The use of oxygenated gasoline was mandated under the Clean Air Act Amendments of 1990 in areas that did not meet the federal ambient air standard for carbon monoxide (CO), an air pollutant associated with potential health risks. CO interferes with the body's ability to utilize oxygen by combining with hemoglobin, which then prevents it from transporting oxygen efficiently to organs of the body. Persons with chronic heart disease are at particular risk for adverse health effects of CO, but other groups, including the elderly, pregnant women, infants, and persons with anemia or cardiopulmonary disease, are also likely to be at increased risk due to CO exposure. Motor vehicle emissions are the primary source of ambient CO levels in most areas. The Clean Air Act requires at least a 2.7% oxygen content for gasoline sold in CO nonattainment areas. Gasoline containing 2.7% oxygen (by weight) is typically achieved by the addition of 15% methyl tertiary butyl ether (MTBE) or 7.5% ethanol (by volume). The higher oxygen content of oxygenated gasoline compared to conventional gasoline is intended to lead to a more complete combustion of the gasoline and therefore to reduced tailpipe emissions of CO. In addition, oxygenates usually displace aromatics in gasoline, such as benzene, toluene and xylene, which are a source of octane. These aromatics have been targeted for reduction in the reformulated gasoline program.
Soon after the oxygenated gasoline program was introduced nationally in the winter of 1992-1993, anecdotal reports of acute health symptoms were received by health authorities in various areas of the country. Such health concerns had not been anticipated but have subsequently focused attention on the possible health risks associated with using oxygenated gasoline. The purpose of this interagency assessment of the potential health risks of oxygenated gasoline is to explore the question of whether evidence from recent health studies of oxygenated gasoline or oxygenates, especially MTBE, and acute illness warrants a reconsideration by EPA of potential health risks of the oxygenated gasoline program during the coming winter months. This assessment is being prepared for the Office of Science and Technology Policy, under the auspices of the Interagency Oxygenated Fuels Assessment Steering Committee. The potential health benefits of the oxygenated program are not addressed in this assessment; such benefits will be considered in the more comprehensive interagency assessment of oxygenated fuels that is currently under way.
A team of scientists from three Federal agencies, the Centers for Disease Control and Prevention (CDC), the National Institute for Environmental Health Sciences (NIEHS), and the Environmental Protection Agency (EPA) was assembled to complete this evaluation. Members of this team were chosen by members of the Interagency Steering Committee or their designates on the basis of the members' expertise in different scientific disciplines and knowledge about issues related to oxygenated gasoline.
This assessment focuses on exposures that occur through inhalation; ingestion of oxygenates from contaminated water supplies may also be important but was considered beyond the scope of this assessment. The data base on ethanol is extensive, but it pertains primarily to ingestion, not inhalation. A full consideration of the relevance of the literature regarding the health effects of ethanol by ingestion to exposure by inhalation from evaporative and combustion mixtures of ethanol in gasoline also was considered beyond the scope of this assessment.
This assessment also attempts to identify areas where the scientific data base is particularly limited. It should also be noted that most of the information available on the health effects of MTBE pertains to MTBE alone rather than the mixture of MTBE and gasoline. This assessment does not attempt to assess the risks or benefits of MTBE-oxygenated gasoline in relation to conventional gasoline.
In addition, a more comprehensive review of the fundamental basis and efficacy of the Environmental Protection Agency's winter oxygenated gasoline program is currently being conducted by a combination of technical and scientific experts from across several Federal agencies, under the coordination of the National Science and Technology Council's Committee on Environment and Natural Resources. This more comprehensive review will consider not only health effects, but also air quality benefits, fuel economy and engine performance, ground water and drinking water quality, and an economic analysis of benefits. In addition, an independent review of the health effects of the use of oxygenates in gasoline is currently being conducted by the Health Effects Institute (HEI) and a panel of experts, in response to a request from the EPA. It is expected that the HEI review will be the core of the health effects section of the more comprehensive interagency review.
Various studies (e.g., Stump et al., 1989, 1990, 1992, 1994; Auto/Oil Air Quality Improvement Research Program 1990, 1991a,b,c; Carter et al., 1991; Kiskis et al., 1989; Reuter et al., 1992) have examined the impact of oxygenates on evaporative and combustion emissions from vehicles. For example, one study (Reuter et al., 1992) of 20 vehicles (1989 models with three-way catalysts) showed that, at 24°C, there was a trend for MTBE to cause a net reduction in the total mass of toxic combustion emissions. In general, emissions of benzene tended to decrease and emissions of formaldehyde tended to increase; 1,3-butadiene was not substantially altered when MTBE was added. However, more testing is needed with different types of vehicles under a variety of conditions (especially lower and higher temperatures) to characterize fully the overall changes in evaporative and combustion emissions associated with oxygenate usage.
Although some models for relating emissions data to ambient air quality are available, these changes in emissions cannot be quantitatively extrapolated to estimate with acceptable precision their impact on the air quality of a city. Moreover, ambient air measurements may provide only a rough (and sometimes misleading) guide to personal exposure levels that a population might encounter during hour-to-hour and day-to-day activities.
Little is known about the atmospheric transport and transformation of the complex mixtures associated with the use of oxygenated fuels or the possible significance of human exposure to these fuels. For example, tertiary butyl formate is a photochemical-transformation product of MTBE, as is formaldehyde (Tuazon et al., 1991). The possibility of increased human exposures to such substances as a result of using oxygenated fuel has not yet been evaluated. Ambient and microenvironmental air monitoring is needed to determine the levels of these and other possibly significant by-products of oxygenated fuels.
Zweidinger (1993) analyzed air samples collected over 8-hour periods in Fairbanks, Alaska; Stamford, Connecticut; and Albany, New York for aldehydes, MTBE, and other volatile organic compounds (VOCs) in conjunction with investigations by CDC (1993a,b,c). The Fairbanks samples were collected by the State of Alaska during three time periods: time period 1 (December 1 through 12, 1992) occurred immediately before the phase out of 15% MTBE in gasoline, and 25 VOC and 35 aldehyde samples were obtained; time period 2 occurred during the phase-out itself (December 18 through 22, 1992), and 31 VOC and 26 aldehyde samples were obtained; and time period 3 occurred after the phase-out of MTBE-oxygenated fuels (February 2 through March 5, 1993) during which 73 samples each of VOCs and aldehydes were obtained. On the basis of the analysis of gasoline samples collected in Fairbanks during time periods 2 and 3, the percentage of MTBE in gasoline decreased from 8.5% to 1% for unleaded regular gasoline and from 14.7% to 5.6% for premium gasoline. The Stamford samples were collected by EPA Region 1 from April 13 through 14, 1993; 30 samples each of VOCs and aldehydes were collected. These samples were obtained in Stamford because the city represented another part of the country where MTBE-oxygenated gasoline was also sold. The Albany samples were collected from May 5 through May 27, 1993, by the New York State Department of Health; 20 samples each of VOCs and aldehydes were collected. Albany represented an area of the country where MTBE was not used as an oxygenate but was present only as an octane enhancer in gasoline. The Fairbanks time-period 1 VOC samples were analyzed by the Oregon Graduate Center, and the Desert Research Institute analyzed the aldehyde samples. All other ambient samples were analyzed by EPA's Atmospheric Research and Exposure Assessment Laboratory.
The samples from each city consisted of samples taken at roadside intersections, the pump islands of gas stations, garage service bays, and in residential neighborhoods. Samples of indoor air and from background sites were also collected. Samples of indoor air and of air from service bay or background sites were not collected in Albany, and no air samples from service bays were collected in Fairbanks during time period 2. In addition, air samples from the interiors of commercial vehicles in Fairbanks were collected early during time periods 1 and 3. Significant differences in ambient temperature and other meteorological conditions existed among the cities where samples were collected. Relatively few samples were collected in a given area, and the samples were collected over a few days only. Therefore, the data cannot be used to describe the air quality quantitatively in any of these cities. Rather, the data can be used to estimate approximate ranges of air quality in the locations sampled.
The highest average concentrations of MTBE (0.345 parts per million [ppm], or 1.28 mg/m³), benzene (0.629 mg/m³), total nonmethane organic carbon (80.5 ppm C), and formaldehyde (0.038 mg/m³) were found in garage service bays. One of the highest average concentrations for a single compound was found to be 1,1,1-trichloroethane (methyl chloroform), which exceeded 38.0 mg/m³ in service bays (Fairbanks, time period 3). Aside from the service bays, gas stations showed next highest concentrations of MTBE (Fairbanks: 0.05 ppm, or 0.194 mg/m³, time period 1; 0.037 ppm or 0.134 mg/m³, time period 2; and 0.006 ppm, or 0.020 mg/m³, time period 3). MTBE concentrations in Stamford gas stations were the lowest (0.004 ppm or 0.013 mg/m³) but were likely the result of sampler location (the Albany average was 0.024 ppm, or 0.086 mg/m³). Whereas the air samplers in Fairbanks and Albany were located on the pump islands, the samplers in Stamford were located at least 15 feet from the islands. In Fairbanks, indoor and outdoor MTBE concentrations were similar and averaged about 0.007 ppm (0.025 mg/m³) for the samples from time period 2, falling to 0.001 ppm (0.0037 mg/m³) in time period 3, with the exception of one home where the average indoor value was 0.02 ppm (0.072 mg/m³). This home had an attached garage, and air samples taken there showed elevated levels of benzene (0.138 mg/m³) and other compounds associated with gasoline. Indoor MTBE concentrations in Stamford averaged 0.0006 ppm (0.002 mg/m³). MTBE concentrations measured inside vehicles in Fairbanks averaged 0.007 ppm (0.024 mg/m³) during time period 1 (not including one sample of 0.067 ppm or 0.241 mg/m³) and averaged 0.005 ppm (0.019 mg/m³) during time period 3 (not including one sample of 0.035 ppm or 0.127 mg/m³).
Formaldehyde concentrations were higher indoors (0.012 to 0.034 mg/m³) than outdoors (0.0025 to 0.025 mg/m³), which is generally the case, and levels appeared typical of those seen in indoor air studies. Benzene levels were higher in Fairbanks (average roadside levels were 0.026 mg/m³ for December and 0.042 mg/m³ during time period 3) than in the other cities (Stamford, 0.003 mg/m³ Albany, 0.0014 mg/m³). A related report by Gordian et al. (1995) noted that the gasoline used in Alaska has the highest concentration of benzene (~ 5%) of any state in the nation; the report also stated that, during this same period in Fairbanks, formaldehyde and benzene levels in indoor and ambient air samples were higher after MTBE was removed from gasoline. However, these findings do not establish a causal relationship between decreased MTBE levels in gasoline and increased formaldehyde and benzene levels in the air.
In response to complaints by residents of Milwaukee, Wisconsin, regarding reformulated gasoline (2.0% oxygen content), the State of Wisconsin (Anderson et al., 1995) recently measured air concentrations of MTBE and other compounds (ETBE, benzene, ethyl benzene, xylenes, toluene). Levels of MTBE were less than 0.001 ppm at most of the sites sampled, including at the North Campus of the University of Wisconsin-Milwaukee where 24-hour and 2-hour samples were obtained at various intersections and gasoline stations. The highest ambient air concentrations of MTBE were obtained at two gasoline stations and a civic center parking ramp (0.00243, 0.00458, and 0.00205 ppm, respectively). The highest value was found at a station without a stage II vapor-recovery system.
The Environmental and Occupational Health Sciences Institute (EOHSI) and the Research Triangle Institute (RTI) completed a study of field measurements of MTBE concentrations inside automobiles during an approximately 1-hour commute and during refueling (Lioy et al., 1993, 1994). Field measurements were collected during April 1993 in Middlesex County, New Jersey, at two stations with full service and stage II vapor-recovery systems; in Westchester County, New York, at three stations with self service and stage II vapor-recovery systems; and in Fairfield County, Connecticut, at five stations with self service and no stage II vapor-recovery system. One new-model automobile (a 1992 Corsica) and one older model automobile, either a 1985 Caprice or a 1986 Monte Carlo, were assigned to each commuter route. The samples were collected from the interior of the front passenger side of the automobile. The number of samples per automobile ranged from 14 to 20 for the commute and from 3 to 5 for the fill-up.
The driver's window was open during the fill-up. The time to complete the fill-up was about 2 minutes, and the total time at the gas station was 5 to 10 minutes. In addition to the measurements taken inside the automobile, a few measurements were collected near the breathing zone of the person refueling the gas tank.
Concentrations of MTBE measured in the cabin interior during the 1-hour commute had a geometric mean of 0.006 ppm (0.021 mg/m³), with a range of 0.001 ppm (0.004 mg/m³) to 0.16 ppm (0.58 mg/m³). The commuter runs in Connecticut had higher geometric mean concentrations (0.007 ppm or 0.023 mg/m³) than those in New Jersey (0.005 ppm or 0.016 mg/m³). One vehicle, the 1987 Caprice, had somewhat higher levels than the other vehicles; inside the older-model automobiles, concentrations were higher, probably reflecting differences in the design and deterioration of the older-model vehicles.
Concentrations during a 5-minute refueling at a self-service station averaged in excess of 0.3 ppm (1 mg/m³) and ranged from about 0.01 ppm to 4.1 ppm (0.036 to 14.7 mg/m³), with the highest values measured in the breathing zone of a person refueling at a station with no stage II vapor-recovery system. Measurements taken inside the cars during refueling generally ranged from 0.01 to 0.1 ppm (0.036 to 0.36 mg/m³) for stations both with and without stage II vapor-recovery systems.
International Technologies Inc. completed a set of field measurements of MTBE concentrations in the personal breathing zone (PBZ) during fill-up, at the pump island, and around the property line of gas stations (Johnson, 1993). This study was done in conjunction with the above EOHSI/RTI study that was conducted at the same 10 gas stations. All concentrations measured for this study, even those for intermittent exposures in the personal breathing zone, were from a 4-hour continuous sample. Average fence-line concentrations, which are typically taken at the apparent property line, ranged from 0.005 to 0.065 ppm (0.018 to 0.234 mg/m³) MTBE (Johnson, 1993). The highest fence-line concentrations ranged from 0.1 to 0.139 ppm (0.36 to 0.5 mg/m³) MTBE. The highest breathing-zone and pump-island concentrations ranged from 0.194 to 2.5 ppm (0.7 to 9 mg/m³) MTBE. Further discussion of this study and other exposure measurements among service station attendants is included in the next section on occupational exposures.
