygenates to meet requirements
of reformulated and winter oxygenate gasoline.
Weight percent of
oxygen
Required Oxygenate Concentration
RFC
Per-gallon
Basis
2.0 %
Averaged Basis
Standard
2.1%
Minimum
1.5%
Winter Oxygenate
2.7%
3.1%
Common Oxygenates and Required Content to Meet Oxygen Requirements (Volume Percent Basis)
Methyl tertiary butyl
ether (MTBE)
Ethyl tertiary butyl ether
(ETBE)
Tertiary amyl methyl
ether (TAME)
Di-isopropyl ether
(DIPE)
Ethanol
11.1%
13.0%
12.4%
13.2%
5.5%
11.7%
13.6%
13.0%
13.8%
5.7%
8.4%
9.7%
9.3%
9.9%
4.1%
15.0%
17.5%
16.8%
17.8%
5.1%
17.3%
20.1%
19.2%
20.4%
5.9%
Approx. Vol % = 100* (required weight fraction oxygen) / (weight oxygen per molecule of oxygenate)
x (gasoline density) / (oxygenate density)
Gasoline density assumed equal to 0.75 g/ml
Figure A-l. Locations in the United States where Federal reformulated gasoline has
been required. The effective date ranges for use of oxygenated additives are given
in the legend.
Conventional gasolines were not required by federal rule to contain oxygenated additives,
but these products could be used to boost a gasoline's octane rating and many oxygenates
115
-------
had been approved by EPA as additives to gasoline. Also, although still considered
conventional gasoline, some states may have renewable fuel standards in place for
gasoline with specific requirements for blending ethanol (oxygenate) into their state's
gasoline pool. Further, samples of premium (i.e., higher octane rated), conventional
gasoline often show higher levels of oxygenates than do samples of regular (lower octane
rated) gasoline from the same location (U.S. EPA, 2005).
3. Oxygenated gasoline
The 1990 Amendments also required states to mandate at least 2.7% oxygen by
weight in gasoline sold in areas of non-attainment of the National Ambient Air Quality
Standards (NAAQS) for carbon monoxide (CO) (Figure A-2). This requirement was
imposed by statute for at least four months of the year when the ambient CO
concentrations were the highest. EPA could reduce this period if a state could
demonstrate that there would be no exceedances of the carbon monoxide NAAQS.
The program began during the late fall of 1992 in 36 areas with one more added
in 1993. Table A-2 lists the twelve areas that remained in the program until the winter of
2005/2006, along with their required oxygen content and effective months.
B. State MTBE Bans
Because of concerns over ground water contamination due to leaks of MTBE-
gasoline, some states began passing bans of MTBE (Figure A-3). One city (Chicago) and
one county (Washoe County, Nevada) also banned MTBE. A few states banned other
ethers and oxygenates (Arizona, California, Minnesota, New Hampshire, Rhode Island,
and Vermont). In the majority of cases, the bans allowed for a relatively low level (0.05
to 1.0 volume %) of MTBE to remain in gasoline (Weaver et al., 2007) and were
implemented between the years 2000 and 2007. One ban has yet to become effective
(New Jersey).
116
-------
Figure A-2. Locations of past (triangles) and present (circles) winter oxygenate
programs.
Winter Oxygenate Fuel Program
in the United States
Atlantic
Ocean
Source: USEPA EPA420-B-05-013 "State
Winter Oxygenated Fuel Program Requirements"
http://www.epa.gov/otaq/regs/fuels/420b05013.pdf
Y Current Cities
Z Past Cities
U.S. EPA Athens, GA
May 2007
httpi/Mww. epa.gov/athens/
research/regsupport/ust .html
Current Winter Oxygenate cities: El Paso, TX; Denver/Boulder & Longmont, CO;
Missoula, MT; Provo/Orem, UT; Las Vegas 8- Reno, NV; Phoenix 8- Tucson, AZ;
Los Angeles, CA; Albuquerque, NM; Portland, OR. Minnesota has adopted a
year-round, statewide oxygen mandate.
Table A-2. Cities implementing the winter oxygenate program through the winter
of 2007-2008 (U.S EPA, 2008c).
