Seasonal variation of the size distribution of urban particulate matter and associated organic pollutants in the ambient air

Atmospheric Environment 43 (2009) 4557–4569

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Seasonal variation of the size distribution of urban particulate matter and associated organic pollutants in the ambient air Loukia P. Chrysikou, Constantini A. Samara* Environmental Pollution Control Laboratory, Department of Chemistry, Aristotle University, 54124 Thessaloniki, Greece

a r t i c l e i n f o

a b s t r a c t

Article history: Received 8 December 2008 Received in revised form 18 June 2009 Accepted 19 June 2009

Size-segregated samples of urban particulate matter (<0.95, 0.95–1.5, 1.5–3.0, 3.0–7.5, >7.5 mm) were collected in Thessaloniki, northern Greece, during winter and summer of 2007–2008, in order to study the size distribution of organic compounds such as polycyclic aromatic hydrocarbons (PAHs), aliphatic hydrocarbons (AHs) including n-alkanes and the isoprenoids pristane and phytane, organochlorine pesticides (OCPs) and polychlorinated biphenyls (PCBs). All organic compounds were accumulated in the particle size fraction <0.95 mm particularly in the cold season. Particulate matter displayed a bimodal normalized distribution in both seasons with a stable coarse mode located at 3.0–7.5 mm and a fine mode shifting from 0.95–1.5 mm in winter to <0.95 mm in summer. Unimodal normalized distributions, predominant at 0.95–1.5 mm size range, were found for most organic compounds in both seasons, suggesting gas-to-particle transformation after emission. A second minor mode at larger particles (3.0– 7.5 mm) was observed for C19 and certain OCPs suggesting redistribution due to volatilization and condensation. Ó 2009 Elsevier Ltd. All rights reserved.

Keywords: n-Alkanes Organochlorine pesticides PAHs Particle size distribution Seasonal pattern Thessaloniki

1. Introduction The particle size distribution affects the particulate matter’s transport in air, the dry or wet deposition from the atmosphere onto natural surfaces and the deposition in the human lungs. So, data about the particle size distribution of organic compounds is vital in order to estimate their inputs into the ecosystems and the human respiratory system (Wu et al., 2005). More importantly small particles penetrate the respiratory system, and cause direct health impact (Pagano et al., 1996), since the threat to humans connected with the inhalation of organic aerosols depends on the size (Hien et al., 2007). Therefore, particle size distribution of organic pollutants has been repeatedly studied in several areas around the world (e.g. Chen et al., 1996; Bi et al., 2005, Wu et al., 2006; Hien et al., 2007, Duan et al., 2007; Karanasiou et al., 2007; Lin et al., 2008). Aliphatic hydrocarbons (AHs) constitute a significant fraction of ambient aerosols ranging between 20 and 80% (Zheng et al., 1997). Particle-bound n-alkanes originate from natural and anthropogenic sources (Kalaitzoglou et al., 2004) and can be allocated using the size distribution of particular species (Karanasiou et al., 2007). Anthropogenic emissions are fossil fuel combustion, biomass

* Corresponding author. Tel.: þ30 310 997805; fax: þ30 310 997747. E-mail address: [email protected] (C.A. Samara). 1352-2310/$ – see front matter Ó 2009 Elsevier Ltd. All rights reserved. doi:10.1016/j.atmosenv.2009.06.033

burning and forest fires, while wind erosion of leaf epicuticular waxes, vegetation debris and degradation products are major natural sources of n-alkanes (Kalaitzoglou et al., 2004). The size distribution of n-alkanes is affected by a number of parameters such as vapour pressure, absorption and adsorption affinities, emission sources, atmospheric processes (Bi et al., 2005). Polycyclic aromatic hydrocarbons (PAHs) are among the aerosol constituents of greatest concern because some of these compounds are carcinogenic. PAHs are formed during the incomplete combustion of organic materials and several anthropogenic activities. The freshly generated PAHs are emitted in the form of gases and ultrafine particles (Richter and Howard, 2000). In the atmosphere, PAHs are partitioned between the gaseous and the particulate phase with the carcinogenic 5- and 6-ring species being mostly associated with particles of Dp < 2.0 mm (Allen et al., 1996; Cecinato et al., 1999; Venkataraman et al., 1999; Bi et al., 2005; Wu et al., 2006; Duan et al., 2005, 2007; Hien et al., 2007). Polychlorinated biphenyls (PCBs) and organochlorine pesticides (OCPs) are known to be toxic to humans and many organisms, and are considered widespread environmental contaminants, due to their properties according to chemical stability, lipophilicity and bioaccumulation (Garcia-Alonso et al., 2002). Most of these compounds exist in all environmental sectors and are susceptible to long-range atmospheric transport from major urban/industrial areas to surrounding terrestrial and aquatic ecosystems with harmful effects on man and wildlife (Hung et al., 2005; Li and

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L.P. Chrysikou, C.A. Samara / Atmospheric Environment 43 (2009) 4557–4569

