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Large PAHs detected in fine particulate matter emitted from light-duty gasoline vehicles
ARTICLE IN PRESS
Atmospheric Environment 41 (2007) 8658–8668 www.elsevier.com/locate/atmosenv
Large PAHs detected in fine particulate matter emitted from light-duty gasoline vehicles Sarah G. Riddlea, Chris A. Jakoberb, Michael A. Robertc, Thomas M. Cahillb,1, M. Judith Charlesb,{, Michael J. Kleemanc, a
Department of Chemistry, University of California, 1 Shields Avenue, Davis, CA 95616, USA Department of Environmental Toxicology, University of California, 1 Shields Avenue, Davis, CA 95616, USA c Department of Civil and Environmental Engineering, University of California, 1 Shields Avenue, Davis, CA 95616, USA b
Received 17 February 2007; received in revised form 13 July 2007; accepted 13 July 2007
Abstract Emission factors of large PAHs with 6–8 aromatic rings with molecular weights (MW) of 300–374 were measured from 16 light-duty gasoline-powered vehicles (LDGV) and one heavy-duty diesel-powered vehicle (HDDV) operated under realistic driving conditions. LDGVs emitted PAH isomers of MW 302, 326, 350, and 374, while the HDDV did not emit these compounds. This suggests that large PAHs may be useful tracers for the source apportionment of gasoline-powered motor vehicle exhaust in the atmosphere. Emission rates of MW 302, 326, and 350 isomers from LDGVs equipped with three-way catalysts (TWCs) ranged from 2 to 10 (mg L1 fuel burned), while emissions from LDGVs classified as low emission vehicles (LEVs) were almost a factor of 10 lower. MW 374 PAH isomers were not quantified due to the lack of a quantification-grade standard. The reduced emissions associated with the LEVs are likely attributable to improved vapor recovery during the ‘‘cold-start’’ phase of the Federal Test Procedure (FTP) driving cycle before the catalyst reaches operating temperature. Approximately 2 (mg g1 PM) of MW 326 and 350 PAH isomer groups were found in the National Institute of Standards and Technology standard reference material (SRM)#1649 (Urban Dust). The pattern of the MW 302, 326, and 350 isomers detected in SRM#1649 qualitatively matched the ratio of these compounds detected in the exhaust of TWC LDGVs suggesting that each gram of Urban Dust SRM contained 5–10 mg of PM originally emitted from gasoline-powered motor vehicles. Large PAHs made up 24% of the total LEV PAH emissions and 39% of the TWC PAH emissions released from gasoline-powered motor vehicles. Recent studies have shown certain large PAH isomers have greater toxicity than benzo[a]pyrene. Even though the specific toxicity measurements on PAHs with MW 4302 have yet to be performed, the detection of significant amounts of MW 326 and 350 PAHs in motor vehicle exhaust in the current study suggests that these compounds may pose a significant public health risk. r 2007 Elsevier Ltd. All rights reserved. Keywords: Gasoline vehicle; Diesel vehicle; Source apportionment; Toxicity
Corresponding author. Tel.: +1 530 752 8386; fax: +1 530 752 7872.
E-mail address: [email protected] (M.J. Kleeman). Present address: Department of Integrated Natural Sciences, Arizona State University at the West Campus, P.O. Box 37100, Phoenix, AZ 85069, USA. { Deceased. 1
1352-2310/$ - see front matter r 2007 Elsevier Ltd. All rights reserved. doi:10.1016/j.atmosenv.2007.07.023
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1. Introduction Internal combustion engines in motor vehicles generate fine particles in their exhaust with diameters o2.5 mm (see for example, Hildemann et al., 1991; Ondov and Wexler, 1998; Kleeman et al., 2000). Composition plays a role in determining the ecosystem impacts and human health effects of these particles. Polycyclic aromatic hydrocarbons (PAHs) are one class of compounds present in both diesel and gasoline motor vehicle exhaust known to contribute to the adverse health effects of vehicular emissions (EPA, 2000). PAHs can originate from unburned fuel (Westerholm et al., 1988; Marr et al., 1999) and lubricating oil (Brandenberger et al., 2005) or they can be formed during the combustion process from simpler organic compounds (Westerholm et al., 1988). Due to their low vapor pressures, PAHs with five or more aromatic rings rapidly condense onto particles once emitted into the atmosphere resulting in a higher proportion of these PAHs in smaller, respirable size particles (Venkataraman and Friedlander, 1994; Venkataraman et al., 1994). Due to their toxicity and atmospheric prevalence, 16 PAHs ranging in size from naphthalene (2-rings, molecular weight (MW ¼ 128) to benzo[g,h,i]perylene (6-rings, MW ¼ 276) have been identified as priority pollutants by the U.S. Environmental Protection Agency (USEPA, 1988). Most of the literature on PAH concentrations from both source and ambient air samples focus on these 16 EPA priority compounds (Benner et al., 1989; Rogge et al., 1993a, b; Allen et al., 1996; Miguel et al., 1998; Cadle et al., 1999; Marr et al., 1999; Schauer et al., 1999, 2002, 2003; Blanchard et al., 2002; Dyke et al., 2003; Zielinska et al., 2004). When additional compounds are examined, they are typically isomers or of similar MW. Larger PAHs (MW4300) exist in the environment (Marvin et al., 1999), but are seldom identified in source and/or environmental samples even though several of these compounds have shown equal if not greater carcinogenic activity when compared with smaller PAHs (Schmidt et al., 1987; Cavalieri et al., 1991, 1988). This is mostly due to the lack of commercially available standards for larger PAHs with MW4302 Da. Coronene (C24H12) has been proposed as potential tracer for gasoline motor vehicle exhaust (Schauer et al., 1996; Cass, 1998) but this compound was recently detected in emissions from heavy-duty diesel vehicles operated under low speed driving
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conditions (Riddle et al., 2007a). Measurements of even larger PAHs in both source and ambient aerosols would further aid source apportionment of ambient particulate matter. In the present study, emission factors are presented for PAHs with MW up to 350 for light-duty gasoline-powered vehicles (LDGV) and heavy-duty diesel-powered vehicles (HDDV) operated under realistic driving conditions. The concentrations of these compounds were also quantified in the National Institute of Standards and Technology (NIST) Urban Dust standard reference material (SRM)#1649. The direct potential toxicity of PAH compounds with MW4300 Da and their usefulness as tracers for gasoline engine exhaust are discussed in the following sections. 2. Methods 2.1. Light-duty gasoline-powered vehicles (LDGV) emissions collection conditions LDGV emission samples were collected at the Haagen-Smit Laboratory in El Monte, CA, during August–September 2002. The emissions sampling conducted in this study utilized an experimental design similar to that described by Schauer et al. (2002). Briefly, vehicles were operated on a stationary chassis dynamometer using the Federal Test Procedure (FTP) driving cycle. Dynamometer resistance was specified to simulate appropriate vehicle inertial load and wind drag. Multiple vehicles within a technology class were composited on one set of sampling media to obtain sufficient mass for chemical analysis. The LDGV technology classes reported in the current study include low emission vehicles (LEV) and three-way catalyst (TWC) equipped vehicles. A complete description of the vehicles tested in the current study is provided in the upper portion of Table S1 in the Supplemental Information. Measured PM1.8 emissions rates and the estimated fuel economies for each vehicle category are shown in Table 1. Thorough details of the emission collection and dilution methods for the LDGV vehicle matrix are discussed by Robert et al. (2007a). 2.2. Heavy-duty diesel-powered vehicle (HDDV) emissions collection conditions HDDV emissions were collected from a 1999 Freightliner truck (see description in Table S1) in
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Table 1 Measured PM1.8 emissions rates and estimated fuel economy for test vehicles used in the current study Vehicle class— driving cycle LDGV LEV— FTP LDGV TWC— FTP HDDV— Idle+Creep HDDV— HHDDT 5-mode
PM1.8 emissions rate (mg km1)
Estimated fuel economy (km L1)
370
9.3
910
10.0
1,500,000 220,000
0.51 2.3
Riverside, CA, during June–July 2003. Emissions from this vehicle were comparable to emissions from three additional heavy-duty diesel vehicles that were tested at the same time (Riddle et al., 2007a; Robert et al., 2007b). The heavy-duty diesel vehicle was driven through a 5-mode heavy heavy-duty diesel truck (HHDDT) transient driving cycle consisting of a 30 min idle, 17 min creep, 11 min transient stage and two cruise stages of 34 and 31 min, with a top speed of 65 mile h1 for the second cruise (Gautam et al., 2002). In addition to the full HHDDT test cycle, emissions were sampled while the truck was operating under only the idle and creep modes of the HHDDT cycle, hereafter referred to as idle+creep. The measured PM1.8 emissions rate and estimated fuel economy for both tests are summarized in Table 1. The inertial weight used for all heavy-duty diesel tests was 56,000 lbs. Only a few heavy-duty dynamometers in the United States were capable of simulating transient driving cycles using such large inertial weight at the time that the tests were conducted. All HDDV results reported in the current study used the mobile chassis dynamometer operated by West Virginia University. A more detailed description of the testing conditions is provided by Robert et al. (2007b). 2.3. Sample collection Tailpipe emissions from both LDGVs and HDDVs were diluted in two stages and then aged for approximately 30–60 s using a dilution sampling system (Hildemann et al., 1989). Dilution ratios were set to balance the need to reproduce the high dilution factors that occur in the atmosphere against the need to concentrate the sample sufficiently for chemical analysis. Typical dilution ratios ranged
