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Traffic source impacts on chlorinated polycyclic aromatic hydrocarbons in PM2.5 by short-range transport
Atmospheric Environment 216 (2019) 116944
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Traffic source impacts on chlorinated polycyclic aromatic hydrocarbons in PM2.5 by short-range transport
T
Ryosuke Oishia, Yuki Imaib, Fumikazu Ikemoric, Takeshi Ohuraa,b,∗ a
Faculty of Agriculture, Meijo University, 1-501 Shiogamaguchi, Nagoya, 468-8502, Japan Graduate School of Agriculture, Meijo University, 1-501 Shiogamaguchi, Nagoya, 468-8502, Japan c Nagoya City Institute for Environmental Sciences, 5-16-8 Toyoda, Nagoya, 457-0841, Japan b
G R A P H I C A L A B S T R A C T
A R T I C LE I N FO
A B S T R A C T
Keywords: ClPAHs Biomarkers PM2.5 Diagnostic ratio Vehicle exhaust
Chlorinated polycyclic aromatic hydrocarbons (ClPAHs) are recognized as ubiquitous hazardous pollutants in the environment, whereas their behavior in local areas remains unclear. Additionally, there is limited information on the sources in local areas. Here, we investigated the seasonal trends of ClPAHs associated with fine particles (PM2.5) at two sites near heavy traffic roads to evaluate the local atmospheric behaviors and sources. The annual mean concentrations of total ClPAHs at the north (site A) and south side (site B) across a heavy traffic road were 12.0 and 19.2 pg/m3, respectively. The higher concentration at site B was further emphasized during the winter season; at that time the site was located beneath the wind from the traffic road. In addition, for individual ClPAHs, the behaviors of 8-chlorofluoranthene (8-ClFluor) and 7-chlorobenz[a]anthracene (7-ClBaA) were consistent with the frequency of the north wind, suggesting that 8-ClFluor and 7-ClBaA have the ability to be indicators of vehicle exhaust. Diagnostic ratios using these specific ClPAHs during high-traffic activities provided the specific values to differentiate the impacts.
1. Introduction Polycyclic aromatic hydrocarbons (PAHs) are recognized as typical hazardous pollutants because some of them act as carcinogenic/mutagenic and endocrine-disrupting compounds (Baek et al., 1991). To evaluate the human health risks for PAHs in more detail, a number of researchers have investigated not only their biological toxicities but ∗
also their environmental behaviors and sources (Bostrom et al., 2002; Kim et al., 2013). PAHs are mainly produced by the incomplete combustion of organic materials. Therefore, the principal sources are present anywhere human activity occurs while the impacts continually vary among countries, areas, seasons, and dates. In urban cities, vehicle exhaust has been frequently identified as a dominant source of atmospheric PAHs, that is, traffic activity can greatly impact human health
Corresponding author. Faculty of Agriculture, Meijo University, 1-501 Shiogamaguchi, Nagoya, 468-8502, Japan. E-mail address: [email protected] (T. Ohura).
https://doi.org/10.1016/j.atmosenv.2019.116944 Received 6 May 2019; Received in revised form 26 August 2019; Accepted 28 August 2019 Available online 31 August 2019 1352-2310/ © 2019 Elsevier Ltd. All rights reserved.
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Fig. 1. Map of the marked sampling sites.
2. Experimental
(Abdel-Shafy and Mansour, 2016). Additionally, not only PAHs but also PAH derivatives such as nitrated PAHs are emitted from diesel engine exhaust (Schuetzle et al., 1982). Chlorinated PAHs (ClPAHs) are a group of PAHs derivatives, some species of which are ubiquitously detected in the environment (Ohura, 2007; Ohura et al., 2014; Sun et al., 2013; Wang et al., 2018). Recently, we investigated the atmospheric behaviors of ClPAHs, including the spatial distributions and temporal trends, followed by a comparison to the concentrations of PAHs (Ohura et al., 2016, 2018, 2019). Throughout these studies, we observed that the compositions of ClPAHs in the air vary significantly compared with PAHs. We concluded that atmospheric PAHs could be produced by similar sources across Japan, whereas the sources of atmospheric ClPAHs changed depending on the site and season. This difference implies that the atmospheric fates of ClPAHs might be short compared with PAHs, so that the distribution of atmospheric ClPAHs over wider regions is limited. To date, numerous researchers have carried out studies on the sources of ambient PAHs. These studies have sought to better understand global air pollution, in which the influences of global long-range transport are investigated, suggesting that the overall contributions have become too large to ignore (Björseth et al., 1979; Keyte et al., 2013; Yang et al., 2007). However, few studies have investigated the behaviors of air pollutants between short distances, that is, short-range transport. In particular, for ClPAHs, there is no information available, to the best of our knowledge. In this study, we survey ClPAHs found in PM2.5 (particulate matter that have diameters less than 2.5 μm) near two sites (ca. 800 m distance) across heavy traffic roads. The differences in the profiles caused by the short-range transport of ClPAHs were evaluated, and we propose new ClPAH diagnostic ratios for identifying the traffic sources.
2.1. Chemicals This study targeted 25 species of ClPAHs with three to five aromatic rings. Most of these ClPAHs were synthesized ab initio; the detailed procedures were reported previously (Ohura et al., 2005). In addition, we analyzed 18 PAHs, including 16 that have been classified as priority pollutants by the U.S. Environmental Protection Agency. The PAH mixture standard was purchased from LGC Labor GmbH (Augsburg, Germany). As petroleum markers, a standard solution of five hopanes and five steranes (NIST SRM 2266) was purchased from Sigma-Aldrich (Saint Louis, MO, USA). Three deuterated PAHs (phenanthrene-d10, fluoranthene-d10, and perylene-d12), purchased from Cambridge Isotope Laboratories, Inc. (Andover, MA, USA), were used as internal standards. The extractions and purifications were performed with analytical grade solvents purchased from Wako Pure Chemical (Osaka, Japan) or Kanto Chemical (Tokyo, Japan). The 25 ClPAHs used in this study were abbreviated as follows: 9chlorophenanthrene (9-ClPhe), 1,9-dichlorophenanthrene (1,9Cl2Phe), 3,9-dichlorophenanthrene (3,9-Cl2Phe), 9,10-dichlorophenanthrene (9,10-Cl2Phe), 3,9,10-trichlorophenanthrene (3,9,10-Cl3Phe), 2-chloroanthracene (2-ClAnt), 9-chloroanthracene (9ClAnt), 9,10-dichloroanthracene (9,10-Cl2Ant), 3-chlorofluoranthene (3-ClFluor), 8-chlorofluoranthene (8-ClFluor), 1,3-dichlorofluoranthene (1,3-Cl2Fluor), 3,4-dichlorofluoranthene (3,4-Cl2Fluor), 3,8-dichlorofluoranthene (3,8-Cl2Fluor), 1-chloropyrene (1-ClPy), dichloropyrene (Cl2Py), trichloropyrene (Cl3Py), tetrachloropyrene (Cl4Py), 6-chlorochrysene (6-ClChry), 6,12-dichlorochrysene (6,12-Cl2Chry), 7-chlorobenz[a]anthracene (7-ClBaA), 7,12-dichlorobenz[a]anthracene (7,12Cl2BaA), tetrachlorofluoranthene (Cl4Fluor), 6-chlorobenzo[a]pyrene (6-ClBaP), dichlorobenzo[a]pyrene (Cl2BaP), and trichlorobenzo[a] pyrene (Cl3BaP). The 16 PAHs used in this study were abbreviated as
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analyses, the initial oven temperature of 100 °C was held for 2 min. The temperature was then increased at 25 °C/min to 200 °C, increased at 5 °C/min to 300 °C, and finally held at 300 °C for 15 min. The PAHs and petroleum markers (hopanes and steranes) were analyzed by the same GC conditions, that is, the initial oven temperature of 70 °C was held for 3 min. The temperature was then increased at 20 °C/min to 240 °C, increased at 5 °C/min to 310 °C, and finally held for 30 min. The injector temperatures for the ClPAH and PAH analyses were 300 and 280 °C, respectively. For both the ClPAH and PAH analyses, the transfer line and ion source temperature was 280 °C. The mass spectrometer was operated in both the selected ion monitoring and electron impact ionization mode. A constant ion current of 200 μA was used, and the electron energy was 70 eV. The target ions of the ClPAHs and PAHs were described elsewhere (Kamiya et al., 2015). The target ions of the petroleum markers for quantification and confidence were m/z 191 and 217, and m/z 355, 357, 371, 383, 385, 387, 397, and 411, respectively. Twelve species of elements were targeted as follows, including abbreviations: vanadium (V), chromium (Cr), manganese (Mn), iron (Fe), nickel (Ni), copper (Cu), zinc (Zn), arsenic (As), selenium (Se), and lead (Pb). These elements were determined using energy dispersive X-ray fluorescence spectrometry (EDXRF, Rigaku Corp., Tokyo, Japan) coupled with fundamental parameter quantification. The analytical conditions and quality controls were confirmed and reported elsewhere (Okuda et al., 2013).