As should be expected, these 4-hour breathing-zone concentrations described above were lower than those reported by the Clayton Environmental Consultant study (Clayton Environmental Consultants, 1991), which collected samples only during the fill-up period (approximately 2 minutes). In the Clayton study, mean MTBE concentrations in the breathing zone for oxygenated fuels containing 12% to 13% MTBE were 3.6 ppm (13 mg/m³), with vapor recovery, and 8.3 ppm (30 mg/m³), without vapor recovery. The absolute range among these MTBE concentrations was 0.089 to 38 ppm (0.32 to 137 mg/m³). Although several stations were monitored, the highest and lowest measurements were made at one station, illustrating the variability of breathing zone exposures. Indeed, a wide range of air concentrations within the breathing zone can be expected. Ambient air concentrations measured at a gas station will be highly dependent on wind speed and direction. In addition, breathing-zone concentrations can be dramatically influenced by one's position in relation to wind direction. Any spilling of fuel while the tank is being filled also can dramatically increase the inhaled concentration.
Personal breathing zone (PBZ) measurements obtained in conjunction with the Wisconsin study of reformulated gasoline (2.0% oxygen content) ranged from 0.070 to 2.93 ppm for MTBE, 0.050 to 0.100 ppm for ETBE, and 0.050 to 0.150 ppm for benzene (Anderson et al., 1995). These samples were collected approximately 1 meter from the fuel nozzle over 15-minute periods. The highest air concentrations of MTBE were measured at a station lacking a Phase II vapor recovery system. In some cases, the concentrations were higher for MTBE at a station using ETBE-oxygenated fuel and were likely due to vapors from the previous fill-up escaping during the refueling. Thus, it may be difficult to ascertain the particular oxygenate to which people are exposed under real-world conditions.
The data on air quality and microenvironments (e.g., at gas pumps, inside cars, in personal garages) are too limited to provide a distribution of MTBE levels for the general population. These data, however, may be used to roughly estimate reasonable worst-case potential exposures, based on certain assumed activity patterns and approximate microenvironmental concentrations. Because of the interest in MTBE, the present evaluation focuses on this compound, even though any potential health effects might result from complex pollutant mixtures of which MTBE is only one component.
Acute as well as long-term exposures are of importance in evaluating potential health risks of oxyfuels. Table 1 illustrates two hypothetical types of people with varying activity patterns and oxyfuel exposure potentials. The concentrations and activity patterns were previously used by Huber (1993) for estimating MTBE exposures and were based on a variety of population-activity studies and microenvironmental measurements of MTBE as described above. Exposure (ppm.hr) is a function of concentration (ppm) and time (hr). The concentrations in Table 1 are based on measurement averages that were rounded up to the next order of magnitude so as to provide inherent conservatism (i.e., overestimation) in the exposure estimates. The two estimates in Table 1 reflect different assumptions about the microenvironmental concentrations and activity patterns represented. For example, the concentrations of MTBE during refueling in Scenarios I and II differ by an order of magnitude. Scenarios I and II each represent hypothetical persons who visit a gasoline station 1.5 times per week, commute an average of 10 hour/week, visit an auto repair shop four times per year for 15 minutes per visit, and spend about 57 hours per week in an office or public building. However, scenario I assumes that the person lives in a house without an attached garage and does not reside near or spend time around a highway or in the vicinity of a filling station.
Hypothetical person II differs from person I in that the individual lives in a house with an attached garage and spends time outside in the vicinity of a gasoline station or heavily used highway. The higher concentrations of MTBE in the residential garage and home assume evaporative emissions from the vehicle in the garage or a small gasoline spill with the garage door closed. Hypothetical person II represents a "reasonable worst case" scenario (Huber, 1993). More extreme cases are conceivable, but they would probably reflect differences in the occurrence or in the duration of activities rather than reflecting higher average concentrations of MTBE in the microenvironments considered. For example, a person who visits a gasoline station every day to refuel his or her car and who works 40 hours per week outside near a highway or other source of high MTBE levels would not likely be exposed to higher concentrations of MTBE within those microenvironments, but that person's overall exposure might be higher than that of hypothetical person II because of the greater cumulative duration of his or her exposure. Currently it is not possible to state what percentages of the general population might be represented by hypothetical persons I and II, but the term "reasonable worst case" is meant to imply that few individuals would be exposed at higher levels. It is not meant to imply a "high average" exposure. The average exposure for the entire exposed population would be expected to be somewhat lower than hypothetical person I and much lower than hypothetical person II.
Time-weighted average exposure levels may be estimated for the oxyfuel season itself or for an entire year, assuming that the oxyfuel season is limited to a few months in an area that is not required to use reformulated gasoline. During the oxyfuel season, the time-weighted average MTBE exposure of a hypothetical person would be one-half the sum of that person's exposure levels divided by one-half the number of hours in a year e.g. (159/2 ÷ 8760/2 = 0.018 ppm for person I, or 309/2 ÷ 8760/2 = 0.0345 ppm for person II). To calculate an annual time-weighted average exposure, an allowance is made for the assumed use of 1.5% MTBE in nonoxygenated gasoline. This assumption may be a high estimate of the average amount of MTBE in nonoxygenated gasoline because of the relatively limited use of premium (high-octane) fuels containing MTBE. Nevertheless, assuming 1.5% MTBE (which is 1/10 of the 15% MTBE concentration typical of oxygenated gasoline) during the non-oxyfuel season and assuming a 6-month oxyfuel season, the annual time-weighted average MTBE exposure level for hypothetical person I would be [(159/8760)*(6/12)] + (0.1*[(159/8760)*(6/12)]) = 0.010 ppm. The annual time-weighted average MTBE exposure level of hypothetical person II would be 0.019 ppm. For a 4-month oxyfuel season, the annual time-weighted average levels would be 0.007 ppm for person I and 0.014 ppm for person II. If reformulated gasoline containing 10% MTBE by volume were used during the remainder of the year, the annual time-weighted average exposure for a 6-month oxyfuel season would be 0.015 ppm for person I and 0.029 ppm for person II.
It can be assumed that a gasoline fill-up scenario, although brief, would result in the highest acute-exposure concentrations. The highest human exposure is expected when one is near evaporative emissions. Thus, exposure would be greatest when handling gasoline. The highest reported MTBE concentration measured at a filling station was 38 ppm (137 mg/m³), although levels as low as 0.089 ppm (0.32 mg/m³) were also measured at the same station, illustrating the variability in fill-up exposures (Clayton Environmental Consultants, 1991). A more typical worst-case scenario for the MTBE concentration in the breathing zone during fill-up would be 10 ppm (36 mg/m³) MTBE for a few minutes (Johnson, 1993; Lioy et al., 1993, 1994; Clayton Environmental Consultants, 1991). However, higher concentrations are possible, especially in the event of an accidental spill.
For purposes of comparison with the 1-hour human experimental exposure studies (at 1.4 and 1.7 ppm [5 and 6 mg/m³]), discussed below, 1-hour time-weighted average concentrations of MTBE were calculated for two exposure scenarios by using high concentration data. The first scenario assumed highest measured values and involved a 2-minute fill-up (38 ppm or 137 mg/m³), a 30-minute commute associated with a fill-up (0.5 ppm or 1.8 mg/m³), and a 28-minute commute (0.076 ppm or 0.275 mg/m³); the average MTBE concentration was 1.56 ppm (5.6 mg/m³) MTBE. The second scenario used the MTBE levels in Table 1. Scenario 2 assumed a 2-minute fill-up (10 ppm or 36 mg/m³), 2 minutes in a personal garage (1 ppm or 3.6 mg/m³), a 30-minute commute (0.1 ppm or 0.36 mg/m³), 10 minutes in a public garage (0.5 ppm or 1.8 mg/m³), and 16 minutes in a public building (0.01 ppm or 0.036 mg/m³); the average concentration was 0.5 ppm (1.8 mg/m³) MTBE.
In summary, selected hypothetical scenarios suggest that the annual time-weighted average daily personal exposure level for a "reasonable worst case" motorist in the general (nonoccupational) population might be on the order of 0.019 ppm MTBE. This level is not simply a "high average" exposure. The annual time-weighted average exposure levels for most people would be somewhat lower, perhaps closer to the estimate of 0.010 ppm MTBE for hypothetical person I, or even an order of magnitude lower than 0.010 ppm, in view of the rounding up of the data on which the estimates were based. In addition, long-term exposure levels are likely to be lower than those found during short-term measurements. That is, hypothetical person II is unlikely to have consistently high exposure levels from day to day, and the trend over a year or a lifetime would be for exposure levels to regress toward the mean. Thus, in estimating health risks in relation to these exposure estimates, it is important to consider the likely period of exposure over which health effects could be induced. For example, a 4 or 6-month oxyfuel season might in itself pose little risk with respect to lifetime cancer risk but could be of significance with respect to potential effects due to acute exposures. Two illustrative acute-exposure estimates for different scenarios yield averages of 0.5 ppm and 1.56 ppm, reflecting in part the highly variable air concentrations that have been measured at gasoline stations. It must be emphasized that these exposure values are not based on, and make no predictions regarding, the nature of population-exposure distributions. Personal exposure data for probabilistic samples are needed before an exposure assessment for MTBE can provide an adequate characterization of population exposures for health risk assessment purposes.
|Scenario I||Scenario II|
(ppm x hr)
(ppm x hr)
|Gas station||1.5/week@10 min||1||13||1||13||13||In vehicle||10 hr/week||0.1||52||0.1||52|
|Auto repair shop||4/year@15 min||1||0.5||1||0.5|
|Public garage||10 min/day||0.5||30||0.5||30|
|Residential garage||2 min/day||12||0.06||12||12|
|Residence||10 hr/day + weekend||4,147||0.01||41.5|
|Office/public buildings||57 hr/week||2,964||0.01||29.6|
|Time-weighted average exposure during an Oxyfuel Season||0.018||0.035|
|Annual time-weighted average exposure, assuming a 6-month oxyfuel season||0.010||0.019|
|Annual time-weighted average exposure, assuming a 4-month oxyfuel season||0.007|
a Concentrations are based on measurement averages that were rounded up to the next order of magnitude.
Since 1990, occupational exposure to MTBE has been assessed under a variety of conditions and for various purposes, most notably in support of evaluations of health complaints associated with exposures. The realm of occupational exposures to MTBE involving gasoline includes exposures to the pure compound during manufacture and transport (for subsequent blending) and exposure to gasoline mixtures during blending, transport, distribution, and sale. Other tangential occupational exposures have been evaluated in jobs where at least a portion of a work shift is spent near an MTBE-blended gasoline source; the most common source is the automobile.
Undoubtedly, the largest occupational group with a significant potential exposure to MTBE is service station attendants. The number of retail automotive service stations has been estimated at 150,000 to 210,000 (ENVIRON, 1990). Even with self service, fuel dispensing undoubtedly continues to be an important source of occupational exposure to MTBE, although no data were available on the number of attendants exposed to reformulated or oxygenated fuels containing MTBE.
In 1990, at the request of the American Petroleum Institute (API), the National Institute for Occupational Safety and Health (NIOSH) evaluated exposure to the gasoline components MTBE, benzene, toluene, and xylene at six retail automotive service stations (NIOSH, 1993a). To reflect MTBE's multiple uses and the range of potential exposures, two facilities were selected to represent its ubiquitous use as an octane enhancer (generally blended at less than 1% of the fuel); two facilities were selected to represent requirements to use MTBE in oxygenated fuel (blended at 12-15 liquid volume percent [LV%] of the fuel), and two with stage II-type vapor-recovery systems were selected to determine the relative effectiveness of these engineering controls.
Of 16 PBZ samples collected from attendants working at stations that sold gasoline with less than 1% MTBE, 15 were below the lowest detectable concentration (LDC) of 0.02 ppm. The one sample above the LDC was reported at 0.16 ppm. At stations using at least 12% MTBE blends, 41 PBZ samples ranged from 0.03 to 3.89 ppm, averaging 0.54. At stations equipped with stage II vapor recovery systems, 15 of the 48 PBZ samples had detectable levels (above the LDC), ranging from 0.02 to 0.73 ppm, averaging 0.18 ppm. Sampling and analysis for MTBE was by the NIOSH Method 1615 (Palassis, 1993). Benzene exposures among this same population averaged approximately 0.07 ppm.
The study included an assessment of factors, such as climatic conditions and work practices, affecting the extent of exposure. Statistical analysis, using step-wise linear regression, indicated that MTBE exposures were most affected by wind velocity, followed by the amount of fuel dispensed by the attendant. The report also noted that benzene exposures were not significantly affected by the amount of MTBE in the fuel.
In 1994, NIOSH conducted a follow-up study to assess short-term, or "peak," exposures among attendants (Cook, 1995). Two service stations in New Jersey were selected, and direct-reading real-time measurements for total hydrocarbons, as a surrogate for peak MTBE exposures, were made in attendants' breathing zones, as were long-term, full-shift determinations of MTBE exposures. A video-monitoring overlay technique was used to determine the predominant sources of peak exposures. Twenty-one of the full-shift samples taken from the PBZs of the station attendants who were evaluated for MTBE exposure ranged from 0.08 to 1.27 ppm, with a geometric mean (GM) of 0.38 ppm. Total hydrocarbon exposures averaged 1.89 ppm, with peaks as high as 327 ppm. NIOSH is currently assessing the magnitude of associated MTBE peak exposures.
Service station attendants' exposures were again assessed by NIOSH (NIOSH, 1993b) as part of a larger effort in support of a CDC epidemiologic investigation of public health complaints reportedly associated with exposure to MTBE in Fairbanks, Alaska. Two samples were obtained from attendants at service stations using MTBE as an octane enhancer (approximately 1% MTBE). The use of MTBE as an oxygenated-fuel had previously been discontinued after reports surfaced of ill health effects associated with its use. MTBE concentrations were reported at 0.03 ppm (the LDC) and 0.15 ppm; both results were within the range of exposure concentrations determined from the previous NIOSH study.
