Environ. Sci. Technol. 1991, 25,964-972
Driver Exposure to Volatile Organic Compounds, CO, Ozone, and NO, under Different Driving Conditions Chang-Chuan Chan,+ HalGk Ozkaynak,z John D. Spengler,* and Linda Sheldon5 Department of Environmental Health, Harvard School of Public Health, 665 Huntington Avenue, Boston, Massachusetts 02 1 15 W The in-vehicle concentrations of 24 gasoline-related volatile organic compounds (VOCs) and three criteria air pollutants, ozone, carbon monoxide, and nitrogen dioxide, were measured in the summer of 1988, in Raleigh, NC. Two four-door sedans of different ages were used to evaluate in-vehicle concentrations of these compounds under different driving conditions. Factors that could influence driver exposure, such as different traffic patterns, car model, vehicle ventilation conditions, and driving periods, were evaluated. Isopentane was the most abundant aliphatic hydrocarbon and toluene was the most abundant aromatic VOC measured inside the vehicles. In-vehicle VOC and CO concentrations were highest for the urban roadway, second highest for the interstate highway, and lowest for the rural road. The median concentration ratio of urban/interstate/rural for each VOC was about 10/6/1. No differences in in-vehicle VOC concentrations were found between morning and afternoon rush hour driving, but higher in-vehicle ozone and NOz concentrations were found during afternoon driving. In-vehicle VOC levels were lowest with the air conditioner on and highest when the vent was open with the fan on. The in-vehicle/car exterior concentraton ratio for VOCs, CO, and NOz was slightly higher than 1. The VOC concentration measured by a pedestrian on the urban sidewalk was lower than the in-vehicle measurements but higher than the fixed-site measurements on urban roadways 50 m from streets. The VOC measurements were positively correlated with the CO measurement and negatively correlated with the ozone measurement.
Introduction Emissions of volatile organic compounds (VOCs), particularly those designated as toxic air pollutants, have become an increasing concern to state and federal agencies. The recently passed amendments to the Clean Air Act greatly expand the list of compounds that will be regulated by statue. Exposures to many aromatic and halogenated VOCs are reported as higher inside homes than out by the EPA’s Total Exposure Assessment Methodology (TEAM) studies (1). Wallace, extrapolating from TEAM studies and applying cancer risk values to several VOCs, estimated a range of 1000-5000 additional cancer deaths per year in the United States (2). While concentrations encountered indoors for many VOCs are higher than concentrations in many other locations, in-vehicle concentrations may be quite important contributors. The TEAM data from California revealed that personal exposure to several VOCs such as benzene, m-lp-xylene, and ethylbenzene was strongly associated with vehicle use ( 3 ) . These four volatile organic compounds, however, are only a subset of the VOCs to which drivers are potentially exposed. About 30 vehicle-related VOCs, including alt Present address: National Taiwan University, College of Medicine, Graduate Institute of Public Health, No. 1Jen-Ai Rd., 1 st Sec., Taipei, Taiwan. 3 Enerev and Environmental Policv Center, John F. Kennedy School oFGovernrnent, Harvard University, 79 J.F.K. Street, Cambridge, MA 02138. Research Triangle Institute, Research Triangle Park, NC 27711.
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Environ. Sci. Technol., Vcl. 25, No. 5, 1991
kanes, alkenes, aldehydes, and aromatics, were commonly found in several different studies, such as tunnel studies, automobile emission tests, and roadway studies. Hampton et al. (4) identified over 300 hydrocarbons, 22 of which were quantified, in the Allegheny Mountain Tunnel, PA. Lonneman et al. (5) quantified 25 VOCs (C2-C,,) in the Lincoln Tunnel. Volatile organic compound emissions from 46 in-use passenger cars were measured by Sigsby et al. in the U.S. EPA’s emission test facilities a t Research Triangle Park, NC (6) by using three well-known driving tests: the federal test procedure (FTP),the crowded urban expressway (CUE), and the New York City cycle (NYCC). Eighty-two VOCs were identified for each testing cycle. The species of VOCs collected were consistently the same among all driving cycles and for most model years of the tested vehicles. Mass emission rates decreased with the later model year cars, and the ratio of hydrocarbons to nitrogen oxides increased dramatically on the lower speed cycles. Twenty-two VOCs were identified by GC/MS analysis of roadway samples collected by Zweidinger et al. (7) on Highway 70 near Raleigh, NC. None of these past studies measured in-vehicle VOCs directly. A study was designed to investigate the in-vehicle exposure to 24 gasoline-related VOCs, particularly benzene and 1,3-butadiene, and three criteria air pollutants, carbon monoxide, nitrogen dioxide and ozone. Factors that could influence the amount of exposure, such as different traffic patterns, car models, vehicle ventilation conditions, and driving periods, were tested. Background VOC concentrations were measured a t fixed-site locations to evaluate the relationships between fixed-site VOC levels and in-vehicle VOC levels. In addition, VOC levels on the sidewalks of urban streets were measured in order to evaluate the pedestrian’s exposure to VOCs. Measurements represent l-h segments of urban, interstate, and rural driving. The urban and interstate sampling was conducted during “rush-hour’’ conditions. Experimental Section Target Pollutants. Twenty-four VOCs, ranging from C4 to CI1, and three criteria pollutants, carbon monoxide, nitrogen dioxide, and ozone, were selected as the target pollutants. The VOCs measured were 16 aliphatic hydrocarbons, 1 olefinic hydrocarbon, and 7 aromatic hydrocarbons. These VOCs are components and/or combustion products of gasoline. Some of them, such as benzene and 1,3-butadiene, are suspected carcinogens. Vehicles Tested. The two automobiles used in this experiment were selected to represent newer (less mileage) and older vehicles. They were a 1987 Mercury Sable four-door sedan with 26856 mi and a 1983 Mercury Marquis four-door sedan with 62856 mi. Both cars were in good condition when leased from a rental car company. Prior to field sampling, physical examination and the federal test procedure (FTP) emission tests were performed on both cars in the mobile emissions test facilities of the U S . EPA a t Research Triangle Park, NC. Since both car bodies were physically intact, both cars were tested without any further repairs. The 1983 car had higher tailpipe emissions for total hydrocarbons (THC),