Control Period
10/1 to 1/31
10/1 to 3/31
11/1 to 2/29
11/2 to 3/31
Area
(Consolidated Metropolitan
Statistical Area)
Reno, NV
El Paso, TX
Las Vegas, NV
Reno, NV
Tucson , AZ
Albuquerque, NM
Missoula, MT
Los Angeles, CA
Phoenix, AZ
117
-------
Figure A-3. States, cities and counties with MTBE bans and maximum permissible
MTBE levels with effective dates from 2000 to 2009.
Total MTBE ban or not
greater than
0.05 vol%
0.3 vol%
0.5 vol%
0.6 vol%
1.0 vol%
Total MTBE ban
Trace amounts only
No ban
C. Effect of regulation on gasoline composition
The effects of these regulations can be seen in historical gasoline composition
data. The National Institute for Petroleum and Energy Research (NIPER) collected data
on gasoline composition beginning in the 1930's. The industry-lead successor (Northup-
Grumman) continues to collect data from roughly 1,000 gasoline samples taken twice a
year. Data have been collected in 174 areas around the country, although the number of
locations with continuous participation is lower (about 35). The samples are analyzed
with industry-standard methods (e.g., Dickson, 2007), but the selection criteria used to
determine which locations to include are not disclosed. The data provide the most
extensive publicly-available, historical record of both conventional and reformulated
gasoline composition throughout the entire United States. These data address issues such
as variability of the vendor's local fuel source and changing fuel composition
requirements over time. The number of samples per city included, however, is limited in
some cases, and the data might under-reflect the true variability. In contrast, a review of
composited data from the entire country shows large scale trends in fuel usage. See U.S.
EPA (2008) for a comprehensive study of nationwide trends in gasoline composition
covering the period from 1995 to 2005. The following section extends this material by
examining data for specific cities for a similar time period.
1. Reformulated Gasoline
118
-------
The MTBE content in premium New York City gasoline is shown in Figure A-4.
New York City entered the oxygenated gasoline program in 1993 and continued until
2000. The figure shows that the MTBE content of premium gasoline increased in the
winter months, consistent with regulatory requirements. In the summer months, between
1993 and 1995, the MTBE content dropped to levels that are consistent with MTBE use
as an octane enhancer. In 1995, the RFG program started and the MTBE content
remained near the required level to attain 2.0% oxygen by weight, year round, although
some of the required oxygen was supplied by TAME. In 2004, MTBE was banned in
New York, and the MTBE level dropped below the state's maximum allowable level of
0.05 vol %. Ethanol replaced the ethers as the oxygenate after this time. In 2006, the
oxygenate requirement of RFG was removed. However, ethanol continued to be used in
the NY RFG market, likely for a number of reasons, including compliance with the RFG
specifications for criteria and toxic pollutants, its value as an octane component, and the
positive economics of blending ethanol into gasoline.
Benzene content in New York City gasoline is shown in Figure A-5. Prior to the
beginning of the RFG program in 1995, the benzene content ranged above 4% in some
samples, and the average was sometimes above 1%. After 1995, the average benzene
level remained below 1%. Some individual samples show benzene content above 1%,
but these samples may comply with the regulation on an averaged basis, as permitted by
the statute.
Since limits on the benzene content of reformulated gasoline were set by
regulation, there is little observable difference in the benzene content of reformulated
gasoline made with MTBE and reformulated gasoline made with ethanol. EPA RFG
survey data for 2002, 2004, and 2006 show that the average MTBE content of RFG had
decreased, the average ethanol content had increased, and the average benzene content
remained nearly the same over this time period (Table A-3).
Data from five RFG-using cities (Boston, Los Angeles, New York City, San
Francisco and Washington, DC) are compared in Figures A-6 and A-7. All contents are
plotted as the average value for premium reformulated gasoline from the Northup-
Grumman data set (Dickson, 2007). Regular reformulated gasoline is, for the most part,
similar because of the RFG regulations. The benzene content of this gasoline (Figure A-
6) varied prior to the implementation of the reformulated gasoline program and, on
average, was as high as 2.5% volume. After the beginning of the RFG program in 1995,
the benzene content dropped below 1% volume in all cities, with few exceptions. Figure
A-7 shows the average MTBE, TAME and ethanol content for the five cities. In each
case, the MTBE content in the gasoline increased so that by 1995 the MTBE content in
these cities was close to the mandated weight percent level. In two cases, New York and
Washington D.C., the large fluctuations in MTBE content are due to wintertime
requirements for oxygenated gasoline that began in 1992. Of the other oxygenates
(ETBE, TAME, DIPE, TEA and methanol) only TAME had widespread and sustained
use in these cities. Its period of usage corresponds to the beginning of the RFG program
through the passage of the MTBE bans, when its usage also declined. This decline
occurred despite the lack of explicit TAME or other ether bans in most states.