follows: naphthalene (Nap), acenaphthylene (Acy), acenaphthene (Ace), fluorine (Flu), phenanthrene (Phe), anthracene (Ant), fluoranthene (Fluor), pyrene (Py), benzo[a]anthracene (BaA), chrysene (Chry), benzo[b]fluoranthene (BbF), benzo[k]fluoranthene (BkF), benzo[e]pyrene (BeP), benzo[a]pyrene (BaP), perylene (Pery), benzo [ghi]perylene (BghiP), dibenzo[a,h]anthracene (DBahA), and indeno [1,2,3-cd]pyrene (IP). The five hopanes used in this study were abbreviated as follows: 17α(H)-22,29,30-trisnorhopane (Tm), 17α(H),21β(H)-30-norhopane (29αβ), 17α(H),21β(H)-hopane (30αβ), 17α(H),21β(H)–22S-homohopane (31αβS), and 17α(H),21β(H)-22Rhomohopane (31αβR). The five steranes used in this study were αββ20R-cholestane, ααα20R-cholestane, αββ20R,24S-methylcholestane, αββ20R,24R-ethylcholestane, and ααα20R,24R-ethylcholestane. 2.2. Sampling sites and procedures The study was carried out in the city of Nagoya (latitude 35° 54′ N, longitude 136° 54′ E; Fig. 1). The detailed information of the city was described elsewhere (Ohura et al., 2016). Samples of PM2.5 were simultaneously collected at two sites (sites A and B) in Nagoya, from April 2012 through April 2013. These sites were located in an industrial area of southeast Nagoya; the distance between sites was ca. 800 m. Additionally, an express highway (ca. 60,000 vehicles/day) and two national roads (ca. 86,000 vehicles/day consist of 79% light- and 31% heavy-duty vehicles for Route 23 and ca. 43,000 vehicles/day consist of 87% light- 13% heavy-duty vehicles for Route 247) were located near the sites. Sampling at site A was performed at rooftop locations (20 m above ground level) at the Nagoya City Institute for Environmental Science (35° 5′ N, 136° 54′ E). Sampling at site B was carried out at ground level in the Motoshio park (35° 5′ N, 136° 55′ E), approximately 100 m from the nearest busy roads. PM2.5 samples were collected on quartz fiber filters (QFFs; 47 mm diameter, 2500 QAT-UP; Pall Corp. NY, USA) through an EPA well-type impactor ninety-six (Peters et al., 2001) for PM2.5, volumes of which were determined by a low-volume air sampler (Partisol Model 2000; Thermo Scientific, MA, USA) operating at a constant flow rate of 16.7 L/min. The sampling details for PM2.5 were described elsewhere (Yamagami et al., 2019). Samples were continuously collected for one week, after which the QFFs were changed during the sampling period. After atmospheric sampling, the QFFs were cut into pieces for the analysis of ClPAHs, PAHs, hopanes, steranes, and trace elements. The QFFs were wrapped in aluminum foil, sealed, and stored in a freezer at −35 °C until extraction.
For quality assurance of the analytical procedures, a recovery test of the target compounds using a sequence of analytical procedures was performed. Recoveries were confirmed to be reasonable to our previous work (Kamiya et al., 2015). Recoveries of the internal standards spiked onto the samples at sites A and B were 100 ± 4% for fluoranthene-d10 and 104 ± 13% for perylene-d12 and 97 ± 7% for fluoranthene-d10 and 119 ± 14% for perylene-d12, respectively. The target ClPAHs were not detected in the procedural blank samples. The limits of detection were estimated by multiplying the standard deviations of the concentrations found in the six repeated analyses of a dilute standard by a solvent (Kamiya et al., 2015). If the sampled air volume was 160 m3, the ClPAH limits of detection ranged from 0.025 pg/m3 for 9-ClPhe to 0.78 pg/m3 for Cl3BaP, and the PAH limits of detection ranged from 0.013 pg/m3 for Phe to 0.10 pg/m3 for BbF. For the petroleum markers, the limits of detection ranged from 0.09 pg/m3 for ααα20R-cholestane to 0.95 pg/m3 for 31αβS.
2.3. Sample extraction and analysis
3. Results and discussion
Samples collected at each site and during each week were treated as a single sample. After air sample collection, a part of each filter was cut using stainless steel scissors and then extracted using a 1:1 mixture of hexane and dichloromethane in an accelerated solvent extraction apparatus (Dionex-ASE 350, Sunnyvale, CA, USA). Before extraction, an internal standard mixture consisting of fluoranthene-d10 (25 ng), and perylene-d12 (25 ng) was spiked onto the filters. The eluate was concentrated to ~200 μL under a gentle stream of N2 at 45 °C. During these operations, the solutions were protected from light to prevent photochemical degradation of the analytes. Each condensed solution was purified using column chromatography with silica gel (Supelclean LC-Si SPE tube, Sigma-Aldrich Co., St. Louis, MO, USA) and was eluted with 13 mL of 10% dichloromethane in n-hexane. Prior to injection, phenanthrene-d10 (25 ng) was added to the residue as a recovery standard. The extracts were analyzed with a JMS-Q1000GC quadrupole mass spectrometer (JEOL, Tokyo, Japan) equipped with a 7890 A gas chromatograph (Agilent Technologies Inc., Santa Clara, CA, USA) fitted with an InertCap 5MS/NP capillary column (30 m long, 0.25 mm i.d., 0.25 μm film thickness; GL Science Inc., Tokyo, Japan). Helium was used as the carrier gas at a flow rate of 1.0 mL/min. For the ClPAH
3.1. Concentrations of pollutants in PM2.5
2.4. Quality assurance
3.1.1. ClPAHs and PAHs Throughout the campaign, most of the target ClPAHs and PAHs were detected from a PM2.5 fraction at each site. For individual ClPAHs, the mean concentration of 6-ClBaP was the highest at both sites. The mean concentrations at sites A and B were 4.76 and 6.14 pg/ m3, respectively (Table S1). The second most abundant species was 7,12-Cl2BaA (1.22 pg/m3) at site A and 1-ClPy (2.02 pg/m3) at site B, which indicates the different behaviors of ClPAHs despite the sites being near each other. This difference might be caused by the presence of different sources of ClPAHs at each site. For the total ClPAHs, the mean concentrations at sites A and B were 12.0 and 19.2 pg/m3, respectively. Overall, the concentrations of ClPAHs at site B tended to be higher (1.6 times on average) than those at site A. Furthermore, these concentrations were a few times lower than that of the mean level surveyed over other Japanese locations. These results suggest that the ClPAHs at the current sites may be influenced by the limited sources. For PAHs, the overall concentrations were higher at site B (from 1.2 to 4.3 times) compared with site A (Table S1). Among them, BbF was 3
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Fig. 2. Temporal concentrations of (A) total ClPAHs, (B) total PAHs, (C) total hopanes, and (D) total sterane at sites A and B, and the variation in the concentration ratios of site B to site A.
city, British Columbia (Ding et al., 2009), whereas they were considerably lower (ca. ~50 times) than those observed in a highly polluted city, New Delhi (Pant et al., 2015).
the most abundant species at both sites, BkF at site A and Chry at site B were second, and Chry at site A and BghiP at site B were third. The orders of magnitude of the PAHs were also different at each site, such as ClPAHs. Such predominant PAHs were observed in the results from a suburban area of Zhengzhou, China, in which various sources such as coal combustion, vehicles, coking plants, and biomass burning were estimated as the main sources (Wang et al., 2015). Therefore, the influential sources of the PAHs might also be different between the sites. The mean concentrations of total PAHs at site A and B were 1.61 and 2.37 ng/m3, respectively. The levels were comparable to those reported from surveys of European areas (Jedynska et al., 2014), and were approximately two times lower than that of the mean level surveyed over other Japanese sites.
