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Environ. Sci. Technol. 2006, 40, 4880-4886

Aerosol-Borne Quinones and Reactive Oxygen Species Generation by Particulate Matter Extracts MYEONG Y. CHUNG, RICK A. LAZARO,† DIANNE LIM, JOSCELYNE JACKSON,‡ JULIE LYON, DORA RENDULIC, AND ALAM S. HASSON* Department of Chemistry, 2555 East San Ramon Avenue M/S SB70, California State University Fresno, Fresno, California 93711

The mass loadings of quinones and their ability to generate reactive oxygen species (ROS) were investigated in total suspended particulate samples collected in Fresno, CA, over a 12-month period. Particles were collected on Teflon filters and were analyzed for the presence of 12 quinones containing one to four aromatic rings by gas chromatography with mass spectrometry. Measured levels are generally greater than mass loadings reported at other locations. The mass loadings were highest during winter months and were strongly anticorrelated with temperature. ROS generation was investigated by measuring the rate of hydrogen peroxide production from the reaction of laboratory standards and ambient samples with dithiothreitol (DTT). ROS generation from ambient samples shows a strong positive correlation with the mass loadings of the three most reactive quinones and may account for all of the ROS formed in the DTT test.

Introduction The adverse health affects associated with elevated levels of particulate matter (PM) have been widely documented (17) and have sparked intensive efforts to understand their origins. Despite the best attempts of researchers, a complete understanding of the causes of these detrimental health effects has proved elusive. Correlations have been made with various parameters including total PM mass loading (6, 8), surface area (9-12), and chemical composition. Studies have documented the potential health effects of an array of chemical constituents of PM, many of which have focused on the role of organics such as polyaromatic hydrocarbons (PAH) (9-12), oxygenated PAH (OPAH) (13, 14), and nitroPAHs (15, 16). Several researchers have investigated the link between reactive oxygen species (ROS) and the potential health impacts of particles (17, 18). ROS (oxygen-centered free radicals and their metabolites, such as hydrogen peroxide, H2O2) may either be present within the ambient particles themselves or the particles may contain chemical species that are capable of generating ROS upon deposition within the lung. Once present, the ROS may be converted to hydroxy radicals (OH) under biological conditions that may then * Corresponding author phone: (559)278-2420; fax: (559)278-4402; e-mail: [email protected]. † Now at Department of Chemistry, University of California, Davis. ‡ Now at School of Pharmacy, University of California, San Francisco. 4880