119
-------
Figure A-4. Historical MTBE usage in New York City. Data points show the
average level of MTBE and TAME, which supplied most of the oxygen for this
location. The New York State MTBE ban had the effect of eliminating both ethers
in 2004, which were then replaced by ethanol. The required oxygen level is given
by the thin line and reflects requirements for oxygenated and reformulated gasoline.
4 i
3
2
CD
D)
X
O
1
Oxygen Content From :
MTBE
MTBE and TAME
Ethanol
Required Level
1988
1992
1996
2000
2004
2008
Date
120
-------
Figure A-5. Benzene concentration in premium, New York City, gasoline. The
average benzene concentration was required to be below 1% by volume after
adoption of the reformulated gasoline program in 1995.
Benzene (vol %), High Octane, New York City, NY
4
m -3
N *>
o>
m
* V*.
: *«**,:*» ::** * * %
1975 1980 1985 1990 1995 2000 2005 2010
Date
Table A-3. Average MTBE, ethanol, and benzene contents (% vol.) for
reformulated gasoline in the United States taken from EPA reformulated gasoline
survey data.
Constituent
MTBE
Ethanol
Benzene
Average Content
2002
7.93
2.20
0.61
Average Content
2004
5.78
4.39
0.65
Average Content
2006
0.008
9.20
0.69
TAME was the second most prevalent ether in these cities and is shown in the
second column of Figure A-7. Varying amounts of TAME were used in these cities, the
most in Boston and the least in Los Angeles. In Boston, for example, an increase in
TAME use prior to the year 2000 corresponds to a decrease in MTBE content. Similar
corresponding peaks and valleys can be seen for New York, Washington D.C. and San
Francisco. MTBE usage declined either due to the state bans (effective in 2003 in
California and 2004 in New York), or due to the removal of the oxygenate mandate in
2006 in states with no MTBE ban (Massachusetts and the District of Columbia for Figure
A-7).
121
-------
Ethanol was used throughout the sampling period at fluctuating levels. Although
never used as consistently as MTBE, its usage appears to have decreased during the
period of intensive MTBE usage. This pattern is visible for Boston, San Francisco and
Los Angeles. Ethanol began to be used again when the New York and California MTBE
bans became effective and later when the oxygenate mandate for RFG was removed.
Although TEA had limited independent usage as an oxygenate, it is a known by-
product of MTBE manufacture. Typically the TEA content of RFG has been 1% or less
of the MTBE content (Figure A-8) for this reason.
Figure A-6. Benzene content in five cities using reformulated gasoline.
3 i
O
-------
Figure A-7. MBTE, TAME and ethanol content
in five reformulated-gasoline using cities.
c
£
to
o
m
<,
>i
;) New York Cit
m
O
ci
ashington
5
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o
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in
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£ 5-
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15
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15 -n
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ss 10 -
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i\ t
y-AV
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nl I
2010 1990 2000 2010 1990 2000 2010
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10 -
5 -
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2010 1990 2000 2010 1990 2000 2010
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2010 1990 2000 2010 1990 2000 2010
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10 -
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15 n 15 n
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~
5 -
- , n
10 -
5 -
_A«.. , n
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1990 2000 2010
Date
1990 2000 2010
Date
1990 2000 2010
Date
123
-------
Figure A-8. Scatter plot of TEA content versus MTBE content
in five cities using reformulated gasoline.