0013-936X/91/0925-0964$02.50/0
0 1991 American Chemical Society
Table I. Comparisons of Emissions Rates between FTP Emissions Tests and the Emissions Rates Calculated by the EPA MOBILE4 Emissions Model for Two Test Vehicles tailpipe emissns, g/mi THC CO NO,
FTP emissions 1983 Mercury Marquis 1987 Mercury Sable MOBILE4
0.61 0.26
8.19 4.79
2.50 0.97
0.68 0.42
7.93 4.80
0.88
emissions
1983,63 8QO mi 1987, 26400 mi
0.74
CO, and NO, than the 1987 car. However, the 1983 car had lower hot soak evaporative emissions for THC (1.97 g/mi) than the 1987 car (2.52 g/mi). Compared to the fleet average tailpipe emissions for respective model year calculated from the EPA MOBILE^, the 1983 car had lower tailpipe THC emissions but higher CO and NO, emissions than the model’s estimated emissions. The 1987 car had lower tailpipe ‘I’HC emissions but higher NO, emissions than the model’s estimated emissions. The 1987 car’s tailpipe CO emissions, however, were equal to the model’s estimated emissions. These comparisons are listed in Table I. Fuel Used. A single tank of Amoco summer-grade gasoline (RVP = 10.5, octane index 87.8) was used as the fuel source for all of the test drives. The use of fuel from a single source for the entire experiment eliminated the confounding effects that different fuel constituents could have on in-vehicle VOC species and concentrations. Since the gasoline vapor pressure will affect running loss emissions, the results of this study may not be applied to other areas where gasoline RVPs are under tighter regulations. Roadways Chosen. Three driving routes around Raleigh, NC, were chosen to represent three distinct traffic patterns. A 1.8-mi route in downtown Raleigh was selected to represent urban traffic, which is characterized by heavy traffic volume, slow speed (25-35 mi/h), and frequent stops (7-8 stops/h). A 12.8-mi section of beltway (Interstate 40) between two ramps near north Raleigh was chosen to represent interstate highway traffic, which is characterized by moderate traffic volume and high speed (50-60 mi/h). A 17-mi loop of paved country roads between Research Triangle Park, Apex, and Jordan Lake, NC, was selected to represent rural driving, which is characterized by little traffic volume and moderate speed (35-45 mi/h). Ventilation. Three ventilation conditions were tested for each vehicle in the study: (1)windows and vent closed, and air conditioning on; (2) windows closed, vent fan on, and air conditioning off; (3) front windows half-opened, vent on, fan on, and air conditioning off. Sampling Time. Samples associated with urban and interstate driving were collected during morning (7:15-8:15 a.m.) and evening (4:30-5:30 p.m.) rush hours. Samples associated with rural driving were collected during the middle of the day (1O:OO a.m. to 2:OO p.m.). All samples were collected over a 1-h period. Sample Types. Samples were collected from four locations: (1)In-vehicle samples were collected from the driver’s breathing zones. (2) Fixed-site samples were collected at a location near the midpoint of each sampling route. These sites were located about 100-300 ft away from the roadways. (3) Car exterior samples were collected by extending a Teflon tube outside the vehicle. The sampling inlet was placed in the middle of the car roof. (4) Sidewalk samples were collected at a pedestrian’s breathing zone while he walked on the sidewalks around the urban route driven by the test cars.
Sampling Schedule. Field sampling was carried out from August 11to September 20,1988. There were nine sampling days for both the urban route and the interstate route, with one morning trip and one evening trip on each sampling day. The two test cars were driven one directly behind the other during each trip. Six rural sampling trips were spread randomly over the 18 sampling days. Another three rural samples were collected on a single day. One fixed-site VOC sample was collected concurrently with each in-vehicle sampling trip. Altogether, 10 VOC samples from the car exterior and 6 pedestrian sidewalk samples were collected. For the 1987 car, both in-vehicle and car exterior CO and NO2 were measured. For the 1985 car, no in-vehicle NO2 was collected. The in-vehicle ozone levels were only measured in one car per trip. The state-run ambient air monitoring data were used to estimate the background CO and ozone concentrations. Sample Collection and Analysis. Six-liter Summa polished stainless steel canisters were used to collect VOC samples. A restrictive orifice was designed to provide a relatively constant flow of 70 mL/min over a 60-min Sampling period, i.e., -4 L of total sample volume. The analyses of VOCs from the canister samples were performed by using a high-resolution gas chromatography/ mass spectrometry (GC/MS) technique (8). The major components of the system were a custom-built cryofocusing canister interface system, a capillary gas chromatographic system (30 m X 0.32 mm DB-624 fused-silica capillary column), a quadruple mass spectrometer (HP5995A), and a data processing system (HP59970C). The quantification of target VOCs was accomplished by using chromatographic peak areas derived from extracted ion profiles. Four Interscan Model 4146 CO monitors and three Interscan Model 1152 NO2 monitors made of electrochemical cells were used to measure CO and NO2. A portable chemiluminescent ozone analyzer (AID Model 560 ozone analyzer) was used to measure in-vehicle ozone. The electrical signals of these samplers were continuously transmitted to Rustrak Ranger data loggers and subsequently analyzed with a personal computer. Quality Assurance and Quality Control. Ten sets of quality control samples were prepared to evaluate the accuracy, precision, and method detection limits of the sampling and analytical system for the target VOCs. Each set contained an unspiked canister to serve as a blank and a spiked canister as a control. Most target VOCs had recovery efficiencies greater than 85%. Ten pairs of canisters were collected and analyzed to assess the precision of the method. The relative mean deviation of these duplicates was within 3% for most VOCs. The analytical system audited by EPA showed very good accuracy (97-103%) and precision (SD = 15%). The CO, ozone, and NOz monitors were calibrated daily by relevant EPA reference methods at the Research Triangle Institute (RTI).