from 125 to 583 (Robert et al., 2007a, b). The emissions were collected after dilution using an annular denuder-filter-polyurethane foam (PUF) sampling train similar to that described by Schauer et al. (1999, 2002), with specific details described by Jakober et al. (2006, 2007). The majority of the sampling hardware and PUF substrates were obtained from University Research Glassware (Chapel Hill, NC). Quartz fiber filters, 47 mm, were obtained from Pall Gelman (Ann Arbor, MI) and baked at 550 1C for a period of at least 12 h to minimize any organic contaminants prior to each testing event. Eight-channel annular denuders were coated with XAD-4 polystyrene resin using a procedure similar to that described by Gundel et al. (1995) and Gundel and Lane (1999). After collection, particle-phase filter samples were stored in baked glass Petri dishes that were covered with baked aluminum foil and wrapped with Teflon tape prior to placement in a desiccator, under organicfree nitrogen and stored at 20 1C until solvent extraction. 2.4. Chemical extraction procedures Quartz filter PM2.5 samples were extracted individually. Filters were spiked with isotopically labeled PAHs and the solvent was allowed to dry before proceeding. Substrates were then placed into screw-cap centrifuge tubes for organic solvent extraction. Each tube was filled with 15 mL of a 1:1 mixture of hexane (Burdick and Jackson trace analysis grade) and dichloromethane (Burdick and Jackson trace analysis grade) then suspended in an ultrasonic cleaning bath for 15 min. The sonication extraction procedure was repeated three times and all of the extracts were combined. Samples were concentrated by nitrogen evaporation to a final volume of 200 mL. For each batch of extracted samples, a method blank was also collected and NIST SRMs#1649 (Urban Dust) and #1650 (Diesel Particulate Matter) were simultaneously extracted to provide further confidence in the methods. 2.5. GC– MS data acquisition methods and data conversion The organic chemical speciation data collected for this project was obtained on a Varian 3400 gas chromatograph (GC) coupled with a Varian 2000 ion-trap mass spectrometer (GC–ITMS). Chemical separation of PAHs was performed on a J&W
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DB-5HT high-temperature capillary column (30 m 0.25 mm i.d., 0.1 mm film thickness, 5% phenylsubstituted polysiloxane, Agilent Technologies, Palo Alto, CA) with helium as the carrier gas at a linear velocity of 37 cm s1. Samples were introduced through a temperature programmable injection port that was held at the initial temperature of 64 1C for 1 min, then ramped at a rate of 20 1C min1 to a temperature of 120 1C, the rate of heating was then increased to 100 1C min1 to a final injector temperature of 375 1C where it was held until the end of the column program. The column was initially at a temperature of 64 1C for the first 7 min to allow the analytes to pass through the injection port and become cryo-focused on the front-end of the analytical column. The column oven temperature was then increased at a rate of 5 1C min1 to a final temperature of 400 1C where it was held for 5 min for a total run time of 79.2 min. The ion trap oven, manifold, and transfer line were operated at 220, 80, and 300 1C, respectively. Analyte identification and quantification was conducted by electron ionization mass spectrometry– mass spectrometry (EI–MS2) analysis. In all cases, the parent ion was the molecular ion for the compound while the product ion was generally [M-2]+, which corresponds to a loss of H2. The mass/charge range monitored during analysis was 100–420. Elution times and target ions for each
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PAH isomer family are shown in Table 2. Analysis data files were processed using Varian Saturn GC–MS Workstation software version 5.51, with chromatographic peak integrations performed manually. The methods used to generate vehicle emission rates are discussed in detail elsewhere (Jakober et al., 2007; Robert et al., 2007a, b) and are summarized here. Briefly, the collected mass measured by GC–ITMS analysis is converted to sampled concentration using six point calibration curves. The sampled concentration was corrected for any background material that was measured to be present in the dilution air used during each vehicle test. The corrected sampled concentration was converted into emitted mass using the ratio of sampled flow to total flow while accounting for primary and secondary dilutions. Emissions can be expressed in units of (mg km1) or (mg L1 fuel burned). Fuel consumption was calculated from measured CO2 emissions using previously reported conversion factors of 2.28 and 2.77 kg CO2 L1 of combusted gasoline and diesel fuel, respectively, were applied to the CO2 measurements (United States Department of Energy, 2004). Appropriate quantification and treatment of recovery efficiencies is a critical step in the measurement of PAH emissions rates from motor vehicles. All results in the current manuscript were
Table 2 GC–ITMS analysis protocol for PAHs Time segment
Elution window (min)
Targeted compounds
Molecular ion (m/z)
Resonant excitation energya (V)
Quantification ionb (m/z)
1 2 3
25.0–30.0 30.0–36.0 36.0–42.0
4 5
42.0–48.0 48.0–52.5
6
52.5–57
7 8
57.0–62.0 62.0–67.0
9
67.0–79.2
MW 178 isomers MW 202 isomers Cyclopenta[c,d]pyrene MW 228 isomers MW 252 isomers MW 276 isomers Dibenz[a,h]anthracene Coronene MW 302 isomers MW 326 isomers MW 350 isomersc MW 352 isomers MW 374 isomersc MW 376 isomersc
178 202 226 228 252 276 278 300 302 326 350 352 374 376
1.5 1.6 1.8 1.6 1.5 1.6 1.6 1.6 1.6 1.6 1.6 1.6 1.6 1.6
152 200 224 226 250 274 276 298 300 324 348 350 372 374
a
The optimal energy was the excitation energy that gave the greatest intensity of a product ion. Most PAHs lose two hydrogen atoms to form the quantification ion, one exception are the MW 178 isomers which lose a –C2H2-group to form the quantification ion. c Elution times for these isomers were based on those of an impure standard, quantification could not be directly performed for these analytes. b
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efficiency of only 3% was observed for dibenzo[a,i]pyrene in the heavy-duty diesel sample operated under the full 5-mode HHDDT driving cycle. All other compounds were recovered with better than 50% efficiency in all tests
corrected for recovery efficiency by dividing the measured PAH concentration by the measured recovery efficiency of the surrogate compound with a similar structure. Recoveries of the surrogate internal standards (IS) spiked onto filters prior to solvent extraction are displayed in Table 3. The obtained IS recoveries are consistent with our expectations based on previous observed recovery of model compounds (data not presented). Generally speaking, recovery of PAH compounds depends on the MW of the PAH and the amount of elemental carbon present in the sample. A recovery
3. Results 3.1. LDGV and HDDV PAH emission factors Mass emissions rates of PAHs (mg PAH L1 fuel burned) for LDGVs and HDDVs vehicles are
Table 3 Recovery of spiked internal standards used to correct emissions from low emission vehicle (LEV FTP), three-way catalyst vehicle (TWC FTP), heavy-duty diesel vehicle idle+creep (HDDV Idle/Creep) and heavy-duty diesel vehicle 5-mode (HDDV 5-mode) tests Compound
d10 d12 d12 d12 d14
Internal standard recovery (%)
Phenanthrene Chrysene Benzo[k]fluoranthene Benzo[g,h,i]perylene Dibenzo[a,i]pyrene
LEV FTP
TWC FTP
HDDV Idle/Creep
HDDV 5-mode
67 93 87 96 99
22 82 86 102 108
62 107 94 100 59
42 82 64 38 3
Compounds were quantified using d10-pyrene as the internal standard. LEV
TWC
99 Frtlnr Idle/Creep
Light PAHs
99 Frtlnr 56 K
Heavy PAHs
90
16 Large PAHs
70
14
0
0 MW 350 isomers
2 MW 326 isomers
10 MW 302 isomers
4
coronene
20
benzo[ghi]perylene
6
indeno[1,2,3cd]pyrene
30
benzo[a]pyrene
8
benzo[e]pyrene
40
benzofluoranthene isomers
10
MW 228 isomers
50
pyrene
12
fluoranthene
60
phenanthrene
Light PAH Emission Factor (ug/L of fuel burned)
80
Heavy and Large PAH Emission Factor (ug/L of fuel burned)
18
Fig. 1. PAH emissions factors (mg L1 fuel burned) for low emission gasoline-powered vehicles (LEV), three-way catalyst gasolinepowered vehicles (TWC), heavy-duty diesel vehicle operated under idle+creep (99 Frtlnr Idle/Creep) and heavy-duty diesel vehicle operated under 5-mode transient cycle (99 Frtlnr 56 K). Note that the scale for the light PAH measurements is given on the left and the scale for the heavy and large PAHs is given on the right side of the figure.