0.1 ,
0.1
4-
O
a
A
TBA-MTBE
Boston
Los Angeles
New York
San Francisco
Washington, DC
"o
c
o
O
<
CQ
o.o H
0.0
H3
+
+
D
+ A-A
+ D
4 8
MTBE Content (Vol %)
2. Conventional Gasoline
Average benzene, MTBE and ethanol levels in regular and premium conventional
gasoline from six cities are shown in Figure A-9. Average benzene contents above 1%
were not uncommon, although the values for Cincinnati, Las Vegas, Kansas City and
Miami have tended to stay below 1% since the early 1990s. Average benzene contents
for Spokane tended to stay above 1% over the same time period, although these appear to
be reduced below the levels for the early 1990s (when these cities were first included in
the sampling). The higher benzene levels in Memphis (above 2% volume) may be due to
characteristics of the local refinery that produces gasoline in this area or its crude oil
sources. In several of these cities (Cincinnati, Kansas City, Miami and Spokane), the
benzene content in regular gasoline tended to be slightly higher than in the premium. In
these cities, the premium gasoline usually contained significantly more MTBE than the
regular gasoline.
For the most part, MTBE was used in premium conventional gasoline at higher
levels than in regular conventional gasoline. The majority of the average MTBE contents
have been 4% or less, which contrasts with the higher levels used in RFG. In 2004,
Kansas and Washington banned MTBE, followed by Ohio in 2005. The MTBE levels in
these cities dropped at these times. Memphis is an example where the average MTBE
level remained below 4% during the period when other states were banning MTBE.
124
-------
Miami, however, experienced an increase in MTBE content. By 2006, however, the
MTBE was eliminated in both of these cities.
The other ethers (ETBE, TAME and DIPE) were used only occasionally in these
cities (Table A-4). ETBE was only seen at low concentrations (less than 0.4% volume)
and was used infrequently. TAME had a maximum concentration of 5.9% volume in
premium Memphis conventional gasoline and a maximum of 5.4% volume in premium
Miami gasoline. Usage in other cities was at concentrations of 1% or lower. TAME was
seen frequently in Miami premium conventional gasoline and less so in Miami regular,
Memphis premium and Cincinnati premium conventional gasoline. Usage in other cities
was infrequent. DIPE was seen most frequently in Memphis premium gasoline, but at a
maximum concentration of 0.6%. Other usage of DIPE was less frequent and at lower
concentrations.
The average ethanol content in premium conventional gasoline showed variability
in the six cities (Figure A-9). Miami had no ethanol usage after the late 1980s, a feature
that is probably related to the relatively high levels of MTBE found in Miami gasoline.
Memphis ethanol usage became negligible after some usage in the late 1980s. Kansas
City had consistent usage at about 2% or less. Las Vegas and Spokane had seasonally
varying usage due to their participation in the winter oxygenate gasoline program. These
levels were similar for premium and regular in Las Vegas, but ethanol was used only in
premium gasoline in Spokane. Cincinnati had a variable amount of ethanol throughout
the time period where data were collected. Methanol and TEA were used infrequently
and at low levels in regular and premium gasoline in the six cities studied (Table A-5).
125
-------
Figure A-9. Benzene, MTBE and ethanol content in six cities using conventional
gasoline. The gray solid lines indicate premium gasoline, and the dashed lines
indicate regular gasoline.
ro
c
c
'o
c
b
4 -i
o
Benzene
O
(/)
ro
«)
c
ro
1990 2000 2010
4 -i
o
«)
ro
D)
0
O
Q.
E
-------
Table A-4. Summary of maximum reported ether (ETBE, TAME, DIPE)
concentration, date, and number of non-zero measurements above 0.1% volume per
number of times measurements were made for each of the six cities using
conventional gasoline.
City
Regular
Max
Cone.
(%vol)
Date
Number
Premium
Max
Cone.
