Traffic-Generated Changes in the Chemical Characteristics of Size-Segregated Urban Aerosols

The road traffic impact on the concentrations of 13 fractions of particulate matter (PM) and their components was assessed. PM was sampled at two points in Katowice (southern Poland), a background point beyond the effects of road traffic, and a near-highway traffic point. The samples were analyzed for organic and elemental carbon, 8 water-soluble ions, 24 elements, and 16 polycyclic aromatic hydrocarbons (PAHs). The traffic emissions (mainly particles from car exhaust) enriched the ultrafine, submicron, and fine PM particles with elemental carbon. The traffic-caused re-suspension of the road and soil dust affecting the concentrations and chemical composition of the coarse PM fraction. However, for each PM fraction, the carcinogenic equivalent ratios, assumed as a measure of the hazard from 16 PAHs in this paper, were similar at the two sampling points. The traffic emissions from the highway appeared to have a weaker influence on the concentrations and chemical composition of PM in a typical urban area of southern Poland than elsewhere in Europe.

The area was a living district surrounded by greenbelts, about 2,000 m from the city center. The meteorological conditions at the point were typical of the region. The second point, HW (Fig. 1), was a traffic point located on the shoulder of the highway A4, about 1,500 m south of the city center and 1,200 m northwest of the point UB. Average traffic density at HW was 30,000 cars per 24 h. To the north of HW, there was a living quarter. To the south, there was an open area of the A4 highway, and further behind A4, about 300 m from the point, there were some public utilities and detached houses of another living quarter. PM was sampled with the use of two thirteen-stage low pressure DLPI impactors (Dekati Ltd, Kangasala, FL).
In Katowice, especially in winter, traffic PM emissions can be obscured by emissions of PM from combustion of fossil fuels or biomass for residential heating . Therefore, the sampling period was selected beyond the major heating season, and the samples were taken within mid-March to mid-June of 2012. There were nine, about one-week long (from 142 to 173 h), pairwise sampletakings.
The mass of the collected PM was determined by weighing the substrates before and after the exposure; a MYA 5.3Y.F micro balance (RADWAG; Radom, PL) was used (1 lg resolution). Before each weighing, the substrates were conditioned for 48 h in the weighing room (relative air humidity 45 % ± 5 %, air temperature 20 ± 2°C). The substrates and impactors were prepared for exposure in a laminar chamber. After weighing, the exposed substrates were stored in a freezer until analysis.
Two kinds of Whatman (GE Healthcare Bio-Sciences Corp.; Piscataway, NJ, USA) substrate filters were used. Alternating between the sample-takings, quartz fiber filters (QMA, ø25 mm, CAT No. 1851-025) and nylon membrane filters (0.2 lm, ø25 mm, Cat No. 7402-002) were used; the former four and the latter five times. Always the same type substrates were used on all impactor stages, with the same type simultaneously used at both sampling points. Altogether, nine samples (four on quartz and five on nylon filters) of each of the 13 PM fractions were taken at each point over the entire measuring period.
At each point, four samples of each fraction were taken on quartz filters. Two equal (1.5 cm 2 ) fragments were cut out from each quartz filter just before the analysis. The PM on one of them was analyzed for organic carbon (OC) and elemental carbon (EC). The remaining fragments, four for each of the 13 fractions and each point, were used to make fraction samples. A fraction sample for a point was made by putting together the four filter fragments containing this PM fraction sample from this point. Each of these 26 fraction samples was analyzed for 16 PAHs. 1   Five samples of each fraction were taken on nylon filters at each sampling point. The PM on each nylon filter was analyzed for elemental composition first (i.e. for Al, Si, Sc, Ti, V, Cr, Mn, Fe, Co, Ni, Cu, Zn, As, Se, Br, Rb, Sr, Mo, Ag, Cd, Sb, Te, Ba, Pb), then it was extracted in water and concentrations of the main ions (Cl -, NO 3 -, SO 4 2-, Na ? , NH 4 ? , K ? , Ca 2? , Mg 2? ) were determined in the extracts. The OC and EC contents of dust were determined with the use of a Lab OC-EC Aerosol Analyzer (Sunset Laboratories Inc.; Portland, OR, USA). Sixteen PAHs 1 in the PM fractions were determined on a Clarus 500 gas chromatograph (PerkinElmer; Waltham, MA, USA). The analytical procedures, equipment, and quality control for OC, EC, and PAH in PM are described in  and .