3.1.3. Trace elements Analyses of the trace elements in the environmental samples offer another useful approach to explore the possible sources. We focused on the occurrences of nine species of anthropogenic trace elements in PM2.5. Among them, Zn was found to be the most abundant species (35.3 ng/m3 at site A and 43.8 ng/m3 at site B, on average), followed by Mn, Cu, and Ni at both sites (Table S1). These four elements accounted for 75.5 and 77.0% of the total mass of elements at sites A and B, respectively. In the case of trace elements, the concentrations also tended to be higher at site B than that at site A. Zn has been associated with industrial emissions and incineration (Duvall et al., 2012; Harrison et al., 1997). Although Mn has often been used as a marker for crustal dust/soil, Cu and Ni are supplied by anthropogenic activities such as fossil fuel burning and industrial activities (Tian et al., 2012; Viana et al., 2008). In particular, Cu has been used as a traffic source marker because it is associated with brake wear (Almeida et al., 2005; Lawrence et al., 2013). Comparing the relationships of the concentrations among trace elements throughout the seasons, significant correlations (p < 0.05) were observed among the various elements (data not shown). For instance, the concentrations of Mn correlated to those of both Zn and Cu at each site. Zn and Cu are not only major additives to lubricating oils but also associated with other transportation activities such as brake and tire wear (Lough et al., 2005).
3.1.2. Biomarkers Hopanes and steranes are natural products, and are considered to be petroleum markers (Simoneit, 1984). In addition, hopane and sterane homologues are relatively stable and have similar atmospheric fate processes. The mean concentrations of total hopanes at sites A and B were 0.252 and 0.420 ng/m3, respectively. Among the hopanes detected, on average 30αβ was the most abundant species at both sites, followed by 29αβ and 31αβR. This similar order of magnitude has been observed in the results of a traffic-related aerosol sample study (Wang et al., 2007). In addition, Cui et al. (2016) reported that the high proportions of 30αβ and 29αβ are attributable to vehicle exhaust emission. Therefore, the main source of the PM2.5 samples used in this study could be considered to be vehicles. Conversely, the mean concentrations of total steranes at sites A and B were 0.0212 and 0.0867 ng/m3, respectively. Among the individual sterane concentrations, αββ20R,24R-ethylcholestane was the most abundant species at both sites, followed by ααα20R,24R-ethylcholestane, and then αββ20R,24Smethylcholestane or αββ20R-cholestane. However, atmospheric steranes have limited information available compared with hopanes; both species have been assigned as traffic markers (Simoneit, 1984). Comparing the sterane concentrations to other sites, the present concentrations were somewhat consistent with the levels found in an urban
3.2. Seasonal trend Fig. 2 shows the annual concentrations of (A) total ClPAHs, (B) total PAHs, (C) total hopanes, and (D) total steranes in PM2.5 at sites A and B, and the variation in the concentration ratios at site B to site A. The concentrations of total ClPAH showed a typical seasonal trend at both sites, a decrease during the warmer season and an increase during the colder season (Fig. 2A). Such seasonal variation is caused by 4
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predominant at both sites. A previous study reported that the homohopane index (31αβS/31αβS+31αβR) with a range of 0.54–0.67 was found in mineral-oil-derived emission sources, whereas the index values with a range of 0.05–0.35 was identified as coal smoke (Schnelle-Kreis et al., 2007). In our samples, the homohopane index, except the summer season, remained at around 0.5 (Fig. S3). These results suggest that the PM2.5 collected during the colder seasons at these sites were adequately considered as coming from vehicle exhaust rather than that of coal combustion.
competitive factors such as meteorological conditions, including wind conditions, ambient temperatures, and hours of daylight, but characteristics of the sources and the physicochemical properties of the substances, which has been frequently observed in global atmospheric PAHs (Shen et al., 2013). Indeed, the trend was also observed in the seasonal variations of the total PAHs in PM2.5 in the present study (Fig. 2B), which were consistent with those of the total ClPAHs. Furthermore, concerning total hopanes (Fig. 2C) and steranes (Fig. 2D), the seasonal variations of the fossil petroleum markers showed a similar pattern to the ClPAHs and PAHs. The total hopanes and steranes decreased during the warmer season and increased during the colder season. Such seasonal variations have been frequently observed in past studies that analyzed ambient particles (Ding et al., 2009; Jedynska et al., 2014). Among the trace elements, Fe and Cu are considered to be indicators of traffic sources. Those two elements showed very different trends compared with other trace elements. The concentrations of V, Cr, and Zn, known as markers of anthropogenic sources, such as industrial activities, tended to decrease during the colder season at both sites, whereas neither Fe nor Cu displayed such variation throughout the seasons or increased during the colder season (Fig. S1). Specifically, the concentration levels at site B were increased during the colder season and were considerably higher than those at site A. These results suggest that the contributions of industrial activities and traffic sources in PM2.5 at these sites changes during the seasons. That is, it implies that PM2.5 could be driven by seasonal factors such as wind direction.
3.5. Diagnostic ratios of ClPAHs for traffic impacts As described above, the concentrations of most air pollutants at site B tended to be notably higher than those at site A during the colder season with a blowing north wind, and were estimated to be strongly influenced by traffic exhaust. For site A, the air pollutants might be impacted less by traffic sources during the summer season with a blowing south wind because of the influence of various industrial sources located in the south area and photodegradation, as showing in Fig. 4. In other words, air pollutants at site B could be directly affected by the traffic sources during the colder season compared with site A. Therefore, the concentrations of pollutants at site B implied that those at site A directly reflect the traffic impacts because the superfluous factors affecting the concentrations in site B were eliminated. Indeed, the determined concentrations of hopanes were frequently increased by northern winds. To evaluate the behaviors of ClPAHs by the traffic impact, we analyzed the data and subtracted the concentrations of individual ClPAH at site A from the corresponding values at site B. The seasonal trends of the subtracted concentrations of typical ClPAHs were roughly categorized into two patterns: i) high concentrations with a north wind, and ii) equal concentrations over the seasons (Fig. 5). At a high frequency of the north wind, the subtracted concentrations indicate the influence of a major road located across the north section of site B. The ClPAHs showing high concentrations with a north wind were 9-ClPhe, 8-ClFluor, 7-ClBaA, and 6-ClBaP, and were strongly affected by the traffic sources. However, 3-ClFluor and 1-ClPy appear to be barely affected by the traffic sources, because the subtracted concentrations were somewhat constant over the season. In addition, 3-ClFluor, 8-ClFluor, and 7-ClBaA are relatively photostable compared with other ClPAHs (Ohura et al., 2008). Based on this evidence, the impacts of the traffic sources could be estimated by the diagnostic ratios combining ClPAHs that are high and low responsible for traffic with photostable. Consequently, [8-ClFluor]/[3-ClFluor] and [7ClBaA]/[3-ClFluor] were the most suitable traffic indicators of ClPAHs. Indeed, the values of [8-ClFluor]/[3-ClFluor] and [7-ClBaA]/[3ClFluor] increased during the traffic impact, and were estimated to be 2.2–3.2 and 4.2–7.0 respectively. Our previous studies on the ambient ClPAHs in urban cities in Japan found the values of [7-ClBaA]/[3ClFluor] ranged from 0.66 to 9.8 (Ohura et al., 2016, 2019). Furthermore, the ratio ranged from 0.16 to 0.91 in the surveys of other Asian countries (Kakimoto et al., 2014; Ma et al., 2013). These results suggest that vehicle exhaust, rather than other sources, might be a limited source of ClPAHs in ambient air. Indeed, the sources of ambient ClPAHs can be specific for each local site compared with those of PAHs. That is, such diagnostic ratios using specific ClPAHs are likely to adapt to the analysis of local sources. To evaluate the potential of the diagnostic ratios, further studies are needed.