(%vol)
Date
Number
ETBE
Cincinnati
Kansas City
Las Vegas
Memphis
Miami
Spokane
0.1
0.1
0.2
0.2
Summer, 1995
Winter, 95/96
Summer, 1994
Summer, 1995
0/21
1/17
1/24
2/24
0/10
0.4
0.1
0.4
Winter, 96/97
Summer, 1994
Winter, 96/97
2/23
0/20
0/20
1/24
2/24
0/20
TAME
Cincinnati
Kansas City
Las Vegas
Memphis
Miami
Spokane
0.2
0.6
0.3
3.3
0.1
Winter, 96/97
Summer, 1997
Summer, 2004
Winter, 04/05
Summer, 1996
4/27
3/21
0/19
2/24
16/25
1/13
0.8
2.3
0.3
5.9
5.4
1.2
Summer, 1994
Summer, 1997
Winter, 94/95
Summer, 1994
Summer, 2004
Summer, 1996
11/26
6/22
2/23
10/26
25/27
2/22
DIPE
Cincinnati
Kansas City
Las Vegas
Memphis
Miami
Spokane
0.3
0.2
0.2
0.3
0.2
0.2
Winter 93/94
Winter 94/95
Summer 1995
Summer, 1995
Winter, 93/94
Summer, 1994
Winter, 94/95
Summer, 1995
Summer, 1995
4/22
3/21
2/17
8/23
4/23
1/10
0.3
0.2
0.1
0.6
0.4
0.1
Winter 94/95
Winter, 94/95
Sumer 1995
Winter, 94/95
Summer 1995
Summer, 1996
Winter, 99/00
Winter, 00/01
Winter 02/03
Winter 94/95
Summer, 1995
2/22
3/20
2/20
11/23
3/23
1/20
127
-------
Table A-5. Summary of maximum reported alcohol (methanol, TEA)
concentration, date, and number of non-zero measurements above 0.1% volume per
number of times measurements were made for each of the six cities using
conventional gasoline.
City
Regular
Max
Cone.
(%vol)
Date
Number
Premium
Max
Cone.
(%vol)
Date
Number
Methanol
Cincinnati
Kansas City
Las Vegas
Memphis
Miami
Spokane
0.2
0.1
~
0.3
Winter 95/96
Winter 85/86
Summer 1986
Winter 86/87
~
Winter 96/97
1/38
3/23
0/20
0/36
3/35
0/18
(a)
0.1
0.1
0.1
0.1
0.1
Winter 87/88
Winter 94/95
Winter 98/99
Winter 85/86
Summer 1986
Winter 86/87
Winter 90/91
Summer 1991
Summer 9 1/92
Winter 85/86
Summer 1996
Summer 1998
Summer 1999
Winter 01/02
Summer 2004
Winter 95/96
1/20
2/23
6/40
6/37
2/29
TEA
Cincinnati
Kansas City
Las Vegas
Memphis
Miami
Spokane
0.3
1.3
0.2
0.5
-
Winter 96/97
Summer 1985
Winter 85/86
Winter 90/91
Winter 96/97
-
2/38
4/23
2/20
0/36
1/35
0/18
0.3
0.1
0.2
0.2
0.5
0.1
Winter 96/97
Winter 86/87
Winter 90/91
Winter 9 1/92
Winter 87/88
Winter 95/96
Winter 1987
10/37
1/20
2/24
7/40
11/37
1/27
(a) data not available for Cincinnati premium
128
-------
References
Dickson, C.L. (2007) Motor Gasolines, Summer 2006, Northrop Grumman, Bartlesville,
Oklahoma, NGMS-246 PPS 2007/1.
U.S. Environmental Protection Agency (2005) Predicted ground water, soil and soil gas
impacts from U.S. gasolines, 2004: First analysis of the autumnal data. Athens,
GA: Office of Research and Development; report no. EPA/600/R-05/032.
U. S. Environmental Protection Agency (2006) Regulatory Announcement, Removal of
reformulated gasoline oxygen content requirement. Washington, DC: Office of
Transportation and Air Quality. EPA420-F-06-035.
U. S. Environmental Protection Agency (2008a) RFG Property Survey Data, "Gasoline
Properties Data" link at http://www.epa.gov/otaq/rfg.htm [9 June, 2008].
U. S. Environmental Protection Agency (2008b) Fuel Trends Report: Gasoline 1995-
2005. Washington, DC: Office of Transportation and Air Quality; report no.
EPA420-R-08-002.
U. S. Environmental Protection Agency (2008c) State Winter Oxygenated Fuel Program
Requirements for Attainment or Maintenance of CO NAAQS. Washington, DC:
Office of Transportation and Air Quality. EPA 420-B-08-006.
Weaver, J. W.; Exum, L. E.; Prieto, L. M. (2008) Historical impacts of federal regulation
on gasoline composition 1990 - 2007. Accepted for publication in Environ.