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presented in Fig. 1. The uncertainty range on each emission factor is based on the %RSD of a similar compound obtained in the model analyte recovery experiment (data not presented). This error does not fully capture the uncertainty associated with sample collection and should be interpreted as a lower bound on uncertainty for the emissions rate. Emissions of light and heavy PAHs from these vehicles have been reported previously (Riddle et al., 2007a, b) while emissions of large PAHs (MW4302) described in this paper are reported for the first time based on the methods described above. Fig. 1 clearly shows that diesel vehicles operating under the full 5-mode HHDDT driving cycle emit the greatest quantity of light PAHs per unit of fuel burned. Engine load had a strong effect on PAH emissions from diesel vehicles, with greater quantities of light PAHs emitted during the full HHDDT driving cycle and greater amounts of heavy PAHs emitted during the idle+creep cycle. The higher engine loads and hotter combustion temperatures in the full 5-mode HHDDT driving cycle appear to break down the heavy PAHs. Heavy-duty diesel vehicles did not emit PAHs with MW 4300 Da (coronene) even under the lightest driving cycle tested (idle+creep). Emissions of light PAHs per unit of fuel consumed by gasoline-powered vehicles were much lower than emissions from heavy-duty diesel vehicles. Emissions of heavy PAHs per unit of fuel consumed were comparable from light-duty gasoline and heavy-duty diesel vehicles. The greatest difference between the vehicle classes was apparent in the large PAH emissions. Gasoline-powered vehicles emitted MW 302, 326, and 350 PAH isomers while heavy-duty diesel vehicles did not emit these compounds. MW 374 PAH isomers were also detected in particles emitted from gasolinepowered vehicles but no pure standard was available for quantification. Emission rates of MW 302, 326, and 350 isomers from gasoline-powered vehicles equipped with TWCs ranged from 2 to 10 (mg L1 fuel burned) while emissions from LEVs were almost a factor of 10 lower. The reduced emissions associated with the LEVs are likely attributable to improved vapor recovery during the ‘‘cold-start’’ phase of the FTP driving cycle before the catalyst reaches operating temperature. Large PAHs with MW of 326 and 350 Da made up 24% of the total LEV PAH emissions and 39% of the TWC PAH emissions. Besides the potential direct toxicity of these compounds, large PAHs with MW4302 Da appear
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to be good candidates for the source apportionment of gasoline-powered motor vehicle exhaust. Coronene has previously been suggested as a marker for gasoline engine exhaust (Schauer et al., 1996; Cass, 1998), but the detection of coronene in the exhaust from diesel-powered vehicles operated under the idle+creep driving cycle highlights the need to quantify large PAHs that are possibly unique to gasoline engines. Relative concentrations of PAHs (mg PAH g1 total PM) in the exhaust of the motor vehicles tested in the current study are summarized in Table 4. These units are useful since they can be directly used in source apportionment studies. Total PM concentrations were measured by Robert et al. (2007a, b) using gravimetric methods. A comparison between Fig. 1 and Table 4 shows that emissions of PAHs from diesel vehicles are generally similar when expressed in either (mg PAH L1 fuel burned) or (mg PAH g1 PM). In contrast, emissions from gasoline vehicles are several orders of magnitude larger when they are normalized by the amount of particulate matter emitted rather than the amount of fuel burned. This reflects the fact that gasolinepowered vehicles emit significantly less particulate matter per liter of fuel burned than heavy-duty diesel vehicles. When PAHs were detected in the diesel exhaust, their relative concentrations were at least 2–3 times lower than analogous concentrations in light-duty gasoline-powered vehicle exhaust. Gasoline-powered vehicles equipped with TWCs produced higher concentrations of large PAHs per unit of total PM emitted than LEVs despite the fact that total PM emissions from TWCs are nearly 2.5 times greater. 3.2. Analysis of NIST SRMs and comparisons with previous studies NIST SRMs#1649 (Urban Dust) and #1650 (Diesel Particulate Matter) were used to monitor the accuracy of the chemical analysis methods during the extraction of PM emission samples. Five and two milligrams NIST SRM#1649 and #1650 were solvent extracted, respectively, using the procedure described above. Internal standard recoveries (Table S3 in the Supplemental Information section) ranged between 80% and 109% for SRM#1649 (Urban Dust) and 32% and 109% for SRM#1650 (Diesel Particulate Matter). Recoveries of light PAHs (those up to and including d12 chrysene) were all 490%; recoveries decreased with
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Table 4 PAH emission factors normalized by total PM emissions rates (mg PAH g1 PM) for gasoline-powered low emission vehicles (LEVs), three-way catalyst vehicles (TWC), heavy-duty diesel vehicles operated under idle+creep driving conditions (HDDV Idle/Creep) and heavy-duty diesel vehicles operated under the full HHDDT 5-mode driving cycle (HDDV 5-mode) Compound
Emission rate (mg PAH g1 PM)a LEV FTP
Light PAHs Phenanthrene Anthracene Fluoranthene Pyrene Cyclopenta[c,d]pyrene MW 228 isomersc
320737 3975 110711 230723 3874
HDDV Idle/Creep
HDDV 5-mode
b
240728
4075
150717
8178 160717 3774 350726
670.6 1572
7877 110711
d d e e e e
TWC FTP
det
Heavy PAHs Benzofluoranthene isomersd Benzo[e]pyrene Benzo[a]pyrene Perylene Indeno[1,2,3-c,d]pyrene Dibenz[a,h]anthracene Benzo[g,h,i]perylene
170716 8778 300727 1071 140713 870.3 180715
700765 190717 600755 2172 320730 2070.7 470738
170.1 670.6 2172
Large PAHs Coronene MW 302 isomers MW 326 isomers MW 350 isomers MW 374 isomers
9478 190716 10078 130711 det
260721 1100788 200716 460737 det
570.4
871 270.1
9577 det 1071 1872 871 det
e f f f f g g g g g g,h,i i
Emissions can be expressed in other units using the information in Table 1. a Analytes that were observed with a signal:noise ratio below 10:1 but above 3:1 are listed as detected (det). b (d) Corrected for recovery of d10 phenanthrene, (e) corrected for recovery of d12 chrysene, (f) corrected for recovery of d12 benzo[k]fluoranthene, (g) corrected for recovery of d14 benzo[g,h,i]perylene, (h) quantification based on the relative response of the MW 352 isomer, and (i) identification based on the retention time and mass spectra of an impure standard. c The MW 228 isomers were the sum of chrysene, triphenylene and benz[a]anthracene. d The isomers were the sum of benzo[b]fluoranthene and benzo[k]fluoranthene.
MW from 76% for d12 benzo[k]fluoranthene to 32% for d14 dibenzo[a,i]pyrene. Difficulties in extracting heavy PAHs from SRM#1650 have been widely reported and are believed to be related to the high amount of carbon present in the sample (Paschke et al., 1992; Benner, 1998; Pineiro-Iglesias et al., 2002). The PAH measurements made in the current study agree with NIST certified values for SRM#1650 (Diesel Particulate Matter) with a regression slope of 0.81 and a correlation coefficient of 0.98, as shown in Fig. S1 in the Supplemental Information. Figs S2 and S3 illustrate that measurements of light and heavy (MW 202–278) PAH emissions from gasoline- and diesel-powered motor vehicles in the current study are comparable to those made in previous studies (Rogge et al., 1993a, b; Cadle et al., 1999; Fraser et al., 2002;
Schauer et al., 2002). A detailed discussion of all the comparisons to certified values and previous measurements is provided in the Supplemental Information. Concentrations of PAHs in SRM#1649 determined in the current study are in strong agreement with the values reported by NIST as shown in Figs. 2 and S1-A. Seven of the 12 PAHs identified fall within the given uncertainty ranges of the certified values. Anthracene measurements in the current study are a factor of three higher than the NIST certified values. Additional compounds that did not fall within the uncertainty bounds (fluoranthene, MW 228 isomers, dibenz[a,h]anthracene, and benzo[g,h,i]perylene) had an average error of 2479%. Some of this discrepancy may be related to the small amount of sample extracted. NIST recommends that 1 g of SRM#1649 be analyzed to obtain the certified values, while only 5 mg was extracted in the
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10
Experimental Value
8665
NIST Certified Value
9
Concentration (ug/g)
Heavy PAHs
Light PAHs
8
Large PAHs
7 6 5 4 3 2 1
§ ∗ MW 350 isomers
MW 326 isomers
∗ MW 302 isomers
benzo[ghi]perylene
dibenz[a,h]anthracene
indeno[1,2,3cd]pyrene
perylene
benzo[a]pyrene
benzo[e]pyrene
benzofluoranthene isomers
MW 228 isomers
pyrene
fluoranthene
anthracene
phenanthrene
coronene
∗
0
* No NIST Certified value exists for this compound § Only two isomers (dibenzo[a,h]pyrene and dibenzo[a,i]pyrene) quantified by NIST. Fig. 2. Measured and certified PAH concentrations (mg kg1) in the NIST Urban Dust SRM#1649. Only two of the MW 302 isomers have values reported by NIST. The MW 326 and 350 PAHs are quantified in the SRM #1649 for the first time in the current study.