Mcdonald, 2005; Alegria et al., 2006). Consequently, these chemicals are a topic of intense research interest, since the spatial distribution of OCPs in the atmosphere has been extensively even in remote regions, with no historical usage such as the Arctic (Li and Mcdonald, 2005). Although, most OCPs are severely restricted or banned in industrialized countries, they are detected in air of urban, rural and even remote areas (Yeo et al., 2003; Scheyer et al., 2005; Li et al., 2007; Yang et al., 2008). Therefore, it is essential to monitor OCPs, in an effort to evaluate the imposition of the restrictions, the atmospheric transport and the input from volatilization residues. Research concerning the size distribution of particle-bound PCBs and OCPs in the atmosphere has shown strong accumulation in the submicron size fraction (Chen et al., 1996; Chrysikou et al., 2009). The aim of the present study is to investigate the size distribution of particle-phase AHs, PAHs, OCPs and PCBs in the urban atmosphere and to examine seasonal variations. For this purpose, the size distribution of target compounds was studied at an urban site in the city of Thessaloniki, northern Greece, during cold and warm period. Seasonal variations were investigated in relation to prevailing meteorological conditions. 2. Experimental 2.1. Area description Thessaloniki (40 620 E, 22 950 N) is one of the most densely populated cities in Greece accounting for approximately 16,000 habitants km2. It is a coastal city surrounded by several residential communities with an extended industrial complex westerly and northwesterly. Oil refining, petrochemical, fertilizer and cement production iron and steel manufacture, non-ferrous metal smelting, etc. are some of the major industrial particle emission sources (estimated total annual emissions 8000 tonnes), whereas traffic and domestic heating are the urban sources respectively (estimated total annual emissions 700 tonnes). The climate in the area is temperate Mediterranean with mean monthly temperature values 4–28  C and relative humidity 47–80%. Prevailing winds (w75%) are weak with SW and NW directions. Frequently occurring calms result to inadequate dispersion of atmospheric pollutants and short-range transport processes (Samara and Voutsa, 2005). 2.2. Sampling and analysis Size-segregated aerosol samples were collected during winter (11–24 December 2007) and summer (3–27 June 2008) from a kerbside site in the centre of Thessaloniki surrounded by commercial shops and residential houses. The sampling system was situated on the roof (w3.0 m above ground level) of an air quality monitoring station located in a small green traffic island adjacent to a busy street with a traffic density w11,000 vehicles hour1. Sampling was performed using a high volume air sampler (Graseby-Andersen Ltd) equipped with a four-stage inertial impactor (Sierra Instruments, model 234). The flow rate of the sampler was calibrated by means of orifice plates provided by the manufacturer. At the operational flow rate (1.1 m3 min1), the sampling system was configured to have effective cut-off diameters (D50) 0.95, 1.5, 3.0, 7.5 mm for the four impaction stages, respectively. The upper cut size was taken to be 30 mm, in accordance with other studies (Hien et al., 2007; Karanasiou et al., 2007). No substantial drop (<5%) in flow rate was observed at the end of each sampling. Particles were acquired over a 24-h period starting at 9:00 a.m. The atmospheric conditions prevailing during the sampling periods are given in Table 1.

Table 1 Meteorological conditions during the sampling periods. Sampling date

Ambient temperature ( C)

RH (%)

Wind speed (m s1)

11/12/2007 13/12/2007 15/12/2007 17/12/2007 19/12/2007 21/12/2007 23/12/2007 3/6/2008 5/6/2008 18/6/2008 21/6/2008 23/6/2008 25/6/2008 27/6/2008

11.8 9.82 1.88 4.32 4.36 4.70 5.93 22.1 23.9 27.1 28.9 29.3 31.6 30.3

77.4 65.8 86.4 79.8 67.4 60.8 78.1 63.3 64.2 66.5 49.7 55.9 49.2 50.7

0.59 2.07 1.79 1.68 1.79 1.19 0.48 1.80 1.50 1.30 1.40 1.60 1.30 1.70

Quartz filters (Environmental Tisch TE-230QZ) were used as impaction substrates (slotted 5.7  5.7 cm) for the collection of the four coarser size fractions and rectangular backup filters (2500QATUP) for the finest size fraction. Before use, filters were baked out at 450  C for 8 h and stored in aluminium foil packages. Filters were weighed using a KERN 870 semi-microbalance with resolution 10 mg after 48-h equilibration at relative humidity 50  5% and temperature 20  0.5  C. Loaded filters were stored in the dark, in aluminium foils, at 20  C until extraction (within 15 days) and analysis. The uncertainty of weighing quartz filters, as estimated according to EN 14907:2005, was equal to 550 mg. Filter portions (1/4 of backup filters, and 1/2 of slotted filters) were ultrasonically extracted using dichloromethane/n-hexane (3:2, v/v). Extracts were concentrated by rotary evaporation to w2 ml and fractionated using glass column chromatography with anhydrous silica and alumina using sequentially different solvents according to Sienra et al. (2005). Fraction 1 eluted with n-hexane contained AHs, OCPs and PCBs, fraction 2 eluted with n-hexane/ dichloromethane (3:2) contained PAHs, while fractions 3 and 4 eluted with dichloromethane and acetone/n-hexane (3:7) contained nitro-PAHs. This paper presents data for the first two fractions only. AHs were determined using a Hewlett Packard-5890 gas chromatograph coupled with a 5791A mass spectrometric detector operated in the Electron Impact Mode. A HP-5MS capillary column (30 m length, 0.25 mm i.d., 0.25 mm film thickness) was used. Analysis was performed in the scan mode according to Kalaitzoglou et al. (2004). Identification was accomplished by using authentic standards including C14, C16–C20, C22, C24, C28, C32, pristane and phytane, and by comparison of the fragmentation pattern of unknown compounds with the NIST mass spectral reference library. The unresolved complex mixture (UCM) of branched and cyclic hydrocarbons was quantified by integrating the total ion current corresponding to the ‘‘hump’’ observed in the chromatograms as described by Kalaitzoglou et al. (2004). PAHs analysis was carried out in the same system in the SIM mode, as described by Chrysikou et al. (2009). Identification was accomplished by using the standard PAH mixture (Ultra Scientific PM-612) containing 16 PAHs (napthalene – Np, acenaphthylene – Acy, acenapthene – Ace, fluoranthene – Fl, phenanthrene – Ph, anthracene – An, fluorene – F, pyrene – Py, benzo[a]anthracene – B[a]An, chrysene – Chry, benzo[b]fluoranthene – B[b]Fl, benzo[k]fluoranthene – B[k]Fl, benzo[a]pyrene – B[a]Py, indeno[1,2,3-cd]pyrene – IPy, dibenz[a,h]anthracene – dB[a,h]An, benzo[ghi]perylene – B[ghi]Pe) plus benzo[e]pyrene – B[e]Py and perylene – Per, and by comparison of the fragmentation pattern of unknown compounds with the NIST mass spectral reference library. Internal standards, 5a-androstane