Forensics.
129
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APPENDIX B. PHASE PARTITIONING OF COMPOUNDS FROM GASOLINE
The dissolution of specific compounds from gasoline depends on at least four
factors: gasoline composition, equilibrium phase partitioning, mass transfer limitations
and geologic heterogeneity. Gasoline composition is described in more detail in
Appendix A.
A. Equilibrium Partitioning
At equilibrium, partitioning of aromatics such as benzene, toluene and xylene has
been found to follow Raoult's law which states that the effective solubility, C, is equal to
the mole fraction of the chemical in gasoline, X, multiplied by the pure phase solubility
of the compound, S. For constituent, i,
From this relationship a partition coefficient, K°, can be calculated from
Co /~io o i
jar = -i- = P l
X° S MW° S?
where p° is the density of the gasoline and MW° is the average molecular weight of the
gasoline. Thus the partition coefficient for chemical (/') is inversely proportional to its
solubility. The approximation that replaces the oil phase concentration and mole fraction
with the density and average molecular weight requires that the gasoline behaves as an
ideal mixture. Cline et al. (1991) showed from the partitioning of 31 gasoline samples
that partitioning generally followed Raoult's law; specifically, their calculated partition
coefficients varied by less than 30% from the experimentally determined partition
coefficients.
Although a value of about 48,000 mg/L is often cited for the solubility of MTBE,
literature data on solubility contain inconsistencies. Data from Peters et al. (2002),
Fischer et al. (2004), and Lide (2000) are roughly consistent with each other (Figure B-
1). These indicate solubility of MTBE decreases as temperature increases. However,
there can also be variability in reported solubility at any given temperature. Some
indication of the possible variability in reported MTBE concentrations are given on the
figure. Montgomery (2000) cites literature which reports 2,2,4-trimethylpentane (i.e.,
iso-octane) solubility ranging from 0.56 mg/L to 2.46 mg/L at 25°C. Compounding this
single-temperature variability is the variability in average temperature of shallow ground
water in the U.S., which ranges from 5°C to 25°C. Thus the reported data give a general
indication of solubility, rather than precisely established values.
Table B-l shows the effective solubility for benzene, ether oxygenates and iso-octane
based on various assumptions of the mole fraction of each compound in gasoline. The
effective solubilities are ranked in the table from highest effective solubility to lowest.
130
-------
Although MTBE has virtually been eliminated as an option for blenders, historical
considerations indicate that winter oxygenated gasoline with MTBE had the highest
effective solubility and iso-octane, with the lowest measured mass fraction in the study,
by the U.S.EPA (2005) had the lowest. The maximum MTBE concentrations are
consistent with field data presented by Tong and Rong (2002) who reported a maximum
MTBE concentration of 1560 mg/L from their study of 90 sites in the Los Angeles area.
The prospects for wide spread ground water contamination generated from iso-
octane in gasoline at these levels (up to 19%) is very low because of these very low
effective solubility. Alkylation produces high-octane gasoline components called
"alkylates," by reacting isobutane with olefins (alkenes). One example of an alkylate is
iso-octane which is also called 2,2,4-trimethylpentane (Owen and Coley, 1995. Similar
behavior is expected with other components of alkylate.
Figure B-l. MTBE solubility data
from Peters et al. (2002), Fischer et al. (2004) and Lide (2000).
B
Tamp cratim (0)
Heermann and Powers (1998) measured ethanol/water partition coefficients
between a one-, two-, three- and eight-component artificial gasoline and water mixture.
Their results show similar partitioning behavior between the three- and eight-component
mixtures implying that these results better represent an actual gasoline, than the few
component mixtures. The measured partition coefficients decreased with increasing
ethanol concentration in the water phase. Their results for these mixtures can be
represented by
131
-------
rw
Jethano
J-o rw
J ethanol J ethanol
where the concentrations are expressed on a volume fraction basis, and the fraction of
ethanol in water, fwethanoi, is less than 0.2. Thus ethanol is partitioned preferentially into
the aqueous phase and the fraction in the water phase is roughly 150 times the fraction in
the gasoline phase.