current study to match the small amount of material present in the emission samples. Only two of the MW 302 isomers have values reported by NIST in this SRM; our measurement includes the sum of all isomers thus it is much larger than the value given by NIST. The far right side of Fig. 2 illustrates concentrations of large PAHs that were detected in the NIST SRM#1649 (Urban Dust) and are reported here for the first time. Approximately 2 (mg g1 PM) were found of MW 326 and 350 PAH isomer groups, with a slightly greater value of 6 (mg g1 PM) detected for the MW 302 isomer group. The pattern of the MW 302, 326, and 350 isomers detected in the Urban Dust SRM qualitatively matches the ratio of these compounds detected in the exhaust of gasoline vehicles equipped with TWCs (see Fig. 1). The source profile information for large PAHs shown in Table 4 combined with the concentration of large PAHs detected in the SRM suggest that each gram of Urban Dust SRM#1649 may contain 5–10 mg
of PM originally emitted from gasoline-powered motor vehicles. 4. Discussion The search for a unique tracer for gasoline engine exhaust vs. diesel engine exhaust is motivated by the detection of coronene (previously suggested as a tracer for gasoline-powered vehicles) in particulate matter emitted from diesel-powered vehicles operating under idle+creep driving cycles (see Table 4). Large PAHs with MW of 326 and 350 Da appear to be uniquely emitted by gasoline-powered vehicles. These large PAH compounds may improve the resolution of source apportionment studies that seek to segregate airborne particulate matter emitted from gasoline-powered vehicles and diesel engine exhaust. It should be noted that the recovery of the larger PAH internal standards declines from samples with a high amount of elemental carbon, such as the diesel truck samples, the diesel
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particulate SRMs and to a lesser degree, the Urban Dust particulate matter SRM. Light and heavy PAHs do not suffer as much from reduced recovery efficiency in the presence of high elemental carbon concentrations. Care must be taken to properly account for recovery efficiency during future source apportionment studies using large PAHs. PAH isomers with MW of 326 and 350 Da are emitted from gasoline-powered motor vehicles at a combined rate of 1–6 (mg L1 fuel burned) depending on the emissions control technology applied to the vehicle. These compounds were quantified in the NIST Urban Dust SRM#1649, proving their abundance in the ambient environment. Although the toxicity of individual MW 326 and 350 PAH isomers are not known, their high potential for mutagenic behavior can be inferred by their structural similarity to smaller, carcinogenic PAHs. Several PAHs are known carcinogens or can form more toxic compounds when metabolized (Mastrangelo et al., 1996; Warshawsky, 1999). The most well known of these compounds is benzo[a]pyrene (5-rings) with high levels of carcinogenic and mutagenic behavior (Bostrom et al., 2002). Recent studies have shown that 6-ring heavy PAHs dibenzo[a,l]pyrene, dibenzo[a,h]pyrene, and dibenzo[a,i]pyrene exhibit much stronger carcinogenic effects than benzo[a]pyrene (Cavalieri et al., 1991, 1988). Allen et al. (1998) identified the size distribution of these toxic PAHs in the atmosphere in Boston, Massachusetts, and another more recent study quantified several of the 84 possible MW302 isomers in different standard reference materials including SRM#1649 (Urban Dust) (Bergvall and Westerholm, 2006). Marvin et al. (2000) identified large PAHs in the sediment of Halmilton Bay Harbor and in the Urban Dust SRM#1649; with accompanying bioassays of these extracts showing the fraction containing the large PAHs exhibiting the greatest toxicity using the Ames Salmonella/microsome assay (Marvin et al., 2000, 1999). The large PAHs with MW of 326 and 350 Da detected in the current study for the first time could pose a significant health risk for populations that are heavily exposed to motor vehicle emissions. Future toxicity measurements are needed to gain more insight into the potential health risks that large PAHs pose to the public. Acknowledgments This research was supported by the California Air Resources Board, Research Division under contracts
#00-318 and #01-306, the National Science Foundation’s Nanophases in the Environment, Agriculture, and Technology-Integrative Graduate Education, Research, and Training (NEAT-IGERT) initiative award #DGE-9972741 and the Coordinating Research Council under project #E-55/59-1.5a. The authors would like to thank Dr. John C. Fetzer (Chevron Research and Technology Company, retired) for the donation of several large PAH standards, the staff at the CARB Hagen-Smit Laboratory in El Monte, CA and the staff at the WVU portable HDDV dynamometer facility in Riverside, CA for their assistance in vehicle testing along with Drs. Roger Atkinson and Janet Arey and the staff of the Air Pollution Research Center at the University of California, Riverside for the use of laboratory and fume hood space during diesel sample collection. Disclaimer: This report was prepared by the University of California at Davis as an account of work partially sponsored by the Coordinating Research Council (CRC). Neither the CRC, members of the CRC, the University of California at Davis nor any person acting on their behalf: (1) makes any warranty, express or implied, with respect to the use of any information, apparatus, method, or process disclosed in this report, or (2) assumes any liabilities with respect to use, or damages resulting from the use or inability to use, any information, apparatus, method, or process disclosed in this report.
Appendix A. Supplementary material Supplementary data associated with this article can be found in the online version at doi:10.1016/ j.atmosenv.2007.07.023.