In 1993, in support of an EPA solicitation for MTBE exposure data, the API contracted for a service station study with ITAir Quality Services (API, 1994) at 10 service stations in the northeastern United States. In addition to measuring MTBE, ITAQS measured benzene, toluene, ethyl benzene, xylene, formaldehyde, and carbon monoxide in attendant and customer breathing zones, near the pump islands, and at perimeter locations. For comparison of sampling or analytical methods, attendant exposures were measured using both charcoal-tube samplers and canisters. Service stations with and without vapor-recovery systems were included in the assessment. The MTBE content of the gasoline ranged from 13.4 to 15.7 LV% at all stations monitored. Of the eight, 4-hour, side-by-side charcoal tube or canister samples collected at vapor-recovery stations, seven were above the analytical detection limit for MTBE (detection limit not specified), ranging from 0.084 to 0.558 ppm and averaging (geometric mean) 0.33 ppm. Two 8-hour PBZ charcoal-tube samples collected at stations without vapor-recovery systems had MTBE concentrations of 0.554 and 1.191 ppm, respectively.
In 1994, the API contracted with NATLSCO for an evaluation of exposure to oxygenated fuel components among attendants and mechanics at sixteen service stations in four geographical areas (API, 1995). For comparison purposes, sampling was conducted during the winter (February - April) oxyfuel season, and summer (July - August) non-oxyfuel season. During the winter, MTBE was present in the fuel of the stations in approximate amounts ranging from 10-17 wt%. During the summer, no oxygenates were detected above the 0.1 wt% analytical limit of detection.
During the "winter phase", 51 long-term (generally greater than six hours) PBZ samples for MTBE ranged from 0.03 to 0.5 ppm, with a geometric mean of 0.2 ppm. Fifty-nine short-term (generally 15 - 20 minutes) PBZ samples ranged from 0.32 - 2.1 ppm, with a geometric mean of 0.6 ppm. During the "summer phase", 53 long-term PBZ samples ranged from 0.03 - 0.42 ppm, with a geometric mean of 0.08. Sixty-one short-term samples ranged from 0.19 - 0.33 ppm, averaging 0.31 ppm. The author notes that summary statistics (range and means) include non-detected results; they were assigned their LDC value and included in range and GM computations. This conservative approach yields overestimates of actual average exposures.
In 1992, the API surveyed member companies to obtain MTBE occupational exposure data (API, 1994). Of the samples collected at service stations, MTBE concentrations of 13 full-shift samples ranged from 0.09 to 34 ppm, averaging 0.77 ppm (geometric mean). The author noted that, in addition to fuel attendants, samples were collected from mechanics, weights and measures inspectors, and people repairing fuel pumps.
As with service station attendants, auto mechanics and auto technicians represent a large population with potential for significant exposure to gasoline. In the API 16 service station study (API, 1995a), 86 long-term (generally greater than six hours), PBZ samples for MTBE ranged from 0.02 to 2.6 ppm, averaging (geometric mean) 0.12 ppm for the "winter phase", and from 0.02 to 0.18 ppm, averaging 0.03 ppm during the "summer phase". Short-term samples (n=88) ranged up to 32 ppm, with winter and summer geometric mean exposures of 1.04 and 0.42 ppm, respectively. As with the attendant sample results, these data include non-detectable values, which overestimates the average exposures.
In the Fairbanks study (NIOSH, 1993b), 17 of 26 PBZ samples from auto mechanics were above the LDC for MTBE (0.03 ppm). Results ranged from less than 0.03 to 0.45 ppm, averaging 0.06 ppm (geometric mean). As previously mentioned, at the time of the NIOSH investigation, MTBE use as an oxygenate had been discontinued; its use was only as an octane enhancer (generally less than 1% MTBE/gasoline blend).
In a similar study in Stamford, Connecticut, (NIOSH, 1993c), 23 of 28 PBZ samples collected from auto mechanics were above the LDC, ranging from less than 0.03 to 12.04 ppm, averaging 0.11 ppm. The MTBE content of five bulk gasoline samples collected from the area ranged from 13 to 17 LV%.
In a third NIOSH study supporting a CDC epidemiologic investigation in Albany, New York, an area not using oxygenated gasoline (NIOSH, 1993d), three of eight PBZ samples from mechanics were above the LDC, ranging from less than 0.03 to 0.14 ppm and averaging 0.03 (geometric mean). The MTBE content of the gasoline was reported to be in concentrations of up to 10%.
In the three NIOSH evaluations of MTBE exposure in support of CDC epidemiologic investigations (NIOSH 1993b, NIOSH 1993c, NIOSH 1993d), researchers evaluated occupations in which people had significant exposure to motor vehicles and traffic (but not gasoline) in an effort to determine the extent of exposure to MTBE. These occupations included 1) service station cashiers or managers, 2) service advisors, 3) parking meter attendants, 4) animal control personnel, 5) truck drivers, and 6) taxi drivers. Of the 26 PBZ samples collected from people in these occupations in Fairbanks, Stamford, and Albany, only one sample, reported at 0.10 ppm from a service station manager, was above the LDC for MTBE. From these same investigations in Fairbanks and Stamford, exposure to MTBE by commuters was assessed. Commuters were defined as people whose occupations required them to spend a significant amount of their time driving. None of 14 PBZ samples from commuters was above the 0.03 ppm LDC. The last major category assessed in these studies was parking lot attendants. Of six PBZ samples from people in this group, one was reported at 0.10 ppm, which is above the LDC for MTBE.
From a survey of 17 member company's occupational exposure data for MTBE, the API published summary statistics for PBZ long-term and short-term samples (API, 1995b), see Table 2. Although additional environmental data were presented in the API report, the data in Table 2 represent samples for which the sampling time was known/available. As expected, due to the numerous contributors, sampling and analyses varied between the OSHA Method 7, NIOSH method 1615, and in-house, industry-developed methods.
|Short-term Samples1||Long-term Samples2|
|Operation||N/ND3||Range (ppm)||GM4||N/ND||Range (ppm)||GM|
|Mfg.; routine maint.||8/1||0.50-7.19||1.12||4/0||0.04-0.70||0.13|
|Blending; neat MTBE||35/1||0-97||4.73||12/5||0.04-87.9||1.73|
|Transport; neat MTBE||66/4||0.30-1050||11.84||10/1||0.03-711.9||0.30|
The major metabolic pathway of MTBE in both animals and humans is its oxidative demethylation, which leads to the formation of t-butyl alcohol (TBA) and formaldehyde. There is evidence that oxidative demethylation of MTBE is catalyzed by the cytochrome P450 enzymes (Brady et al., 1990). TBA was identified in the blood and expired air of rats which received 14C-MTBE (radiolabeled at the central butyl carbon) by oral, intravenous, dermal, or inhalation exposure (Bio-Research Laboratories Report # 38843; Exxon 1988). Secondary metabolism of TBA results in the formation of 2-methyl-1,2-propanediol, which is further metabolized to (alpha-hydroxy isobutyric acid (Bio-Research Reports # 38843-38844, 1990). Both metabolites were identified in the urine of rats which received 14C-MTBE (Bio-Research Report # 38843-38844, 1990). In vitro studies with rat liver microsomes also indicated that TBA may undergo oxidative demethylation to produce formaldehyde (Cederbaum and Cohen, 1980). A negligible portion of the administered MTBE dose (1% or less) to the rats was converted to 14C02 (Table 3).
In another study (API 1984), significantly larger amounts of 14C02 (ca. 7.0% of dose) were reported to have been expired after intraperitoneal administration of 232 mg 14C-MTBE/kg (radiolabeled at the methyl and the central butyl carbons). Formaldehyde was not reported in rats which received 14C-MTBE by oral, intravenous, dermal, or inhalation exposure (Bio-Research Laboratories Reports # 38842-38845, 1990). However, it was reported that 14C-formic acid accounted for most of the MTBE metabolites which were excreted in the urine and feces (approximately 3 and 1 % of dose, respectively) of rats that received 232 mg 14C-MTBE/kg by intraperitoneal administration (API, 1984).
In vitro metabolism studies revealed that incubating MTBE (5 mM) with rat hepatic microsomes resulted in the formation of relatively equimolar amounts of TBA and formaldehyde (Brady et al., 1990). Additional evidence for formaldehyde formation from MTBE was reported in studies which investigated the in vitro metabolism of TBA using rat hepatic microsomes (Cederbaum and Cohen, 1980). Results of these studies showed that oxidative demethylation of TBA by rat microsomes resulted in the formation of formaldehyde.
The disposition and toxicokinetics of 14C-MTBE and its major metabolite, TBA, were compared in mature adult male and female F344 rats after oral, intravenous, inhalation, and dermal exposure (Bio-Research Laboratories Reports #38842-45, 1990). 14C-MTBE (radiolabeled at the central butyl carbon) was administered to groups of rats in a dosing vehicle of 0.9% saline at a single dose of 40 mg/kg body weight using gavage, intravenous, or dermal (occluded application) routes. A higher dose of 400 mg/kg body weight was similarly administered to rats using oral or dermal application. A summary of the disposition of MTBE is presented in Table 3.
The major routes of MTBE elimination after oral, intravenous, intraperitoneal or dermal exposure were via the lungs as expired organic volatiles and via the kidneys as urinary metabolites (Table 3). Quantitative analysis of charcoal-trapped expired organic volatiles showed that the unchanged parent compound accounted for most of the organic volatiles exhaled by these rats. The portions of MTBE dose expired as unchanged parent compound and TBA were relatively higher after oral compared to intravenous exposure (Table 3). Elimination of MTBE via the lungs was also demonstrated in earlier studies conducted after MTBE was administered to Fischer rats orally or intravenously (Exxon, 1988). Additionally, exhalation of unchanged MTBE via the lungs was reported in mice as the major elimination pathway of MTBE after intraperitoneal administration at 50, 100, and 500 mg/kg (Yoshikawa et al., 1994). Pulmonary elimination of parent MTBE in mice ranged from 23% to 69% of the administered dose (Yoshikawa et al., 1994). In a study where MTBE was administered to CD rats intraperitoneally, a surprising 91-92% of the dose was eliminated in the expired air as parent MTBE (Table 3).
TBA exhalation accounted for 3% or less of the administered dose after all 3 routes of administration (Table 3). Two urinary metabolites, 2-methyl-1,2-propanediol and (alpha-hydroxy isobutyric acid, were identified in rats which received MTBE (Bio-Research Laboratories Reports #38843-38844, 1990). These 2 metabolites are most likely formed as a result of secondary metabolism of TBA. Fecal elimination of MTBE-derived radioactivity was minimal and accounted for approximately 1% of the administered dose regardless of the route of administration (Table 3).
The toxicokinetic parameters of MTBE and TBA depend on the dose and route of administration (Table 4). While the half-life (t1/2) of MTBE ranged from 0.45 to 0.62 hour after the low oral and intravenous dose, it ranged from 0.79 to 0.93 hour after administration of the high doses (Table 4). On the other hand, the t1/2 of MTBE after intraperitoneal and dermal application varied from 0.8-1.0 and 0.9-2.3 hours, respectively (Table 4). The t1/2 of inhaled MTBE was unaffected by dose or repetitive exposure to MTBE and ranged from 0.51-0.63 hour (Table 4). Maximum plasma concentration (Cmax) was higher among males than females after oral administration of MTBE (Table 4), and the difference was statistically significant only after intravenous administration. The total plasma clearance (CL) of MTBE was relatively similar in all animals receiving similar doses of MTBE regardless of the route of administration and ranged from 273 to 481 ml/hr. TBA, the main metabolite of MTBE, was detected in the expired air; however, products of TBA's secondary metabolism (2-methyl-1,2-propanediol and alpha-hydroxy isobutyric acid) were detected in the urine (Bio-Research Laboratories Report #38843, 1990).
In another study, MTBE metabolism and toxicokinetics were investigated in F344 rats of both sexes using a 6-hour single exposure to 400 ppm (calculated doses in males and females are 215 and 293 mg/kg, respectively) or 8000 ppm (the calculated doses in males and females were 4220 and 5840 mg/kg, respectively) using nose-only inhalation (Bio-Research Reports Laboratories #38844-38845, 1990). Using the same route of exposure, the toxicokinetics of MTBE was also investigated using repeat 6-hour daily exposure to 400 ppm (calculated doses in males and females were 245 and 344 mg/kg, respectively) after 14 days of 6-hour daily exposures to unlabeled MTBE (Bio-Research Laboratories Reports #38844-38845, 1990).
In contrast with routes of elimination when 14C-MTBE was administered orally, dermally, intravenously, or intraperitoneally, the main route of post exposure elimination of MTBE-derived radioactivity after nose-only inhalation was in the urine; the second major route of elimination was exhalation of organic volatiles (Table 3). However, the percentage of dose eliminated via each of the 2 routes (lungs and the kidneys) may actually be different since exhalation of MTBE and TBA during the 6-hour exposure to MTBE was not accounted for in these studies. Minimal changes in the disposition of MTBE were observed after repeat nose-only inhalation exposure to 400 ppm. In both exposure regimens, approximately 60-70% of the dose was eliminated in the urine compared to 15-20 % of the dose exhaled in the expired air. Furthermore, the ratio of MTBE/TBA exhaled in the expired air was lower in animals which received MTBE via nose-only inhalation compared to oral, intravenous, or dermal exposure (Table 3). Significant changes in the disposition of MTBE were observed after exposure to 8000 ppm. A decline in the portion of the dose eliminated in the urine in association with an increase in the exhalation of MTBE was observed (Table 3). In addition, the ratio of MTBE/TBA exhalation increased after exposure to 8000 ppm as a result of increased exhalation of unchanged MTBE (Table 3). Further, significant differences in the AUCs of the 2 doses were observed (Table 4).
In summary, studies of the metabolism, disposition, and toxicokinetics of MTBE in animals demonstrated that:
|Route and Dose|
|Route & Dose|
Cmax = maximum plasma concentration; AUC = area under the curve; T1/2 = half life; CL = total clearance; NA = data not available.
To date, three controlled human exposure studies on MTBE have been conducted. These three studies had similar designs and were aimed at evaluating internal dose levels resulting from known controlled exposure and the pharmacokinetics of uptake and decay.