Table B-l. Solubility estimated from the SPARC calculator20 and effective
solubility for benzene, ether oxygenates and isooctane. The results are ranked from
highest effective solubility to lowest, illustrating the combined impact of solubility
and the content of each chemical in gasoline.
Chemical
MTBE
MTBE
TAME
DIPE
TAME
ETBE
MTBE
DIPE
ETBE
Benzene
MTBE
MTBE
Benzene
Benzene
TAME
Benzene
TAME
Iso-octane
Iso-octane
Iso-octane
Condition
Winter Oxygenate
(wo)
RFC
wo
wo
RFC
wo
Avg CG 93 octane
RFC
RFC
Max
Avg CG 87 octane
Min
Avg CG 87 octane
RFC
RFC 93 octane
Avg CG 93 octane
RFC 87octane
Max
Avg
Min
Solubility3
(mg/L)
13700
13700
3860
3030
3860
2350
13700
3030
2350
1600
13700
13700
1600
1600
3860
1600
3860
0.3
0.3
0.3
Molecular
Weight
(g/mol)
88.15
88.15
102.18
102.18
102.18
102.18
88.15
102.18
102.18
78.11
88.15
88.15
78.11
78.11
102.18
78.11
102.18
114.23
114.23
114.23
Volume
fraction
0.1490
0.1100
0.1720
0.1720
0.1275
0.1720
0.0333b
0.1275
0.1275
0.0500
0.0052C
0.0050
0.0145d
0.0100
0.0034e
0.0082'
0.0024s
0.196911
0.05621
0.00191
Approximate
Mole fraction (k)
0.1277
0.0943
0.1778
0.1681
0.1318
0.1709
0.0285
0.1246
0.1267
0.03446
0.0045
0.0043
0.0130
0.0089
0.0035
0.0073
0.0025
0.2039
0.0582
0.0020
Effective Solubility
(mg/L)
1750.
1292.
686.5
509.3
508.9
401.7
391.1
377.6
297.8
71.4
61.1
58.7
20.8
14.3
13.4
9.48
8.92
0.062
0.018
0.0006
a For background see Hilal et al., 2004
b Average MTBE content in conventional Georgia 93 Octane Gasoline (U.S. EPA, 2005)
0 Average MTBE content in conventional Georgia 87 Octane Gasoline (U.S. EPA, 2005)
d Average benzene content in conventional Georgia 87 Octane Gasoline (U.S. EPA, 2005)
e Average TAME content in reformulated Virginia 93 Octane Gasoline (U.S. EPA, 2005)
f Average benzene content in conventional Georgia 93 Octane Gasoline(U.S. EPA, 2005)
8 Average TAME content in reformulated Virginia 87 Octane Gasoline(U.S. EPA, 2005)
hMaximum Isooctane content for reformulated and conventional gasoline (U.S. EPA, 2005)
'Maximum Iso-octane content for reformulated and conventional gasoline (U.S. EPA, 2005)
1 Maximum Iso-octane content for reformulated and conventional gasoline (U.S. EPA, 2005)
k Assumed 105 g/mol average molecular weight of gasoline.
1. Co-solvency and Partitioning
20Estimates made from the SPARC Performs Automated Reasoning in Chemistry calculator, see
http://ibmlc2.chem.uga.edu/sparc/
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The presence of a water-miscible component of gasoline, such as ethanol or
methanol, has the potential to increase the solubility of other, less soluble compounds in
gasoline like benzene. For methanol-gasoline blends with MTBE contents up to 15%,
Poulson et al. (1992) reported no increases in benzene, toluene, ethylbenzene, xylenes
(collectively, BTEX) concentrations. They also found that when methanol was present in
gasoline at 10% or less, there was no appreciable effect on BTEX solubility, if the
gasoline to water ratio was 1 to 10.
At higher methanol content, benzene solubility increased dramatically. For 85%
methanol, two effects are significant. First, there is less benzene in the gasoline (as
would be true for E85 also) but the co-solvency effect was found to produce higher initial
benzene concentrations in water when the gasoline to water ratio was higher (1 to 1
versus 10 to 1). As the source is depleted of benzene, then the concentrations dropped to
values slightly below those of gasoline containing no methanol, due to the lower benzene
content in the methanol-containing gasoline.