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Pineiro-Iglesias, M., Lopez-Mahia, P., Vazquez-Blanco, E., Muniategui-Lorenzo, S., Prada-Rodriguez, D., 2002. Problems in the extraction of polycyclic aromatic hydrocarbons from diesel particulate matter. Polycyclic Aromatic Compounds 22 (2), 129–146. Riddle, S.G., Robert, M.A., Jakober, C.A., Hannigan, M.P., Kleeman, M.J., 2007a. Size distribution of trace organic species emitted from heavy duty diesel vehicles. Environmental Science and Technology 41 (6), 1962–1969. Riddle, S.G., Robert, M.A., Jakober, C.A., Hannigan, M.P., Kleeman, M.J., 2007b. Size distribution of trace organic species emitted from light-duty gasoline vehicles. Environmental Science and Technology, submitted for publication. Robert, M.A., Jakober, C.A., Kleeman, M.J., 2007a. Size and composition distribution of particulate matter 1. Light-duty gasoline vehicles. Journal of the Air and Waste Management Association, in press. Robert, M.A., Jakober, C.A., VanBergen, S., Kleeman, M.J., 2007b. Size and composition distribution of particulate matter 2. Heavy-duty diesel vehicles. Journal of the Air and Waste Management Association, in press. Rogge, W.F., Hildemann, L.M., Mazurek, M.A., Cass, G.R., Simoneit, B.R.T., 1993a. Sources of fine organic aerosol. 2. Noncatalyst and catalyst-equipped automobiles and heavyduty diesel trucks. Environmental Science and Technology 27 (4), 636–651. Rogge, W.F., Mazurek, M.A., Hildemann, L.M., Cass, G.R., Simoneit, B.R.T., 1993b. Quantification of urban organic aerosols at a molecular-level-identification, abundance and seasonal-variation. Atmospheric Environment 27 (8), 1309–1330. Schauer, J.J., Rogge, W.F., Hildemann, L.M., Mazurek, M.A., Cass, G.R., Simoneit, B.R.T., 1996. Source apportionment of airborne particulate matter using organic compounds as tracers. Atmospheric Environment 30 (22), 3837–3855. Schauer, J.J., Kleeman, M.J., Cass, G.R., Simoneit, B.R.T., 1999. Measurement of emissions from air pollution sources. 2. C-1 through C-30 organic compounds from medium duty diesel trucks. Environmental Science and Technology 33 (10), 1578–1587. Schauer, J.J., Kleeman, M.J., Cass, G.R., Simoneit, B.R.T., 2002. Measurement of emissions from air pollution sources. 5. C-1–C-32 organic compounds from gasoline-powered motor
vehicles. Environmental Science and Technology 36 (6), 1169–1180. Schauer, C., Niessner, R., Poschl, U., 2003. Polycyclic aromatic hydrocarbons in urban air particulate matter: decadal and seasonal trends, chemical degradation, and sampling artifacts. Environmental Science and Technology 37 (13), 2861–2868. Schmidt, W., Grimmer, G., Jacob, J., Dettbarn, G., Naujack, K.W., 1987. Polycyclic aromatic-hydrocarbons with massnumber 300 and 302 in hard-coal flue-gas condensate—partial identification by spectrometry and independent synthesis. Fresenius Zeitschrift Fur Analytische Chemie 326 (5), 401–413. United States Department of Energy (US DOE), 2004. Documentation for emissions of greenhouse gases in the United States 2002. Office of Integrated Analysis and Forecasting. US Department of Energy, Washington, DC. USEPA, 1988. Second supplement to compendium of methods for the determination of toxic organic compounds in ambient air. Atmospheric Research and Exposure Assessment Laboratory, Research Triangle Park, NC, pp. TO-13–TO-97. Venkataraman, C., Friedlander, S.K., 1994. Size distributions of polycyclic aromatic-hydrocarbons and elemental carbon. 2. Ambient measurements and effects of atmospheric processes. Environmental Science and Technology 28 (4), 563–572. Venkataraman, C., Lyons, J.M., Friedlander, S.K., 1994. Sire distributions of polycyclic aromatic-hydrocarbons and elemental carbon. 1. Sampling, measurement methods, and source characterization. Environmental Science and Technology 28 (4), 555–562. Warshawsky, D., 1999. Polycyclic aromatic hydrocarbons in carcinogenesis. Environmental Health Perspectives 107, 317–319. Westerholm, R.N., Alsberg, T.E., Frommelin, A.B., Strandell, M.E., Rannug, U., Winquist, L., Grigorladis, V., Egeback, K.E., 1988. Effect of fuel polycyclic aromatic hydrocarbon content on the emissions of polycyclic aromatic hydrocarbons and other mutagenic substances from a gasoline-fueled automobile. Environmental Science and Technology 22, 925–930. Zielinska, B., Sagebiel, J., Arnott, W.P., Rogers, C.F., Kelly, K.E., Wagner, D.A., Lighty, J.S., Sarofim, A.F., Palmer, G., 2004. Phase and size distribution of polycyclic aromatic hydrocarbons in diesel and gasoline vehicle emissions. Environmental Science and Technology 38 (9), 2557–2567.
Atmospheric Environment 41 (2007) 8658–8668 www.elsevier.com/locate/atmosenv
Large PAHs detected in fine particulate matter emitted from light-duty gasoline vehicles Sarah G. Riddlea, Chris A. Jakoberb, Michael A. Robertc, Thomas M. Cahillb,1, M. Judith Charlesb,{, Michael J. Kleemanc, a
Department of Chemistry, University of California, 1 Shields Avenue, Davis, CA 95616, USA Department of Environmental Toxicology, University of California, 1 Shields Avenue, Davis, CA 95616, USA c Department of Civil and Environmental Engineering, University of California, 1 Shields Avenue, Davis, CA 95616, USA b
Received 17 February 2007; received in revised form 13 July 2007; accepted 13 July 2007
Abstract Emission factors of large PAHs with 6–8 aromatic rings with molecular weights (MW) of 300–374 were measured from 16 light-duty gasoline-powered vehicles (LDGV) and one heavy-duty diesel-powered vehicle (HDDV) operated under realistic driving conditions. LDGVs emitted PAH isomers of MW 302, 326, 350, and 374, while the HDDV did not emit these compounds. This suggests that large PAHs may be useful tracers for the source apportionment of gasoline-powered motor vehicle exhaust in the atmosphere. Emission rates of MW 302, 326, and 350 isomers from LDGVs equipped with three-way catalysts (TWCs) ranged from 2 to 10 (mg L1 fuel burned), while emissions from LDGVs classified as low emission vehicles (LEVs) were almost a factor of 10 lower. MW 374 PAH isomers were not quantified due to the lack of a quantification-grade standard. The reduced emissions associated with the LEVs are likely attributable to improved vapor recovery during the ‘‘cold-start’’ phase of the Federal Test Procedure (FTP) driving cycle before the catalyst reaches operating temperature. Approximately 2 (mg g1 PM) of MW 326 and 350 PAH isomer groups were found in the National Institute of Standards and Technology standard reference material (SRM)#1649 (Urban Dust). The pattern of the MW 302, 326, and 350 isomers detected in SRM#1649 qualitatively matched the ratio of these compounds detected in the exhaust of TWC LDGVs suggesting that each gram of Urban Dust SRM contained 5–10 mg of PM originally emitted from gasoline-powered motor vehicles. Large PAHs made up 24% of the total LEV PAH emissions and 39% of the TWC PAH emissions released from gasoline-powered motor vehicles. Recent studies have shown certain large PAH isomers have greater toxicity than benzo[a]pyrene. Even though the specific toxicity measurements on PAHs with MW 4302 have yet to be performed, the detection of significant amounts of MW 326 and 350 PAHs in motor vehicle exhaust in the current study suggests that these compounds may pose a significant public health risk. r 2007 Elsevier Ltd. All rights reserved. Keywords: Gasoline vehicle; Diesel vehicle; Source apportionment; Toxicity
Corresponding author. Tel.: +1 530 752 8386; fax: +1 530 752 7872.
E-mail address: [email protected] (M.J. Kleeman). Present address: Department of Integrated Natural Sciences, Arizona State University at the West Campus, P.O. Box 37100, Phoenix, AZ 85069, USA. { Deceased. 1
1352-2310/$ - see front matter r 2007 Elsevier Ltd. All rights reserved. doi:10.1016/j.atmosenv.2007.07.023
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1. Introduction Internal combustion engines in motor vehicles generate fine particles in their exhaust with diameters o2.5 mm (see for example, Hildemann et al., 1991; Ondov and Wexler, 1998; Kleeman et al., 2000). Composition plays a role in determining the ecosystem impacts and human health effects of these particles. Polycyclic aromatic hydrocarbons (PAHs) are one class of compounds present in both diesel and gasoline motor vehicle exhaust known to contribute to the adverse health effects of vehicular emissions (EPA, 2000). PAHs can originate from unburned fuel (Westerholm et al., 1988; Marr et al., 1999) and lubricating oil (Brandenberger et al., 2005) or they can be formed during the combustion process from simpler organic compounds (Westerholm et al., 1988). Due to their low vapor pressures, PAHs with five or more aromatic rings rapidly condense onto particles once emitted into the atmosphere resulting in a higher proportion of these PAHs in smaller, respirable size particles (Venkataraman and Friedlander, 1994; Venkataraman et al., 1994). Due to their toxicity and atmospheric prevalence, 16 PAHs ranging in size from naphthalene (2-rings, molecular weight (MW ¼ 128) to benzo[g,h,i]perylene (6-rings, MW ¼ 276) have been identified as priority pollutants by the U.S. Environmental Protection Agency (USEPA, 1988). Most of the literature on PAH concentrations from both source and ambient air samples focus on these 16 EPA priority compounds (Benner et al., 1989; Rogge et al., 1993a, b; Allen et al., 1996; Miguel et al., 1998; Cadle et al., 1999; Marr et al., 1999; Schauer et al., 1999, 2002, 2003; Blanchard et al., 2002; Dyke et al., 2003; Zielinska et al., 2004). When additional compounds are examined, they are typically isomers or of similar MW. Larger PAHs (MW4300) exist in the environment (Marvin et al., 1999), but are seldom identified in source and/or environmental samples even though several of these compounds have shown equal if not greater carcinogenic activity when compared with smaller PAHs (Schmidt et al., 1987; Cavalieri et al., 1991, 1988). This is mostly due to the lack of commercially available standards for larger PAHs with MW4302 Da. Coronene (C24H12) has been proposed as potential tracer for gasoline motor vehicle exhaust (Schauer et al., 1996; Cass, 1998) but this compound was recently detected in emissions from heavy-duty diesel vehicles operated under low speed driving