In 1993, EPA and CDC collaborated to examine the internal dose concentrations of MTBE and its metabolite TBA resulting from exposing one male and one female volunteer for 1 hour to a 1.39 ppm MTBE concentration in a controlled environmental chamber (Gerrity et al., 1993; Prah et al., 1994; Buckley et al., 1995). MTBE and TBA levels in blood and urine and MTBE in the breath of these two subjects were determined both during the exposure phase and for 8 hours after the subjects were removed from the exposure. These subjects had significantly different body mass - the male weighed 102.5 kg and the female weighed 66.5 kg.
In both subjects, blood MTBE levels showed a rapid increase during the exposure phase but did not reach a steady-state plateau after 1 hour. Once the subjects were removed from the chamber, the blood MTBE levels decreased rapidly indicating that mo st of the internal dose quickly leaves the body, but in one subject these levels did not return to pre-exposure levels even after 7 hours. TBA levels in blood also rose quickly after the exposure began, but there was a much longer plateau period that extended over the entire 7-hour evaluation. Breath measurements on these subjects did not return to baseline after 7 hours in either case (Buckley et al., 1995).
MTBE levels measured in urine were of similar concentration and followed a similar course to blood MTBE concentrations with a rapid rise and decay. TBA levels in urine increased to a plateau, which remained constant for a number of hours, but the rise in these levels appeared to be delayed for blood TBA. Levels of MTBE in the breath also showed a similar response to levels in blood and urine with a rapid increase upon the initiation of exposure. The rate of increase slowed with time, but did not reach a plateau during this 1-hour exposure period. Thus, these results show that there is a rapid equilibration among blood, breath, and urine levels of MTBE. The metabolism of MTBE to TBA is a rapid process, and even though a large portion of MTBE is exha led unchanged, a fraction is metabolized to form TBA. Excretion of unmetabolized MTBE should also be considered as a significant elimination mechanism since the urine levels of this compound rapidly increase upon exposure.
The MTBE blood level reached after a 1-hour exposure to 1.39 ppm MTBE in air for the male subject was 8.2 ug/L and for the female subject, 14.1 ug/L. Thus, the response of MTBE blood levels was 1.7 times higher in the second subject than in the first. The breath levels in the second subject were also much higher than in the first subject. After 1 hour, MTBE levels in the breath were 43.0 ug/L in the first subject and 69.5 ug/L in the second, so that the second subject's breath MTBE level was 1.6 times higher than the first subject's breath MTBE level. This figure is in agreement with the blood results. These limited findings suggest a dependence of the internal dose of MTBE on body mass that differed in the two subjects by a ratio of 1.5. In contrast, the level of TBA in blood was similar for these subjects, reaching maximum levels of 9.5 ug/L and 10.4 ug/L.
Evaluation of the blood and breath MTBE concentrations after removal from exposure indicates that the decay process over time is multiexponential. Fitting the blood and breath to a three-exponential model yields residence times of 10.5 minutes, 75.1 minutes and 31 hours; and 3.3 minutes, 33 minutes and 8 hours, respectively. Because of the limited number of samples taken during the decay period, there is a wide uncertainty in these residence times, but the presence of a multiexponential decay is clear. This finding explains how MTBE can be eliminated rapidly from the body immediately after exposure stops and still not return to baseline 7 hours after the subjects were removed from exposure to MTBE. Previous studies have discovered this same finding with other VOCs (Pellizzari et al., 1992). In these previous studies, the long-term exponential decay was given as evidence of the deposition of VOCs into deeper body stores, most likely in adipose tissue. The determination of a long-term exponential decay for MTBE concentration suggests that, upon exposure, a significant portion of this compound also deposits in deeper body stores. The studies performed by the EPA and CDC cited earlier were limited to short-term exposures to people who had been exposed recently to MTBE. A slowly eliminated component suggests that, with repeated exposures, the internal dose of MTBE in the body may remain elevated over a long time period. There will be short-term spikes in the internal dose of MTBE as each exposure occurs, but this compound will give a steady-state level that depends on the frequency of exposure events and the exposure levels during these events. Pre-exposure levels of MTBE will be indicative of long-term exposure integration, and samples taken immediately after or during exposure will show short-term levels. The slow excretion of TBA suggests that this compound will accumulate even more than MTBE, so that a substantial internal dose level of TBA may result even when exposure to MTBE occurs at lower levels or less often.
In 1993, investigators at Yale University (Cain et al., 1993) measured blood MTBE and TBA levels in four subjects before, during, and after 1 hour of exposure to 1.7 ppm of MTBE. Measurements were made for 90 minutes after the end of exposure. This study also showed a rapid rise of MTBE blood concentrations during the exposure period with failure to reach equilibrium after 1 hour. MTBE concentrations reached an average of 17.1 ug/L after 1 hour and decayed after the subjects were removed from the exposure. The rate of decay appears to be slower than that reported in the EPA/CDC study in which the half-life was reported to be approximately 40 minutes; however, after 90 minutes, the MTBE levels were still a factor of 9 above the pre-exposure level. Peak MTBE blood concentrations were equivalent to the second EPA/CDC subject after differences in exposure level are taken into account.
Swedish researchers (Nihlen et al., 1994; Johanson et al., 1995) have performed a series of controlled chamber MTBE-exposure experiments at concentrations that were significantly higher than those described in the first two studies and that lasted for 2 hours. They measured levels of MTBE and TBA in blood, urine, inhaled air, and exhaled air during and after exposure of 10 healthy male volunteers to 5, 25, and finally, 50 ppm MTBE vapor during light physical exercise. They monitored blood levels for 24 hours (48 hours at 50 ppm) after the cessation of exposure. Low uptake, high post-exposure exhalation, and low blood clearance indicated the slow metabolism of MTBE.
Blood concentrations of MTBE reached a maximum concentration of approximately 1.2, 6.5, and 12.5 uM, which are equivalent to 106, 570, and 1100 ug/L for the 5, 25, and 50 ppm MTBE exposures. These levels are approximately twice those expected from the previous studies. The differences can be explained at least in part by the longer exposure time and by the subjects' light exercise during exposure. The pharmacokinetics of MTBE in blood was determined to require a three-exponential fit of the data with half-lives of 10 minutes, 1.5 hours, and 19 hours. These results are in good agreement with the EPA/CDC results and support the indication that, although most of the blood concentration of MTBE is quickly eliminated, long-term elevated levels of MTBE may occur with repeated exposure. The area under the concentration curves of blood MTBE and blood TBA were linearly related to the MTBE exposure level, suggesting that the toxicokinetics are linear up to at least 50 ppm. In contrast to MTBE however, blood TBA continued to increase during the 2-hour MTBE exposure, then leveled off and started to decline about 6 hours later. The post-exposure half-life of TBA in blood was 10 hours.
The study by Nihlen et al. 1994 also found detectable levels of MTBE and TBA in urine as did the EPA/CDC study. Less than 1% of the absorbed dose of MTBE was excreted as TBA in urine within 24 hours; this finding may indicate further metabolism of TBA in humans as has been shown in rats (Bio-Research Laboratories Report #38845, 1990). Urinary MTBE levels decreased rapidly after exposure ended, and TBA concentrations changed during the uptake and elimination phases with a rapid increase, long plateau phase, and a half-life of 7-9 hours. The presence of MTBE and TBA in urine and the longer half-life of MTBE in urine compared to blood suggest that urine may be a useful matrix for determining internal-dose levels of MTBE and TBA. Since people are more likely to be willing to give a urine specimen than a blood sample, it is probable that subject participation would increase. However, since urinary volume output is variable, a single measurement of TBA or MTBE may not be as meaningful as a similar measurement in blood. Use of MTBE or TBA in urine must be fully investigated before this matrix can be accepted as a sensitive measure of MTBE exposure.
For workers in Phase 1, the median pre-shift concentration of MTBE in blood was 1.15 ug/L (range 0.1 - 27.8 ug/L), with an increase to a median post-shift concentration of 1.80 ug/L (range 0.2 - 37.0 ug/L). This difference was statistically significant. Among workers in Phase 2, median pre-shift MTBE levels were 0.21 ug/L (range < 0.05 - 4.35 ug/L) and the median postshift concentration was 0.25 ug/L (range < 0.05 - 1.44 ug/L). The results show that blood MTBE levels increased during the work-shift while MTBE was being used as a fuel oxygenate, and blood levels were lower after the cessation of the oxygenated fuels program.
In Phase 1, the blood MTBE levels among commuters changed from 0.18 ug/L before they commuted to 0.83 ug/L afterward. In Phase 2, the blood MTBE levels before the commute (0.09 ug/L) and after the commute (0.10 ug/L) were not significantly different (t-test). Since many people commute to work each day, the increase in MTBE internal dose levels among commuters suggests that a large proportion of the public in locations where MTBE is present in gasoline receive a measurable exposure to this compound.
It is significant that blood MTBE levels were not below detection limits for either workers or commuters before the work shift or morning commute (the period of MTBE exposure). The median blood MTBE level in 15 participants in the Third National Health and Nutrition Examination Survey was below the detection limit of 0.05 ug/L (Ashley, 1993). Elevated levels measured before exposure support the results of controlled human exposure studies because they suggest that MTBE levels in the body have not returned to baseline even after 16 hours. Thus, daily exposures may yield elevated levels for a time much longer than the estimated half-life.
In the Fairbanks study, there was a significant correlation between the difference between pre- and post-shift blood MTBE levels among workers and in workplace air concentrations. For these workers, the median air concentration of MTBE was 0.1 ppm and blood levels showed a statistically significant mean difference between pre- and post-exposure samples of 0.65 ug/L. Among some of the workers whose exposure was greater, air levels of 0.55 ug/L yielded a difference between pre- and post-exposure samples of 10 ug/L. This finding is in agreement with the EPA/CDC controlled-chamber exposure study in which air levels of 1.4 ppm MTBE in the chamber yielded 11 ug/L MTBE in the blood. For the workers, the internal dose levels of MTBE were higher, since the workers were exposed for 8 hours rather than 1 hour as in the controlled exposure studies. Blood levels do not increase linearly as time increases, because the internal dose levels begin to plateau after about 1 hour. Thus, the internal dose levels of MTBE are a function of both exposure time and exposure concentration and not just their product as exposure is normally defined.
The measurements of blood MTBE levels among car-repair workers in Stamford, Connecticut, is in excellent agreement with the levels found among the workers in Fairbanks. Thus, this segment of the population receives a substantial MTBE exposure even in a region of the country that does not have Alaska's extreme conditions. Internal dose levels of MTBE among commuters were lower than those found in Fairbanks. This difference may be due to the differences in driving habits during the periods when these studies occurred. The first phase of the Fairbanks investigation took place during the winter when commuters are more likely to not open their car windows, thus reducing the free flow of fresh air into the vehicle. In the Stamford study, samples were collected at the beginning of spring, when ventilation through the vehicle may have reduced the extent of exposure. Since the oxygenated fuels program is chiefly in effect during the winter months, the weather may have had a significant effect on the extent of exposure of commuters.
Studies in humans of MTBE metabolism suggest that, with repeated exposures, levels of both MTBE and TBA may be elevated in biological fluids for a time much longer than the half-life with TBA doing so to a larger extent as a result of the longer residence times for this metabolite. Exposure will be integrated over a period of days for MTBE and possibly longer for TBA when repeat exposure to MTBE occurs. Thus, there will be repeated spikes in the internal dose immediately after exposure followed by a steady-state level between exposure events.
By carefully timing sample collection with exposure events, it is possible to evaluate exposure and to gain insight into the duration of either short-term or long-term MTBE levels. Internal-dose concentrations found among workers and commuters indicate that both highly exposed populations and the commuting public experienced increases in internal-dose levels of MTBE when the oxygenated fuels program was in effect.
In response to considerable media coverage and public concern about oxygenated fuel containing MTBE in Missoula, Montana, during the 1992-1993 winter season, the Missoula City-County Health Department conducted a survey of local physicians in January 1993 to screen for possible health effects (Missoula City-County Health Department, 1993). A total of 18 physicians' offices were contacted by telephone. Twelve of the offices had received complaints from patients specifically regarding oxygenated fuels; the symptoms reported included many of the same acute effects that had been reported in Alaska, as well as others, including an exacerbation of symptoms among people with asthma. The next winter, the oxygenate used was ethanol rather than MTBE, and public concern over this issue essentially disappeared (E. Leahy, Missoula City-County Health Department, personal communication, 1995).
Several other state health departments have also received some health complaints related to oxygenated gasoline from citizens similar to those received in Alaska. Passive reliance on reports of citizen complaints to health departments, however, is a highly insensitive and potentially misleading measure of health effects related to gasoline and its components. State health departments do not routinely conduct surveillance for complaints related to gasoline or other environmental exposures. In addition, citizens would not necessarily report health complaints of a nonspecific and noninfectious nature, such as a headache, to the health department, even if they believed symptoms were caused by gasoline or some other environmental agent. If citizens did report such symptoms to other organizations, such as to a poison control center, local medical providers, or the local environmental control agency, these reports easily might not be communicated to health authorities.
For example, in the winter of 1992-1993, the Colorado Department of Public Health and Environment received three complaints about the health effects of, or odors related to, oxygenated gasoline (Livo, 1995). That same winter, a private citizen in Colorado Springs actively collected reports from more than 50 people about health complaints as a result of exposure to MTBE (Pat McCord, personal communication, 1993). In the absence of concerted efforts to identify and verify health complaints in the community, analyses of the number of complaints received by different health departments or trends in the number of complaints received over time are difficult to interpret.
A recurrent hypothesis put forth by many of the people who have expressed concerns about the health effects of MTBE, and one that EPA acknowledges (USEPA, 1993; 1994), is that some people may be more sensitive than others to MTBE. To explore this possibility further, investigators in New Jersey interviewed 14 people with multiple chemical sensitivities (MCS) and 5 people with chronic fatigue syndrome (CFS) (Fiedler et al., 1994). The investigators also interviewed six otherwise healthy people as control subjects, but comparisons between groups composed of so few people are difficult to interpret. Qualitatively, people with MCS and CFS reported symptoms previously associated with MTBE with apparently greater frequency than they reported other symptoms and more frequently than control subjects, but these symptoms were reported to develop in locations such as shopping malls where exposures to MTBE were presumed to be fairly low.