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initiate cell injury. Studies investigating in-vitro effects of ambient particulate matter and diesel exhaust particles are consistent with ROS as a mediator of cell damage (19-21). While the potential toxicity of aerosol particles is generally measured by in-vivo or in-vitro tests, Li et al. (22) have reported a chemical assay involving the measurement of dithiothreitol (DTT) consumption from its reaction with particle components such as quinones, which generates ROS (23) as shown in Scheme 1. Li et al. report that the data from this DTT assay exhibit a linear relationship (R2 > 0.9) with the HO-1 (cellular heme oxygenase 1) immunoblotting assay used to investigate ROS production from primary bronchial epithelial cells. Field measurements have shown that ambient particulate matter may contain mass loadings of ROS about 101 ng m-3 (24-27), which is high enough to cause lung cell damage (28). Other work has focused on the role of persistent free radicals, which may be present within particulate matter at levels as high as 1017 radicals/g and which appear to be stable indefinitely (19). These are believed to include oxygencentered free radicals such as semiquinone radicals. Despite being present at lower levels than ROS in ambient PM, semiquinone radicals may ultimately be more damaging to cells because of their ability to catalytically generate hydrogen peroxide in biological environments via redox cycling between semiquinone radicals, hydroquinones, and quinones (19, 20, 29). Hydrogen peroxide may then be converted to OH radicals (which are far more damaging to cells than H2O2) in the presence of Fe2+ via the Fenton reaction. Fe3+, which is formed in the Fenton reaction, can then regenerate Fe2+ by oxidizing hydroquinones to quinones. The amount of iron present in the particles thus impacts the potential toxicity of the particles by converting a less reactive ROS species (hydrogen peroxide) to a more reactive form (OH). However, the 1:1 stoichiometry of the Fenton reaction means that the presence of iron does not directly affect the total quantity of ROS present. UV photolysis is also an important source of radicals in the atmosphere but not in biological environments. The effects of quinones and hydroquinones present in PM may exceed those of both ROS and persistent free radicals. Since these species undergo the same redox cycling reactions as semiquinone radicals, they may generate disproportionately large quantities of ROS in biological environments. Further, they are likely to be present within ambient particles at significantly higher levels than persistent free radicals. Numerous field studies have attempted to catalog the diverse range of organic compounds found in ambient aerosols. Despite analyzing dozens of compounds, the identified organics generally comprise only a few percent of the total organic mass present. While the lists of target compounds in these studies often include a few quinones, field measurements specifically targeting these compounds are sparse in the literature. Quinone mass loadings measured in some studies of OPAH are summarized in Table 1. These data were generally obtained by gas chromatography with mass spectrometry (GC-MS) analysis of filter samples, often following separation of the organic components into fractions (nonpolar, semipolar, etc.) by high-performance liquid chromatography (HPLC). Some additional measurements have been performed using less traditional techniques such as laser desorption/resonance enhanced multiphoton ionization followed by mass spectrometry (30, 31) and liquid chromatography with mass spectrometry (LC-MS) (32). Recently, Cho et al. (33) reported a new derivatization technique for the analysis of quinones by GC-MS. The method involves 10.1021/es0515957 CCC: $33.50

 2006 American Chemical Society Published on Web 07/19/2006

SCHEME 1. Formation of Hydrogen Peroxide in the Reaction of Dithiothreitol with Quinones

TABLE 1. Summary of Some Previous Quinone Mass Loading Measurements quinone 1,2-naphthoquinone 1,4-naphthaquinone acenaphthenequinone phenanthrenequinone 5,12-naphthacenequinone 9,10-anthraquinone

methyl anthraquinone dimethyl anthraquinone benz[a]-9,10-anthraquinone benz[a]-7,12-anthraquinone chrysenequinone benzo[a]pyrene-6,12-quinone benzo[a]pyrene-3,6-quinone benzo[a]pyrene-1,6-quinone a

measured concentration/ng m-3 reference 0-0.06 n.s.a 0-0.17 0-7.1 0.427 0-0.57 0.323 0.11-0.54 n.s. 0.1-6.2 0.22-1.89 0-1.14 0.59 0.02-0.20 0.42-5.4 0.07-0.50 0-0.24 n.s. n.s. 0.29-1.05 0.042-0.13 0-0.31 n.s. 0-1.2 0-18 0.0096 0-0.05 0-0.25 0-0.22

33 44 33 34 44 33 44 45 46 35 45 37 47 33 34 45 37 46 46 45 36 37 46 34 34 44 32 32 32

n.s. ) not specified.

the conversion of the carbonyl moieties to their diacetylated derivatives by reaction with acetyl acetate in the presence of zinc. The authors report improvements in detection limits of 1-2 orders of magnitude for 1,2-naphthoquinone, 1,4naphthoquinone, and phenanthraquinone. Reported quinone mass loadings are typically well below 1 ng m-3, and the species are predominantly found in fine particles (Table 1). A California Air Resources Board study (34) measured mass loadings for over 100 organic compounds, including four quinones, in Fresno during 20002001. The quinone levels reported are significantly higher than data from other locations. Mass loadings of acenaphthenequinone, anthraquinone, and benzo[a]anthraquinone are highest during winter, while very high levels of chrysenequinone are reported for summer. Literature data do