Results and Discussion Table I1 summarizes the valid results for 24 VOCs and three criteria pollutants, for both vehicles in the three driving locations, measured at four sampling locations: in-vehicle, fixed-site, car exterior, and sidewalk. In this table, the VOC concentrations of measurements below the method detection limits were set equal to of method detection limits of respective VOCs. There were substantial differences in the concentrations of these target pollutants among these four sampling locations. VOC and CO concentrations were higher in vehicles than at fixed sites. The fixed-site ozone level was higher than the invehicle level. The in-vehicle ozone levels, however, were Environ. Sci. Technol., Vol. 25, No. 5, 1991 965
Table 11. Concentrations of VOCs and Criteria Pollutants by Sampling Location in Raleigh, NC, 198P concn, pg/m3 chemicals
min
isobutane n-butane 1,3-butadiene isopentane n-pentane 2-methylpentane hexane cyclohexane benzene 2,2,4-trimethylpentane n-heptane 2,3,3-trimethylpentane 2,3-dimethylhexane to 1u en e n-octane 2,3-dimethylheptane ethylbenzene mlp-xylene n-nonane o-xylene 1,3,5-trimethylbenzene n-decane 1,2,4-trimethylbenzene n-undecane carbon monoxide, ppm nitrogen dioxide, ppb ozone, ppb
0.2 3.9 1.2 4.9 2.2 1.8 0.9 0.1 0.8 1.1 0.4 0.3 0.1 3.7 0.2 0.2 0.8 2.2 0.2 1.0 0.4 0.2 1.8 0.2 1.0 8.0 2.0
isobutane n-butane 1,3-butadiene isopentane n-pentane 2-methylpentane hexane cyclohexane benzene 2,2,4-trimethylpentane n-heptane 2,sJ-trimethylpentane 2,3-dimethylhexane toluene n-octane 2,3-dimethylheptane ethylbenzene mlp-xylene n-nonane o-xylene 1,3,5-trimethylbenzene n-decane 1,2,4-trimethylbenzene n-undecane carbon monoxide, ppm ozone, ppb
0.2 1.5 1.2 1.1 0.3 0.5 0.1 0.1 0.0 0.1 0.4 0.3 0.1 0.2 0.2 0.1 0.3 0.2 0.2 0.2 0.2 0.5 0.2 1.7 16.0
isobutane n-butane 1,3-butadiene isopentane n-pentane 2-methylpentane hexane cyclohexane benzene 2,2,4-trimethylpentane n-heptane 2,3,3-trimethylpentane 2,3-dimethylhexane 1o1u en e n-octane 2,3-dimethylheptane ethylbenzene
0.2 14.0 1.2 13.8 10.6 7.0 4.4 0.1 6.8 7.7 2.3 2.6 0.1 24.5 1.0 0.2 5.2
max
mean
SD
281.7 588.4 17.2 510.8 682.3 417.4 154.5 12.0 42.8 95.3 25.1 21.4 7.8 118.9 5.0 0.9 21.8 76.3 4.8 27.6 12.0 6.7 39.0 7.2 32.0 196.0 86.0
9.1 54.3 3.3 68.7 30.8 25.0 13.6 1.2 11.6 17.0 1.3 5.3 2.0 46.5 1.9 0.3 8.8 30.5 0.6 11.4 4.7 1.5 15.6 1.9 11.3 87.3 15.4
32.9 84.7 2.4 68.8 77.1 47.4 18.5 1.4 6.9 13.1 3.3 3.4 1.4 27.3 1.0 0.2 4.9 17.2 1.0 6.2 2.7 1.2 8.7 1.4 5.1 42.6 18.0
3.7 14.8 1.2 16.4 6.5 5.1 3.0 0.3 2.4 3.7 1.3 1.3 0.4 11.5 0.6 0.2 2.1 6.8 0.2 2.7 1.0 0.5 3.6 0.5 3.4 70.0
43.4 186.8 1.2 170.1 51.9 22.1 27.9 1.3 8.5 13.5 4.0 4.2 1.6 46.8 2.1 0.2 8.2 27.8 1.3 10.4 4.4 2.0 13.7 2.3 5.5 123.0
3.9 16.7 1.2 17.6 6.1 4.5 3.3 0.2 1.9 2.7 0.9 0.9 0.3 8.2 0.4 0.2 1.5 5.0 0.3 1.9 0.9 0.4 2.7 0.4 2.9 52.8
8.3 34.1 0.0 30.9 8.8 5.3 5.2 0.3 1.6 2.6 0.7 0.8 0.3 8.1 0.4 0.0 1.5 5.0 0.3 1.9 0.8 0.4 2.5 0.4 1.0 28.1
(C) Car Exterior Measurementsbte 0.2 0.2 10.9 26.0 28.9 47.2 1.2 2.6 3.1 46.0 49.7 63.2 14.8 17.4 22.8 13.4 17.1 22.0 8.5 11.5 7.8 1.0 1.0 1.3 9.6 11.1 14.9 13.1 15.1 16.5 3.1 4.1 4.7 4.1 4.8 5.6 2.2 1.3 2.0 40.1 49.0 58.8 1.6 2.0 2.4 0.2 0.2 0.2 7.5 9.8 11.5
56.0 362.9 6.9 417.3 122.5 82.1 31.7 3.3 17.2 28.0 7.8 7.4 2.9 67.6 2.7 0.5 12.1
8.5 64.5 2.9 83.3 28.2 22.7 11.3
17.3 105.5 1.9 118.6 33.4 21.4 7.7 0.8 3.5 5.7 1.5 1.5 0.8 13.3 0.6 0.1 2.4
25%
median
75%
(A) In-Vehicle Measurements*,' 0.2 0.5 9.9 25.8 36.1 55.6 1.2 2.9 4.5 39.2 52.6 78.2 15.0 19.4 27.7 12.7 18.3 22.6 6.5 9.2 14.5 0.7 1.0 1.3 8.3 10.7 14.4 10.4 14.6 18.6 2.7 3.8 4.9 3.3 4.7 6.2 1.3 1.8 2.4 30.8 43.1 55.4 1.3 1.7 2.3 0.2 0.2 0.4 5.9 10.9 8.3 19.6 29.4 37.8 0.2 0.2 0.2 7.5 11.0 14.2 2.9 4.5 5.7 0.9 1.2 1.9 9.2 19.7 14.5 2.4 1.0 1.6 9.0 11.0 14.0 61.0 81.0 119.0 3.0 11.0 16.0
(B) Fixed-Site Measurementsb*d
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Environ. Sci. Technol., Vol. 25, No. 5, 1991
0.7 3.6 1.2 3.6 2.2 1.2 0.7 0.1 1.1 0.8 0.4 0.3 0.1 2.2 0.2 0.2 0.5 1.3 0.2 0.6 0.5 0.2 0.8 0.2 2.3 26.0
1.7 6.9 1.2 8.3 3.4 2.8 1.6 0.1 1.6 2.2 0.4 0.5 0.2 6.9 0.2 0.2 1.4 4.4 0.2 1.8 0.8 0.2 2.5 0.2 2.8 49.0
1.2
11.7 15.6 4.2 4.9 1.8 47.8 1.9 0.2 9.2
~~
Table I1 (Cont,inued) chemicals
min
25 %
median
concn. ug/m3 75%
max
mean
SD
mlp-xylene n-nonane o-xylene 1,3,5-tIimethylbenzene n-decane 1,3,4-trimethylbenzene n-undecane carbon monoxide, ppm nitrogen dioxide, ppb
17.8 0.2 6.6 3.1 0.2 9.3 0.7 6.0 9.0
25.6 0.2 9.7 3.9 1.1 13.2 1.0 9.0 41.3
33.5 0.2 12.4 5.4 1.2 17.3 1.6 10.0 61.0
39.5 0.2 14.3 5.9 1.5 20.2 1.9 14.0 96.5
43.5 0.2 16.6 6.9 3.4 23.0 2.2 22.0 183.0
31.8 0.2 11.9 5.1 1.4 16.7 1.5 11.7 71.2
8.6 0.0 3.2 1.3 0.8 4.6 0.5 4.0 39.3
isobutane n-butane 1,3-butadiene isopentane n-pentane 2-methylpentane hexane cyclohexane benzene 2,2,4-trimethylpentane n-heptane 2,3,3-trimethylpentane 2,3-dimethylhexane toluenle n-octane 2,3-dirnethylheptane ethylbenzene mlp-xylene n-nonane o-xylene 1,3,5-trimethylbenzene n-decane 1,2,4-trimethylbenzene n-undecane