Refinery workers who are members of the Oil, Chemical, and Atomic Workers Union and exposed to MTBE at several refineries have reported many of the same symptoms, as well as others not previously associated with MTBE or gasoline, that were reported among motorists in Fairbanks when MTBE was first introduced as an oxygenate (Medlin, 1995). The most common complaints were headaches, sinus problems, fatigue, and shortness of breath.
The workers also completed questionnaires that asked about the occurrence of 15 different health symptoms, the 7 that had been frequently reported by citizens and 8 others. Questions focused on symptoms that had occurred for the first time or with greater frequency since the beginning of the oxygenated season and which the respondent did not attribute to having a cold or the flu. The questionnaires were administered in person to the workers at the end of the work shift. By necessity, the workers studied were few and were not randomly selected, so the actual prevalence figures for specific symptoms may not reflect the prevalence of these symptoms among other workers who were similarly exposed. Certain symptoms, especially headache, eye irritation, and a burning sensation of the nose and throat, were much more common when the oxygenated program was in operation than after the program had been suspended. This investigation, however, could not determine whether exposure to MTBE in oxygenated gasoline actually was the cause of increased symptom reporting.
After the investigations in Fairbanks and with the support of the EPA, CDC undertook two subsequent investigations in collaboration with state health officials in Stamford, Connecticut (CDC, 1993b), and in Albany, New York (CDC, 1993c). The follow-up investigations were intended to duplicate the methods used in Fairbanks in a city that had used oxygenated gasoline but had not experienced media publicity and had not received complaints from local citizens (Stamford), and in a city that had not participated in the oxygenated fuel program (Albany). One important distinction between the investigations in Fairbanks and these two subsequent investigations is that they occurred at different times of the year under different climatic conditions.
As in the Fairbanks investigations, the investigations in Stamford and Albany included workers, such as auto mechanics and gasoline service station workers, who had opportunities for exposure to gasoline and motor vehicle emissions. The questionnaires were designed to be nearly identical to those used in Fairbanks and also focused on symptoms that the respondent did not attribute to a cold or the flu. Air and blood measurements of MTBE and other gasoline components from these two investigations are discussed in the exposure-profile section.
Because of time and resource constraints, both the Stamford and Albany investigations relied on convenience samples of relatively small numbers of workers in different job categories. In both locations, the occupationally exposed workers were predominantly men. Despite these and other limitations, these investigations provided good comparative data on exposures and doses and gave a qualitative indication of the magnitude of symptoms experienced by some workers in these two cities.
The small numbers of workers and the nonrandom nature of their selection, however, limit the interpretation of specific prevalence figures. Qualitatively, the prevalence of the most common symptoms, such as headache and cough, occurring over the last month, were not appreciably higher among men who worked around cars and gasoline in Stamford than men with similar occupations in Albany, where exposure to MTBE was generally much lower. In addition, the prevalence of these symptoms was lower among workers in Stamford than among workers in similar job categories in Fairbanks during the oxygenated fuel season.
In Stamford, job category alone was not a good indicator of exposure; both blood and air measurements revealed that people who worked as mechanics and at gasoline stations had highly variable exposures to MTBE (White et al., 1995). Blood measurements were available for only 30 workers; of these workers, the 8 with the highest blood levels of MTBE were significantly more likely to report 1 or more of the 7 key symptoms on that day than were the remaining 22 workers (OR = 21.0, 95% CI = 1.8 - 540). These workers with the highest exposures to MTBE were also more heavily exposed to gasoline. There was no comparable reference group of workers in Stamford exposed to gasoline not containing MTBE.
In April and May 1993, researchers at EOHSI interviewed garage workers in two parts of New Jersey: the northern part of the state near New York City where oxygenated fuel was still in use and where exposures to MTBE were higher, and the southern part, near Philadelphia, where use of oxygenated fuels had been discontinued at the end of February (Mohr et al., 1994). These garage workers were all employed by the State of New Jersey, had similar backgrounds and training, and worked under similar conditions. Workers were asked about the prevalence of symptoms during the last 30 days and at the beginning of the work shift and again at the end of the work shift. The questionnaire asked about the same symptoms that had been included in other investigations, but the wording was different and did not ask the respondent to identify symptoms not attributed to colds.
Overall, this investigation did not report major differences in symptom reporting between these two groups of workers in New Jersey with presumably different levels of exposure to MTBE. Comparisons between people who pumped gasoline in the north and south, matched by age, also did not demonstrate a difference in the prevalence of symptoms, but this analysis was made on the basis of information from only 11 workers in each area. The authors noted that this investigation was conducted at the end of the oxygenated fuel season; any change in the occurrence of symptoms, if caused by exposure to oxygenated fuels, would probably have been more noticeable at the beginning of the oxygenated fuel season.
Two telephone surveys also were conducted in Fairbanks during the oxygenated fuel season (December 1992) and after the program had been suspended in February 1993 (CDC, 1993a). Only 41 people were interviewed in December; the most prevalent symptoms they reported were eye irritation, headache, and a burning sensation of the nose or throat, each reported by a third of the people interviewed. In February, 100 people were interviewed, and the prevalence of headaches and other symptoms was much lower. Given the quasi-random sampling and small number of people interviewed, the estimate of the prevalence of different symptoms is fairly crude in both surveys.
In Anchorage, the medical officer examined records for outpatient visits among state employees, retirees, and dependents in Anchorage and Fairbanks for the winter of 1992-1993, when oxygenated gasoline was in use, compared with such visits the previous two winters (Gordian et al., 1995). The report of this analysis is brief and difficult to evaluate fully. The authors report that there had been an epidemic of headaches in January 1992 due to a viral illness. Therefore, a comparison in visits for headaches between the winter of 1992-1993, when oxygenated fuels were first in use, and the previous winter would have been unlikely to measure an increase in headaches. Compared with the winter of 1990-1991, however, the winter of 1992-1993 showed both Anchorage and Fairbanks reportedly experiencing a 40%-50% increase in visits for headaches, but the confidence interval for this estimate was fairly wide and certainly does not rule out chance as an explanation for this increase. As the authors point out, however, the data that were used for this analysis were limited in their ability to measure the sort of transient health symptoms that had been reported as related to MTBE exposure.
Nonoccupationally exposed motorists had also been included among the people interviewed in both the Stamford and Albany investigations (CDC, 1993b; CDC, 1993c). Since both investigations relied on convenience samples rather than on random samples of motorists, the people interviewed may not have been representative of the larger population in either city. In particular, unemployed and retired people were not included among the motorists interviewed. In both cities, the most common symptoms were headaches and cough, and the reported prevalence of these and other symptoms was fairly similar among motorists in both cities.
During the winter of 1994-1995, the Alaska Department of Health conducted a weekly random telephone survey of 100 adult residents in Anchorage, for 16 consecutive weeks, to identify possible health problems related to the use of ethanol as an oxygenate in fuels (Egeland and Ingle, 1995). Responses were grouped by time period to correspond to periods when ethanol containing gasoline was not in use, was in use, or was being phased in or phased out. The survey questionnaire asked about the presence of symptoms over the previous week, and persons who reported illnesses over the past week that could have accounted for symptoms (fever, sweats or chills, or cold or flu) were excluded. The results indicated a much lower prevalence of symptoms than had been reported 2 years earlier when MTBE was used in gasoline, and the reported prevalence of symptoms remained fairly similar across the different time periods.
Approximately 500 people in each location were interviewed in February and March 1995 by using a questionnaire that was based in part on the questionnaires that had previously been used in Alaska. In particular, the questionnaire asked whether the respondent had experienced any unusual health symptoms unrelated to a cold or the flu. Overall, the survey found that people in Milwaukee reported a higher prevalence of unusual symptoms than did residents of other areas of the state or of Chicago. Several other findings are of relevance in evaluating this increase. First, every symptom was elevated in Milwaukee, not just symptoms that had previously been associated with gasoline or chemical solvents. Second, the symptom prevalences were not elevated in Chicago compared with such prevalences in areas of Wisconsin where reformulated gasoline was not used, although Chicago was also using reformulated gasoline. Third, although people were asked to report unusual symptoms, the symptoms reported in Milwaukee were more likely to be associated with having had a cold or the flu, smoking cigarettes, or being aware of reformulated gasoline than were symptoms reported by people in Chicago or the rest of Wisconsin.
Although well conducted, the study has many limitations. It was a cross-sectional survey of symptoms but did not measure exposure. Participation rates were not high, particularly in Chicago. The number of people interviewed, although substantially greater than in other surveys of health effects and oxygenated gasoline, still did not provide sufficient statistical power to measure modest elevations in symptoms or to look carefully at specific population subgroups.
The study underwent peer review by the Environmental Committee of the Association of State and Territorial Health Officers (ASTHO) and other scientists not associated with the study (Anderson et al., 1995a). This committee concluded that the study was well conducted, had limitations, and that it had not ruled out the possibility that some people might have a greater sensitivity to reformulated gasoline. All but one of 16 committee members agreed with the statement that the study did "not support a conclusion that reformulated gasoline is associated with widespread or serious adverse health effects."
The same telephone survey questionnaire was subsequently administered to 1339 persons who had called Wisconsin government agencies with health concerns regarding reformulated gasoline (health contacts) between January 26 and March 17, 1995, had provided names and telephone numbers, and agreed to complete the questionnaire when later contacted again (Anderson et al., 1995b). Compared to residents of Milwaukee who had participated in the random telephone survey, the health contacts were slightly older, more likely to be white, male, own a car, and commute more than one hour per day, and they were twice as likely to be retired. In addition, the health contacts were less likely to smoke cigarettes and were more likely to have been diagnosed as having allergies, although they were not more likely to have been diagnosed as having asthma. None of these factors had been observed to be significant predictors of unusual health complaints in the random telephone survey. Health contacts also were more likely to have seen various news stories about MTBE than other Milwaukee residents.
Taken together, these studies suggest that most people do not experience adverse health effects from MTBE in gasoline, but the studies cannot rule out the possibility that some people do experience more acute symptoms from exposure to oxygenated gasoline than to conventional gasoline. Many basic questions, such as the relative importance of individual characteristics, exposure situations, and factors other than oxygenates for the occurrence of various health symptoms remain. Thus, a causal association between acute health effects and exposure to MTBE or other oxygenates in gasoline in a relatively smaller proportion of persons has not been demonstrated but cannot be ruled out on the basis of the limited epidemiologic studies that have been conducted to date. More definitive epidemiologic studies are needed to assess fully the connection between exposure to oxygenates in gasoline and acute health effects.
In EPA's investigation (Prah et al., 1994), 37 healthy, nonsmoking subjects (18 women, 19 men) from 18 through 35 years of age were studied. Each subject was exposed once for 1 hour to clean air and to 1.4 ppm "pure" MTBE in air on different days. The temperature and relative humidity in the chamber were maintained at 24°C and 40%, respectively. The endpoints selected for the EPA study were based on the observation that the symptomatic reports from Alaska resembled the types of symptoms associated with low-level organic solvent exposure. The endpoints for the EPA MTBE study can be divided into four categories:
1. Indicators of symptomatic response, including headache, nasal irritation, throat irritation, cough, eye irritation, odor-strength perception, and dizziness (measured before and during exposure):
2. Indicators of behavioral response (measured before and at the end of exposure):
3. Indicators of upper airway inflammation (measured before, immediately after, and 20 hours after exposure):
4. Indicators of eye inflammation:
Before exposure testing, each subject had a determination of his or her individual odor threshold for MTBE in water. Of the subjects in the study, 76% correctly detected the presence of the odor of MTBE in water at a concentration of 0.18 ppm. These data compare reasonably well with data from another study reporting a detection threshold of 0.13 ppm (Clark, 1993). Thus, it can be assumed that most of the subjects studied had normal odor thresholds for MTBE.
There was no significant effect of MTBE on the overall qualitative rating of air quality or on reporting of headache and nasal irritation symptoms using either the computerized questionnaire or the EOHSI questionnaire. The neurobehavioral test battery also showed no effect of MTBE exposure. None of the markers of nasal and eye inflammation showed a statistically significant difference in the response to MTBE exposure from that of clean air. MTBE exposure also had no statistically significant effect on eye redness or on tear-film breakup times. The primary hypothesis tested in this protocol was that MTBE would cause changes in the reporting of symptoms of headache, nasal irritation, air-quality perception, and odor strength perception. Power calculations performed on the symptom data showed that (at p<0.05), there was adequate power to detect a 0.5 point change (on a five-point scale) with 90% power for odor-detection level, headache, and nasal irritation. A 0.25 point change in nasal irritation could have been detected with 80% power. Thus, the study had adequate statistical power to detect MTBE-related changes in symptoms if they had been present.
Investigators at Yale University (Cain et al., 1995) replicated the EPA study (Prah et al., 1994). A total of 43 subjects (22 men, 21 women) from the ages of 18 through 34 years participated. All of the endpoints studied by the EPA investigators were also studied by the Yale investigators, although slightly different methods for measuring eye redness, tear-film breakup times, and eye inflammation were used. The MTBE exposure concentration was slightly higher in the Yale study (1.7 ppm). In addition to single exposures to clean air and to MTBE for 1 hour at 75°F, each subject in the Yale study also underwent a 1-hour exposure to a complex mixture of 16 VOCs commonly found in gasoline (C4, C5, and C6 saturates; and C4 and C5 olefins). This exposure to a surrogate gasoline served as a positive control for the MTBE exposure. In the pilot phase of investigation, the surrogate gasoline was found to have no detectable odor. Consequently, isopropyl mercaptan (the odorant used in natural gas) was added to provide an unpleasant odor. The total VOC concentration of the atmosphere was 7.1 ppm. Besides the addition of the VOC exposure and the slightly higher MTBE exposure concentration, the only other major difference between the EPA and Yale studies was that, in the latter study, exposures of a given individual were separated by only 3 days, as opposed to 1 week in the EPA study (in which each person received both an air and an MTBE exposure).