3.3. Influence of wind direction We investigated the influence of the ClPAH and PAH concentrations by the wind direction at each site. Fig. 3A shows the wind direction frequency and mean wind speed at each sampling campaign. The mean wind speed was 2.5 m/s, and there were roughly two principal classifications of the wind direction over the sampling period. The data suggests that the southeast and northwest winds frequently occurred during the warmer and colder seasons, respectively. Fig. 3B–E shows contour plots with a polar axis of total ClPAH or total PAH concentrations at the rate of frequent wind direction. These plots indicate that the ClPAH and PAH concentrations were elevated by highly frequent northwest wind rather than by southeast wind. Interestingly, the concentrations of ClPAHs (Fig. 3C) and PAHs (Fig. 3E) at site B were more emphasized than those at site A (Fig. 3B and D) when a northwest wind blew. Conversely, the southeast wind contributed less to the concentrations of ClPAHs and PAHs at both site. These findings suggest that there are principal sources related to ClPAHs and PAHs from the northwest, especially, high-impact sources will be present between sites A and B. Therefore, ClPAH and PAH concentrations at site B might be influenced by various sources such as not only near automobile (Source II) but away factories and other anthropogenic activities (Source I) located on the windward side, so that the concentrations at site B were higher than those at site A (Fig. 4). 3.4. Source analysis Figure S2A shows the concentration relationships between the ΣClPAHs and ΣPAHs at each site. Significant correlations (p < 0.01) were observed at both sites, suggesting that the ClPAHs could be produced by the same PAH sources extending to the surroundings at these sites. Indeed, the concentrations of the ΣClPAHs and ΣPAHs at site A showed significant correlations to those at site B (Fig. S2B). These data indicate that the current PM2.5 level was produced by similar sources. To identify the sources, we analyzed the behaviors of the petroleum marker hopanes. For certain hopanes, the concentration ratios of the constitutional isomers are used to identify the origins of the fossil sources. According to Simoneit (1999), the proportions of 29αβ and 30αβ in the hopanes tend to be dominant in vehicle exhaust emissions. In our study, the proportions of 29αβ and 30αβ were relatively
4. Conclusion In this study, we investigated ClPAHs in PM2.5 at two sampling sites in heavy traffic areas throughout the seasons. A major traffic road was between the two sites; thus, the differences of the air pollutant concentrations between the sites could be expressive of the behavior of short-range transport. The concentrations of ClPAHs at the two sites were regularly varied by the wind direction, evidence that the ClPAHs 5
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Fig. 3. (A) Wind direction frequency and wind speed in the sampling area during the campaign. The circle sizes and colors represent the wind speed (m/s) and seasons: spring (gray), summer (red), autumn (blue), winter (green). Contour maps of total ClPAH (B, C) and total PAH (D, E) concentrations in frequency of each wind direction at site A (B, D) and site B (C, E). These rotations of axis 0 represent the north direction. (For interpretation of the references to color in this figure legend, the reader is referred to the Web version of this article.)
interests or personal relationships that could have appeared to influence the work reported in this paper.
are caused by vehicle exhaust. Specifically, the atmospheric behaviors of 8-ClFluor and 7-ClBaA were unique among the ClPAHs, which were consistent with those of hopanes. We proposed new diagnostic ratios using these specific ClPAHs to estimate traffic sources. Applying the ratios to other surveyed sites, it suggests that the contributions of traffic sources to ClPAH pollution would be slightly high.
Acknowledgments This work was supported in part by the Ministry of Education, Culture, Sports, Science and Technology, via a Grant-in-Aid for Scientific Research (C) (No. 26340015), and a Grant-in-Aid for Scientific Research (B) (No. 16H05625 and 17H01865).
Declaration of competing interest The authors declare that they have no known competing financial 6
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Source apportionment of traffic emissions of particulate matter using tunnel measurements. Atmos. Environ. 77, 548–557. Lough, G.C., Schauer, J.J., Park, J.-S., Shafer, M.M., DeMinter, J.T., Weinstein, J.P., 2005. Emissions of metals associated with motor vehicle roadways. Environ. Sci. Technol. 39, 826–836. Ma, J., Chen, Z., Wu, M., Feng, J., Horii, Y., Ohura, T., Kannan, K., 2013. Airborne PM2.5/ PM10-associated chlorinated polycyclic aromatic hydrocarbons and their parent compounds in a suburban area in Shanghai, China. Environ. Sci. Technol. 47, 7615–7623. Ohura, T., 2007. Environmental behavior, sources, and effects of chlorinated polycyclic aromatic hydrocarbons. Sci. World J. 7, 372–380. Ohura, T., Amagai, T., Makino, M., 2008. Behavior and prediction of photochemical degradation of chlorinated polycyclic aromatic hydrocarbons in cyclohexane. Chemosphere 70, 2110–2117. Ohura, T., Horii, Y., Yamashita, N., 2018. Spatial distribution and exposure risks of ambient chlorinated polycyclic aromatic hydrocarbons in Tokyo Bay area and network approach to source impacts. Environ. Pollut. 232, 367–374. Ohura, T., Kamiya, Y., Ikemori, F., 2016. Local and seasonal variations in concentrations of chlorinated polycyclic aromatic hydrocarbons associated with particles in a Japanese megacity. J. Hazard Mater. 312, 254–261. Ohura, T., Kitazawa, A., Amagai, T., Makino, M., 2005. Occurrence, profiles, and photostabilities of chlorinated polycyclic aromatic hydrocarbons associated with particulates in urban air. Environ. Sci. Technol. 39, 85–91. Ohura, T., Sakakibara, H., Watanabe, I., Shim, W.J., Manage, P.M., Guruge, K.S., 2014. Spatial and vertical distributions of sedimentary halogenated polycyclic aromatic hydrocarbons in moderately polluted areas of Asia. Environ. Pollut. 196, 331–340. Ohura, T., Suhara, T., Kamiya, Y., Ikemori, F., Kageyama, S., Nakajima, D., 2019. Distributions and multiple sources of chlorinated polycyclic aromatic hydrocarbons in the air over Japan. Sci. Total Environ. 649, 364–371. Okuda, T., Fujimori, E., Hatoya, K., Takada, H., Kumata, H., Nakajima, F., Hatakeyama, S., Uchida, M., Tanaka, S., He, K.B., Ma, Y.L., Haraguchi, H., 2013. Rapid and simple determination of multi-elements in aerosol samples collected on quartz fiber filters by using EDXRF coupled with fundamental parameter quantification technique. Aeosol Air Qual. Res. 13, 1864–1876. Pant, P., Shukla, A., Kohl, S.D., Chow, J.C., Watson, J.G., Harrison, R.M., 2015. Characterization of ambient PM2.5 at a pollution hotspot in New Delhi, India and inference of sources. Atmos. Environ. 109, 178–189. Peters, T.M., Vanderpool, R.W., Wiener, R.W., 2001. Design and calibration of the EPA PM2.5 well impactor ninety-six (WINS). Aerosol Sci. Technol. 34, 389–397. Schnelle-Kreis, J., Sklorz, M., Orasche, J., Stölzel, M., Peters, A., Zimmermann, R., 2007. Semi volatile organic compounds in ambient PM2.5. Seasonal trends and daily resolved source contributions. Environ. Sci. Technol. 41, 3821–3828. Schuetzle, D., Riley, T., Prater, T., Harvey, T., Hunt, D., 1982. Analysis of nitrated polycyclic aromatic hydrocarbons in diesel particulates. Anal. Chem. 54, 265–271.