L.P. Chrysikou, C.A. Samara / Atmospheric Environment 43 (2009) 4557–4569

for AHs and a solution containing napthalene-d8, acenapthene-d10, phenanthrene-d10, chrysene-d12, perylene-d12 for PAHs, were added prior to instrumental analysis. Quantification of AHs and PAHs was performed by the linear regression method using fivepoint calibration curves established between the authentic standard/internal standard concentration ratios and corresponding peak area ratios. PCBs and OCPs were determined in a gas chromatograph (Hewlett Packard-5890) equipped with an electron capture detector using a DB-5 analytical column (length 50 m, 0.32 mm id, 0.17 mm film thickness) as described by Chrysikou et al. (2009). A standard mixture (Pesticide Mix 33, Dr. Ehrenstorfer), was used for calibration containing seven PCBs congeners (#28, #52, #101, #118, #138, #153 and #180) and 19 OCPs (hexachlorobutadiene, dichlobenil, hexachlorobenzene, quintozene, heptachlor, aldrin, isobenzan, isodrin, heptachlor-exo-epoxide, heptachlor-endo-epoxide, aendosulfan, p,p0 -DDE, dieldrin, endrin, p,p0 -DDD, p,p0 -DDT, a-HCH, b-HCH, g-HCH). 2.3. Data validation The recovery efficiency of the method was evaluated by the analysis of filters spiked with low and high concentrations of standard compounds. Most of the compounds provided high recoveries with mean values ranging between 72 and 97% for PAHs, 65 and 100% for AHs, 57 and 108% for OCPs and 82 and 110% for PCBs, respectively. No correction of concentrations for the recovery of target analytes was made. The recoveries of the extraction process were in the range of 71–97% for PAHs, 65–100% for AHs, 68– 110% for OCPs and 75–90% for PCBs, respectively. Field and laboratory blanks were routinely analyzed for quality control. Blank levels of individual analytes were normally very low and in most cases not detectable. Method detection limits (MDLs) were determined by measuring a minimum of six aliquots of a prepared solution with concentration five times the lowest standard solution used in the calibration curve. The MDLs for PAHs, AHs, OCPs, PCBs were 0.89–11.7 ng mL1, 6.14–15.5 ng mL1, 0.15– 7.85 pg mL1, and 0.36–10.6 pg mL1, respectively. Method accuracy and precision were evaluated by analysis of the NIST certified reference material 1649a-urban dust, containing several of the target analytes (13 PAHs, 7 PCBs, p,p0 -DDT, p,p0 -DDD, p,p0 -DDE, hexachlorobenzene and heptachlor). Percent recoveries of PAHs, OCPs, and PCBs from the CRM were found in the range of 78–146%, 79–86%, and 64–97%, respectively, and measured concentrations agreed sufficiently with certified values (within 20% for PAHs, 25% for OCPs and 15% for PCBs). The estimated precision was 3– 15% for PAHs, 0.9–5.9% for n-alkanes, 2.9–5.2% for OCPs and 3.7– 12% for PCBs, respectively. 3. Results and discussion Summary data concerning the total concentrations (sum of 5 size fractions) of particle mass (PM) and organic pollutants determined in the sampling periods are presented in Table 2. 3.1. Particle mass In both seasons, the highest mean PM concentration was obtained in the particle fraction <0.95 mm, accounting for 62% and 36% of the total PM in cold and warm period, respectively (Fig. 1). The mean normalized (dC/Cd log Dp vs log Dp) size distribution of PM is presented in Fig. 2. For both seasons, a bimodal distribution is evident with two peaks at the fine and the coarse size range, which for a kerbside site might be primarily attributed to traffic emissions and traffic-induced road dust resuspension, respectively

4559

(Duan et al., 2005; Wu et al., 2006). A contradiction between the distribution of percent concentrations (Fig. 1) and normalized concentrations (Fig. 2) is apparent and should be attributed to the low fractionation resolution of particles <0.95 mm. As seen in Fig. 2, the fine mode peak localized at fraction 0.95– 1.5 mm in winter shifted to <0.95 mm in summer. Particles <0.95 mm originate through gas-to-particle processes of young aerosols (Wu et al., 2006), and the shift to larger sizes in winter might be due to aerosol growth during longer accumulation times in winter under conditions of higher RH as compared to summer (Table 1). According to Vaeck and Van Cauwenberghe (1985), the distribution shift toward larger particle size in the accumulation mode may reflect long-term ageing that however is less likely for a traffic-impacted site. On the contrary, a shift of all particle fractions to larger sizes in summer was reported by Barrero Mazquiara´n and Canto´n Ortiz de Pinedo (2007) and attributed to a higher incidence of resuspension and abrasion processes. In general, bimodal mass size distribution with two distinct modes located at fine and coarse particles has been found in several locations (e.g. 0.08–0.61 mm and 4.9–10.0 mm, Barrero Mazquiara´n and Canto´n Ortiz de Pinedo, 2007; 0.43–2.1 mm and 9.0–10.0 mm, Duan et al., 2007; 0.4–0.7 mm and 4.7–5.7 mm, Hien et al., 2007; 0.95–1.5 mm and 3.0–7.5 mm; Chrysikou et al., 2009). The size distribution may vary as a result of the different sampling location, the emission sources and the weather conditions. Moreover, literature data are not easily comparable, due to the variant fractionation resolution used in different studies. The normalized concentrations of PM in the various size fractions exhibited significant seasonal difference (P < 0.05) only in fractions 0.95–1.5 mm and 1.5–3.0 mm with higher values in winter suggesting the presence of sources emitting larger particles and/or atmospheric conditions favouring particle growth in this season. The lack of seasonal differences in fraction <0.95 mm might be due to the close vicinity of the sampling site to fresh traffic emissions. Seasonal differences in coarse particles have been attributed to increased washout coefficients for PM sizes in 3–10 mm (Chate and Pranesha, 2004; Lin et al., 2008). However, in the present study, no rainfalls occurred in any of the sampling campaigns resulting in insufficient wash-out for PM. The mass median diameter (MMD) for total PM and associated organic compounds (Fig. 3) was calculated from cumulative distribution percentages for each sampling day, so that a linear relationship (R2 > 0.95) was exhibited between the cumulative mass fractions and the logarithm of the cut-off diameters. It is noteworthy that MMDs are only rough estimates particularly for species whose mass was largely on <0.95 mm size range. The MMDs for total PM in the size range <0.95  7.5 mm averaged 1.26 mm and 1.18 mm in winter and in summer, respectively, without significant seasonal differences. These values are in fair consistency with the average MMD value found in another urban site of Thessaloniki during cold season for the same size distribution resolution (Chrysikou et al., 2009) verifying that particle pollution in the centre of the city is severe, strongly related to potential health risks of inhaled particles. Lin et al. (2008), found MMD for PM in the 0.01–10 mm size range in a heavily trafficked roadside in southern Taiwan to be around 1.2 mm. Relatively lower MMD (0.98 mm) has been reported for PM in the <0.49–10 mm size range in the urban atmosphere of Guangzhou, China (Bi et al., 2005). Literature MMDs for particles emitted from gasoline-powered cars, motorcycles, and resuspendable road dust are 0.572, 0.385 and 62.5 mm, respectively (Yang et al., 1999). In comparison to fresh combustion emissions, it is expected that ambient air particles are coarser as a result of coagulation and surface growth from secondary aerosol gas-toparticle conversion (Bi et al., 2005).