The elemental composition of PM was determined by means of energy dispersive X-ray fluorescence (EDXRF). An Epsilon 5 instrument (PANalytical B.V.; Almelo, NL), calibrated with thin-layer single-element standards (Micromatter; Vancouver, BC,CA) was used to measure total concentrations of the elements. To control the performance of the analytical procedure, samples of reference material (SRM2873, NIST, Gaithersburg, MD, USA) were measured weekly. The recoveries were between 85 % and 120 % of the certified values (except 52 % and 39 % recoveries of V and Co) and the X-ray tube and detector drifts were monitored monthly. The detection limits were from 0.18 ng/cm 2 (Se) to 19.6 ng/cm 2 (Si).
The water extracts of PM were made by ultrasonizing the substrates containing the samples in 25 cm 3 of deionized water for 60 min at 15°C, then shaking for about 12 h (18°C, 60 r/min). The ion content of extracts was determined using an ion chromatograph (Metrohm AG; Herisau, CH). The method was validated against the CRM Fluka products nos. 89316 and 89886; the standard recoveries were 92 %-109 % of the certified values, and the detection limits were: 10 ng/cm 3 for NH 4 ? , 18 ng/cm 3 for Cl -, SO 4 2and K ? , 27 ng/cm 3 for NO 3 and Na ? , and 36 ng/cm 3 for Ca 2? and Mg 2? .

Results and Discussion
The average concentrations for the sampling period are presented in Table 1. In Katowice, the core PM mass consisted of particles with D p between 0.17 and 1.6 lm (PM 0.17-1.6 ). They contributed 72 % to the total PM mass at UB and 64 % at HW (Table 1). The density function of the PM mass size distribution assumes the absolute maximum between 0.4 and 1.0 lm at both sampling points (Fig. 2). It is clear that the concentrations of PM 0.17-1.6 were not significantly affected by road traffic.  The concentrations of PM 0.17-1.6 and of its chemical components did not differ between the points except for (usually) traffic-related Fe, Co, Ni, and Mo (Pant and Harrison 2013), and for EC.
In general, the fractionated EC concentrations were higher at HW than at UB. Within PM 0.17-1.6 , the concentrations of PM 0.17-0.26 -, PM 0.26-0.4 -, PM 0.4-0.65 -, PM 0.65-1 -, PM 1-1.6 -bound EC were from 40 % to [300 % higher at HW than at UB. The concentrations of PM 0.03-0.06 , PM 0.06-0.108 , PM 0.108-0.17 -related EC were from 4 to 8 times greater at HW than at UB. The striking difference in the behavior of the concentrations of PM-related EC between HW and UB can be seen by comparing the density functions of the mass distribution (relative to particle size) of EC bound to the nuclei and accumulation modes at HW and UB (Fig. 2).
Also, the concentrations of PM 0.03-0.17 and of PM 0.03-0.17 -bound OC were higher at HW than at UB (averages of the concentrations of PM 0.03-0.06 , PM 0.06-0.108 , PM 0.108-0.17 , and of the related OC were higher by 84 % and 54 %, respectively). The concentrations of PM 0.03-0.17related Fe, Cu, and Zn were also elevated at HW. The density functions for some PM-related metals (e.g. Zn, Fig. 2) had maxima in the interval 0.03-0.17 lm at HW; these maxima did not exist at UB. Great differences in the concentrations of PM 0.03-0.06 -, PM 0.06-0.108 -, and PM 0.108-0.17 -bound EC, OC, Fe, Cu, and Zn between the points are due to the chemical composition of car exhaust: car exhaust consists mainly of ultrafine particles of EC, OC (incomplete fuel combustion), substances synthesized from residues of fuel combustion, and lubricant or fuel additives (metals and their compounds, e.g.) (Geller et al. 2006;Maricq 2007).