0.2 13.8 1.2 26.2 4.8 8.7 5.6 0.6 4.5 8.1 2.2 2.5 0.1 25.1 1.1 0.2 4.9 16.3 0.2 6.3 2.5 0.7 8.9 0.2
22.0 113.8 1.2 136.1 37.2 26.1 11.3 1.1 8.9 14.3 4.0 4.1 1.7 39.5 1.9 0.2 7.8 26.1 4.6 10.0 4.2 3.8 14.6 3.2
8.0 40.5 1.2 52.8 15.6 13.9 7.4 0.7 6.8 9.9 2.9 3.1 1.1 30.8 1.4 0.2 5.9 19.9 0.9 7.6 3.2 1.5 10.9 1.4
7.4 36.7 0.0 41.3 11.3 6.5 2.2 0.2 1.5 2.3 0.7 0.5 0.5 5.9 0.3 0.0 1.0 3.6 1.8 1.3 0.6 1.1 2.1 1.0
(D) Sidewalk Measurements*?f 5.2 22.6 1.2 31.9 8.9 9.1 5.7 0.6 5.9 8.7 2.4 2.8 1.0 27.0 1.2 0.2 5.2 17.3 0.2 6.7 2.8 0.9 9.4 1.1
6.4 29.6 1.2 39.8 13.7 12.5 6.7 0.7 7.1 9.2 2.7 3.1 1.3 28.3 1.3 0.2 5.6 19.0 0.2 7.2 3.1 1.3 10.3 1.2
7.8 33.5 1.2 43.1 15.5 14.5 8.2 0.8 7.4 9.8 3.2 3.1 1.3 36.8 1.5 0.2 6.1 21.6 0.2 8.1 3.4 1.3 11.9 1.3
a Includes data from two vehicles and three roadways. bSample size is equal to the number of valid hourly averaging samples a t each sampling location. ‘Sample size: 77 for VOCs; 70 for CO; 35 for NOz; 30 for ozone. dSample size: 44 for VOCs; 16 for CO; 37 for ozone. eSample size: 10 for VOCs; 81 for CO; 75 for NOe. /Sample size: 6 for VOCs.
only 2090 of the ambient ozone concentrations measured at an ambient monitoring station. Low in-vehicle ozone is certainly expected because ozone reacts rapidly with by exhaust NO, which comprises about 90-959’0 of NO, emitted by vehicles. On the average, the in-vehicle VOC concentrations were -6 times higher than the VOC levels measured at the fixed sites. The in-vehicle CO level was -4.5 times higher than the ambient CO measurements reported by a state-operated monitoring station. These in-vehicle/fixed-site concentration ratios are plotted in Figure 1. Although the absolute concentration of each VOC at these four sampling locations differed, their relative abundances (i.e., ratios of each VOC concentration to the sum of‘the 24 VOC concentrations) were the same a t all four locations. Isopentanc?, n-butane, and n-pentane were the three most abundant aliphatic VOCs inside the vehicles, with median concentrations of 52.6, 36.1, and 19.4 pg/m3, respectively. Among the aromatic VOCs, toluene, m-/pxylene, and 1,2,4-trimethylbenzene were the three most abundant species. Median in-vehicle concentrations for these compounds were 43.1, 29.4, and 14.5 pg/m3, respectively. The median in-vehicle concentrations of two carcinogenic VOCs, 1,3-butadiene and benzene, were 2.9 and 10.7 ll.g/m3, respectively. The median in-vehicle concentrations of the three criteria pollutants were 11 ppb for ozone, 11.0 ppm for CO, and 81.0 ppb for NOz. The driver’s exposure to ozone was 3 times lower than the time-matched fixed-site measurements, which averaged 50.6 ppb. The driver’s exposure
to CO was 4.5 times higher than the fixed-site level, which was 3.0 ppm. One important feature of VOCs and criteria pollutants measured in this study was the large standard deviations for their measured concentrations. The maximum in-vehicle concentrations of some pollutants, such as isopentane, n-pentane, and hexane, measured 10-20 times higher than their mean concentrations. In the case of 1,3-butadiene and benzene, respective maximum 1-h in-vehicle concentrations were as high as 17.2 and 42.8 pg/m3 for each compared with median values of 2.9 and 10.7 pg/m3. The concentrations of the VOC measurements below method detection limits were set equal to l/z of method detection limits in the data analysis. Table I1 shows less variation in three VOC concentrations, 1,3-butadiene, 2,3-dimethylheptane, and n-nonane, than others because there were more samples below the method detection limits of these three VOCs. For example, there were 30% of in-vehicle samples, 30% of car exterior samples, and 100% of fixed-site and sidewalk samples below the method detection limit of 1,3-butadiene. Therefore, the values of these three compounds should be interpreted more cautiously than others. Car Effect. In-vehicle VOC, CO, and ozone concentrations were not found to be significantly different (p > 0.05) between the two experimental cars. NOz levels were not compared because only one car had in-vehicle measurements. The similarity of in-vehicle concentrations of the target compounds in both cars might indicate that pollutant concentrations were more strongly influenced by Environ. Sci. Technol., Vol. 25, No. 5, 1991
967
Table 111. Median Concentrationsn of the Target Pollutants Measured Inside Vehicles and at Fixed Sites for Each Driving Routebfc concn, pg/m3 chemicals isobutane n- butane 1,3-butadiene isopentane n-pentane 2-methylpentane hexane cyclohexane benzene 2,2,4-trimethylpentane n-heptane 2,3,3-trimethylpentane 2,3-dimethylhexane toluene n-octane 2,3-dimethylheptane ethylbenzene mlp-xylene n-nonane o-xylene 1,3,5-trimethylbenzene n-decane 1,2,4-trimethylbenzene n-undecane totald ozone, ppb (n) carbon monoxide in ppm (n) nitrogen dioxide in ppb (n)