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Table 2 Totala concentrations of PM and organic pollutants during winter and summer. Winter (N ¼ 7)

Summer (N ¼ 7)

Mean

Median

Min

Max

Mean

Median

Min

Max

272

229

153

589

172

189

45.0

268

Np Acy Ace F Ph An Fl Py B[a]An Chry B[b]Fl B[k]Fl B[a]Py B[e]Py Per IPy dB[a,h]An B[ghi]Pe

3

ng m ng m3 ng m3 ng m3 ng m3 ng m3 ng m3 ng m3 ng m3 ng m3 ng m3 ng m3 ng m3 ng m3 ng m3 ng m3 ng m3 ng m3

Nd Nd Nd Nd 0.75 1.25 2.65 2.76 2.47 3.73 1.54 0.83 1.06 3.64 0.47 1.50 Nd 1.81

Nd Nd Nd Nd 1.01 1.10 3.06 3.21 2.18 3.68 1.23 0.75 0.88 2.48 0.46 0.95 Nd 1.15

Nd Nd Nd Nd 0.03 0.18 0.58 0.49 0.52 0.91 0.50 0.60 0.21 0.43 0.05 0.20 Nd 0.20

Nd Nd Nd Nd 1.25 2.32 4.82 5.13 4.60 6.40 3.27 1.55 2.68 12.2 1.55 4.60 Nd 5.98

0.73 Nd Nd 0.21 0.20 0.35 0.55 0.30 0.39 0.92 0.65 0.60 0.12 0.93 0.56 0.57 0.43 0.53

0.71 Nd Nd 0.22 0.18 0.33 0.52 0.28 0.36 0.80 0.71 0.63 0.12 0.89 0.67 0.63 0.46 0.50

0.60 Nd Nd 0.20 0.12 0.24 0.50 0.22 0.30 0.70 0.40 0.50 0.05 0.61 0.19 0.34 0.29 0.29

0.97 Nd Nd 0.22 0.30 0.54 0.64 0.39 0.56 1.78 0.83 0.87 0.21 1.49 0.78 0.70 0.56 0.95

C14 C16 C17 C18 C19 C20 C22 C24 C28 C32 Pristane Phytane

ng m3 ng m3 ng m3 ng m3 ng m3 ng m3 ng m3 ng m3 ng m3 ng m3 ng m3 ng m3

3.63 3.64 3.93 6.87 10.7 17.7 23.3 22.9 14.9 15.3 1.45 4.42

3.75 4.05 4.28 7.49 11.8 22.1 21.1 25.8 13.6 15.0 1.66 4.76

2.45 2.08 1.82 4.34 5.71 7.29 9.01 9.43 0.93 10.4 0.61 2.86

5.18 5.26 5.26 8.80 16.7 26.9 41.4 42.9 29.8 23.5 1.99 2.83

4.22 2.72 Nd 1.35 2.33 17.8 3.35 6.69 8.67 12.3 Nd 1.40

4.19 3.26 Nd 1.43 0.05 20.1 2.83 6.81 8.75 12.8 Nd 1.17

3.58 0.10 Nd 0.58 0.05 7.28 2.22 3.63 0.75 9.80 Nd 0.29

4.76 4.19 Nd 1.98 15.5 24.4 5.27 8.77 13.4 15.9 Nd 3.17

Hexachlorobutadiene Dichlobenil Hexachlorobenzene a-HCH b-HCH g-HCH Quintozene Heptachlor Aldrin Isobenzan Isodrin Heptachlor-epoxide a-Endosulfan Dieldrin Endrin p-p’-DDE p-p’-DDD p-p’-DDT

pg m3 pg m3 pg m3 pg m3 pg m3 pg m3 pg m3 pg m3 pg m3 pg m3 pg m3 pg m3 pg m3 pg m3 pg m3 pg m3 pg m3 pg m3