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conditions (Riddle et al., 2007a). Measurements of even larger PAHs in both source and ambient aerosols would further aid source apportionment of ambient particulate matter. In the present study, emission factors are presented for PAHs with MW up to 350 for light-duty gasoline-powered vehicles (LDGV) and heavy-duty diesel-powered vehicles (HDDV) operated under realistic driving conditions. The concentrations of these compounds were also quantified in the National Institute of Standards and Technology (NIST) Urban Dust standard reference material (SRM)#1649. The direct potential toxicity of PAH compounds with MW4300 Da and their usefulness as tracers for gasoline engine exhaust are discussed in the following sections. 2. Methods 2.1. Light-duty gasoline-powered vehicles (LDGV) emissions collection conditions LDGV emission samples were collected at the Haagen-Smit Laboratory in El Monte, CA, during August–September 2002. The emissions sampling conducted in this study utilized an experimental design similar to that described by Schauer et al. (2002). Briefly, vehicles were operated on a stationary chassis dynamometer using the Federal Test Procedure (FTP) driving cycle. Dynamometer resistance was specified to simulate appropriate vehicle inertial load and wind drag. Multiple vehicles within a technology class were composited on one set of sampling media to obtain sufficient mass for chemical analysis. The LDGV technology classes reported in the current study include low emission vehicles (LEV) and three-way catalyst (TWC) equipped vehicles. A complete description of the vehicles tested in the current study is provided in the upper portion of Table S1 in the Supplemental Information. Measured PM1.8 emissions rates and the estimated fuel economies for each vehicle category are shown in Table 1. Thorough details of the emission collection and dilution methods for the LDGV vehicle matrix are discussed by Robert et al. (2007a). 2.2. Heavy-duty diesel-powered vehicle (HDDV) emissions collection conditions HDDV emissions were collected from a 1999 Freightliner truck (see description in Table S1) in
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Table 1 Measured PM1.8 emissions rates and estimated fuel economy for test vehicles used in the current study Vehicle class— driving cycle LDGV LEV— FTP LDGV TWC— FTP HDDV— Idle+Creep HDDV— HHDDT 5-mode
PM1.8 emissions rate (mg km1)
Estimated fuel economy (km L1)
370
9.3
910
10.0
1,500,000 220,000
0.51 2.3
Riverside, CA, during June–July 2003. Emissions from this vehicle were comparable to emissions from three additional heavy-duty diesel vehicles that were tested at the same time (Riddle et al., 2007a; Robert et al., 2007b). The heavy-duty diesel vehicle was driven through a 5-mode heavy heavy-duty diesel truck (HHDDT) transient driving cycle consisting of a 30 min idle, 17 min creep, 11 min transient stage and two cruise stages of 34 and 31 min, with a top speed of 65 mile h1 for the second cruise (Gautam et al., 2002). In addition to the full HHDDT test cycle, emissions were sampled while the truck was operating under only the idle and creep modes of the HHDDT cycle, hereafter referred to as idle+creep. The measured PM1.8 emissions rate and estimated fuel economy for both tests are summarized in Table 1. The inertial weight used for all heavy-duty diesel tests was 56,000 lbs. Only a few heavy-duty dynamometers in the United States were capable of simulating transient driving cycles using such large inertial weight at the time that the tests were conducted. All HDDV results reported in the current study used the mobile chassis dynamometer operated by West Virginia University. A more detailed description of the testing conditions is provided by Robert et al. (2007b). 2.3. Sample collection Tailpipe emissions from both LDGVs and HDDVs were diluted in two stages and then aged for approximately 30–60 s using a dilution sampling system (Hildemann et al., 1989). Dilution ratios were set to balance the need to reproduce the high dilution factors that occur in the atmosphere against the need to concentrate the sample sufficiently for chemical analysis. Typical dilution ratios ranged
from 125 to 583 (Robert et al., 2007a, b). The emissions were collected after dilution using an annular denuder-filter-polyurethane foam (PUF) sampling train similar to that described by Schauer et al. (1999, 2002), with specific details described by Jakober et al. (2006, 2007). The majority of the sampling hardware and PUF substrates were obtained from University Research Glassware (Chapel Hill, NC). Quartz fiber filters, 47 mm, were obtained from Pall Gelman (Ann Arbor, MI) and baked at 550 1C for a period of at least 12 h to minimize any organic contaminants prior to each testing event. Eight-channel annular denuders were coated with XAD-4 polystyrene resin using a procedure similar to that described by Gundel et al. (1995) and Gundel and Lane (1999). After collection, particle-phase filter samples were stored in baked glass Petri dishes that were covered with baked aluminum foil and wrapped with Teflon tape prior to placement in a desiccator, under organicfree nitrogen and stored at 20 1C until solvent extraction. 2.4. Chemical extraction procedures Quartz filter PM2.5 samples were extracted individually. Filters were spiked with isotopically labeled PAHs and the solvent was allowed to dry before proceeding. Substrates were then placed into screw-cap centrifuge tubes for organic solvent extraction. Each tube was filled with 15 mL of a 1:1 mixture of hexane (Burdick and Jackson trace analysis grade) and dichloromethane (Burdick and Jackson trace analysis grade) then suspended in an ultrasonic cleaning bath for 15 min. The sonication extraction procedure was repeated three times and all of the extracts were combined. Samples were concentrated by nitrogen evaporation to a final volume of 200 mL. For each batch of extracted samples, a method blank was also collected and NIST SRMs#1649 (Urban Dust) and #1650 (Diesel Particulate Matter) were simultaneously extracted to provide further confidence in the methods. 2.5. GC– MS data acquisition methods and data conversion The organic chemical speciation data collected for this project was obtained on a Varian 3400 gas chromatograph (GC) coupled with a Varian 2000 ion-trap mass spectrometer (GC–ITMS). Chemical separation of PAHs was performed on a J&W
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DB-5HT high-temperature capillary column (30 m 0.25 mm i.d., 0.1 mm film thickness, 5% phenylsubstituted polysiloxane, Agilent Technologies, Palo Alto, CA) with helium as the carrier gas at a linear velocity of 37 cm s1. Samples were introduced through a temperature programmable injection port that was held at the initial temperature of 64 1C for 1 min, then ramped at a rate of 20 1C min1 to a temperature of 120 1C, the rate of heating was then increased to 100 1C min1 to a final injector temperature of 375 1C where it was held until the end of the column program. The column was initially at a temperature of 64 1C for the first 7 min to allow the analytes to pass through the injection port and become cryo-focused on the front-end of the analytical column. The column oven temperature was then increased at a rate of 5 1C min1 to a final temperature of 400 1C where it was held for 5 min for a total run time of 79.2 min. The ion trap oven, manifold, and transfer line were operated at 220, 80, and 300 1C, respectively. Analyte identification and quantification was conducted by electron ionization mass spectrometry– mass spectrometry (EI–MS2) analysis. In all cases, the parent ion was the molecular ion for the compound while the product ion was generally [M-2]+, which corresponds to a loss of H2. The mass/charge range monitored during analysis was 100–420. Elution times and target ions for each
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PAH isomer family are shown in Table 2. Analysis data files were processed using Varian Saturn GC–MS Workstation software version 5.51, with chromatographic peak integrations performed manually. The methods used to generate vehicle emission rates are discussed in detail elsewhere (Jakober et al., 2007; Robert et al., 2007a, b) and are summarized here. Briefly, the collected mass measured by GC–ITMS analysis is converted to sampled concentration using six point calibration curves. The sampled concentration was corrected for any background material that was measured to be present in the dilution air used during each vehicle test. The corrected sampled concentration was converted into emitted mass using the ratio of sampled flow to total flow while accounting for primary and secondary dilutions. Emissions can be expressed in units of (mg km1) or (mg L1 fuel burned). Fuel consumption was calculated from measured CO2 emissions using previously reported conversion factors of 2.28 and 2.77 kg CO2 L1 of combusted gasoline and diesel fuel, respectively, were applied to the CO2 measurements (United States Department of Energy, 2004). Appropriate quantification and treatment of recovery efficiencies is a critical step in the measurement of PAH emissions rates from motor vehicles. All results in the current manuscript were
Table 2 GC–ITMS analysis protocol for PAHs Time segment
Elution window (min)
Targeted compounds
Molecular ion (m/z)
Resonant excitation energya (V)
Quantification ionb (m/z)
1 2 3
25.0–30.0 30.0–36.0 36.0–42.0
4 5
42.0–48.0 48.0–52.5
6
52.5–57
7 8
57.0–62.0 62.0–67.0
9
67.0–79.2