not clearly show if aerosol-phase quinones are mostly primary or secondary in origin. Two studies (35, 36) report a correlation between OPAH levels and NOx and O3 concentrations at sites in the United States and North Africa. These data are consistent with a significant secondary source of OPAH from photooxidation processes. On the other hand, Fraser et al. (37) measured OPAH levels that were lower downwind of the primary PAH source and noted that OPAH levels did not increase during the afternoon when photochemical oxidation processes occur most rapidly. The authors concluded that levels of many of the OPAH compounds detected were consistent with known emission rates from primary sources. The San Joaquin valley in central California is a region with poor air quality. The air basin is officially classified by the Environmental Protection Agency (EPA) as being in extreme non-attainment for ozone and serious non-attainment for PM. Incidences of respiratory-related diseases are also high among valley residents. For example, the childhood asthma rate is among the highest in California (38, 39). While the levels of criteria pollutants are high in central California, the air quality characteristics are similar to other regions where incidences of asthma are lower (38, 39). The link between air quality and health effects within this region is therefore not clear-cut. Levels of PAH, which are routinely monitored at California Air Resources Board monitoring locations are, however, significantly higher at many sites within the San Joaquin Valley (40). Given that quinones and hydroquinones are likely to be cogenerated with PAH or photochemically formed from PAH in the atmosphere, mass loadings of several different quinones may also be high. The high loadings of four quinones reported previously in Fresno (34) are consistent with this picture. In this work, levels of 12 quinones contained within total suspended particles (TSP) have been measured periodically at a site in Fresno over a 12-month period. Additionally, a chemical test based on the assay of Li et al. (22) was utilized to investigate the relationship between ROS formation and the level of quinones in these samples.

Experimental Section Samples were collected at a site in Fresno, CA, a city with a population of about 450 000 in the center of the San Joaquin valley. Sampling equipment was located on the roof of the Engineering East building on the campus of California State University Fresno, about 20 m above ground level in a largely residential part of the city. Prevailing winds were from the northwest on all sampling days. VOL. 40, NO. 16, 2006 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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Forty-seven millimeter diameter, 2-µm pore-size Teflon filters (Gellman) were cleaned by sonication with five 40-mL aliquots of dichloromethane (Sigma-Aldrich) for 3 min each and were then baked under reduced pressure for 2 h. The final aliquot of dichloromethane was retained and analyzed as a blank sample for both the derivatized and nonderivatized methods described below. Cleaned filters were stored at -4 °C until used. Ambient air was drawn through cleaned filters at a flow rate of 52 L min-1 for sampling periods of 24-72 h. Following collection, organic compounds present were extracted from the filters by sonication with three 10-mL portions of dichloromethane. These extracts were combined and then were divided into two fractions (A and B). Fraction A was concentrated to a final volume of 0.5 ( 0.01 mL by evaporation of the solvent by gently blowing nitrogen onto the surface of the sample. Since deuterated internal standards were not used in this work, the final volumes of the concentrated samples were carefully determined. Part of the concentrated sample was then analyzed by GC-MS, and the remainder was retained for the DTT assay described below. Quinone Analysis. Quinones present in fraction B and filter blanks were converted to their diacetylated derivatives using the method of Cho et al. (33). One tenth of a gram of zinc and 0.2 mL of acetyl acetate (Sigma-Aldrich) were heated with fraction B to 80 °C in sealed vials for 30 min. The vials were removed and shaken every 5 min, and an additional 0.1 g zinc was added after 15 min. Following the reaction, 0.5 mL of water and 3 mL of pentane (Sigma-Aldrich) were added to the vial, and the mixture was centrifuged for 10 min. The organic fraction was then removed and was evaporated to dryness. The residue was redissolved in 0.5 ( 0.01 mL dichloromethane and was analyzed by GC-MS. Fraction A and the filter blank were analyzed by GC-MS using two selected ion monitoring (SIM) methods. Five microliters of the sample was injected into the GC-MS (Agilent 6890/5973) in the splitless mode, with the inlet at 280 °C. The column (Agilent, DB5) was held at 100 °C for 4 min and was then ramped at 5 °C min-1 to 310 °C and was held for 5 min for a total run time of 51 min. Quinones were identified and quantified in fraction A using the SIM program for underivatized quinones by comparison to mixtures of authentic standards prepared in the laboratory. Retention times and m/z ratios used to identify specific compounds in this SIM method are given in the Supporting Information section. A positive identification of a compound requires that peaks are present in fraction A at a level greater than 3 times the signal for the corresponding filter blank. Chromatograms were recorded for the derivatized filter blank, fraction A, and derivatized fraction B using a SIM method for derivatized quinones. Retention times and m/z ratios of the ions used in this method are given in the Supporting Information section. A positive identification of a derivatized quinone requires that the signal obtained from fraction B is 3 times greater than the corresponding signal in both the derivatized filter blank and fraction A. Signals were calibrated using freshly prepared standard mixtures (Sigma-Aldrich) dissolved in dichloromethane injected directly into the GC-MS. During the field measurements, standards were analyzed regularly to check for changes in signal response. Calibration curves for the diacetylated derivatives were generated by subjecting the standard solutions to the same procedure described for fraction B above. To determine the recovery efficiency of quinones during extraction and concentration, filters were spiked with known quantities of quinones and were extracted and analyzed as described above for the ambient filter samples. Initially, the calibration mixture consisted of seven quinone compounds. However, in November 2004, five compounds were added to this set of standards. One of these, 2-amino anthraquinone, was not observed in any of the samples and 4882