148 Nd Nd Nd 2.66 6.69 12.2 8.07 5.90 2.18 8.48 23.9 3.49 Nd 5.00 2.82 5.97 1.91

150 Nd Nd Nd 0.92 6.83 13.1 5.48 5.76 1.46 6.38 18.5 2.37 Nd 3.28 Nd 3.66 Nd

99 Nd Nd Nd Nd 4.09 Nd 3.16 2.96 Nd Nd Nd Nd Nd 1.29 Nd 1.90 Nd

207 Nd Nd Nd 8.76 8.68 35.5 21.2 9.48 8.29 29.6 74.7 12.5 Nd 8.92 7.52 13.9 6.78

Nd Nd Nd 8.13 12.9 9.69 5.90 23.9 13.0 22.1 Nd 26.9 28.5 11.3 Nd Nd 5.45 Nd

Nd Nd Nd 7.85 7.30 6.31 4.76 18.4 13.5 24.3 Nd 25.3 27.9 3.75 Nd Nd 6.01 Nd

Nd Nd Nd 5.40 Nd 3.85 Nd 7.97 7.01 2.95 Nd 10.8 0.53 Nd Nd Nd 0.87 Nd

Nd Nd Nd 12.0 38.9 19.9 14.9 44.2 18.9 53.5 Nd 48.1 64.2 41.4 Nd Nd 10.9 Nd

PCB-28 PCB-101 PCB-118 PCB-153 PCB-138 PCB-180

pg m3 pg m3 pg m3 pg m3 pg m3 pg m3

4.00 Nd 3.05 1.52 Nd Nd

1.77 Nd Nd Nd Nd Nd

Nd Nd Nd Nd Nd Nd

14.0 Nd 12.9 10.6 Nd Nd

10.6 1.77 5.69 5.85 4.23 67.8

5.37 2.7 6.48 3.69 Nd 86.4

Nd Nd Nd Nd Nd Nd

25.7 3.39 9.28 16.0 16.0 165

mg m3

PM b

a b

Sum of size fractions <0.95 mm, 0.95–1.5 mm, 1.5–3.0 mm, 3.0–7.5 mm, >7.5 mm. PAHs abbreviations are given in text Nd: not detected.

3.2. PAHs Most targeted PAHs were quantifiable in all size fractions, excepting Ace, Acy in both seasons and Np, F, dB[a,h]An in wintertime samples (Table 2). The mean total PAH concentrations (SPAH) associated with total PM in winter and in summer were 24.6 ng m3 and 8.54 ng m3, respectively, in accordance with the values previously found in other sites within Thessaloniki (30.8 ng m3 in PM10, Manoli et al., 2004; 22.6 ng m3 in TSP, Chrysikou et al., 2009), in

Athens (32.9 ng m3 in PM10, Mantis et al., 2005), and other urban and rural Greek locations (0.78–24.8 ng m3 in TSP, Kalaitzoglou et al., 2004). The levels are similar to those found in Veszpre´m, Hungary (26 ng m3 sum of 14 PAHs in TSP, Kiss et al., 1998), and in Prague, Czech Republic (55.1 ng m3 in PM10, Saarnio et al., 2008). The lower PAH concentrations in summer reflect changes in prevailing meteorology (i.e. greater dispersion, photolytic and thermal decomposition) and the absence of seasonal emissions (e.g. residential heating facilities, cold start of vehicles) (Mantis et al., 2005).

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80

different volatility of PAHs, since low molecular weight species are absorbed to fine aerosol and distribute in coarse particles by rapid volatilization and condensation. On the contrary, for the higher molecular weight PAHs, due to the lower vapour pressures, the time required for this repartitioning process is much greater (Bi et al., 2005), therefore, they tend to remain in fine particles initially emitted (Duan et al., 2007). Moreover, the distribution differences of PAHs with respect to particle size could result from chemical affinities between PAHs and different size particles, different emission sources and different PAH reactivity on photooxidation (Allen et al., 1996; Wu et al., 2006). In contrast to PM, the size distribution pattern of PAHs was similar in the two seasons (Fig. 5), in agreement with Aceves and Grimalt (1993). Significant seasonal variations were found for the normalized concentrations of PAHs according to the Mann–Whitney U-test (P < 0.05) with Np, Fl and Per exhibiting higher intensities in all size fractions in summer, whereas, Py, B[a]An, Chry, and B[a]Py in the finer size fractions (<0.95 mm and 0.95–1.5 mm) in winter. Increased distribution of PAHs in coarse particles (2.5–10 mm) was observed during the warm months at a receptor site, near Los Angeles, assuming repartition as a result of the longer ageing process in these seasons, and/or mixing of local emissions with long-range transported aerosol (Miguel et al., 2004). Hien et al. (2007), did not found distinct differences in the size distribution of 5- and 6-ring PAHs between the rainy (summer) and the dry (winter) season, however, in the rainy season, the intensities of dominant peaks in fine particle size (0.4–0.7 mm) were significantly higher than those in the dry season with an exception of B[ghi]Pe. As for 4-ring PAHs, dry season samples showed predominant distribution in particles 4.4–5.8 mm. These seasonal variations were attributed to the different meteorological conditions in the two seasons, mainly wind direction/speed and sunshine hours. A shift to larger particles was observed from summer to winter, due to higher dispersion of aerosols in summer (Kiss et al., 1998; Duan et al., 2007). The MMDs of individual PAHs (Fig. 3) ranged between 0.46 and 2.34 mm in winter and 0.94 and 2.43 mm in summer, respectively, with statistically significant difference found only for B[a]An and B[a]Py that presented higher MMDs in summer (P < 0.05). The MMDs of SPAH was 1.72 mm in winter and 1.56 mm in summer, higher than the values reported for the engine exhausts of gasolinepowered cars (0.45 mm) and of motorcycles (0.35 mm), respectively (Yang et al., 1999). The values are also higher than those found at an urban site of Guangzhou, China (0.33–0.84 mm for total PAHs in the <0.49–10 mm size range, Bi et al. (2005) or at a heavily trafficked roadside in southern Taiwan (w0.3 mm for total PAHs in the 0.01– 10 mm size range, Lin et al., 2008). In general, the traffic volume in Thessaloniki constitutes of 73% gasoline-fuelled passengers cars, 14% diesel-fuelled tracks, buses and taxis and 13% motorcycles, with diesel-traffic accounting for about 40% of total transports. Some PAHs (Ph, An) exhibited MMDs larger than total PM in both seasons. On the contrary, the MMDs of Py, B[a]An, Chry, B[b]Fl and B[a]Py were lower than total PM in winter but higher in summer. Lin et al. (2008), found the 50% cumulative fraction of total PAHs in the 0.01–10 mm size range to be lower than the corresponding fraction of PM.