When MTBE exposure was compared with clean air exposure, Cain et al. (1995) found essentially the same results as found in the EPA study (i.e., MTBE exposure had no statistically significant effects on symptoms, the neurobehavioral test battery, nasal inflammation, eye inflammation, eye redness, or tear-film breakup times). Results of statistical power calculations on the Yale symptom data by EPA (House, 1993) were similar to those of the EPA study. Likewise, for the objective measures, means of the MTBE and control groups were similar and standard deviations were small. Thus, the design of the study was sufficiently robust to have confidence in the negative outcome. When the VOC exposure was compared with clean air exposure, it was found that the VOC exposure caused an increase in inflammatory cells in the nasal lavage on the day following exposure (Cain et al., 1995). This finding was consistent with previous work done at the EPA laboratories (Koren et al., 1992).
More recently, a human experimental study was conducted at the Swedish National Institute of Occupational Health (Nihlen et al., 1994; Johanson et al., 1995), in which 10 healthy male volunteers, ages 23-51 years old, were exposed during light exercise for 2-hour periods to three successive concentrations of MTBE -- 5, 25, and 50 ppm. Researchers measured eye-blinking frequency, conjunctival epithelial damage, eye redness, tear-film breakup time, acoustic rhinometry, nasal/mouth peak expiratory flow, and inflammation markers in nasal lavage, along with subjective ratings of "discomfort, irritative symptoms, and CNS effects." Other than a slight, marginally significant increase in the indicator "nasal swelling," which was not concentration-related, no significant irritant or subjective effects of MTBE exposure were found. However, subjective ratings indicating detection of a solvent odor were highly significant.
Further experimental studies are needed to determine whether measurable symptoms can be induced in human volunteers with self-described sensitivity to oxyfuels. If such effects can in fact be induced through exposures to gasoline-oxygenate (e.g., MTBE) mixtures under controlled conditions, then it may be possible to identify the variables contributing to the occurrence of such reactions. If indicated, these studies might be extended to investigate the effects of combustive as well as evaporative oxyfuel emissions, although ethical and other considerations may make it difficult or impossible to administer any but relatively low exposure levels of such mixtures. Experimental studies could also be devoted to quantifying the concentration-response relationship for any effects. Such studies could also examine variables such as odor, temperature, chemical sensitivity, smoking, alcohol use, and nonphysiological factors that might contribute to the perception of acute symptoms in some individuals.
Evaluation of acute neurotoxic effects has been conducted in rats following a single 6-hour exposure by inhalation to 0, 800, 4000, and 8000 ppm MTBE (Gill, 1989). Evaluation of neurotoxic effects was conducted within 1 hour after exposure and at 6 and 24 hours to examine the transient nature of the effect. According to clinical observation methods, there was no evidence of toxicity after exposure. When behavioral function was evaluated using a systematic screening battery of tests known as the Functional Observational Battery (FOB), distinct signs of toxicity were evident. Within 1 hour, a concentration-related increase in the incidence or severity of ataxia and gait change was observed in the groups exposed to 4000 ppm and 8000 ppm MTBE. Exposure to 8000 ppm produced labored respiration, lacrimation, decreased muscle tone, decreased rectal temperature, decreased treadmill performance, and an increased hind limb splay. In a few of the animals that received the high dose, there was evidence of decreased pupil size, loss of pupil response to light, decreased startle response, and decreased reflex response to toe pinch. All biologically relevant findings were restricted to the 1-hour post-exposure test session with no effects continuing to the 6-hour evaluation period. Automated measurements of motor activity were conducted during the hours immediately after exposure. Activity measurements were of high counts and large variability suggesting that the specific apparatus used had a photocell placement and sensitivity that measured movements other than just ambulatory. In male rats exposed to 8000 ppm MTBE, a decrease in mean activity level was evident within the initial 10 minutes of the test session, followed by a transient increase and a subsequent decrease during the 5-hour test session. For animals receiving the two lower doses (800 and 4000 ppm), mean activity was elevated relative to control levels during the initial 10 minutes, reflecting either a low-dose stimulant effect or an exaggerated recovery from anesthetic effects similar to what is seen with the biphasic response to other substances with sedative properties. Initially, females showed a similar pattern of depressed activity. However, within 1 hour, activity had returned to control levels. These changes in behavioral functioning, detected by both the FOB and motor activity, are consistent with transient CNS sedation.
Within 1 hour of cessation of exposure, these behavioral alterations were no longer evident. Brain weights were unaltered by MTBE exposure. Lower doses (0, 250, 500, 1000 ppm for 6 hours/day, 5 days/week for 13 weeks) resulted in no clinical signs except a dose-related increase in the depth of anaesthesia (Greenough et al., 1980).
After the 13-day dose-ranging study was completed, Dodd and Kintigh (1989) evaluated neurotoxicity after 13 weeks of repeated exposure to MTBE vapors. Male and female rats were exposed for 6 hours/day 5 days/week for 13 weeks to MTBE vapors of 0, 800, 4000, and 8000 ppm. During the first 3 weeks of exposure, body-weight gains decreased. To minimize the influence of the sedation effects associated with acute exposure, behavioral functioning of the animal was evaluated after a significant post-exposure period for each test interval; 19 hours for motor activity and 50 hours for FOB. Each time they were examined (at 4, 8, and 13 weeks), male rats in the group that received the 4000 ppm MTBE dose showed a decrease in hind-limb grip strength. By week 8, mean motor activity over 90 minutes was decreased in male rats in the group that had received the 8000 ppm dose, whereas female rats that had been given the 4000 ppm MTBE dose exhibited increased activity at weeks 8 and 13. The transient nature of these effects and the lack of a dose-response relationship suggest that exposure to MTBE for 13 weeks at concentrations of 8000 ppm or below did not produce toxicologically significant neurobehavioral changes in the Fischer 344 rat. Significant decreases in absolute mean brain weight were reported in both male (5%) and female (3%) rats exposed to 8000 ppm MTBE. During exposure, MTBE produced a significant decrease in body weight gain over a period of substantial growth. If the brain weight decrease was evaluated relative to the decrease in body weight gain, the relative change in brain weight was not different from that of control rats. A slight decrease in brain length (0.5 mm) was reported to be significant in male rats exposed to 8000 ppm MTBE. No changes were seen in brain width and females showed no effect on either brain parameter. Given the inherent variability associated with excising the brain from the cranium, and the limit of resolution and accuracy of measurement techniques, the slight decrease in brain length, while given numerical significance, is below the level of concern for biological significance. In the determination of statistical significance for brain parameters, the lack of adjustment for multiple comparison greatly influenced the number of values that reached the critical significance levels were not corrected for the number of comparisons performed thus, increasing the likelihood of type 1 statistical errors and inflating the level of significance. Histological examination of various brain regions and peripheral nerves failed to indicate evidence of morphological alterations in nervous system tissue after 13 weeks of exposure (Dodd and Kintigh, 1989).
Robinson and co-workers (1990) examined the oral toxicity of MTBE and found thatadult rats exhibited profound anesthesia immediately after oral exposure to MTBE (1428 mg/kg). Within 2 hours, all animals recovered normal motor and sensory functions. No additional clinical effects were seen during the 14 days of dosing or during the 90 days when the rats received oral doses of 1200 mg/kg MTBE.
Studies examining the effects of MTBE exposure during gestation report similar anaesthetic effects of hypoactivity and ataxia in mice dams at doses of 4000 and 8000 ppm (Bushy Run Research Center, #52-526, 1989). No adverse effects have been reported in rats receiving lower doses, (API, #32-30235, 1984) or in mice receiving doses up to 2500 ppm (API, #32-30237, 1984). In rabbits, doses of 4000 and 8000 ppm produced a decrease in body weight in pregnant dams; however, no other clinical signs were noted (Bushy Run Research Center #51-628, 1989).
Exposure to MTBE has been reported to transiently affect muscle creatine-kinase activity with an inhibition early in exposure followed by an increase in activity by the 15 week of exposure to 300 ppm (Savolainen et al., 1985). Although levels of MTBE and tertiary-butanol in the brain were similar in these animals, no effect was seen on brain succinate dehydrogenase, creatine kinase, or acetylcholinesterase activities.
The existing data on MTBE neurotoxicity suggests a low potential for producing neurotoxicity.
As a result of the positive results with L5178Y cells, MTBE was tested at similar concentrations for its ability to induce chromosome aberrations and sister chromatid exchanges (SCE) in Chinese hamster ovary cells in vitro, with and without exogenous metabolic activation (ARCO, 1980; Brusick, 1979). Two different grades of MTBE were tested, 99% and 99.9% pure. No induction of chromosome aberrations were seen with either grade, with or without exogenous metabolic activation (ARCO, 1980; Brusick, 1979). In the SCE test the 99.9% sample showed a positive response without activation; this result was not repeatable and the testing laboratory concluded that the overall response was negative (ARCO, 1980; Brusick, 1979). The 99% sample was negative under all test conditions (ARCO, 1980; Brusick, 1979).
MTBE did not damage DNA of primary rat hepatocytes in culture as measured by induction of unscheduled DNA synthesis (Cinelli, et al., 1992; Life Science Research, 1989c). A structure-activity relationship analysis resulted in a prediction that MTBE would not be mutagenic in Salmonella or produce chromosome aberrations or sister chromatid exchanges in cultured mammalian cells (Rosenkranz and Klopman, 1991).
In vivo, inhalation of MTBE did not produce chromosome aberrations in bone marrow cells of male or female rats exposed to 0, 800, 4000, and 8000 ppm, 6 hours/day for 5 days (Bushy Run Research Center, 1989), or micronuclei in the bone marrow cells of male or female mice exposed to 0, 400, 3000, and 8000 ppm, 6 hours/day for 2 days (Bushy Run Research Center, 1993). Intraperitoneal administration of 0.04, 0.133, or 0.4 ml MTBE/kg to male rats as single, acute doses, or as 5 consecutive doses, did not produce an increase in chromosome aberrations in femoral bone marrow cells (ARCO, 1980). Oral administration of 0, 1, 10, 100, and 1000 mg/kg to male and female CD-1 mice for 3 weeks did not produce mutations at the hypoxanthine-guanine phosphoribosyl transferase locus in lymphocytes (Ward et al., 1995). MTBE did not induce sex-linked recessive lethal mutations in the fruit fly, Drosophila melanogaster, when administered at 0, 0.03, 0.15, and 0.30% in its food (Hazleton Laboratories America, Inc., 1989)
Formaldehyde produced DNA damage in monkey and rat cells treated in vivo. Mixed results were obtained in other in vivo genetic tests in rodents and humans (IARC, 1995). Mutations were detected in the p53 gene isolated from nasal tumors of F344 rats following inhalation exposure to 15 ppm formaldehyde for 2 years (Recio, et al., 1992). Chromosome aberrations, sister chromatid exchanges, and micronuclei were induced in rat and mouse cells treated in vivo in some studies but not in others. Similarly, chromosome aberrations, sister chromatid exchanges, and micronuclei were found in humans exposed to formaldehyde in work environments in some studies but not others. Many of these differences could be the result of different treatment regimens and test protocols or of differences in the performance of the test laboratories. Formaldehyde produced sperm-morphology changes in rats but not in mice or humans (IARC, 1987, 1995).
Most in vivo rodent genetic toxicity studies used only single-dose treatment. However, a few inhalation studies used multiple dose treatments. Dallas et al. (1992) exposed male Sprague-Dawley rats to 0, 0.5, 3, or 15 ppm formaldehyde, 6 hours/day, 5 days/week, for 1 or 8 weeks. Dose-related increases in chromosome aberrations were seen in pulmonary macrophages; the responses at 15 ppm were significantly increased at both treatment times. There were no chromosome aberrations produced in bone marrow cells at either treatment time. In a similar study, Kligerman et al. (1984) treated male and female F344 rats with formaldehyde by inhalation at 0, 0.5, 6, or 15 ppm for 6 hours/day for 5 days. No increases in sister chromatid exchanges or chromosome aberrations were seen in peripheral blood lymphocytes.
Groups of 60 male and 60 female Sprague-Dawley rats, 8 weeks of age, were administered MTBE (> 99 % purity) in olive oil by gavage at doses of 0 (vehicle control), 250, or 1000 mg/kg body weight, 4 days per week for 104 weeks. Animals were kept under observation for their natural lifetime. Survival was higher in the 1000 mg/kg dose group of males than in control rats and was reduced in both treatment groups of females. Histopathological evaluations revealed a significant dose-related increase in the combined incidence of lymphomas and leukemia in female rats (2/60 in the control rats, 6/60 in the low-dose group, and 12/60 in the high-dose group), but not in male rats. Dysplastic proliferation of lymphoreticular tissue was also increased in treated female rats (1 control rat, 15 in the low-dose group, and 9 in the high-dose group). It should be noted that McConnell et al. (1986) recommended against combining mononuclear cell leukemias (MNCL) with lymphomas in F344 rats. In that strain of rat, MNCL occurs at a high spontaneous rate. For Sprague-Dawley rats, however, it is probably reasonable to combine leukemia and lymphoma, especially when the leukemias are of lymphocytic origin. Interstitial cell (Leydig cell) tumors of the testes were significantly increased in the high dose group of male rats (2/60 control rats, 2/60 in the low-dose group, and 11/60 in the high dose group). An increase in uterine sarcomas was observed in the low dose group of female rats but not in the high dose group (1 control rat, 5 in the low-dose group, 0 in the high-dose group). Decreases in the incidence of fibromas and fibroadenomas of the mammary gland (40 in the control group, 27 in the low-dose group, 16 in the high-dose group), pituitary gland adenomas (22 in the control group, 16 in the low-dose group, and 13 in the high-dose group), and adrenal gland pheochromocytomas (18 in the control group, 11 in the low-dose group, and 10 in the high-dose group) in treated female rats may have been due to increased early mortality in the dosed groups.