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is therefore not included in the data. The sampling periods during which each quinone was analyzed are given in the Supporting Information section. DTT Analysis. The DTT analysis carried out in this work is different from the method reported in the literature (22). Here, formation of H2O2 was measured rather than DTT consumption. The reactivities of individual quinones were first investigated in the lab by mixing (1-10) × 10-7 M quinone solutions with 1 × 10-4 M DTT in a phosphate buffer (pH ) 7.4) at 25 °C. The concentration of H2O2 was determined at 5-10 min intervals by withdrawing 0.1 mL of the mixture and by analyzing each sample using an HPLC-fluorescence technique, described in detail previously (41). H2O2 was plotted as a function of time, and the initial reaction rate was determined from the slope of a fit to the points at time t ) 0. The DTT analysis for ambient samples was performed by mixing 0.5 mL of fraction A with 1 × 10-4 M DTT in the presence of the phosphate buffer at 25 °C. The rate of H2O2 formation at time t ) 0 was then determined as described above for the individual quinones.

Results and Discussion Quinone Analysis. In general, a quinone is thermodynamically more stable than the corresponding semiquinone radical or hydroquinone. As part of this analysis, the sample concentration and analysis procedures described above were carried out for hydroquinone standards. In all cases, most of the hydroquinone was converted to the quinone during analysis. Therefore, while mass loadings of quinones are reported here, it is likely that the levels measured are really the sum of quinone, semiquinone, and hydroquinone mass loadings. Without derivatization, detection limits in the ambient samples were about 0.1 ng m-3 under the given sampling conditions (see Supporting Information section). As reported by Cho et al. (33), derivatization with acetyl acetate may significantly improve the sensitivity of the analysis. This was found to be the case here, but only five compounds were found to be completely converted to their diacetate derivatives during the analysis. Acenaphthenequinone, methyl anthraquinone, dimethyl anthraquinone, 5,12-naphthacenequinone, and benz[a]anthracene-7,12-dione and anthraquinone only partially reacted, and the derivatization did not improve the detection limits. In this work, derivatized compounds were used to quantify quinone mass loadings for 1,2-naphthoquinone, 1,4-naphthoquinone, 2,6-ditertiarybutyl-1,4-benzoquinone, phenanthraquinone, and chrysenequinone. For all of the quinones containing three or more aromatic rings, mean recoveries from three replicates of spiked filters were 100 ( 20%. Recoveries of 1,2-naphthoquinone, 1,4naphthoquinone, and acenaphthenequinone were lower but were constant within the statistical uncertainties. Recoveries of all of derivatized quinones, including 1,2- and 1,4naphthoquinone, were 100%. Field data for acenaphthenequinone have been corrected for incomplete recovery from the filters, while the remaining quinones were assumed to be recovered with 100% efficiency. The variability in recovery is the dominant source of error in the determination of ambient quinone mass loadings leading to a 20% standard deviation in these values. Presumably, the lower recoveries resulted from volatilization of the lighter compounds during sample concentration. During sampling, compounds may adsorb to or desorb from the filter and may undergo chemical transformation on the filter. Two tests were carried out to investigate some of these possible artifacts. First, samples were collected with two filters arranged in series. No quinones were detected on the back filter, suggesting that adsorption of gas-phase quinones onto the filter was not significant. Second, two sets of samplers were used to collect particles simultaneously.