40

3.3. AHs

100 winter summer

80

%

60 40 20 0 <0.95

0.95-1.5

1.5-3.0

3.0-7.5

>7.5

Dp (µm) Fig. 1. Average mass percentage of PM (Error bars show standard deviation).

As illustrated in Fig. 4, in winter all PAHs exhibited highest proportions (from around 30% up to 80%) on particles <0.95 mm, with decreasing association at increasing particle size fractions. In summer, the percent distribution on the fraction <0.95 mm was lower presenting a shift to larger particles. It has been reported that the shorter the distance to the source, the higher the portion of PAHs associated with fine particles (Schnelle-Kreis et al., 2001). So, it can be supposed that the high proportion of PAHs found on particles <0.95 mm in the present study originates from fresh traffic emissions, near the sampling site. A different view to the seasonal size distribution of PAHs was obtained by evaluation of their normalized size distributions (Fig. 5). In both sampling periods, all PAHs presented unimodal size distribution with one predominant peak at 0.95–1.5 mm. The peak at this size range is unexpected since it is known that PAHs are associated primarily with submicron particles as a result of gas-toparticle transformation after emission (Bi et al., 2005). Further fractionation below 0.95 mm would perhaps reveal occurrence of PAHs at smaller size ranges, in better agreement with literature data. For instance, Kiss et al. (1998) also reported one peak at 0.5– 1.0 mm for several PAHs. Bi et al. (2005), found one peak at 0.49– 0.95 mm for the >4-ring PAHs and only for Ph and retene a bimodal pattern with a second peak at 3.0–7.2 mm. A single mode for B[ghi]Pe peaked between 0.18 and 0.32 mm was found by Kleeman et al. (2008). Finally, Schnelle-Kreis et al. (2001) observed one major peak in 0.04–0.14 mm range close to a crossroad that shifted to larger particles when winds flowed from a built up area and from a green park (0.14–0.49 mm and 0.49–1.72 mm, respectively). In other studies, bimodal distribution has been found in fine and coarse particles (Duan et al., 2007; Hien et al., 2007) indicating the

200 winter summer 160

dC/dlogDp

4561

120

0 0.01

0.1

1

10

100

Dp (µm) Fig. 2. Mean size distribution of particles collected during winter and summer.

Targeted AHs were determined in all size fractions in both seasons with the exception of C17 and pristane that were detected only in winter-time samples. The UCM of branched and cyclic hydrocarbons was identified only in the size range <0.95 mm in a few winter samples. UCM ranged between 91.6 ng m3 and

4562

L.P. Chrysikou, C.A. Samara / Atmospheric Environment 43 (2009) 4557–4569

MMD (µm)

10.00

winter summer

1.00

F Ph An Fl B[ Py a] A Ch n B[ ry b B[ ]Fl k B[ ]Fl a] B[ Py e] Py Pe dB r [a IP B[ ,h]Ay gh n i]P ΣP e AH C1 C14 6 C1 7 C1 C18 9 C2 C20 2 C2 4 C2 Pr C 3 8 ist 2 Ph an yt e an e Σ α- NA H Q β CH ui - H nt C oz H e H e γ - ne pt H C ac H h I A lo α- sob ldr r en e in do nza su n En lfan p- dr p'- in DD D

Np

PM

0.10

Fig. 3. Mean MMDs for PM, PAHs, n-alkanes, OCPs for the sampling periods (Error bars show standard deviation).

302 ng m3 with maximum centered at C24, suggesting contamination from vehicular emissions (Azevedo et al., 1999). The mean concentration of total n-alkanes (SNA) in total PM (Table 2) were 129 ng m3 in winter and 61 ng m3 in summer. These concentrations are relatively higher than those measured in

<0.95μm

0.95-1.5μm

smaller cities of Greece (59–80 ng m3 in TSP, Kalaitzoglou et al., 2004), in Errenteria, Spain (62 ng m3 in TSP, Barrero Mazquiara´n and Canto´n Ortiz de Pinedo, 2007), but lower than those reported for Athens (446 ng m3 in TSP, Karanasiou et al., 2007) or Guangzhou, China (233–1037 ng m3 in TSP, Bi et al., 2005).

1.5-3.0μm

3.0-7.5μm

>7.5μm

100 WINTER 80

%

60

40

20

<0.95μm

0.95-1.5μm

3.0-7.5μm

I dB Py [a ,h ]A n B[ gh i]P e

Pe r

y

e] Py B[

Fl

a] P

k]

1.5-3.0μm

B[

B[

C hr y B[ b] Fl

An a]

Py

B[

Fl

An

Ph

F

e

y

Ac

Ac

N p

0

>7.5μm

80 SUMMER

40

20

Fig. 4. Average mass percentage of PAHs in winter and in summer (Error bars show standard deviation).

IP y

Pe r

Py e] B[

Py a] B[

Fl k[ B[

B[ b] Fl

y C hr

n a] A B[

Py

Fl

An

Ph

F

0 N p

%

60

L.P. Chrysikou, C.A. Samara / Atmospheric Environment 43 (2009) 4557–4569

3-ring PAHs (winter)

3-ring PAHs (summer)

0.8

Ph

0.6

An

0.4 0.2 0.0 0.01

0.1

1

10

Ph

0.4

dC/dlogDp

dC/dlogDp

1.0

F 0.2 0.1 0 0.01

100

An

0.3

0.1

1

Dp (µm)

2.0 1.0 10

dC/dlogDp

0.8 0.4 0 0.01

100

0.1

1

5-ring PAHs (winter)

5-ring PAHs (summer)

B[k]Fl 2.0

B[a]Py B[e]Py

1.0

1

10

2

B[b]Fl B[a]Py

1.2

B[e]Py

0.8 0.4 0 0.01

100

0.1

1

10

100

Dp (µm)

6-ring PAHs (winter)

6- ring PAHs (summer) IPy

1.2

B[ghi]Pe

0.8 0.4

dC/dlogDp

dC/dlogDp

100

1.6

Dp (µm)

1.6

10

Dp (µm)