Fig. 4. Schematic model of the influence of air pollutants in the present sampling sites from the north wind.
Fig. 5. Frequency of north and south winds and typical ClPAH concentrations at site A subtracted from site B throughout the campaign.
Appendix A. Supplementary data Supplementary data to this article can be found online at https:// doi.org/10.1016/j.atmosenv.2019.116944. References Abdel-Shafy, H.I., Mansour, M.S.M., 2016. A review on polycyclic aromatic hydrocarbons: source, environmental impact, effect on human health and remediation. Egypt. J. Pet. 25, 107–123. Almeida, S.M., Pio, C.A., Freitas, M.C., Reis, M.A., Trancoso, M.A., 2005. Source apportionment of fine and coarse particulate matter in a sub-urban area at the Western
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and results. J. Aerosol Sci. 39, 827–849. Wang, G., Kawamura, K., Zhao, X., Li, Q., Dai, Z., Niu, H., 2007. Identification, abundance and seasonal variation of anthropogenic organic aerosols from a mega-city in China. Atmos. Environ. 41, 407–416. Wang, J., Li, X., Jiang, N., Zhang, W., Zhang, R., Tang, X., 2015. Long term observations of PM2.5-associated PAHs: comparisons between normal and episode days. Atmos. Environ. 104, 228–236. Wang, L., Li, C., Jiao, B., Li, Q., Su, H., Wang, J., Jin, F., 2018. Halogenated and parent polycyclic aromatic hydrocarbons in vegetables: levels, dietary intakes, and health risk assessments. Sci. Total Environ. 616–617, 288–295. Yamagami, M., Ikemori, F., Nakashima, H., Hisatsune, K., Osada, K., 2019. Decreasing trend of elemental carbon concentration with changes in major sources at Mega city Nagoya, Central Japan. Atmos. Environ. 199, 155–163. Yang, X.-Y., Okada, Y., Tang, N., Matsunaga, S., Tamura, K., Lin, J.-M., Kameda, T., Toriba, A., Hayakawa, K., 2007. Long-range transport of polycyclic aromatic hydrocarbons from China to Japan. Atmos. Environ. 41, 2710–2718.
Shen, H., Huang, Y., Wang, R., Zhu, D., Li, W., Shen, G., Wang, B., Zhang, Y., Chen, Y., Lu, Y., Chen, H., Li, T., Sun, K., Li, B., Liu, W., Liu, J., Tao, S., 2013. Global atmospheric emissions of polycyclic aromatic hydrocarbons from 1960 to 2008 and future predictions. Environ. Sci. Technol. 47, 6415–6424. Simoneit, B.R.T., 1984. Application of molecular marker analysis to reconcile sources of carbonaceous particulates in tropospheric aerosols. Sci. Total Environ. 36, 61–72. Simoneit, B.R.T., 1999. A review of biomarker compounds as source indicators and tracers for air pollution. Environ. Sci. Pollut. Res. 6, 159–169. Sun, J.-L., Zeng, H., Ni, H.-G., 2013. Halogenated polycyclic aromatic hydrocarbons in the environment. Chemosphere 90, 1751–1759. Tian, H.Z., Lu, L., Cheng, K., Hao, J.M., Zhao, D., Wang, Y., Jia, W.X., Qiu, P.P., 2012. Anthropogenic atmospheric nickel emissions and its distribution characteristics in China. Sci. Total Environ. 417–418, 148–157. Viana, M., Kuhlbusch, T.A.J., Querol, X., Alastuey, A., Harrison, R.M., Hopke, P.K., Winiwarter, W., Vallius, M., Szidat, S., Prévôt, A.S.H., Hueglin, C., Bloemen, H., Wåhlin, P., Vecchi, R., Miranda, A.I., Kasper-Giebl, A., Maenhaut, W., Hitzenberger, R., 2008. Source apportionment of particulate matter in Europe: a review of methods
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Traffic source impacts on chlorinated polycyclic aromatic hydrocarbons in PM2.5 by short-range transport
T
Ryosuke Oishia, Yuki Imaib, Fumikazu Ikemoric, Takeshi Ohuraa,b,∗ a
Faculty of Agriculture, Meijo University, 1-501 Shiogamaguchi, Nagoya, 468-8502, Japan Graduate School of Agriculture, Meijo University, 1-501 Shiogamaguchi, Nagoya, 468-8502, Japan c Nagoya City Institute for Environmental Sciences, 5-16-8 Toyoda, Nagoya, 457-0841, Japan b
G R A P H I C A L A B S T R A C T
A R T I C LE I N FO
A B S T R A C T
Keywords: ClPAHs Biomarkers PM2.5 Diagnostic ratio Vehicle exhaust
Chlorinated polycyclic aromatic hydrocarbons (ClPAHs) are recognized as ubiquitous hazardous pollutants in the environment, whereas their behavior in local areas remains unclear. Additionally, there is limited information on the sources in local areas. Here, we investigated the seasonal trends of ClPAHs associated with fine particles (PM2.5) at two sites near heavy traffic roads to evaluate the local atmospheric behaviors and sources. The annual mean concentrations of total ClPAHs at the north (site A) and south side (site B) across a heavy traffic road were 12.0 and 19.2 pg/m3, respectively. The higher concentration at site B was further emphasized during the winter season; at that time the site was located beneath the wind from the traffic road. In addition, for individual ClPAHs, the behaviors of 8-chlorofluoranthene (8-ClFluor) and 7-chlorobenz[a]anthracene (7-ClBaA) were consistent with the frequency of the north wind, suggesting that 8-ClFluor and 7-ClBaA have the ability to be indicators of vehicle exhaust. Diagnostic ratios using these specific ClPAHs during high-traffic activities provided the specific values to differentiate the impacts.
1. Introduction Polycyclic aromatic hydrocarbons (PAHs) are recognized as typical hazardous pollutants because some of them act as carcinogenic/mutagenic and endocrine-disrupting compounds (Baek et al., 1991). To evaluate the human health risks for PAHs in more detail, a number of researchers have investigated not only their biological toxicities but ∗
also their environmental behaviors and sources (Bostrom et al., 2002; Kim et al., 2013). PAHs are mainly produced by the incomplete combustion of organic materials. Therefore, the principal sources are present anywhere human activity occurs while the impacts continually vary among countries, areas, seasons, and dates. In urban cities, vehicle exhaust has been frequently identified as a dominant source of atmospheric PAHs, that is, traffic activity can greatly impact human health
Corresponding author. Faculty of Agriculture, Meijo University, 1-501 Shiogamaguchi, Nagoya, 468-8502, Japan. E-mail address: [email protected] (T. Ohura).
https://doi.org/10.1016/j.atmosenv.2019.116944 Received 6 May 2019; Received in revised form 26 August 2019; Accepted 28 August 2019 Available online 31 August 2019 1352-2310/ © 2019 Elsevier Ltd. All rights reserved.
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Fig. 1. Map of the marked sampling sites.
2. Experimental
(Abdel-Shafy and Mansour, 2016). Additionally, not only PAHs but also PAH derivatives such as nitrated PAHs are emitted from diesel engine exhaust (Schuetzle et al., 1982). Chlorinated PAHs (ClPAHs) are a group of PAHs derivatives, some species of which are ubiquitously detected in the environment (Ohura, 2007; Ohura et al., 2014; Sun et al., 2013; Wang et al., 2018). Recently, we investigated the atmospheric behaviors of ClPAHs, including the spatial distributions and temporal trends, followed by a comparison to the concentrations of PAHs (Ohura et al., 2016, 2018, 2019). Throughout these studies, we observed that the compositions of ClPAHs in the air vary significantly compared with PAHs. We concluded that atmospheric PAHs could be produced by similar sources across Japan, whereas the sources of atmospheric ClPAHs changed depending on the site and season. This difference implies that the atmospheric fates of ClPAHs might be short compared with PAHs, so that the distribution of atmospheric ClPAHs over wider regions is limited. To date, numerous researchers have carried out studies on the sources of ambient PAHs. These studies have sought to better understand global air pollution, in which the influences of global long-range transport are investigated, suggesting that the overall contributions have become too large to ignore (Björseth et al., 1979; Keyte et al., 2013; Yang et al., 2007). However, few studies have investigated the behaviors of air pollutants between short distances, that is, short-range transport. In particular, for ClPAHs, there is no information available, to the best of our knowledge. In this study, we survey ClPAHs found in PM2.5 (particulate matter that have diameters less than 2.5 μm) near two sites (ca. 800 m distance) across heavy traffic roads. The differences in the profiles caused by the short-range transport of ClPAHs were evaluated, and we propose new ClPAH diagnostic ratios for identifying the traffic sources.