FIGURE 1. (a) Mean mass loadings of quinones measured during summer 2004, winter 2004-2005, and spring 2005. Asterisks denote quinones not analyzed in summer samples. (b) Quinone mass loadings for all samples. The first sampler collected particles on a single filter for 48 h, while the second sampler collected two 24-h samples on separate filters during the same period. The total mass loading of quinones collected by the two samplers was the same, indicating that desorption of quinones from the filters was not significant. While no direct test for chemical transformations on the filters was performed, quinone levels did not positively correlate with ozone concentrations (see below) which might be expected if quinones are generated on the filters from the oxidation of PAH. The 29 samples collected have been divided into three sets: summer (June and July 2004), winter (November 2004February 2005), and spring (April and May 2005). Average

mass loadings within each data set and values for each sample are shown in Figure 1 and are reported in Table 2 along with the standard deviation of each mean value. The large standard deviations reflect the high sample-to-sample variations of the quinone mass loadings. The quinone mass loadings in each individual sample can be found in the Supporting Information section. Comparing the loadings obtained with those reported in the literature (Table 1), it can be seen that levels of many compounds are high compared to other regions. These levels are largely consistent with the mass loadings reported for Fresno previously (34). Levels of the most prevalent compounds measured were compared to ambient ozone concentrations and meteorological parameters. No correlation was observed between quinone levels and TSP mass, wind speed, or wind direction. However, with the exception of 5,12-naphthacenequinone and anthraquinone, the levels of each quinone are anticorrelated with temperature and may be anticorrelated with ozone levels, as shown in the Supporting Information section. The anticorrelation between quinone mass loading and temperature can likely be explained by volatilization from the particles at elevated temperatures. This effect has been observed previously (33, 36) and provides another possible explanation of the elevated quinone mass loadings during the winter months. Given that quinones may be secondary products of tropospheric photochemistry, a positive correlation between quinone mass loadings and ozone concentrations might be expected. However, the counterintuitive anticorrelation with ozone is likely the result of the strong positive correlation between ozone and temperature, as shown in the Supporting Information section. Thus, the total (gas + particle phase) concentration of quinones may be greater on high-ozone days, but the elevated temperatures result in a smaller fraction of this total being present in the particles. To gain further insight into the origins of the particulatephase quinones, the relationship between the mass loadings of the six most abundant compounds was investigated and is shown in the Supporting Information section. The levels of five of these quinones (1,2-naphthoquinone, 1,4-naphthoquinone, anthraquinone, phenanthraquinone, and chrysenequinone) mostly correlate well with each other (0.22 e R2 e 0.89 and 0.05 g P g 1 × 10-4). Although far from conclusive, this suggests that these species may have originated from the same sources. In contrast, the mass loadings of 5,12-naphthacenequinone generally correlated poorly with these species (0.01 e R2 e 0.19 and 0.72 g P g 0.07), possibly indicating a different set of sources. As discussed above, the mass loadings of 5,12-naphthacenequinone do not follow the same trends as other