Surprisingly, despite the significant PAH content of exhaust OC, the ambient concentrations of fine particlebound P PAH did not differ significantly between the points except for the PM 0.06-0.108 -bound P PAH, higher at HW than at UB by 90 %. The density function of the PMbound P PAH distribution had the maximum in 0.06-0.108 lm. However, some from among the 16 PAHs had significantly higher concentrations at HW than at UB. The concentration of PM-bound BaP at HW was almost twice that for UB (4.3 and 2.3 ng/m 3 , respectively). The significantly higher PM 0.03-1.6 -bound BaP concentration at HW than at UB allows for consideration of BaP as a marker for exhaust emissions in urban areas of Upper Silesia in the periods of lower municipal emissions.
The concentrations of coarser PM (D p [ 4.4 lm) and of some its components (OC, EC, Na ? , K ? , Cl -, Ca 2? , Al, Si, Fe, Cu, Zn, Mo) also were higher at HW. For example, the average concentrations of PM 4.4-6.8 -, PM 6.8-10 -, and PM 10-40 -related Na ? and K ? were almost four times greater at HW than at UB, while that for Ca 2? was approximately two-fold greater.
The PM mass size distribution was bimodal at UB. At HW, in 6.8-10 lm, the third peak for its density function appeared. Also the maxima of the density functions for Na ? , K ? , Cl -, and Ca 2? occurred in this interval. The coarse ambient particles containing OC, EC, Na ? , K ? , Cl -, Ca 2? , Al, Si, Fe, Cu, Zn, Mo are probably a mixture of the particles of worn tires, brake linings, chassis, other car parts, road surface, soil, etc. (Wahlin et al. 2006;Thorpe and Harrison 2008;Pant and Harrison 2013). Such particles, involved in the process of alternate lifting (by passing cars, wind, etc.) and deposition, can stay within the road vicinity for a long time.
The chemical mass closure of PM was checked at both points. There were six groups of identified chemical PM components, and the seventh one, unidentified matter (UM); their masses together made the total PM mass.  Table 1, i.e. the analytically determined PM-bound chemical elements (including K ? , Ca 2? , and Mg 2? ), were used to define crustal matter (CM) and trace elements (TE). The contents of CM and TE depended on both the PM fraction and the measuring point. These PM-bound elements were divided into two groups based on their enrichment factors EF . EF expresses how anthropogenic an element is: the closer an element's EF value is to 1, the smaller the anthropogenic contribution to the element's concentration in air.
The elements with EF B 20 were assumed crustal (of natural origin). The chemical compounds containing an element with EF B 20 that were known to contribute significantly to the PM mass were included in CM; if there were no such compounds (known) for an element, or the element was probable to occur in numerous compounds, only the element was included in CM. CM included CO 3  (Marcazzan et al. 2001).
The chemical elements with EF [ 20 were assumed to be anthropogenic; these many-source origin elements occurred in many PM-bound compounds that were not determined because of their multiplicity. These elements alone were put in TE. TE consisted of K ? (only for PM 0.4-0.65 and PM 0.06-0.108 at UB), Sc, Ti, V, Cr, Mn, Fe (Fe at both points for PM 0.4-1.6 only), Co, Ni, Cu, Zn, As, Se, Br, Rb (Rb at both points for PM 0.17-1.6 only), Sr (only for PM 0.4-1.6 and for PM 0.03-0.26 at HW and for PM 0.03-1 at UB), Mo, Ag, Cd, Sb, Te, Ba (Ba only for PM 0.03-2.5 and PM 6.8-10 at both points), and Pb. The mass [TE] of TE was assumed to be the sum of all the analytically determined masses of the elements from TE. In Fig. 3, the mass distributions of the PM component groups (relative to particle size) at both sites are presented. Within particular PM fractions, the identified matter accounted for the mass in 52 %-98 % at HW and in 55 %-96 % at UB. For all fractions except PM 4.4-6.8 , the mass of identified compounds was higher at UB than at HW.
The shares of UM in twelve PM fractions were greater, and in ultrafine PM much greater, at HW than at UB (Fig. 3). At HW, the UM shares in PM 0.06-0.65 and in particles with D p [ 10 lm were greatest; whereas at UB, they were greatest in PM 0.108-0.65 and PM 1.6-4.4 . UM consists of the organic compounds and nitrates that evaporate during transportation and storage of PM samples, unidentified compounds, and water (Tsyro 2005). Also, inaccurate identification of compounds in OM, CM and TE affects UM. For example at HW, considering the high UM content of 12 fractions, [OM] should probably have been computed by multiplying [OC] A by a coefficient greater than 1.4 (Turpin and Lim 2001).