in-vehicle (n = 77) urban n = 34 interstate n = 35 9.0 57.3 2.9 78.7 27.9 23.6 16.2 1.3 13.8 21.1 5.2 6.8 2.7 59.1 2.3 0.4 11.3 39.5 14.7 5.8 1.9 20.4 2.3 424.1 7 (11) 13 (30) 83 (14)
27.8 3.1 44.1 16.4 14.9 7.7 0.9 9.5 12.0 3.2 3.7 1.3 32.4 1.5 0.2 6.5 22.3
8.6 3.6 1.o 11.8 1.0 233.5 9 (14) 11 (34) 75 (17)
rural n = 8
urban n = 18
fixed-site (n = 44)' interstate n = 18
rural n = 8
2.2 8.4
3.7 14.8
1.2 5.0
1.0 2.9
10.2 5.2 3.2 2.2
4.7 2.5 1.9 1.1
3.0 1.8 0.5 0.6
1.6 1.6
0.7 0.5
4.5
1.8
1.2 3.7
17.5 7.0 5.2 3.4 0.3 2.4 3.8 1.4 1.3 0.4 12.1 0.6 0.2 2.2 7.0
0.9 3.0
0.4 0.8
1.5 0.7 0.3 2.9 0.9
2.7 1.1 0.5 3.7 0.4
1.3 0.7
0.4
1.7
0.6
91.3 41.5 (14) 2.8 (16) NA
31.5 49 (17) NA NA
14.8 71.5 NA NA
1.5 2.4 0.7 1.0 0.2 5.2
53.2 42 (5) 4 (6) 1102 (4)
"The median concentrations of VOCs that had fewer than 50% of samples above method detection limits are not listed in this table. *Includes data from two vehicles. n, sample size. Sum of 24 VOCs. 'NA, no data available.
ambient roadway concentrations that entered into the passenger compartment. Both of these two cars were Mercury built and had similar ventilation and exhaust systems. Since both cars were driven together under the same ventilation situation during each sampling trip, they were exposed to very similar roadway concentrations of VOCs, CO, and ozone each time. The possibility remains that there can be substantial differences between vehicles. For example, the contribution of VOCs from car materials, faulty exhaust systems, and engine and carburetor evaporative emissions are expected. Roadway Effect. In-vehicle VOC concentrations for the three driving routes were very different. The highest concentrations were found during urban driving, the second highest during interstate driving, and the lowest during rural driving. The fixed-site VOC measurements showed patterns similar to the in-vehicle VOC measurements. The median in-vehicle and fixed-site concentrations of the target compounds on each roadway are shown in Table 111. On the average, the urban/interstate/rural in-vehicle VOC concentration ratio was about 10/6/1. Such distinct differences in VOC levels between roadways may best be explained by the difference in air circulation and average traffic density on the three roadways. Although the VOC levels differed, depending on the type of roadway driven, the relative abundance of each VOC on each roadway (Le., ratios of the median concentration of each VOC to the sum of the median concentrations of the 24 VOCs) was relatively stable across the three roadways. These 24 targeted VOCs, therefore, appear to be related to vehicle traffic and reflect exposure to gasoline-related VOCs during driving. For the criteria air pollutants, in-vehicle ozone, CO, and NO2, concentrations on urban and interstate roadways 968
Environ. Sci. Technol., Vol. 25, No. 5, 1991
Table IV. Median In-Vehicle Concentrations of the Seven Target Aromatic VOCs Measured at Urban and Interstate Routes for Each Driving Period concn. ~ a / m ~ urban interstate evening morning evening morning (n = 18) (n = 16) (n = 17) (n = 18) benzene toluene ethylbenzene mlp-xylene o-xylene 1,3,5-trimethylbenzene 1,2,4-trimethylbenzene
11.6 19.0 9.5 33.3 12.8 5.0
15.9 75.8 13.9 46.0 17.1 7.1
10.8 38.5 7.4 25.7 9.5 3.7
9.1 30.7 6.0 20.6 7.8 3.5
17.3
24.1
13.0
11.0
Includes data from two vehicles.
were not found to be significantly different. In-vehicle CO levels measured on the rural roadway were significantly lower than those measured on the other two routes. Since the rural samples were collected at midday, it would be more appropriate to compare the levels of the two photochemical pollutants, ozone and NOz, by time of day rather than roadway type. Driving Time Effect. For the urban and interstate rush hour driving, the in-vehicle sum of VOC and CO levels was not found to depend on the time of measurement; i.e., the levels during morning driving were approximately the same as the levels during evening driving. However, there was one statistically significant interaction effect noted between roadway and driving time for seven aromatic VOCs. The in-vehicle concentrations of these aromatic
12.0
__
10.0
I
M
I I
8.0
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I
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I
6.0
4.0
__
2.0
-
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1
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Flgure 1 R a t m of median concentrations between in-vehicle and fixed-site measurements for 24 VOCs. CO, and ozone in Raleigh. NC. 1988. 0
.5. 0
jao.
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3 5. 160.
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Flgure 2. lnvehlcle concentrations of nnrogen dioxide and ozone during morning, midday, and evening driving tests.