Groups of 50 male and 50 female CD-1 mice, 6 to 7 weeks of age, were exposed to filtered air containing 0 (control), 400 ppm, 3000 ppm, or 8000 ppm MTBE (99% purity) [0, 1,440, 10,800, or 28,800 mg/m3], 6 hours/day, 5 days/week, for 18 months. Mortality was increased, and mean survival time was decreased only in the high-exposure group of male mice. Mean survival time was 510 days for control male mice and 438 days for high-exposure male mice. An increased incidence of hepatocellular carcinomas was observed in the 8000 ppm exposure group of males (2/49 control mice, 4/50 in the low- exposure group, 3/50 in the mid-exposure group, and 8/49 in the high-exposure group). An increased incidence of hepatocellular adenomas and carcinomas in the high-exposure group of males compared with the incidence in the control group was not significant by the Fisher's Exact Test (12 control mice, 12 mice in the low-exposure group, 12 in the mid-exposure group, and 16 in the high-exposure group); however, this method of analysis does not adjust for differences in survival between the control and exposure groups. In female mice, exposure to 8000 ppm MTBE produced a significant increase in the incidence of hepatocellular adenomas and carcinomas (2/50 in the control group, 2/50 in the low-exposure group, 2/50 in the mid-exposure group, and 11/50 in the high-exposure group). Liver tumors were predominantly adenomas but did include one hepatocellular carcinoma in the 400 ppm exposure group and one in the 8000 ppm exposure group. A significant decrease in the incidence of hepatocellular degeneration was observed in the high exposure group of female mice (34 in the control group, 24 in the low-exposure group, 25 in the mid-exposure group, and 17 in the high-exposure group). Because the duration of this study was only 18 months, it is not possible to determine whether MTBE induces late-developing tumors in the low- or mid-exposure groups or at other sites in mice, or whether more hepatocellular adenomas would have progressed to carcinomas. The National Toxicology Program (NTP) uses 2-year treatment periods to evaluate the chronic toxicity and carcinogenic potential of chemicals in mice and rats.
Groups of 50 male and 50 female F344 rats, 6 to 7 weeks of age, were exposed tofiltered air containing 0 (control), 400 ppm, 3000 ppm, or 8000 ppm MTBE (99% purity), 6 hours/day, 5 days/week, for 2 years. Because of the increased mortality of males exposed to 3000 and 8000 ppm, surviving rats in these exposure groups were killed after 97 weeks and 82 weeks of exposure, respectively. Mean survival time of male rats was significantly reduced in all of the exposure groups compared with that of rats in the control group (632 days for control rats, 617 days for the low-exposure group, 587 days for the mid-exposure group, and 516 days for the high-exposure group). Body weights of males and females in the 8000 ppm exposure group were less than those of control rats. Exposure-related increases in the severity of chronic nephropathy were observed in all exposure groups of males and in the 3000 and 8000 ppm exposure groups of females. Nephropathy is an age-related disease characterized by degeneration and atrophy of the tubular epithelium, dilation of tubules with the formation of hyalin and granular casts, regeneration of tubular epithelium, glomerulosclerosis, and interstitial inflammation and fibrosis. In a previous 13-week inhalation study of MTBE in F344 rats at exposures of 0, 800, 4000, or 8000 ppm (Dodd and Kintigh, 1989), an increase in hyaline droplet formation was observed in the renal proximal tubules of male rats in the high-exposure group; however, this change was not accompanied by other lesions characteristic of alpha-2u-globulin nephropathy, including degeneration and regeneration of the renal tubular epithelium and accumulation of granular casts within tubular lumens.
Chronic progressive nephropathy was considered to be a main cause of the early deaths of males in the 3000 and 8000 ppm exposure groups. Immunohistochemical staining of kidney sections for alpha-2u-globulin in F344 rats exposed to 0 (control), 800, 4000, or 8000 ppm MTBE for 13 weeks did not provide evidence of an exposure-related increase in the intensity or staining area of this protein (percentage of cortex staining for alpha-2u-globulin: 24% control, 48% low-exposure, 36% mid-exposure, and 44% high exposure) (Swenberg and Dietrich, 1991). Furthermore, proteinaceous casts localized at the junction of the proximal tubules and the thin loop of Henle did not stain positively for alpha-2u-globulin. These results indicate that accumulation of alpha-2u-globulin was not causally related to the elevated mortality in male rats resulting from increased severity of chronic nephropathy in the MTBE exposure groups. Alpha-2u-globulin is a low molecular weight protein synthesized in the liver of male rats but not by hepatocytes of female rats or by mice of either sex. The finding that the severity of chronic nephropathy was also increased in female rats exposed to MTBE does not support the hypothesis that alpha-2u-globulin alone accounted for the kidney lesions observed in these studies.
Incidences of renal tubule adenomas and carcinomas were increased in exposed males (one control rat, none in the low-exposure group, eight in the mid-exposure group, and three in the high-exposure group). Three of the eight kidney tumors in the 3000 ppm exposure group were renal tubule carcinomas. No carcinomas were observed in the other treatment groups or in the control group. In addition, preneoplastic adenomatous tubular cell hyperplasia was diagnosed in the kidney of two male rats in the low-exposure group. The latter change represents an early stage in the continuum of renal lesions that can progress from atypical hyperplasia to adenoma to carcinoma (Hard, 1986). One renal tubule adenoma was detected in the mid-exposure group of females. An exposure-related increase in incidence of interstitial cell adenomas of the testes was also observed (32/50 in the control group, 35/50 in the low-exposure group, 41/50 in the mid-exposure group, and 47/50 in the high-exposure group). These tumors occur at a high spontaneous rate in this strain of rat. The level of significance of testicular tumors should be reanalyzed using statistical methods that adjust for differences in survival between the control and exposure groups.
At the 15-month evaluation, dose-related increases in kidney weights were observedin males and females. Nephropathy was present in all control and exposed rats but was generally more severe in the treatment groups. A renal tubule adenoma was observed in one male rat from the high-dose group. At the end of the 2-year study, survival was significantly lower for the males in high-dose group (1/50) and females in the high-dose group (12/50) than for rats in the control groups (10/50 for males and 28/50 for females). Incidences of focal renal tubule hyperplasia and renal tubule adenoma were increased in exposed males. Hyperplasia was found in 3 of 50 controls rats, 7 of 50 in the low-dose group, 6 of 50 in the mid-dose group, and 6 of 50 in the high-dose group. Renal tubule adenomas were found in one control male rat, in three in the low-dose group, in four in the mid-dose group, and in three in the high-dose group.
In the high-dose group, a renal tubule carcinoma was observed in a male rat that also had a renal tubule adenoma. The severity of nephropathy was increased significantly in males in the high-dose group and in all of the exposure groups of females. Renal tubule hyperplasia was seen in one female rat in the high dose group. Because of the observed increase in uncommon proliferative lesions of the renal tubules in exposed males, six to eight additional sections per kidney from all control and treated males were prepared and examined microscopically. The extended evaluation revealed significant increases in the incidence of renal tubule hyperplasia in the high-dose group and of renal tubule adenoma in the mid-dose group. Renal tubule hyperplasia for the standard and extended evaluations combined was found in 14 control rats, 20 rats in the low-dose group, 17 in the mid-dose group, and 25 in the high-dose group; renal tubule adenomas or carcinomas were found in 8 control rats, 13 rats in the low-dose group, 19 rats in the mid-dose group, and 13 rats in the high-dose group. Four renal tubule carcinomas were seen in this study; two were found in the low-dose group, one in the mid-dose group, and one in the high-dose group.
In the preliminary 13-week drinking water study of TBA in rats (at concentrations up to 40 mg/ml), a treatment-related increase in hyaline droplet accumulation was observed in the renal proximal tubules of male rats but not in female rats. In addition, an increase in renal tubular cell replication was observed in male rats, but only at exposures that exceeded the carcinogenic doses used in the 2-year study of TBA (Takahashi et al., 1993).
The incidences of follicular cell hyperplasia and adenoma of the thyroid gland were increased in treated males. Hyperplasia was found in 5 of 60 control male mice, in 18 of 59 the low-dose group, in 15 of 59 in the mid-dose group, and in 18 of 57 in the high-dose group. Adenomas were found in 1 control male mouse, in no mice in the low dose group, in 4 of 59 in the mid-dose group, and in 1 in the high-dose group. In treated females, hyperplasia was seen in 19 of 58 control mice, 28 of 60 in the low dose-group, 33 of 59 in the mid-dose group, and 47 of 59 in the high-dose group. Adenomas were found in two control female mice, three mice in the low-dose group, two in the mid-dose group, and nine in the high-dose group. In addition, a follicular cell carcinoma was present in one male mouse in the high-dose group. The tumor response in female mice in the high-dose group was statistically significant.
Incidences of chronic inflammation and transitional epithelial hyperplasia of theurinary bladder were increased in males and females in the high-dose groups compared with the control groups. There was no evidence of progression of these lesions to urinary bladder neoplasia.
Squamous cell carcinomas of the nasal cavity were induced in male and female F344 rats exposed to 14.3 ppm formaldehyde for 24 months (in males: 0/118 control rats, 0/118 at 2.0 ppm, 1/119 at 5.6 ppm, and 51/117 at 14.3 ppm; in females: 0/114 control rats, 0/118 at 2.0 ppm, 0/116 at 5.6 ppm, and 52/115 at 14.3 ppm) (Kerns et al., 1983). Five additional nasal cavity tumors were observed in rats exposed to 14.3 ppm. In a lifetime inhalation study, exposure of 100 male Sprague-Dawley rats to 14.3 ppm formaldehyde (6 hours/day, 5 days/week) produced 38 squamous cell carcinomas of the nasal cavity and 10 polyps or papillomas; neither of these lesions was observed in control rats (Sellakumar et al., 1985). There were no increases in tumor incidence beyond the respiratory tract. Exposure of Wistar rats to 10 ppm formaldehyde for 28 months (6 hours/day, 5 days/week) produced an increased incidence of nasal tumors in rats in which the nasal mucosa had been severely injured by bilateral intranasal electrocoagulation (17/58 versus 1/54 in control rats with damaged noses); in rats with undamaged noses, the incidence of nasal tumors was 1/26 in rats exposed to 10 ppm formaldehyde and 0/26 in control rats (Woutersen et al., 1989).
Exposure of male Syrian golden hamsters to 10 ppm formaldehyde (5 hours/day, 5 days/week for life) or to 30 ppm formaldehyde (for 5 hours once per week for life) did not induce tumors of the nasal cavity or respiratory tract (Dalbey, 1982).
On the basis of the kidney tumor data from the 2-year inhalation study of MTBE in male F344 rats (Chun et al., 1992), the estimated upper-bound unit cancer risk for lifetime inhalation exposure to MTBE using the linearized multistage model is 6 x 10-4 per ppm (2 x 10-7 per ug/m3) and the human cancer ED10 benchmark dose is 330 ppm. Because of reduced survival in the male rat exposure groups, upper bound unit cancer risk estimates based on the kidney tumor data were calculated using survival-adjusted tumor rates. Even with this adjustment, the highest exposure group (8000 ppm) had to be excluded from the analysis because it failed a goodness-of-fit test. Eliminating this group from the analysis is acceptable for low dose risk estimations because the mortality was greatest in this exposure group, the tumor response in this group was not significant (probably because of early mortality), and the mid exposure group (3000 ppm) had a significant tumor response. Preliminary estimates of cancer potency had been made using a Weibull time-to-tumor model to account for differences in survival. However, because predictions of tumor incidence from this model were inconsistent with the experimental data from the lowest exposure group, use of this model was discontinued in favor of the linearized multistage model.
Based on the lymphoma/leukemia data from the 2-year oral exposure study (Belpoggi et al., 1995), the estimated upper-bound unit cancer risk for lifetime oral exposure to MTBE using the linearized multistage model is 4 x 10-3 per mg/kg body weight/day and the human cancer ED10 benchmark dose is 38 mg/kg body weight/day.
The inhalation upper bound cancer unit risks of MTBE are slightly lower than those of fully vaporized conventional gasoline which has been listed by EPA as a probable human carcinogen based on animal carcinogenicity data. In both cases (MTBE and fully vaporized conventional gasoline), estimates of unit cancer risk were made using the linearized multistage model and based on the induction of liver tumors in mice and kidney tumors in rats from inhalation exposures. If the lymphoma/leukemia data from the gavage study are used to estimate unit cancer risk for inhalation exposure (i.e., assuming equivalent total exposures by inhalation and gavage result in similar internal doses of MTBE and that differences in dose rate do not affect the tumor response), then the estimated upper bound unit cancer risk (4 x 10-3 per ppm) is similar to that of fully vaporized conventional gasoline. It should be noted, however, that human exposures to nonoxygenated gasoline vapors may be very different than exposures to fully vaporized conventional gasoline. The estimated upper bound unit cancer risk for MTBE using the lymphoma/leukemia data is approximately an order of magnitude lower than that of benzene, a constituent of gasoline that is classified as a known human carcinogen. The unit cancer risk of MTBE was more than 100 times less than that of 1,3-butadiene, a carcinogenic emission product of incomplete fuel combustion.
The experimental studies provide sufficient evidence for the carcinogenicity of MTBE in animals. There are no studies on the carcinogenicity of MTBE in humans. Using IARC criteria for evaluating evidence of carcinogenicity, MTBE could be classified as possibly carcinogenic to humans (Group 2B); using EPA criteria of 1986, MTBE could be classified as a probable human carcinogen (Group B2). Other data (e.g., tissue dosimetry of formaldehyde resulting from exposure to MTBE) could influence the overall evaluation. Epidemiologic studies suggest a causal relationship between exposure to formaldehyde and nasopharyngeal cancer (IARC, 1995). Because all known human carcinogens that have been adequately tested produced positive carcinogenic results in experimental animals, it is plausible and prudent to regard agents that have sufficent evidence of carcinogenicity in animals as presenting a carcinogenic risk to humans (IARC, 1995). The existing data on MTBE indicate that this chemical should be regarded as posing a potential carcinogenic risk to humans, while recognizing that the estimated upper bound cancer unit risks of MTBE are similar to or slightly less than those of fully vaporized conventional gasoline.
No studies have been reported on the carcinogenicity of ETBE, TAME, or TBF.
|Upper bound unit|
|Mouse||liver||inhalation||6 x 10-4 per ppmb|
2 x 10-7 per ug/m3
|Rat||kidney||inhalation||6 x 10-4 per ppmb|
2 x 10-7 per ug/m3
|oral||4 x 10-3 per mg/kg/d||38 mg/kg/day|
b For comparison, estimated upper bound unit cancer risks for fully vaporized conventional gasoline are noted: 2 x 10-3 per ppm based on induction of liver tumors in mice and 4 x 10-3 per ppm based on induction of kidney tumors in rats (EPA, 1987).