The EC concentrations were higher at HW, and the EC mass distribution differed from that at UB (Table 1; Fig. 2). Therefore, the EC contributions to particular PM fractions, especially to finer ones, at HW were higher than at UB (Fig. 3). The fractional mass contributions of OM to PM at both points were close, and only for the sub-fractions of PM 2.5-6.8 were they slightly higher (by 5 % in average) at HW.
Although the concentrations of the components of SIA (SO 4 2-, NO 3 -, NH 4 ?; Table 1) at both points were comparable, the distributions of SIA and its mass contributions to the fractions differed slightly between the points. The contributions to PM 0.03-0.108 and coarser particles (D p [ 4.4 lm) were on average greater by 4 and 5 % at UB, respectively.
NaCl was more abundant at HW than at UB only in PM 4.4-10 . This was due to the previously mentioned differences in the ambient concentrations of Na ? and Clbetween the points. Similarly, the elevated concentrations of the components of CM, especially of K ? , Ca 2? , Al, Si, and Fe, caused a slightly greater share of CM in PM 1.6-10 at HW (CM contribution to the PM mass at UB was greater than at HW by no more than 2 %). The mass contribution of TE to PM 0.65-1.6 was greater by about 2 % at HW than at UB. The NaCl, CM, and TE contributions to fine, especially to ultrafine, PM were higher at UB than at HW because ultrafine PM had noticeably lower concentrations at UB (Fig. 3). NaCl and TE contain elements coming from energy production, traffic or industry at UB, and mainly from traffic (both exhaust and non-exhaust emissions) at HW (Pant and Harrison 2013;Kumar et al. 2013;Rogula-Kozłowska et al. 2014). Also CM in ultrafine PM, despite being ''crustal,'' probably contains particles coming from exhaust gases and furnaces, including particles containing metal oxides from condensation (nuclei mode) (Geller et al. 2006;Maricq 2007).
The differences in the chemical composition of PM between the points were mainly in ultrafine and coarse PM. The distributions of the mass among the component groups differed extremely for PM 0.1 (in fact PM 0.03-0.108 ) between the two points (Fig. 3). PM 1 and PM 2.5 (in fact PM 0.03-1 and PM 0.03-2.5 ) differed between the points only in their EC content; PM 10 -in the EC and, greater by about 1 % at HW, CM contents (Fig. 3). The concentration of PM 0.1 was two times greater at HW than at UB (1.5 and 0.7 lg/m 3 ), while the concentrations of PM 2.5 and PM 10 at HW were greater by no more than 5 % (for PM 10 they were 28.6 lg/m 3 at HW and 27.3 lg/m 3 at UB).
So, although the number of cars per person grows closer to the European average, traffic emissions are not as significant a PM source in the Silesian Agglomeration as they are in other European regions (e.g. Viana et al. 2008). In Katowice, PM 0.17-1.6 contributes most to the PM mass, and traffic does not affect this PM fraction much. The concentrations and the chemical composition of this PM fraction within the Silesian Agglomeration are formed by municipal emissions, big power plants, and industry. The presented results show that human health is not more at risk near a big highway than some distance away. At both sampling points the carcinogenic equivalent ratios (CEQ) for particular PM fractions were close, both CEQ for total PM were equal to 4.2 lg/m 3 ( Table 1). As elsewhere in Poland, the concentration, chemical composition, and toxicity of PM in the Silesian Agglomeration depend on the season of the year; i.e., upon meteorological conditions and emissions from heat and power production facilities (Pastuszka et al. 2003, Majewski et al. 2011Rogula-Kozłowska et al. 2014).
The present pioneering research shows that some differences in the chemical characteristics of PM 4.4-10 between two sites in such a specific area as the Silesian Agglomeration can be found only by investigating particular fractions of PM 4.4-10 . However, the experiment should rather be considered as a pilot research and should be repeated on a wider scale at other traffic-affected sites, such as crossroads in a city center, street canyons, etc., to generalize the conclusions.