MW 178 isomers MW 202 isomers Cyclopenta[c,d]pyrene MW 228 isomers MW 252 isomers MW 276 isomers Dibenz[a,h]anthracene Coronene MW 302 isomers MW 326 isomers MW 350 isomersc MW 352 isomers MW 374 isomersc MW 376 isomersc
178 202 226 228 252 276 278 300 302 326 350 352 374 376
1.5 1.6 1.8 1.6 1.5 1.6 1.6 1.6 1.6 1.6 1.6 1.6 1.6 1.6
152 200 224 226 250 274 276 298 300 324 348 350 372 374
a
The optimal energy was the excitation energy that gave the greatest intensity of a product ion. Most PAHs lose two hydrogen atoms to form the quantification ion, one exception are the MW 178 isomers which lose a –C2H2-group to form the quantification ion. c Elution times for these isomers were based on those of an impure standard, quantification could not be directly performed for these analytes. b
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efficiency of only 3% was observed for dibenzo[a,i]pyrene in the heavy-duty diesel sample operated under the full 5-mode HHDDT driving cycle. All other compounds were recovered with better than 50% efficiency in all tests
corrected for recovery efficiency by dividing the measured PAH concentration by the measured recovery efficiency of the surrogate compound with a similar structure. Recoveries of the surrogate internal standards (IS) spiked onto filters prior to solvent extraction are displayed in Table 3. The obtained IS recoveries are consistent with our expectations based on previous observed recovery of model compounds (data not presented). Generally speaking, recovery of PAH compounds depends on the MW of the PAH and the amount of elemental carbon present in the sample. A recovery
3. Results 3.1. LDGV and HDDV PAH emission factors Mass emissions rates of PAHs (mg PAH L1 fuel burned) for LDGVs and HDDVs vehicles are
Table 3 Recovery of spiked internal standards used to correct emissions from low emission vehicle (LEV FTP), three-way catalyst vehicle (TWC FTP), heavy-duty diesel vehicle idle+creep (HDDV Idle/Creep) and heavy-duty diesel vehicle 5-mode (HDDV 5-mode) tests Compound
d10 d12 d12 d12 d14
Internal standard recovery (%)
Phenanthrene Chrysene Benzo[k]fluoranthene Benzo[g,h,i]perylene Dibenzo[a,i]pyrene
LEV FTP
TWC FTP
HDDV Idle/Creep
HDDV 5-mode
67 93 87 96 99
22 82 86 102 108
62 107 94 100 59
42 82 64 38 3
Compounds were quantified using d10-pyrene as the internal standard. LEV
TWC
99 Frtlnr Idle/Creep
Light PAHs
99 Frtlnr 56 K
Heavy PAHs
90
16 Large PAHs
70
14
0
0 MW 350 isomers
2 MW 326 isomers
10 MW 302 isomers
4
coronene
20
benzo[ghi]perylene
6
indeno[1,2,3cd]pyrene
30
benzo[a]pyrene
8
benzo[e]pyrene
40
benzofluoranthene isomers
10
MW 228 isomers
50
pyrene
12
fluoranthene
60
phenanthrene
Light PAH Emission Factor (ug/L of fuel burned)
80
Heavy and Large PAH Emission Factor (ug/L of fuel burned)
18
Fig. 1. PAH emissions factors (mg L1 fuel burned) for low emission gasoline-powered vehicles (LEV), three-way catalyst gasolinepowered vehicles (TWC), heavy-duty diesel vehicle operated under idle+creep (99 Frtlnr Idle/Creep) and heavy-duty diesel vehicle operated under 5-mode transient cycle (99 Frtlnr 56 K). Note that the scale for the light PAH measurements is given on the left and the scale for the heavy and large PAHs is given on the right side of the figure.
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presented in Fig. 1. The uncertainty range on each emission factor is based on the %RSD of a similar compound obtained in the model analyte recovery experiment (data not presented). This error does not fully capture the uncertainty associated with sample collection and should be interpreted as a lower bound on uncertainty for the emissions rate. Emissions of light and heavy PAHs from these vehicles have been reported previously (Riddle et al., 2007a, b) while emissions of large PAHs (MW4302) described in this paper are reported for the first time based on the methods described above. Fig. 1 clearly shows that diesel vehicles operating under the full 5-mode HHDDT driving cycle emit the greatest quantity of light PAHs per unit of fuel burned. Engine load had a strong effect on PAH emissions from diesel vehicles, with greater quantities of light PAHs emitted during the full HHDDT driving cycle and greater amounts of heavy PAHs emitted during the idle+creep cycle. The higher engine loads and hotter combustion temperatures in the full 5-mode HHDDT driving cycle appear to break down the heavy PAHs. Heavy-duty diesel vehicles did not emit PAHs with MW 4300 Da (coronene) even under the lightest driving cycle tested (idle+creep). Emissions of light PAHs per unit of fuel consumed by gasoline-powered vehicles were much lower than emissions from heavy-duty diesel vehicles. Emissions of heavy PAHs per unit of fuel consumed were comparable from light-duty gasoline and heavy-duty diesel vehicles. The greatest difference between the vehicle classes was apparent in the large PAH emissions. Gasoline-powered vehicles emitted MW 302, 326, and 350 PAH isomers while heavy-duty diesel vehicles did not emit these compounds. MW 374 PAH isomers were also detected in particles emitted from gasolinepowered vehicles but no pure standard was available for quantification. Emission rates of MW 302, 326, and 350 isomers from gasoline-powered vehicles equipped with TWCs ranged from 2 to 10 (mg L1 fuel burned) while emissions from LEVs were almost a factor of 10 lower. The reduced emissions associated with the LEVs are likely attributable to improved vapor recovery during the ‘‘cold-start’’ phase of the FTP driving cycle before the catalyst reaches operating temperature. Large PAHs with MW of 326 and 350 Da made up 24% of the total LEV PAH emissions and 39% of the TWC PAH emissions. Besides the potential direct toxicity of these compounds, large PAHs with MW4302 Da appear
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to be good candidates for the source apportionment of gasoline-powered motor vehicle exhaust. Coronene has previously been suggested as a marker for gasoline engine exhaust (Schauer et al., 1996; Cass, 1998), but the detection of coronene in the exhaust from diesel-powered vehicles operated under the idle+creep driving cycle highlights the need to quantify large PAHs that are possibly unique to gasoline engines. Relative concentrations of PAHs (mg PAH g1 total PM) in the exhaust of the motor vehicles tested in the current study are summarized in Table 4. These units are useful since they can be directly used in source apportionment studies. Total PM concentrations were measured by Robert et al. (2007a, b) using gravimetric methods. A comparison between Fig. 1 and Table 4 shows that emissions of PAHs from diesel vehicles are generally similar when expressed in either (mg PAH L1 fuel burned) or (mg PAH g1 PM). In contrast, emissions from gasoline vehicles are several orders of magnitude larger when they are normalized by the amount of particulate matter emitted rather than the amount of fuel burned. This reflects the fact that gasolinepowered vehicles emit significantly less particulate matter per liter of fuel burned than heavy-duty diesel vehicles. When PAHs were detected in the diesel exhaust, their relative concentrations were at least 2–3 times lower than analogous concentrations in light-duty gasoline-powered vehicle exhaust. Gasoline-powered vehicles equipped with TWCs produced higher concentrations of large PAHs per unit of total PM emitted than LEVs despite the fact that total PM emissions from TWCs are nearly 2.5 times greater. 3.2. Analysis of NIST SRMs and comparisons with previous studies NIST SRMs#1649 (Urban Dust) and #1650 (Diesel Particulate Matter) were used to monitor the accuracy of the chemical analysis methods during the extraction of PM emission samples. Five and two milligrams NIST SRM#1649 and #1650 were solvent extracted, respectively, using the procedure described above. Internal standard recoveries (Table S3 in the Supplemental Information section) ranged between 80% and 109% for SRM#1649 (Urban Dust) and 32% and 109% for SRM#1650 (Diesel Particulate Matter). Recoveries of light PAHs (those up to and including d12 chrysene) were all 490%; recoveries decreased with
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Table 4 PAH emission factors normalized by total PM emissions rates (mg PAH g1 PM) for gasoline-powered low emission vehicles (LEVs), three-way catalyst vehicles (TWC), heavy-duty diesel vehicles operated under idle+creep driving conditions (HDDV Idle/Creep) and heavy-duty diesel vehicles operated under the full HHDDT 5-mode driving cycle (HDDV 5-mode) Compound
Emission rate (mg PAH g1 PM)a LEV FTP
Light PAHs Phenanthrene Anthracene Fluoranthene Pyrene Cyclopenta[c,d]pyrene MW 228 isomersc
320737 3975 110711 230723 3874
HDDV Idle/Creep
HDDV 5-mode
b
240728
4075
150717
8178 160717 3774 350726
670.6 1572
7877 110711
d d e e e e
TWC FTP
det
Heavy PAHs Benzofluoranthene isomersd Benzo[e]pyrene Benzo[a]pyrene Perylene Indeno[1,2,3-c,d]pyrene Dibenz[a,h]anthracene Benzo[g,h,i]perylene
170716 8778 300727 1071 140713 870.3 180715
700765 190717 600755 2172 320730 2070.7 470738
170.1 670.6 2172
Large PAHs Coronene MW 302 isomers MW 326 isomers MW 350 isomers MW 374 isomers
9478 190716 10078 130711 det
260721 1100788 200716 460737 det
570.4
871 270.1
9577 det 1071 1872 871 det
e f f f f g g g g g g,h,i i
Emissions can be expressed in other units using the information in Table 1. a Analytes that were observed with a signal:noise ratio below 10:1 but above 3:1 are listed as detected (det). b (d) Corrected for recovery of d10 phenanthrene, (e) corrected for recovery of d12 chrysene, (f) corrected for recovery of d12 benzo[k]fluoranthene, (g) corrected for recovery of d14 benzo[g,h,i]perylene, (h) quantification based on the relative response of the MW 352 isomer, and (i) identification based on the retention time and mass spectra of an impure standard. c The MW 228 isomers were the sum of chrysene, triphenylene and benz[a]anthracene. d The isomers were the sum of benzo[b]fluoranthene and benzo[k]fluoranthene.