I
VOCs during morning and afternoon driving for each driving route are listed in Table IV. On the urban roadway, in-vehicle aromatic VOC levels were higher in the evening. But, on the interstate roadway, in-vehicle aromatic VOC levels were higher in the morning. This may be due to different dispersion or turbulence effects at each roadway and different hot-soak evaporative emissions of individual VOCs during each driving time. The effect of different driving time on the measurements of in-vehicle criteria pollutants was more apparent. As shown in the Figure 2, the in-vehicle NO, and ozone concentrations were higher during evening driving than during morning driving, whereas the in-vehicle ozone level was the highest during midday driving. This finding seems to
1Wpe"me
Toluulc
mlpxylsnc
F W e 3. Comparisons of iwvehick cmcenbatbns of CO and selected VOCs between three ventilation conditions in Rakigh. NC: 1988.
be in agreement with the diurnal patterns of amhient ozone and NO2 concentrations found in most US.cities. Ventilation Effect. The effect of different vehicle ventilation modes on selected in-vehicle VOC and CO levels is shown in Figure 3. There were no significant Environ. Sci. Technd., VoI. 25, No. 5, 1991 96s
Table V. Concentration Ratios of In-Vehicle/Car Exterior for VOCs and CO in Raleigh, NC, 1988 chemicals
n
min
25%
median
75 9c
max
mean
SD
isobutane n-butane 1,3-butadiene isopentane n-pentane 2-methylpentane hexane cyclohexane benzene 2,2,4-trimethylpentane n- heptane 2,3,3-trimethylpentane 2,3-dimethylhexane toluene n-octane 2,3-dimethylheptane ethylbenzene m-lp-xylene n-nonane o-xylene 1,3,5-trimethylbenzene n-decane 1,2,4-trimethylbenzene n-undecane carbon monoxide nitroeen dioxide
10 10 10 10 10 10 10 10 10 10 10 10 10 10 10 10 10 10 10 10 10 10 10 10 70 32
0.02 0.99 0.44 0.99 0.86 0.64 0.85 0.81 0.75 0.91 0.93 0.96 0.05 0.73 0.79 1.00 0.70 0.66 1.00 0.64 0.54 0.58 0.51 0.60 0.10 0.20
0.99 1-00 0.55 1.09 1.11 1.09 0.99 1.00 0.88 1.03 1.00 1.02 1.00 0.97 1.00 1.00 0.94 0.95 1.00 0.94 0.89 0.91 0.91 0.90 0.64 0.88
1.00 1.10 1.21 1.11 1.13 1.16 1.13 1.13 1.08 1.11 1.05 1.11 1.14 1.07 1.07 1.00 1.06 1.06 1.00 1.06 1.03 1.00 1.04 1.11 1.10 1.04
1.00 1.19 2.33 1.22 1.14 1.24 1.25 1.17 1.12 1.16 1.13 1.18 1.20 1.14 1.18 1.00 1.15
1.08 1.62 14.33 2.33 1.18 1.46 1.30 8.00 2.49 1.63 1.60 1.89 20.00 1.66 1.85 4.50 1.63 1.60 14.00 1.55 1.46 1.52 1.34 4.22 3.56 3.32
0.91 1.14 2.58 1.31 1.09 1.11 1.10 1.76 1.16 1.12 1.10 1.17 2.96 1.09 1.17 1.35 1.05 1.04 3.15 1.03 0.98 1.01 0.97 1.49 1.09 1.18
0.31 0.19 4.21 0.45 0.10 0.24 0.15 2.20 0.49 0.20 0.19 0.27 6.00 0.24 0.32 1.11 0.26 0.26 4.66 0.25 0.26 0.27 0.25 1.11 0.59 0.63
1.14 1.00 1.12 1.08 1.09 1.10 1.42 1.43 1.31
Table VI. Correlation Coefficients for In-Vehicle VOCs and Criteria Pollutants Measured at Raleigh, NC, 1988O isopentane
benzene
toluene
m-Jp-xylene
co
n02
ozone
1.000 (n = 77) 0.855*** (n = 77) 0.744*** (n = 77) 0.744*** (n = 77) 0.372** (n = 70) 0.312 (n = 35) -0.353 (n = 30)
1.000 (n = 77) 0.903*** (n = 77) 0.827*** (n = 77) 0.458*** (n = 70) 0.214 (n = 35) -0.547** (n = 30)
1.000 (n = 77) 0.957*** (n = 77) 0.461*** (n = 70) 0.165 (n = 35) -0.423* (n = 30)
1.000 (n = 77) 0.429*** (n = 70) 0.112 (n = 35) -0.432* (n = 30)
1.000 (n = 70) 0.155 (n = 35) -0.195 (n = 29)
1.000 (n = 35) 0.622** (n = 21)
1.000 (n = 30)
isopentane benzene toluene m-lp-xylene
eo NO2 ozone "Statistical significance:
*, p < 0.05; **, p < 0.01; ***, p < 0.001.
differences in in-vehicle VOC concentrations between the windows closed/vent and fan on mode and windows opened mode. This result can be explained by the high air-exchange rate of a moving vehicle operating in either ventilation mode. An air-exchange rate as high as 36 air exchanges/h (ACH) was calculated by Hayes (9), based on infiltration data on an automobile reported by Peterson and Sabersky (IO). In-vehicle VOC levels were lowest when air conditioning was used. Air conditioning reduces the contribution of running loss emissions to in-vehicle VOC concentrations. In contrast, the lowest in-vehicle CO levels were measured when the car windows were open. However, the concentration difference for CO between the different ventilation modes was only 1 ppm. Penetration Effect. The penetration of pollutants from the car exterior into the car was determined by simultanfously measuring the pollutants inside and outside the experimental cars. These inside vs outside car ratios (I/O ratios) for VOCs, CO, and NO2 were calculated by taking the ratio between inside and outside measurements for each matched sample. The distribution of these 1/0 ratios is summarized in Table V. The median 1/0 ratio was -1.1 for CO and most VOCs and 1.04 for NOz. The slightly higher in-vehicle concentration appears systematic 970
Environ. Sci. Technol., Vol. 25, No. 5, 1991
and may be explained by the different heights of the inlets between the exterior sampling lines and the ports of vehicle's intake mechanism. The air in the car compartments came mostly from the front hood and the side windows. The engine running loss emissions and tailpipe exhausts, therefore, may have a more direct impact on the in-vehicle concentrations than the car exterior concentrations. Pedestrian Exposure. In order to evaluate the VOC exposure of urban pedestrians, six pedestrian sidewalk samples were collected. As shown in Table 11, the median pedestrian sidewalk concentrations of 1,3-butadiene and benzene were 1.2 and 7.1 pg/m3, respectively. Isopentane was the most abundant aliphatic VOC with a median concentration of 39.8 pg/m3, and toluene was the most abundant aromatic VOC with a median concentration of 28.3 pg/m3. There was a clear VOC concentration gradient decreasing from in-vehicle to sidewalk to fixed-site locations in this data set. This concentration trend is shown in Figure 4 for selected VOCs. The median concentration ratio of in-vehicle/pedestrian/fix site was about 101512 for benzene, toluene, and m-/p-xylene. This result indicates that we would underestimate in-vehicle VOC levels as well as the urban pedestrian's VOC exposures by using fixed-site monitoring data only.