B[b]Fl

0.1

1.2

Dp (µm)

3.0

0.0 0.01

dC/dlogDp

3.0

Fl Py Chry B[a]An

1.6

dC/dlogDp

dC/dlogDp

4.0

1

100

4-ring PAHs (summer) Fl Py Chry B[a]An

0.1

10

Dp(µm)

4-ring PAHs (winter)

0.0 0.01

4563

0.6

IPy

0.5

dB[a,h]An

0.4

B[ghi]Pe

0.3 0.2 0.1

0

0.01

0.1

1

10

0 0.01

100

0.1

1

Dp (µm)

ΣPAHs (winter)

6.0 4.0

0.6 0.4 0.2

2.0 0.0 0.01

100

ΣPAHs (summer)

0.8

dC/dlogDp

dC/DlogDp

8.0

10

Dp (µm)

0.0 0.1

1

10

100

0.01

0.1

Dp (µm) Fig. 5. Mean size distribution of PAHs during winter and summer.

1

Dp (µm)

10

100

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L.P. Chrysikou, C.A. Samara / Atmospheric Environment 43 (2009) 4557–4569

<0.95μm

0.95-1.5μm

1.5-3.0μm

3.0-7.5μm

>7.5μm

80 WINTER

60

%

40

20

<0.95μm

0.95-1.5μm

1.5-3.0μm

3.0-7.5μm

C 32

C 28

C 24

C 22

C 20

C 19

e Ph yt an

C 18

e Pr is ta n

C 17

C 16

C 14

0

>7.5μm

80 WINTER

%

60

40

20

C 32

C 28

C 24

C 22

C 20

C 19

e an yt Ph

C 18

C 16

C 14

0

Fig. 6. Average mass percentage of AHs in winter and in summer (Error bars show standard deviation).

In winter, the highest proportions of all AHs (40–55%) was found on particles <0.95 mm, demonstrating equal distribution in larger fractions (Fig 6). Similar to PAHs, in summer, there is a shift of AHs to larger particles, particularly for C19 and C20. The normalized size distributions of individual AHs are shown in Fig. 7. Similar to PAHs, most AHs demonstrated unimodal distribution in particles 0.95–1.5 mm, possibly due to gas-to-particle transformation after their emission, without difference between odd and even homologues. Karanasiou et al. (2007), reported predominant peaks in the particle size fraction 1.2–1.8 mm for C18, C24, C28 and C32 in the warm season. Bi et al. (2005), found unimodal distribution for C22 and C25 centered at 0.49–0.95 mm, and bimodal distribution for C17 (0.95–1.5 mm and 7.2–10 mm), C29, C31, C33 (0.49–0.95 mm and 7.2–10.0 mm) in winter. The distribution pattern of AHs did not exhibit distinct seasonal variations in agreement with Aceves and Grimalt (1993), only the C22 maximum in summer shifted to larger particles (1.5–3.0 mm) and a second peak was identified for C19 on particles 3.0–7.5 mm, attributable possibly to volatilization/condensation.The normalized concentrations of several AHs (C18, C19, C20, C22, C24, phytane) in most size fractions were significantly higher in the cold season (P < 0.05), suggesting major fossil fuel contamination in winter. The calculated MMDs for individual AHs (Fig. 3) ranged from 0.69 mm to 1.64 mm in winter in agreement with the values reported

for the <0.49–10 mm size range (0.31–1.24 mm, Bi et al., 2005). In summer, MMDs were relatively higher (0.75 mm to 2.7 mm) particularly for C18, C19, C20 appearing to concentrate in larger particles. Only C22, C24 and C28 exhibited lower MMDs than total PM reflecting condensation mechanisms (Van Vaeck and Van Cauwenberghe, 1985; Aceves and Grimalt, 1993). 3.4. OCPs and PCBs Several targeted OCPs and PCBs were not detected in any of the particle fractions; moreover, some of them were detected in certain fractions only. As shown in Table 2, concentrations of OCPs and PCBs in the sum of particle fractions were higher in summer in consistency with other studies (Yeo et al., 2003; Scheyer et al., 2005). In summer, due to the higher ambient temperatures, the presence of these compounds is attributed to volatilization from soil, water, and vegetation (Simcik et al., 1999; Scheyer et al., 2005). Current concentrations were considerably lower than those found at another urban site of Thessaloniki during winter 2006 (Chrysikou et al., 2009). The highest proportions of OCPs were observed on particles <0.95 mm in both seasons, however increased proportions on larger particle sizes were obvious for some compounds in summer (Fig. 8). Similar seasonal trend was also observed for PCBs (Fig. 9).

L.P. Chrysikou, C.A. Samara / Atmospheric Environment 43 (2009) 4557–4569

dC/dlogDp

winter summer

4 2 0

0.01

1 Dp (µm)

10

100

1

dC/dlogDp

dC/dlogDp

winter

2

1 Dp (µm)

10

dC/dlogDp

C19

100

8

summer

6 4 2

1 Dp (µm)

10

winter summer

2

0.1

1 Dp (µm)

10

C20

100

winter

16

summer

12 8 4

0.01

100

0.1

1 Dp (µm)

10

100

C24

C22

winter

20

summer

15 10 5

winter

25

dC/dlogDp

dC/dlogDp

100

0

0.1

25

20

summer

15 10 5 0

0

1 Dp (µm)

10

0.01

100

0.1

summer

12 8 4

1 Dp (µm)

10

C32

winter

16

dC/dlogDp

0.1

C28 dC/dlogDp

C18

winter

0

100

winter

16

summer

12 8 4 0

0

0.1

1

10

0.01

100

0.1

1

10

100

Dp (µm)

Dp (µm)

Phytane

Pristane winter

1.5 1 0.5

winter

4

dC/dlogDp

2

dC/dlogDp

10

4

0.01

dC/dlogDp

0.1

10

summer

3 2 1 0

0

0.01

1 Dp (µm)

0

0

0.01

0.1

6

3

0.01

2

0.01

C17

0.01

winter summer

0

0.1

4

0.01

C16

4

dC/dlogDp

C14

6

4565

0.1

1

10

100

0.01

0.1

Dp (µm) Fig. 7. Mean size distribution of AHs during winter and summer.