2.1. Chemicals This study targeted 25 species of ClPAHs with three to five aromatic rings. Most of these ClPAHs were synthesized ab initio; the detailed procedures were reported previously (Ohura et al., 2005). In addition, we analyzed 18 PAHs, including 16 that have been classified as priority pollutants by the U.S. Environmental Protection Agency. The PAH mixture standard was purchased from LGC Labor GmbH (Augsburg, Germany). As petroleum markers, a standard solution of five hopanes and five steranes (NIST SRM 2266) was purchased from Sigma-Aldrich (Saint Louis, MO, USA). Three deuterated PAHs (phenanthrene-d10, fluoranthene-d10, and perylene-d12), purchased from Cambridge Isotope Laboratories, Inc. (Andover, MA, USA), were used as internal standards. The extractions and purifications were performed with analytical grade solvents purchased from Wako Pure Chemical (Osaka, Japan) or Kanto Chemical (Tokyo, Japan). The 25 ClPAHs used in this study were abbreviated as follows: 9chlorophenanthrene (9-ClPhe), 1,9-dichlorophenanthrene (1,9Cl2Phe), 3,9-dichlorophenanthrene (3,9-Cl2Phe), 9,10-dichlorophenanthrene (9,10-Cl2Phe), 3,9,10-trichlorophenanthrene (3,9,10-Cl3Phe), 2-chloroanthracene (2-ClAnt), 9-chloroanthracene (9ClAnt), 9,10-dichloroanthracene (9,10-Cl2Ant), 3-chlorofluoranthene (3-ClFluor), 8-chlorofluoranthene (8-ClFluor), 1,3-dichlorofluoranthene (1,3-Cl2Fluor), 3,4-dichlorofluoranthene (3,4-Cl2Fluor), 3,8-dichlorofluoranthene (3,8-Cl2Fluor), 1-chloropyrene (1-ClPy), dichloropyrene (Cl2Py), trichloropyrene (Cl3Py), tetrachloropyrene (Cl4Py), 6-chlorochrysene (6-ClChry), 6,12-dichlorochrysene (6,12-Cl2Chry), 7-chlorobenz[a]anthracene (7-ClBaA), 7,12-dichlorobenz[a]anthracene (7,12Cl2BaA), tetrachlorofluoranthene (Cl4Fluor), 6-chlorobenzo[a]pyrene (6-ClBaP), dichlorobenzo[a]pyrene (Cl2BaP), and trichlorobenzo[a] pyrene (Cl3BaP). The 16 PAHs used in this study were abbreviated as
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analyses, the initial oven temperature of 100 °C was held for 2 min. The temperature was then increased at 25 °C/min to 200 °C, increased at 5 °C/min to 300 °C, and finally held at 300 °C for 15 min. The PAHs and petroleum markers (hopanes and steranes) were analyzed by the same GC conditions, that is, the initial oven temperature of 70 °C was held for 3 min. The temperature was then increased at 20 °C/min to 240 °C, increased at 5 °C/min to 310 °C, and finally held for 30 min. The injector temperatures for the ClPAH and PAH analyses were 300 and 280 °C, respectively. For both the ClPAH and PAH analyses, the transfer line and ion source temperature was 280 °C. The mass spectrometer was operated in both the selected ion monitoring and electron impact ionization mode. A constant ion current of 200 μA was used, and the electron energy was 70 eV. The target ions of the ClPAHs and PAHs were described elsewhere (Kamiya et al., 2015). The target ions of the petroleum markers for quantification and confidence were m/z 191 and 217, and m/z 355, 357, 371, 383, 385, 387, 397, and 411, respectively. Twelve species of elements were targeted as follows, including abbreviations: vanadium (V), chromium (Cr), manganese (Mn), iron (Fe), nickel (Ni), copper (Cu), zinc (Zn), arsenic (As), selenium (Se), and lead (Pb). These elements were determined using energy dispersive X-ray fluorescence spectrometry (EDXRF, Rigaku Corp., Tokyo, Japan) coupled with fundamental parameter quantification. The analytical conditions and quality controls were confirmed and reported elsewhere (Okuda et al., 2013).
follows: naphthalene (Nap), acenaphthylene (Acy), acenaphthene (Ace), fluorine (Flu), phenanthrene (Phe), anthracene (Ant), fluoranthene (Fluor), pyrene (Py), benzo[a]anthracene (BaA), chrysene (Chry), benzo[b]fluoranthene (BbF), benzo[k]fluoranthene (BkF), benzo[e]pyrene (BeP), benzo[a]pyrene (BaP), perylene (Pery), benzo [ghi]perylene (BghiP), dibenzo[a,h]anthracene (DBahA), and indeno [1,2,3-cd]pyrene (IP). The five hopanes used in this study were abbreviated as follows: 17α(H)-22,29,30-trisnorhopane (Tm), 17α(H),21β(H)-30-norhopane (29αβ), 17α(H),21β(H)-hopane (30αβ), 17α(H),21β(H)–22S-homohopane (31αβS), and 17α(H),21β(H)-22Rhomohopane (31αβR). The five steranes used in this study were αββ20R-cholestane, ααα20R-cholestane, αββ20R,24S-methylcholestane, αββ20R,24R-ethylcholestane, and ααα20R,24R-ethylcholestane. 2.2. Sampling sites and procedures The study was carried out in the city of Nagoya (latitude 35° 54′ N, longitude 136° 54′ E; Fig. 1). The detailed information of the city was described elsewhere (Ohura et al., 2016). Samples of PM2.5 were simultaneously collected at two sites (sites A and B) in Nagoya, from April 2012 through April 2013. These sites were located in an industrial area of southeast Nagoya; the distance between sites was ca. 800 m. Additionally, an express highway (ca. 60,000 vehicles/day) and two national roads (ca. 86,000 vehicles/day consist of 79% light- and 31% heavy-duty vehicles for Route 23 and ca. 43,000 vehicles/day consist of 87% light- 13% heavy-duty vehicles for Route 247) were located near the sites. Sampling at site A was performed at rooftop locations (20 m above ground level) at the Nagoya City Institute for Environmental Science (35° 5′ N, 136° 54′ E). Sampling at site B was carried out at ground level in the Motoshio park (35° 5′ N, 136° 55′ E), approximately 100 m from the nearest busy roads. PM2.5 samples were collected on quartz fiber filters (QFFs; 47 mm diameter, 2500 QAT-UP; Pall Corp. NY, USA) through an EPA well-type impactor ninety-six (Peters et al., 2001) for PM2.5, volumes of which were determined by a low-volume air sampler (Partisol Model 2000; Thermo Scientific, MA, USA) operating at a constant flow rate of 16.7 L/min. The sampling details for PM2.5 were described elsewhere (Yamagami et al., 2019). Samples were continuously collected for one week, after which the QFFs were changed during the sampling period. After atmospheric sampling, the QFFs were cut into pieces for the analysis of ClPAHs, PAHs, hopanes, steranes, and trace elements. The QFFs were wrapped in aluminum foil, sealed, and stored in a freezer at −35 °C until extraction.