MW from 76% for d12 benzo[k]fluoranthene to 32% for d14 dibenzo[a,i]pyrene. Difficulties in extracting heavy PAHs from SRM#1650 have been widely reported and are believed to be related to the high amount of carbon present in the sample (Paschke et al., 1992; Benner, 1998; Pineiro-Iglesias et al., 2002). The PAH measurements made in the current study agree with NIST certified values for SRM#1650 (Diesel Particulate Matter) with a regression slope of 0.81 and a correlation coefficient of 0.98, as shown in Fig. S1 in the Supplemental Information. Figs S2 and S3 illustrate that measurements of light and heavy (MW 202–278) PAH emissions from gasoline- and diesel-powered motor vehicles in the current study are comparable to those made in previous studies (Rogge et al., 1993a, b; Cadle et al., 1999; Fraser et al., 2002;
Schauer et al., 2002). A detailed discussion of all the comparisons to certified values and previous measurements is provided in the Supplemental Information. Concentrations of PAHs in SRM#1649 determined in the current study are in strong agreement with the values reported by NIST as shown in Figs. 2 and S1-A. Seven of the 12 PAHs identified fall within the given uncertainty ranges of the certified values. Anthracene measurements in the current study are a factor of three higher than the NIST certified values. Additional compounds that did not fall within the uncertainty bounds (fluoranthene, MW 228 isomers, dibenz[a,h]anthracene, and benzo[g,h,i]perylene) had an average error of 2479%. Some of this discrepancy may be related to the small amount of sample extracted. NIST recommends that 1 g of SRM#1649 be analyzed to obtain the certified values, while only 5 mg was extracted in the
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10
Experimental Value
8665
NIST Certified Value
9
Concentration (ug/g)
Heavy PAHs
Light PAHs
8
Large PAHs
7 6 5 4 3 2 1
§ ∗ MW 350 isomers
MW 326 isomers
∗ MW 302 isomers
benzo[ghi]perylene
dibenz[a,h]anthracene
indeno[1,2,3cd]pyrene
perylene
benzo[a]pyrene
benzo[e]pyrene
benzofluoranthene isomers
MW 228 isomers
pyrene
fluoranthene
anthracene
phenanthrene
coronene
∗
0
* No NIST Certified value exists for this compound § Only two isomers (dibenzo[a,h]pyrene and dibenzo[a,i]pyrene) quantified by NIST. Fig. 2. Measured and certified PAH concentrations (mg kg1) in the NIST Urban Dust SRM#1649. Only two of the MW 302 isomers have values reported by NIST. The MW 326 and 350 PAHs are quantified in the SRM #1649 for the first time in the current study.
current study to match the small amount of material present in the emission samples. Only two of the MW 302 isomers have values reported by NIST in this SRM; our measurement includes the sum of all isomers thus it is much larger than the value given by NIST. The far right side of Fig. 2 illustrates concentrations of large PAHs that were detected in the NIST SRM#1649 (Urban Dust) and are reported here for the first time. Approximately 2 (mg g1 PM) were found of MW 326 and 350 PAH isomer groups, with a slightly greater value of 6 (mg g1 PM) detected for the MW 302 isomer group. The pattern of the MW 302, 326, and 350 isomers detected in the Urban Dust SRM qualitatively matches the ratio of these compounds detected in the exhaust of gasoline vehicles equipped with TWCs (see Fig. 1). The source profile information for large PAHs shown in Table 4 combined with the concentration of large PAHs detected in the SRM suggest that each gram of Urban Dust SRM#1649 may contain 5–10 mg
of PM originally emitted from gasoline-powered motor vehicles. 4. Discussion The search for a unique tracer for gasoline engine exhaust vs. diesel engine exhaust is motivated by the detection of coronene (previously suggested as a tracer for gasoline-powered vehicles) in particulate matter emitted from diesel-powered vehicles operating under idle+creep driving cycles (see Table 4). Large PAHs with MW of 326 and 350 Da appear to be uniquely emitted by gasoline-powered vehicles. These large PAH compounds may improve the resolution of source apportionment studies that seek to segregate airborne particulate matter emitted from gasoline-powered vehicles and diesel engine exhaust. It should be noted that the recovery of the larger PAH internal standards declines from samples with a high amount of elemental carbon, such as the diesel truck samples, the diesel
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particulate SRMs and to a lesser degree, the Urban Dust particulate matter SRM. Light and heavy PAHs do not suffer as much from reduced recovery efficiency in the presence of high elemental carbon concentrations. Care must be taken to properly account for recovery efficiency during future source apportionment studies using large PAHs. PAH isomers with MW of 326 and 350 Da are emitted from gasoline-powered motor vehicles at a combined rate of 1–6 (mg L1 fuel burned) depending on the emissions control technology applied to the vehicle. These compounds were quantified in the NIST Urban Dust SRM#1649, proving their abundance in the ambient environment. Although the toxicity of individual MW 326 and 350 PAH isomers are not known, their high potential for mutagenic behavior can be inferred by their structural similarity to smaller, carcinogenic PAHs. Several PAHs are known carcinogens or can form more toxic compounds when metabolized (Mastrangelo et al., 1996; Warshawsky, 1999). The most well known of these compounds is benzo[a]pyrene (5-rings) with high levels of carcinogenic and mutagenic behavior (Bostrom et al., 2002). Recent studies have shown that 6-ring heavy PAHs dibenzo[a,l]pyrene, dibenzo[a,h]pyrene, and dibenzo[a,i]pyrene exhibit much stronger carcinogenic effects than benzo[a]pyrene (Cavalieri et al., 1991, 1988). Allen et al. (1998) identified the size distribution of these toxic PAHs in the atmosphere in Boston, Massachusetts, and another more recent study quantified several of the 84 possible MW302 isomers in different standard reference materials including SRM#1649 (Urban Dust) (Bergvall and Westerholm, 2006). Marvin et al. (2000) identified large PAHs in the sediment of Halmilton Bay Harbor and in the Urban Dust SRM#1649; with accompanying bioassays of these extracts showing the fraction containing the large PAHs exhibiting the greatest toxicity using the Ames Salmonella/microsome assay (Marvin et al., 2000, 1999). The large PAHs with MW of 326 and 350 Da detected in the current study for the first time could pose a significant health risk for populations that are heavily exposed to motor vehicle emissions. Future toxicity measurements are needed to gain more insight into the potential health risks that large PAHs pose to the public. Acknowledgments This research was supported by the California Air Resources Board, Research Division under contracts
#00-318 and #01-306, the National Science Foundation’s Nanophases in the Environment, Agriculture, and Technology-Integrative Graduate Education, Research, and Training (NEAT-IGERT) initiative award #DGE-9972741 and the Coordinating Research Council under project #E-55/59-1.5a. The authors would like to thank Dr. John C. Fetzer (Chevron Research and Technology Company, retired) for the donation of several large PAH standards, the staff at the CARB Hagen-Smit Laboratory in El Monte, CA and the staff at the WVU portable HDDV dynamometer facility in Riverside, CA for their assistance in vehicle testing along with Drs. Roger Atkinson and Janet Arey and the staff of the Air Pollution Research Center at the University of California, Riverside for the use of laboratory and fume hood space during diesel sample collection. Disclaimer: This report was prepared by the University of California at Davis as an account of work partially sponsored by the Coordinating Research Council (CRC). Neither the CRC, members of the CRC, the University of California at Davis nor any person acting on their behalf: (1) makes any warranty, express or implied, with respect to the use of any information, apparatus, method, or process disclosed in this report, or (2) assumes any liabilities with respect to use, or damages resulting from the use or inability to use, any information, apparatus, method, or process disclosed in this report.
Appendix A. Supplementary material Supplementary data associated with this article can be found in the online version at doi:10.1016/ j.atmosenv.2007.07.023.
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