TABLE 2. Mean Mass Loadings and Standard Deviations for Samples Collected in Summer 2004, Winter 2004-2005, and Spring 2005 summer samples (n ) 13) compound 1,4-naphthoquinone 1,2-naphthoquinone phenanthraquinone 2,6-dtb benzoquinone chrysenequinone acenaphthenequinone 2-methyl anthraquinone 2,3-dimethyl anthraquinone benz[a]anthraquinone anthraquinone naphthacenequinone a

winter samples (n ) 10)

spring samples (n ) 6)

mean mass standard deviation mean mass standard deviation mean mass standard deviation loading/ng m-3 of mean/ ng m-3 loading/ng m-3 of mean/ng m-3 loading/ng m-3 of mean/ng m-3 n.d.a n.d. n.d. n.d. n.d. n.d. n.m. n.m. n.m. 0.21 4.5

(0.22 (9.5

0.15 1.1 1.1 0.54 0.48 0.45 0.92 0.54 1.0 0.47 0.93

(0.07 (0.84 (0.77 (0.74 (0.21 (0.57 (0.63 (0.28 (0.71 (0.46 (1.6

0.06 0.20 0.31 n.d. 0.35 n.d n.d. 0.06 0.05 0.01 0.06

(0.05 (0.25 (0.13 (0.10 (0.11 (0.09 (0.04 (0.10

n.d ) not detected; n.m. ) not measured.

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FIGURE 2. (a) H2O2 formation from the reaction of phenanthraquinone with 1 × 10-4 M DTT. The initial rate of H2O2 formation (d[H2O2]/dt)t)0 is determined from the slope of this plot at the origin. (b) The variation in the initial rate of H2O2 formation with quinone concentration for phenanthraquinone and 1,4-naphthoquinone. Pseudo-first-order rate coefficients are obtained from the slopes of these plots. (c) Comparison of the predicted and measured H2O2 production rate for ambient samples (see text for details). quinones. Additionally, within the summer data set, 5,12naphthacenequinone mass loadings are often found to vary by more than an order of magnitude from day to day. To better understand these data, 5-day back trajectory calculations were performed for the summer data using the NOAA Hysplit model (42). The trajectories fell into two general categories as illustrated in the Supporting Information 4884

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section. “Summer A” air masses were found to pass directly over the San Francisco Bay before reaching Fresno. Summer A trajectories were found to include all of the samples with high 5,12-naphthacenequinone levels. “Summer B” trajectories followed paths that did not take the air masses over the San Francisco Bay. The summer A mean for 5,12naphthacenequinone (10 ng m-3) is more than an order of