A mouse study conducted at the Bushy Run Research Center (Bushy Run Research Center, Project Report No. 52-526, 1989; Neeper-Bradley, 1990) involved higher levels of MTBE than those in the Conaway et al. study. Pregnant CD1 females were exposed via inhalation to target-dose levels of 0, 1000, 4000, and 8000 ppm on Gds 6-15. Food and water consumption and maternal weights were recorded throughout dosing. Maternal toxicity was observed at 4000 and 8000 ppm with decreased food consumption, reductions in body weight and weight gain, labored respiration, and lacrimation with other signs of toxicity observed at 8000 ppm and hypoactivity and ataxia observed at 4000 and 8000 ppm. The uterine contents of females killed on Gd 18 were evaluated for number of implantation sites, early and late resorptions, and dead and live fetuses. Ovarian corpus luteum counts were also made. All fetuses were weighed and examined for external malformations. Approximately one-half of the live fetuses were examined for thoracic-abdominal and cranial malformations using modified methods of Staples or Wilson, respectively. All fetuses were processed for skeletal staining with Alizarin Red. Fetal weight changes and skeletal variations were observed at 4000 and 8000 ppm, but they were attributed to maternal toxicities. An increased incidence of cleft palate was observed at 8000 ppm, and the possibility that this condition might also be related to maternal toxicity was discussed. The number of viable implants was signific antly decreased only at 8000 ppm. This decrease was due to an increase in the number of late resorptions and dead fetuses rather than to early resorptions.
Approximately one-half of the live fetuses were examined for thoracic-abdominal and cranial malformations using modified methods of Staples or Wilson, respectively. All fetuses were processed for skeletal staining with Alizarin Red. There were no significant changes in the number of corpus luteums, implants, resorptions, and dead or live fetuses, or in the incidence of malformations.
Biles et al. (1987) report a single-generation inhalation study of MTBE in rats. Exposures were at 0, 300, 1300, and 3400 ppm, 6 hours/day, 5 days/week. Males were exposed for 12 weeks, females for 3 weeks, and then mated while being exposed. These FO rates generated two litters: both were nursed until 21 days after birth, then killed, and the pups examined. MTBE exposure did not change any endpoint related to fertility or mating in the FO rats. There was a 4% reduction in pup viability in the second litter in the top two dose groups (reduced from a control value of 99% to 95% for both groups); this reduction is not considered to be of biological significance. Pup post-mortem evaluations found no differences between groups. There were no body weight differences between the groups of FO rats, a finding which suggesting that a maximum tolerated dose was not reached. No differences between the groups of FO rats were noticed at necropsy or microscopically or at necropsy. No sperm measures were taken, nor was vaginal cyclicity evaluated. In sum, this study found no change in body weight, and no effects in any reproductive or fertility endpoint were measured.
For MTBE, there are no studies of long-term human exposure, a fact which is not surprising since the use of MTBE is fairly recent. There are experimental studies in which animals were exposed to high MTBE doses in order to optimize the ability to detect a carcinogenic response. For inhalation exposure, both sexes of rats and mice were tested, and kidney tumors and testicular tumors were seen in male rats and liver tumors in male and female mice. For gavage exposure, MTBE caused leukemia or lymphoma in female rats and testicular tumors in male rats. Two metabolites of MTBE, TBA and formaldehyde, have also been tested in animals and show carcinogenic activity, with some responses paralleling those seen with MTBE and some responses being at different sites. The manner in which MTBE causes tumors to arise is not known. Increased tumor incidences were not observed at low doses, possibly because of the small group sizes used in the experiments. Such results are not sufficient to rule out a possible low dose hazard, given our lack of understanding about how MTBE causes tumors in animals.
The positive animal studies raise the concern that MTBE may pose a human health hazard. In the absence of good epidemiologic evidence or knowledge about how the agent likely reacts in humans, the positive animal evidence (in two species by two routes of exposure) and other supporting factors provide evidence that MTBE could pose a human health hazard under some conditions of exposure. Currently, a low-dose hazard can not be ruled out. Based on animal data, MTBE is considered to be either a possible or probable human carcinogen.
It may be useful to ask the "what if" question, i.e., if MTBE is a human carcinogen, what might be the impact on an exposed human population? While we have only established that MTBE has potential to be a human hazard, quantitative human risk estimates can be developed if we make the assumption that it is a human carcinogen. Public health conservative assumptions are used to formulate these risk estimates, which are by design almost always high enough so as to not likely underestimate the true risk. In fact, the true risks, which we are rarely able to identify, may be lower than the risk estimates and could even be negligible.
Estimations of cancer risk for populations exposed to MTBE are based on estimates of lifetime human exposure, on extrapolation models, and on several estimates of cancer potency for this chemical. Human exposures are estimated from measurements of MTBE in ambient air, from measurements of MTBE concentrations in specific microenvironments (e.g., inside automobiles), and from estimates of the distribution of time spent in each environment. A reasonable worst case exposure scenario for a typical motorist in the general population that uses oxygenated fuels containing MTBE might result in annual time-weighted average exposures of 0.014 ppm MTBE, based on a 4-month oxygenated fuel season, 0.019 ppm based on a 6-month oxygenated fuel season, or 0.029 ppm based on 6 months of use of oxygenated gasoline and 6 months of use of reformulated gasoline. These exposure scenarios do not necessarily represent the worst possible case, but are expected to be well above the average exposure for the entire population of MTBE-exposed motorists. The largest occupational group with significant potential exposure to MTBE is service station attendants; their estimated time-weighted average exposure to MTBE, including an 8-hour time-weighted average occupational exposure (5 days per week for 40 years working lifetime) of 0.6 ppm during the 6-month oxygenated fuel season and 0.44 ppm during the 6-month reformulated gasoline season, is 0.10 ppm.
Cancer potency estimates for MTBE were made based on mouse liver tumor data from inhalation exposure, rat kidney tumor data from inhalation exposure, and lymphoma/leukemia data from oral exposure (Table 5). Alpha-2u-globulin nephropathy was not considered to be applicable to the kidney tumor response in male rats because:
(ii) Immunohistochemical staining of kidney sections for alpha-2u-globulin in male rats exposed to MTBE for 13 weeks at concentrations up to 8000 ppm did not show an exposure related increase in staining of this protein. Staining was equivalent at the 400, 3000, and 8000 ppm exposures, yet only the two higher concentrations produced kidney tumors and increased the rate of early mortality.
(iii) Exposure to MTBE or its metabolite TBA increased the severity of chronic nephropathy in both male and female rats, yet alpha-2u-globulin is not synthesized in the liver of female rats. Furthermore, kidney tumors were induced in the mid-dose group of male rats exposed to TBA, but the severity of nephropathy in that group was not different from the severity of nephropathy in control male rats. Thus, other factors must be involved in MTBE- or TBA-induced nephropathy and renal carcinogenicity in male rats.
Estimations of human cancer risk associated with lifetime inhalation exposure (70 years) to MTBE are presented in Table 6, using reasonable worst-case exposure scenarios, estimates of upper-bound unit cancer risk, and maximum likelihood estimates (MLE) of excess cancer risk for specific exposures. For lifetime environmental inhalation exposure to MTBE from use of oxygenated gasoline and reformulated gasoline containing this ether (based on a reasonable worst-case exposure scenario of 0.029 ppm), the upper bound excess cancer risk is 2x10-5 using the mouse liver tumor data or the rat kidney tumor data, and 1x10-4 using the rat lymphoma/leukemia data. The MLEs of excess cancer risk for lifetime environmental exposure to this concentration of MTBE are 5x10-15 using the mouse liver tumor data, 8x10-10 using the rat kidney tumor data, and 8x10-5 using the rat lymphoma/leukemia data. For a working lifetime plus environmental exposure to MTBE among service station attendants, the estimated upper bound excess cancer risks were 6x10-5 using the mouse liver tumor data or the rat kidney tumor data, and 5x10-4 using the rat lymphoma/leukemia data. The MLEs of excess cancer risk for service station attendants were 2x10-13 using the mouse liver tumor data, 1x10-8 using the rat kidney tumor data, and 3x10-4 using the rat lymphoma/leukemia data. The MLE values are highly sensitive to small changes in the experimental tumor incidence data. For example, if the renal tubule adenoma in the control male rat had been present instead in the low exposure group, the MLE values based on the rat kidney tumor data would have been more than 3-4 orders of magnitude lower than the MLE values shown in Table 6 (see Appendix A). In this example, the estimated upper-bound excess cancer risk values change by less than 50%. Because of the instability in the MLEs at low exposures, they are not considered reliable for estimation of human risk. While the true risk from low exposures is unknown, the 95% upper confidence limit provides a stable and plausible upper bounds on the excess risk.
Sources of uncertainty in these risk assessments include potential differences in sensitivity between laboratory animals and humans, interindividual differences in sensitivity among the exposed human population, the relationship between total dose versus dose rate on the tumor response, the adequacy of the exposure characterizations (especially with respect to the distribution of exposures in the environment and the workplace and the influence of changes in climate), and the adequacy of the models that were used to perform low-dose extrapolations and estimate cancer potency. Research that addresses these uncertainties could either raise or lower the current estimates of risk.
In estimating excess cancer risks due to exposure to MTBE, several assumptions were made, including the following:
(ii) Humans and laboratory animals share similar sensitivities to the carcinogenic effects of MTBE at equivalent doses.
(iii) Tumor data from a gavage study of MTBE can be used to estimate unit cancer risk for inhalation exposure to MTBE because equivalent total exposures by inhalation and gavage result in similar internal doses of MTBE.
(iv) Differences in dose rate and metabolism from gavage and inhalation exposures do not affect the tumor response.
(v) MTBE-induced nephropathy does not affect the kidney tumor dose-response.
(vi) At low doses, the carcinogenic effects of MTBE are proportional to time-weighted average lifetime daily exposures in both experimental animals and in humans.
(vii) The models used are appropriate for low-dose extrapolations of the carcinogenic effects of MTBE.
(viii) The carcinogenic effects of MTBE occur independently of the effects of other hazardous agents to which humans may be exposed. The effect of addition of MTBE to gasoline on the air concentrations of other chemical carcinogens that are released as evaporative or combustion emissions (e.g., benzene, 1,3-butadiene, and formaldehyde) is beyond the scope of this assessment.
Thus, depending on the validity of the assumptions used in making these estimates, the actual cancer risks may be lower than those given in this document and could even be nearly zero. It is not known whether the cancer risk of oxygenated gasoline containing MTBE is significantly different from the cancer risk of conventional gasoline. The estimated upper bound cancer unit risks of MTBE are similar to or slightly lower than those of fully vaporized conventional gasoline, which has been listed by EPA as a "probable human carcinogen" based on animal carcinogenicity data. However, because of a lack of health data on the nonoxygenated gasoline vapors to which humans are actually exposed, it is not possible to estimate the population cancer risk to conventional gasoline. The comparative risk among oxygenated and nonoxygenated gasoline types has not been established.
The data were generally inadequate to evaluate the health risks of oxygenates other than MTBE, a factor which makes other oxygenates and gasoline mixtures to which they are added all the more important to investigate further.
Table 6. Estimated Upper-Bound Excess Inhalation Cancer Risks and Maximum Likelihood Estimates (MLE) of Excess Cancer Risk Based on Reasonable Worst Case Time-weighted Lifetime (70 Years) Exposure Estimates to MTBE and on Cancer Potency Estimates Derived from Carcinogenicity Data for MTBE in Rats and Mice
|Mouse Liver Tumors||Rat Kidney Tumors||Rat Lymphomas/Leukemiaa|
|4-month oxyfuel season|
|0.014||8 x 10-6||5 x 10-16||9 x 10-6||2 x 10-10||7 x 10-5||4 x 10-5|
|6-month oxyfuel season|
|0.019||1 x 10-5||1 x 10-15||1 x 10-5||4 x 10-10||9 x 10-5||6 x 10-5|
|6-month oxyfuel and 6-month reformulated gasoline|
|0.029||2 x 10-5||5 x 10-15||2 x 10-5||8 x 10-10||1 x 10-4||8 x 10-5|
|service station attendantsb|
|0.10||6 x 10-5||2 x 10-13||6 x 10-5||1 x 10-8||5 x 10-4||3 x 10-4|
Note 1. The actual risks are likely to be somewhat lower than the
upper bound calculated risks and could even be nearly zero.
Note 2. The MLE values are highly sensitive to small changes in the tumor incidence data. For example, if the renal tubule adenoma in the control male rat had been present instead in the low exposure group, the MLE values based on the rat kidney tumor data would have been more than 3-4 orders of magnitude lower than the MLE values shown in this table (see Appendix A). In this example, the estimated upper-bound excess cancer risk values change by less than 50%. Because of their instability, the MLEs are not considered reliable.
a These cancer risk estimates are derived using a cancer potency
estimate from the gavage carcinogenicity study of MTBE in rats with the
assumption that the observed tumor response in gavage-treated animals
infers an inhalation cancer risk.
b Exposure for service station attendants include an 8-hour time-weighted average occupational exposure (5 days per week for 40 years working lifetime) of 0.6 ppm during the 6-month oxygenated fuel season and 0.44 ppm during the 6-month reformulated gasoline season.
In addition, a need exists to more carefully evaluate the carcinogenicity data on MTBE in terms of its relative potency, as well as data on subchronic health effects of MTBE and other oxygenates. It is expected that the more comprehensive review now being undertaken by the Health Effects Institute will be able to more thoroughly address these and other complex issues than was possible in this assessment.
Continued use of oxygenated gasoline will require a continuing evaluation of potential health effects of the oxygenates, their metabolites, and any emission or atmospheric degradation products. Most of the research to date has focused on MTBE, and comparatively little information exists for other oxygenates, such as ETBE or TAME, or some of the atmospheric degradation products, such as t-butyl formate. In addition, there is a differential availability of data by compound, with some endpoints measured for some compounds but not others.
Specifically, further information is needed in the following areas: 1) more data on human exposures; 2) more data on the pharmacokinetics of MTBE and its metabolites (TBA and formaldehyde) in humans and animals; 3) epidemiologic and experimental human-subject research on acute health symptoms related to oxygenates in gasoline mixtures; 4) increased research on the mechanisms of carcinogenicity and on the dose-response relationships between exposure to oxygenated fuels and cancer risk; and 5) better data on potential human exposures and health risks associated with the gasoline mixtures to which oxygenates are added.
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