For quality assurance of the analytical procedures, a recovery test of the target compounds using a sequence of analytical procedures was performed. Recoveries were confirmed to be reasonable to our previous work (Kamiya et al., 2015). Recoveries of the internal standards spiked onto the samples at sites A and B were 100 ± 4% for fluoranthene-d10 and 104 ± 13% for perylene-d12 and 97 ± 7% for fluoranthene-d10 and 119 ± 14% for perylene-d12, respectively. The target ClPAHs were not detected in the procedural blank samples. The limits of detection were estimated by multiplying the standard deviations of the concentrations found in the six repeated analyses of a dilute standard by a solvent (Kamiya et al., 2015). If the sampled air volume was 160 m3, the ClPAH limits of detection ranged from 0.025 pg/m3 for 9-ClPhe to 0.78 pg/m3 for Cl3BaP, and the PAH limits of detection ranged from 0.013 pg/m3 for Phe to 0.10 pg/m3 for BbF. For the petroleum markers, the limits of detection ranged from 0.09 pg/m3 for ααα20R-cholestane to 0.95 pg/m3 for 31αβS.
2.3. Sample extraction and analysis
3. Results and discussion
Samples collected at each site and during each week were treated as a single sample. After air sample collection, a part of each filter was cut using stainless steel scissors and then extracted using a 1:1 mixture of hexane and dichloromethane in an accelerated solvent extraction apparatus (Dionex-ASE 350, Sunnyvale, CA, USA). Before extraction, an internal standard mixture consisting of fluoranthene-d10 (25 ng), and perylene-d12 (25 ng) was spiked onto the filters. The eluate was concentrated to ~200 μL under a gentle stream of N2 at 45 °C. During these operations, the solutions were protected from light to prevent photochemical degradation of the analytes. Each condensed solution was purified using column chromatography with silica gel (Supelclean LC-Si SPE tube, Sigma-Aldrich Co., St. Louis, MO, USA) and was eluted with 13 mL of 10% dichloromethane in n-hexane. Prior to injection, phenanthrene-d10 (25 ng) was added to the residue as a recovery standard. The extracts were analyzed with a JMS-Q1000GC quadrupole mass spectrometer (JEOL, Tokyo, Japan) equipped with a 7890 A gas chromatograph (Agilent Technologies Inc., Santa Clara, CA, USA) fitted with an InertCap 5MS/NP capillary column (30 m long, 0.25 mm i.d., 0.25 μm film thickness; GL Science Inc., Tokyo, Japan). Helium was used as the carrier gas at a flow rate of 1.0 mL/min. For the ClPAH
3.1. Concentrations of pollutants in PM2.5
2.4. Quality assurance
3.1.1. ClPAHs and PAHs Throughout the campaign, most of the target ClPAHs and PAHs were detected from a PM2.5 fraction at each site. For individual ClPAHs, the mean concentration of 6-ClBaP was the highest at both sites. The mean concentrations at sites A and B were 4.76 and 6.14 pg/ m3, respectively (Table S1). The second most abundant species was 7,12-Cl2BaA (1.22 pg/m3) at site A and 1-ClPy (2.02 pg/m3) at site B, which indicates the different behaviors of ClPAHs despite the sites being near each other. This difference might be caused by the presence of different sources of ClPAHs at each site. For the total ClPAHs, the mean concentrations at sites A and B were 12.0 and 19.2 pg/m3, respectively. Overall, the concentrations of ClPAHs at site B tended to be higher (1.6 times on average) than those at site A. Furthermore, these concentrations were a few times lower than that of the mean level surveyed over other Japanese locations. These results suggest that the ClPAHs at the current sites may be influenced by the limited sources. For PAHs, the overall concentrations were higher at site B (from 1.2 to 4.3 times) compared with site A (Table S1). Among them, BbF was 3
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Fig. 2. Temporal concentrations of (A) total ClPAHs, (B) total PAHs, (C) total hopanes, and (D) total sterane at sites A and B, and the variation in the concentration ratios of site B to site A.
city, British Columbia (Ding et al., 2009), whereas they were considerably lower (ca. ~50 times) than those observed in a highly polluted city, New Delhi (Pant et al., 2015).
the most abundant species at both sites, BkF at site A and Chry at site B were second, and Chry at site A and BghiP at site B were third. The orders of magnitude of the PAHs were also different at each site, such as ClPAHs. Such predominant PAHs were observed in the results from a suburban area of Zhengzhou, China, in which various sources such as coal combustion, vehicles, coking plants, and biomass burning were estimated as the main sources (Wang et al., 2015). Therefore, the influential sources of the PAHs might also be different between the sites. The mean concentrations of total PAHs at site A and B were 1.61 and 2.37 ng/m3, respectively. The levels were comparable to those reported from surveys of European areas (Jedynska et al., 2014), and were approximately two times lower than that of the mean level surveyed over other Japanese sites.
3.1.3. Trace elements Analyses of the trace elements in the environmental samples offer another useful approach to explore the possible sources. We focused on the occurrences of nine species of anthropogenic trace elements in PM2.5. Among them, Zn was found to be the most abundant species (35.3 ng/m3 at site A and 43.8 ng/m3 at site B, on average), followed by Mn, Cu, and Ni at both sites (Table S1). These four elements accounted for 75.5 and 77.0% of the total mass of elements at sites A and B, respectively. In the case of trace elements, the concentrations also tended to be higher at site B than that at site A. Zn has been associated with industrial emissions and incineration (Duvall et al., 2012; Harrison et al., 1997). Although Mn has often been used as a marker for crustal dust/soil, Cu and Ni are supplied by anthropogenic activities such as fossil fuel burning and industrial activities (Tian et al., 2012; Viana et al., 2008). In particular, Cu has been used as a traffic source marker because it is associated with brake wear (Almeida et al., 2005; Lawrence et al., 2013). Comparing the relationships of the concentrations among trace elements throughout the seasons, significant correlations (p < 0.05) were observed among the various elements (data not shown). For instance, the concentrations of Mn correlated to those of both Zn and Cu at each site. Zn and Cu are not only major additives to lubricating oils but also associated with other transportation activities such as brake and tire wear (Lough et al., 2005).
3.1.2. Biomarkers Hopanes and steranes are natural products, and are considered to be petroleum markers (Simoneit, 1984). In addition, hopane and sterane homologues are relatively stable and have similar atmospheric fate processes. The mean concentrations of total hopanes at sites A and B were 0.252 and 0.420 ng/m3, respectively. Among the hopanes detected, on average 30αβ was the most abundant species at both sites, followed by 29αβ and 31αβR. This similar order of magnitude has been observed in the results of a traffic-related aerosol sample study (Wang et al., 2007). In addition, Cui et al. (2016) reported that the high proportions of 30αβ and 29αβ are attributable to vehicle exhaust emission. Therefore, the main source of the PM2.5 samples used in this study could be considered to be vehicles. Conversely, the mean concentrations of total steranes at sites A and B were 0.0212 and 0.0867 ng/m3, respectively. Among the individual sterane concentrations, αββ20R,24R-ethylcholestane was the most abundant species at both sites, followed by ααα20R,24R-ethylcholestane, and then αββ20R,24Smethylcholestane or αββ20R-cholestane. However, atmospheric steranes have limited information available compared with hopanes; both species have been assigned as traffic markers (Simoneit, 1984). Comparing the sterane concentrations to other sites, the present concentrations were somewhat consistent with the levels found in an urban
3.2. Seasonal trend Fig. 2 shows the annual concentrations of (A) total ClPAHs, (B) total PAHs, (C) total hopanes, and (D) total steranes in PM2.5 at sites A and B, and the variation in the concentration ratios at site B to site A. The concentrations of total ClPAH showed a typical seasonal trend at both sites, a decrease during the warmer season and an increase during the colder season (Fig. 2A). Such seasonal variation is caused by 4
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predominant at both sites. A previous study reported that the homohopane index (31αβS/31αβS+31αβR) with a range of 0.54–0.67 was found in mineral-oil-derived emission sources, whereas the index values with a range of 0.05–0.35 was identified as coal smoke (Schnelle-Kreis et al., 2007). In our samples, the homohopane index, except the summer season, remained at around 0.5 (Fig. S3). These results suggest that the PM2.5 collected during the colder seasons at these sites were adequately considered as coming from vehicle exhaust rather than that of coal combustion.