magnitude higher than the summer B mean (0.3 ng m-3), and the two values are statistically different at the 90% confidence limit. This suggests that transport from the San Francisco Bay region may be a significant source of this compound during summer. A similar day-to-day variability was reported for chrysenequinone levels during summer 2001 (34). In this work, underivatized chrysenequinone and 5,12naphthacenequinone have similar retention times and virtually identical MS fragmentation patterns. The previous study did not monitor 5,12-naphthacenequinone levels, and so it is possible that chrysenequinone was misidentified in the CRPAQS study. DTT Analysis. H2O2 formation was measured from the reaction of the individual quinones monitored in this study with DTT at pH ) 7.4 and 25 °C. Of the 12 quinones analyzed, only three (phenanthraquinone, 1,2-naphthoquinone, and 1,4-naphthoquinone) were found to generate H2O2 at rates statistically higher than in control experiments. Representative experimental data are shown in Figure 2. For all three “active” quinones, the initial rate of H2O2 formation (d[H2O2]/ dt)t)0 increases linearly with quinone concentration. The pseudo-first-order rate coefficients obtained for the quinones in the presence of 1 × 10-4 M DTT are 12.5 ( 0.5 min-1, 1.3 ( 0.2 min-1, and 0.7 ( 0.1 min-1 for phenanthraquinone, 1,4-naphthoquinone, and 1,2-naphthoquinone, respectively. Samples collected during the winter and spring sampling periods were subjected to DTT analysis. Since this technique was developed during fall 2004, summer 2004 samples could not be analyzed. Further, because of calibration problems with the HPLC system during the analysis of the spring samples, it was decided that only the first data point from this period should be reported, although the remaining five data points did fit the observed trend. Thus, only data for 11 of the 29 total samples collected are reported. No correlation between total quinone mass loading and the rate of H2O2 production is observed (R2 ) 0.05, P )0.52). This is not surprising given that the experiments with individual quinones demonstrate that the majority of quinones detected in the samples do not generate H2O2 in this test. Thus, H2O2 production is found to correlate positively with all three active quinones: 1,2-naphthoquinone (R2 ) 0.73, P ) 8 × 10-4), 1,4-naphthoquinone (R2 ) 0.39, P ) 0.04), and phenanthraquinone (R2 ) 0.76, P ) 4 × 10-4). Previous studies have shown correlations between PAH levels and the ability of particles to generate ROS (22, 43). It is possible that quinones, which are likely to be positively correlated with PAH, are responsible for the observed ROS formation in these studies. As discussed in the Introduction, the presence of iron in PM affects particles’ ability to generate OH radicals in biological environments. However, the reactions of iron do not affect the production rate of ROS in these experiments. Hydrogen peroxide is initially formed as shown in Scheme 1 and may subsequently react with Fe(II) to form OH. In the DTT analysis, the initial rate of H2O2 formation is measured, which does not depend on the iron concentration. Clearly, the iron mass loadings within the particles will affect the rate of OH production by the particles, but these measurements are beyond the scope of this study. Since the concentrations of the active quinones and the pseudo-first-order rate coefficients with DTT are both known, it is possible to compare the rate of H2O2 production expected in the mixture to the rate actually observed. Phenanthraquinone has the largest rate coefficient, and so this species is responsible for the majority of the expected H2O2 production. The comparison is shown in Figure 2c. The measured and predicted rates correlate strongly, and the measured rate is 25% lower than the predicted rate. The uncertainties in the quinone mass loadings and the pseudofirst-order rate coefficients combine to give a 20% uncertainty in the predicted rate (one standard deviation). The standard

deviation in the measured rate is 22%. The predicted and measured rates are therefore not statistically different within the mutual uncertainties. The quinones monitored can thus account for all ROS production in the DTT test within the uncertainties of the measurements. However, because these mutual uncertainties are fairly large, it is still possible that additional ROS-forming chemical species in the sample are present. Phenanthraquinone appears to be a good chemical marker for ROS production, and its levels may therefore correlate with adverse health effects associated with particles.

Acknowledgments Some additional experiments were performed by Tariq Hanifi (CSUF). The authors thank the College of Science and Mathematics, California State University Fresno and the Petroleum Research Fund for financial support of this work. R.L. is grateful to the McNair program for the award of a fellowship. A.S.H. thanks Tim Tyner (UCSF Fresno) for helpful discussions.

Supporting Information Available GC-MS parameters and ambient detection limits for monitored quinones, quinone mass loadings in samples collected, correlation between mass loadings of the six most prevalent quinones and ambient temperature, correlation between mass loadings of the six most prevalent quinones and peak ozone levels, correlation between peak ozone levels and temperature, correlation between mass loadings of the six most prevalent quinones, and representative 5-day back trajectory calculations for summer A samples that passed directly over the San Francisco Bay area and summer B samples that bypassed the San Francisco Bay area. This material is available free of charge via the Internet at http:// pubs.acs.org.

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Received for review August 11, 2005. Revised manuscript received April 2, 2006. Accepted June 15, 2006. ES0515957