Table VII. Correlation Coefficients for Fixed-Site VOCs and Ozone Data Measured at Raleigh, NC, 1988' isopentane
isopentane
benzene
toluene
Loo0 (n = 44) 0.921*** (n = 44) 0.920*** (n = 44) -0.386' (n = 37)
1.ooO (n = 44) 0.985*** (n = 44) -0.376* (n = 37)
m-lp-
xylene
ozone
1.000 ( n = 44)
benzene
0.565*** (n = 44)
toluene
0.517*** (n = 44) 0.459.. (n = 44) -0.086 (n = 37)
rn-/p-xylene ozone (I
Statistical significance:
1.MlO (n = 44) 4).432** 1.oo0 (n = 37) (n = 37)
*, p < 0.05: **, p < 0.01; I**,p < 0.001.
Modeling. The correlation matrix of selected in-vehicle and fixed-site measurements for VOCs, CO, ozone, and NOpis presented in Table VI and VII. All of the selected VOCs were highly correlated ( r = 0.62-0.96). The correlation of VOCs within the aromatic group was even higher ( r = 0.83-0.96). This implies that certain representative VOCs may be chosen as surrogates to assess total VOC exposures involving vehicle emissions. The in-vehicle CO concentrations had only moderate correlations with the measured VOCs ( r = 0.37-0.46). This implies that the extrapolation of CO commuter exposure models to the study of commuters' VOC exposures would be ill-advised. Ozone was found to be negatively correlated with the measured VOCs ( r = -0.35 to -0.54) and positively correlated with N O p ( r = 0.62) for the in-vehicle measurements. Similar correlations were found between fixed-site measurements of ozone, VOCs, and NOz. Since in-vehicle VOC measurements are costly and time-consuming, it would be more practical if we can accurately estimate in-vehicle exposures using fixed-site monitoring data only. Table VI11 presents the predicted coefficients from simple linear regression models applied to a selected numher of VOCs. The in-vehicle VOC levels (Yij)are estimated by using fixed-site measurements (X,) and a categorical variable, roadway (Ai), and a random variable, plus an error term (Error). Yjjis equal to the sum of A , B X , and Error (Y..= Ai + SX, + Error). The linear regression models fittea to the data had consistent slope estimates ( B ) ,large intercepts (Ai), and moderate error terms (Error). Overall, these models account for about 50-63% variation in the in-vehicle VOC measurements. The slope for aliphatic VOCs was -0.7 while that for aromatic VOCs was -1.8. The different slopes estimated for aliphatic and aromatic VOCs by these models could be the result of different depletion and conversion rates for these two classes of VOCs moving from the middle of the roadways to the fixed-site monitoring stations as the result of photochemical reaction and atmospheric dis-
min~xylene
Tolvcne
1SOpc"taX
~
Figure 4. Comparisons of VOC measurements in in-vehicle, car exterior, sidewalk. and fixed-site for selected VOCs.
persion. The intercepts estimated for these models represent about 50-6070 of predicted in-vehicle concentrations, e.g., 8.9 4 m 3 out of 16.6 pg/m3 on urban mutes and 7.9 pg/m3 out of 13.2 pg/m3 on interstate routes for benzene in Table VIII. This implies that a sizable fraction of vehicle-related VOC exposure is associated with roadway type and traffic conditions. The error terms in the models were either from daily variation of VOC concentrations on roadways or from the unrepresentative location of the fixed-site monitoring. This indicates that models for driver's exposure to auto-related VOCs can be improved with more vehicle measurements, better characterization of the day-to-day fluctuation of ambient VOC levels, and relocation of fixed-site monitors to roadways rather than set back from roadsides. Besides the statistical approach of modeling commuter exposure to VOCs discussed above, there are other methods of estimating commuter exposure, which are not discussed in this paper, such as modeling pollutant dispersion on the roadways. Since the outside-car to insidecar ratios were close to 1 for most compounds, it is possible that commuter exposure to VOCs can he estimated by using the existing roadway dispersion models or their modification as well as the linear regression models derived in this paper.