competitive factors such as meteorological conditions, including wind conditions, ambient temperatures, and hours of daylight, but characteristics of the sources and the physicochemical properties of the substances, which has been frequently observed in global atmospheric PAHs (Shen et al., 2013). Indeed, the trend was also observed in the seasonal variations of the total PAHs in PM2.5 in the present study (Fig. 2B), which were consistent with those of the total ClPAHs. Furthermore, concerning total hopanes (Fig. 2C) and steranes (Fig. 2D), the seasonal variations of the fossil petroleum markers showed a similar pattern to the ClPAHs and PAHs. The total hopanes and steranes decreased during the warmer season and increased during the colder season. Such seasonal variations have been frequently observed in past studies that analyzed ambient particles (Ding et al., 2009; Jedynska et al., 2014). Among the trace elements, Fe and Cu are considered to be indicators of traffic sources. Those two elements showed very different trends compared with other trace elements. The concentrations of V, Cr, and Zn, known as markers of anthropogenic sources, such as industrial activities, tended to decrease during the colder season at both sites, whereas neither Fe nor Cu displayed such variation throughout the seasons or increased during the colder season (Fig. S1). Specifically, the concentration levels at site B were increased during the colder season and were considerably higher than those at site A. These results suggest that the contributions of industrial activities and traffic sources in PM2.5 at these sites changes during the seasons. That is, it implies that PM2.5 could be driven by seasonal factors such as wind direction.
3.5. Diagnostic ratios of ClPAHs for traffic impacts As described above, the concentrations of most air pollutants at site B tended to be notably higher than those at site A during the colder season with a blowing north wind, and were estimated to be strongly influenced by traffic exhaust. For site A, the air pollutants might be impacted less by traffic sources during the summer season with a blowing south wind because of the influence of various industrial sources located in the south area and photodegradation, as showing in Fig. 4. In other words, air pollutants at site B could be directly affected by the traffic sources during the colder season compared with site A. Therefore, the concentrations of pollutants at site B implied that those at site A directly reflect the traffic impacts because the superfluous factors affecting the concentrations in site B were eliminated. Indeed, the determined concentrations of hopanes were frequently increased by northern winds. To evaluate the behaviors of ClPAHs by the traffic impact, we analyzed the data and subtracted the concentrations of individual ClPAH at site A from the corresponding values at site B. The seasonal trends of the subtracted concentrations of typical ClPAHs were roughly categorized into two patterns: i) high concentrations with a north wind, and ii) equal concentrations over the seasons (Fig. 5). At a high frequency of the north wind, the subtracted concentrations indicate the influence of a major road located across the north section of site B. The ClPAHs showing high concentrations with a north wind were 9-ClPhe, 8-ClFluor, 7-ClBaA, and 6-ClBaP, and were strongly affected by the traffic sources. However, 3-ClFluor and 1-ClPy appear to be barely affected by the traffic sources, because the subtracted concentrations were somewhat constant over the season. In addition, 3-ClFluor, 8-ClFluor, and 7-ClBaA are relatively photostable compared with other ClPAHs (Ohura et al., 2008). Based on this evidence, the impacts of the traffic sources could be estimated by the diagnostic ratios combining ClPAHs that are high and low responsible for traffic with photostable. Consequently, [8-ClFluor]/[3-ClFluor] and [7ClBaA]/[3-ClFluor] were the most suitable traffic indicators of ClPAHs. Indeed, the values of [8-ClFluor]/[3-ClFluor] and [7-ClBaA]/[3ClFluor] increased during the traffic impact, and were estimated to be 2.2–3.2 and 4.2–7.0 respectively. Our previous studies on the ambient ClPAHs in urban cities in Japan found the values of [7-ClBaA]/[3ClFluor] ranged from 0.66 to 9.8 (Ohura et al., 2016, 2019). Furthermore, the ratio ranged from 0.16 to 0.91 in the surveys of other Asian countries (Kakimoto et al., 2014; Ma et al., 2013). These results suggest that vehicle exhaust, rather than other sources, might be a limited source of ClPAHs in ambient air. Indeed, the sources of ambient ClPAHs can be specific for each local site compared with those of PAHs. That is, such diagnostic ratios using specific ClPAHs are likely to adapt to the analysis of local sources. To evaluate the potential of the diagnostic ratios, further studies are needed.
3.3. Influence of wind direction We investigated the influence of the ClPAH and PAH concentrations by the wind direction at each site. Fig. 3A shows the wind direction frequency and mean wind speed at each sampling campaign. The mean wind speed was 2.5 m/s, and there were roughly two principal classifications of the wind direction over the sampling period. The data suggests that the southeast and northwest winds frequently occurred during the warmer and colder seasons, respectively. Fig. 3B–E shows contour plots with a polar axis of total ClPAH or total PAH concentrations at the rate of frequent wind direction. These plots indicate that the ClPAH and PAH concentrations were elevated by highly frequent northwest wind rather than by southeast wind. Interestingly, the concentrations of ClPAHs (Fig. 3C) and PAHs (Fig. 3E) at site B were more emphasized than those at site A (Fig. 3B and D) when a northwest wind blew. Conversely, the southeast wind contributed less to the concentrations of ClPAHs and PAHs at both site. These findings suggest that there are principal sources related to ClPAHs and PAHs from the northwest, especially, high-impact sources will be present between sites A and B. Therefore, ClPAH and PAH concentrations at site B might be influenced by various sources such as not only near automobile (Source II) but away factories and other anthropogenic activities (Source I) located on the windward side, so that the concentrations at site B were higher than those at site A (Fig. 4). 3.4. Source analysis Figure S2A shows the concentration relationships between the ΣClPAHs and ΣPAHs at each site. Significant correlations (p < 0.01) were observed at both sites, suggesting that the ClPAHs could be produced by the same PAH sources extending to the surroundings at these sites. Indeed, the concentrations of the ΣClPAHs and ΣPAHs at site A showed significant correlations to those at site B (Fig. S2B). These data indicate that the current PM2.5 level was produced by similar sources. To identify the sources, we analyzed the behaviors of the petroleum marker hopanes. For certain hopanes, the concentration ratios of the constitutional isomers are used to identify the origins of the fossil sources. According to Simoneit (1999), the proportions of 29αβ and 30αβ in the hopanes tend to be dominant in vehicle exhaust emissions. In our study, the proportions of 29αβ and 30αβ were relatively
4. Conclusion In this study, we investigated ClPAHs in PM2.5 at two sampling sites in heavy traffic areas throughout the seasons. A major traffic road was between the two sites; thus, the differences of the air pollutant concentrations between the sites could be expressive of the behavior of short-range transport. The concentrations of ClPAHs at the two sites were regularly varied by the wind direction, evidence that the ClPAHs 5
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Fig. 3. (A) Wind direction frequency and wind speed in the sampling area during the campaign. The circle sizes and colors represent the wind speed (m/s) and seasons: spring (gray), summer (red), autumn (blue), winter (green). Contour maps of total ClPAH (B, C) and total PAH (D, E) concentrations in frequency of each wind direction at site A (B, D) and site B (C, E). These rotations of axis 0 represent the north direction. (For interpretation of the references to color in this figure legend, the reader is referred to the Web version of this article.)
interests or personal relationships that could have appeared to influence the work reported in this paper.
are caused by vehicle exhaust. Specifically, the atmospheric behaviors of 8-ClFluor and 7-ClBaA were unique among the ClPAHs, which were consistent with those of hopanes. We proposed new diagnostic ratios using these specific ClPAHs to estimate traffic sources. Applying the ratios to other surveyed sites, it suggests that the contributions of traffic sources to ClPAH pollution would be slightly high.
Acknowledgments This work was supported in part by the Ministry of Education, Culture, Sports, Science and Technology, via a Grant-in-Aid for Scientific Research (C) (No. 26340015), and a Grant-in-Aid for Scientific Research (B) (No. 16H05625 and 17H01865).
Declaration of competing interest The authors declare that they have no known competing financial 6
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Fig. 4. Schematic model of the influence of air pollutants in the present sampling sites from the north wind.
Fig. 5. Frequency of north and south winds and typical ClPAH concentrations at site A subtracted from site B throughout the campaign.
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