Heermann and Powers (1998) developed a model to estimate the effect of ethanol
on the gasoline/water partition coefficient of BTEX. The model was comprised of a
linear relationship for low ethanol volume fractions and a log-linear model for higher
concentrations. For ethanol volume contents from 0% to 25%, variation in measured
BTEX partition coefficients was best represented by a linear model (where the partition
coefficients varied by as much as 49%). At around 25% ethanol volume content, a log-
linear model represents the data better. Above this point the partition coefficients varied
by orders-of-magnitude, indicating a strong co-solvency effect, similar to that observed
by Poulson et al. (1992) for methanol.
B. Mass-Transfer Limitations
Seagren et al. (1999) report on a study where the flow of water past a pool of non-
aqueous phase liquid (NAPL) was evaluated for the establishment of equilibrium
between concentrations in the aqueous and non-aqueous phases. Vertical dispersion was
the main mechanism that was presumed to move a contaminant from the NAPL pool to
the aquifer. These authors concluded that for most of the conditions studied, the local
equilibrium model performed as well as or better than a non-equilibrium partitioning
model. The effects were quantified through usage of the modified Sherwood number
multiplied by the Stanton number21. Local equilibrium is appropriate when the product
of these two quantities is greater than 400. They note, however, that there are
uncertainties in model parameters, particularly, ki and Dz for field sites. Conditions
where local equilibrium may not be achieved include sites with high seepage velocity and
dispersion coefficient, and/or low pool length and small mass transfer coefficient. Arey
and Gschwend (2005) presented an analysis of the effects of varying gasoline
composition on drinking water well impacts. From their analysis they conclude that
partitioning from a NAPL pool into flowing ground water achieves equilibrium for
2 IThe modified Sherwood number is represented by Sfr = Lx k^fDz and the Stanton number by St = kj/Vx,
where Lx is the length of the pool, kj is the mass transfer coefficient, Dz is the vertical diffusion coefficient
and vx is the seepage velocity.
133
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typical conditions (defined in this study as pool length of 10 m, ground water velocity of
0.4 m/d, vertical dispersivity of 0.002 m, and mass transfer coefficient of 1 m/d).
C. Geologic Heterogeneity
Although the maximum observed MTBE concentration reported by Tong and
Rong (2002) in their study of 90 Los Angeles LUST sites corresponds closely to the
predicted effective solubility of MTBE, the majority of their sites have lower maximum
concentrations. These differences could be due to well placement and well screen effects
(U.S. EPA, 2004), lower maximum MTBE concentrations in Los Angeles gasoline
depending on the release date, or reduced flushing of the gasoline due to geologic
heterogeneity. Wilson et al. (2001) presents field results from a North Carolina site with
detailed vertical profiling of the total petroleum hydrocarbons (TPH or fuel) distribution
along with the distribution of hydraulic conductivity (Figure B-2). The significance of
these results is that the TPH is co-located with the lowest hydraulic conductivity values.
Ground water flow through this zone will be very low as well. Transport of MTBE or
other contaminants away from this zone will be limited by the diffusion rate from the low
conductivity zone to the highly conductive aquifer below. In this case, limited diffusion
and high velocities below the contaminated zone may result in mass transfer limited
partitioning from the TPH to the underlying ground water as indicated by the Seagren et
al. (1999) local equilibrium analysis described above.
Since the fining upward geologic sequence is typical of alluvial depositional
systems, there may be numerous LUST sites where these phenomena limit observed
concentration. Although detailed characterization, as was done by Wilson et al. (2001) of
the Tong and Rong (2002) sites is not available, the distribution of TPH within
impermeable zones and mass transfer limitations may explain their results.
134
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Figure B-2. Distribution of total petroleum hydrocarbons (TPH or fuel)
and hydraulic conductivity at a North Carolina site.
Total Petroleum Hydrocarbon mg.kai
0 10000 ""(lOC'i'i M) 000 40000
li
Table
TPH-C
Depth jj .
(feet)
25 -
H v d r a u 11 c C o n d u c 11 v it v
Depth
I
S
i 0.01 0.02 0,03 0,04
Hydraulic Conductivity (cm/sec)
135
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References
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anticipating widespread contamination of community water supply wells by
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U.S. Environmental Protection Agency (2004) On-line tools for assessing petroleum
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