Table VIII. Linear Models for Predicting In-Vehicle VOC Concentrations Using Fixed-Site Measurements' chemicals n-butane isopentane benzene toluene rn-/p-xylene
conditions (i)
[A;]
[Bl
error
72
j
mean Xij
max X;j
min Xjj
interstate urban interstate urban interstate urban interstate urban interstate urban
27.1 44.0 41.4 64.5 7.9 8.9 28.0 40.5 19.5 26.9
0.7 0.7 0.8 0.8 1.6 1.6 1.8 1.8 1.8 1.8
12.5 12.5 16.7 16.7 2.9 2.9 16.0 16.0 10.4 10.4
0.500 0.500 0.627 0.627 0.505 0.505 0.598 0.598 0.563 0.563
35 28 35 29 35 32 35 34 35 34
6.1 15.8 6.4 23.4 1.5 3.0 4.9 14.6 3.0 8.9
16.3 41.4 17.7 116.6 3.6 8.5 12.8 46.8 8.4 27.8
1.5 6.7 2.1 8.7 0.1 2.3 1.1 6.8 0.7 4.3
'Linear regression model: Yii = Ai
+ BXij + error. Units, @g/m3.All parameters in the models are statistically significant at p < 0,001. Environ. Sci. Technol., VoI. 25. No. 5, 1991 971
Conclusion
We have successfully demonstrated that 6-L Summa polished stainless steel canister samplers coupled with GC f MS analysis were able to collect and quantify 24 invehicle and ambient VOC concentrations over a 1-h sampling period. We also identified certain vehicle characteristics and driving conditions affecting in-vehicle VOC levels. Two experimental cars, a 1987 Mercury Sable and a 1985 Mercury Marquis, showed no differences in in-vehicle VOC, CO, NO,, and ozone concentrations despite their difference in year model. However, since these two cars were relatively new and with low mileage, it is necessary to test older models to see the effect of car model on in-vehicle VOC concentrations. Roadway types had the greatest effect on in-vehicle levels of VOCs and CO. The highest in-vehicle levels of VOCs and CO were measured during urban driving and the lowest measured during rural driving. Presumably this reflects different traffic density as well as different ambient atmospheric dispersion and turbulence on the three types of roadway. We also found that the lack of high correlation observed between VOC and CO concentrations could limit the extrapolation of CO commuter exposure models to the modeling of commuter exposures to VOCs. We found that driving time had the most profound effect on in-vehicle levels of NO, and ozone. Evening driving showed higher in-vehicle NOz and ozone concentrations than morning driving. The highest in-vehicle ozone was measured during midday driving. This was in agreement with the diurnal cycle of the ambient photochemical pollutants. I t is, therefore, highly recommended more data be collected on the temporal changes of in-vehicle NO, and ozone concentrations in the areas with high ambient concentrations of these two pollutants. Use of air conditioning was shown to result in the lowest in-vehicle VOC levels. The 1/0 ratios of the target pollutants were calculated to be slightly higher than l. This is thought to be a more direct contribution of tailpipe and engine running loss emissions into the passenger compartment. In general, fixed-site measurements underestimated the in-vehicle VOC and CO levels but overestimated the in-vehicle ozone levels. In the urban environment, in-vehicle VOC levels were found to be higher than sidewalk VOC levels. Although this study provided valuable information, additional monitoring studies with more cars are also recommended. In particular, studies in other metropolitan areas with more congested traffic, different dispersion conditions, and longer commuting time are recommended. These traffic conditions in Raleigh, NC, are different from those in larger cities, such as New York, Boston, Washington, Chicago, and Los Angeles. A subset of surrogate VOCs could be selected for future studies since they were all highly correlated in this study. Simple linear regression models, with moderate explanatory power (r2 = 0.5O-O.63),were fitted to the data collected in this study to estimate the in-vehicle VOC levels by using the fixed-site measurements. Since these models are derived from one city only during the summer, the repetition of the same
972
Environ. Sci. Technol., Vol. 25, No. 5, 1991
experiment in winter and in other cities is essential in order to verify their general representativeness. Acknowledgments
We especially thank Drs. L. Wallace and W. Nelson of U S . EPA and Dr. P. Kinney, currently a t NYU Medical Center a t Tuxedo, for their help in designing this study. Additionally, we thank Dr. D. Harlos, currently a t ESE Inc., and Mr. D. Whitaker and Mr. R. Shores of RTI for their assistance in carrying out this field study. Lastly, we thank reviewers of this paper for their helpful and resourceful comments. Registry No. Isobutane, 75-28-5; n-butane, 106-97-8; 1,3-butadiene, 106-99-0; isopentane, 78-78-4; n-pentane, 109-66-0; 2methylpentane, 107-83-5;hexane, 110-54-3; cyclohexane, 110-82-7; benzene, 71-43-2; 2,2,4-trimethylpentane, 540-84-1; n-heptane, 142-82-5; 2,3,3-trimethylpentane, 560-21-4; 2,3-dimethylhexane, 584-94-1; toluene, 108-88-3; n-octane, 111-65-9; 2,3-dimethylheptane, 3074-71-3; ethylbenzene, 100-41-4; p-xylene, 106-42-3; m-xylene, 108-38-3; n-nonane, 111-84-2; 0-xylene, 95-47-6; 1,3,5trimethylbenzene, 108-67-8; n-decane, 124-18-5; 1,2,4-trimethylbenzene, 95-63-6; n-undecane, 1120-21-4;carbon monoxide, 630-08-0; nitrogen dioxide, 10102-44-0; ozone, 10028-15-6. L i t e r a t u r e Cited Wallace, L. A. The TEAM Study: Summary and Analysis; EPA 600/6-87/002a; N T I S P B 88-1000060; U S . EPA: Washington, DC, 1987; Vol. I. Wallace, L. A. Cancer Risk from Organic Chemicals in the Homes; Proceedings of Environmental Risk Management, APCA, Pittsburgh, PA, November 1986; 1986; p p 14-24. Wallace, L. A. Risk Anal. 1990, 10, 59-64. Hampton, C. V.; Pierson, W. R.; Harvey, T. M.; Updegrove, W. S.; Marano, R. S. Enuiron. Sci. Technol. 1982, 16, 287-298. Lonneman, W. A.; Seila, R. L.; Meeks, S.A. Enuiron. Sci. Technol. 1986,20, 790-796. Sigsby, J. E.; Tejada, S.; Ray, W.; Lang, J. M.; Duncan, J. W. Environ. Sci. Technol. 1987, 21, 466-475. Zweidinger, R. B.; Sigsby, J. E.; Tejada, S. B.; Stump, F. D.; Dropkin, D. L.; Ray, W. D.; Duncan, J. D. Enuiron. Sci. Technol. 1988, 22, 956-962. Sheldon, L. S.; Whitaker, D. A. VOC Commuter Exposure Study Vol. 2: Sampling and Analysis Protocol, Work Plan. U S . Contract 68-02-4544; Research Triangle Institute, Research Triangle Park, NC, 1988. Hayes, S. R. J. Air Pollut. Control Asssoc. 1989, 39, 1453-1461. Petersen, G. A,; Sabersky, R. M. J. Air Pollut. Control ASSSOC.1975,25, 1028-1302. Received for review March 23,1990. Revised manuscript received July 31,1990. Accepted January 4,1991. Although this research was supported by a U.S. Environmental Protection Agency Cooperative Agreement CR-813526-01-2 and a n EPA contract (Contract 68-02-4544), it has not been subjected to required peer a n d administrative review a n d does not necessarily reflect the views of the agency, and no official endorsement should be inferred.