1 Dp (µm)

10

100

4566

L.P. Chrysikou, C.A. Samara / Atmospheric Environment 43 (2009) 4557–4569

<0.95μm

0.95-1.5μm

1.5-3.0μm

3-7.5μm

>7.5μm

100 WINTER 80

%

60 40 20

pp' -D D T

p' -D D D

rin

p-

p-

p' -D

En d

D

E

fa n

e αEn d

po

os ul

xi d

rin -e or H

ep

ta ch l

H

Is od

za n

rin Is ob

en

Al d

or

H C

ta ch l ep

γH

H C H nt oz en e ui

β-

Q

H e bu xac ta hlo di r ne o

0

<0.95μm

0.95-1.5μm

1.5-3.0μm

3-7.5μm

>7.5μm

100 SUMMER 80

%

60 40 20

D -D p-

p'

dr En

D

in

n lfa su do En

α-

H

ep

ta

ch

lo

Is

r-e

ob

po

en

xi

za

de

n

in dr Al

r ta ep H

γ-

H

ch

C

lo

H

e en oz Q

ui

nt

β-

α-

H

H

C

C

H

H

0

Fig. 8. Average mass percentage of OCPs in winter and in summer (Error bars show standard deviation).

The mean normalized size distributions are illustrated in Fig. 10 only for compounds detectable in all size fractions. As seen, most compounds demonstrated one peak in the size fraction 0.95– 1.5 mm, in accordance with previous findings (Chrysikou et al., 2009), while in summer, the maximum of aldrin and p-p0 -DDD shifted on particles <0.95 mm. Bimodal distribution was observed for aldrin, endrin and g-HCH in winter and for a-endosulfan and pp0 -DDD in summer, with a second peak in fraction 3.0–7.5 mm. The second peak may imply adsorption from the gas phase to larger particles (Allen et al., 1996). Normalized concentrations of OCPs in individual size fractions did not present significant seasonal differences with the exception

<0.95 m

0.95-1.5 m

1.5-3.0 m

3.0-7.5 m

of heptachlor in particles 0.95–1.5 mm and aldrin in particles 1.5– 3.0 mm that were higher in summer (P < 0.05). The MMDs of individual OCPs (Fig. 3) were in general larger than total PM indicating that, after entering the ambient air, these species were cooled and condensed onto larger particles by adsorption (Chen et al., 1996). 3.5. Correlations with meteorological conditions Table 3 shows the Pearson’s correlation coefficients for normalized concentrations of measured organics with meteorological parameters (T, RH, WS).

>7.5 m

<0.95 m

100

WINTER

1.5-3.0 m

3.0-7.5 m

>7.5 m

SUMMER

80

80

60

60

%

%

0.95-1.5 m

100

40

40

20

20 0

0 PCB-28

PCB-118

PCB-153

PCB-28 PCB-101 PCB-118 PCB-153 PCB-138 PCB-180

Fig. 9. Average mass percentage of PCBs in winter and in summer (Error bars show standard deviation).

L.P. Chrysikou, C.A. Samara / Atmospheric Environment 43 (2009) 4557–4569

winter

160

hexachloro butadiene

20

120

Quintozene

15

dC/dlogDp

dC/dlogDp

winter

80 40 0 0.01

0.1

1

10

Heptachlor

10 5 0 0.01

100

γ-HCH

0.1

Dp (µm)

1

Endrin

6 4 2

10

β-HCH γ-HCH

20

dC/dlogDp

dC/dlogDp

8

1

100

summer Aldrin

0.1

10

Dp (µm)

winter

0 0.01

4567

15 10 5 0 0.01

100

α-HCH

0.1

Dp (µm)

1

10

100

Dp (µm) summer

summer

Heptachlor Aldrin

40

Quintozene

20 15 10 5

dC/dlogDp

dC/dlogDp

25

0 0.01

0.1

1

10

20 10 0 0.01

100

Heptachlorepoxide

30

0.1

Dp (µm)

summer

6 4 2 1

100

10

100

Dp (µm)

Isobenzan α -Endosulfan

40

dC/dlogDp

dC/dlogDp

8

0.1

10

summer p-p'-DDD

10

0 0.01

1

Dp (µm)

30 20 10 0 0.01

0.1

1

10

100

Dp (µm)

Fig. 10. Mean size distribution of OCPs during winter and summer.

For PM, statistically significant negative correlation with T was found on particles 0.95–1.5 mm implying increased emission and/or formation rates of particles in this size range in the cold season. Significant negative correlation was also found on particles >1.5 mm with WS suggesting increased dispersion of coarse particles with wind. Significant negative correlations with T were found for Ph, Fl and B[a]An in particle size fractions 0.95–1.5 mm suggesting increased emission and/or formation rates of particle-bound PAHs at this size range in the cold season. On the contrary, positive correlations were found for Np in size fractions 1.5–3.0 mm and >7.5 mm, revealing redistribution on larger particles with increased temperature. Significant negative correlations with T were also found for C18, C19,

C24, C28 and phytane, probably for the same reasons as those stated for PAHs. Correlations of PAHs with RH were in general poor. According to Wu et al. (2006), an effect of RH on the repartitioning of higher PAHs is expected for RH values >99% not occurring during the present study (Table 1). Significant negative correlations with RH were found for some OCPs, particularly for aldrin and endrin in size fractions 1.5–3.0 mm and 3.0–7.5 mm, respectively, consistent with the second peak exhibited by these compounds in winter. Finally, Fl and Py in the 3.0–7.5 mm and C24 in the 0.95–1.5 mm particle size fractions displayed decreased distribution with WS indicating local emissions from traffic. On the contrary, B[a]An in

4568

L.P. Chrysikou, C.A. Samara / Atmospheric Environment 43 (2009) 4557–4569

Table 3 Pearson correlation coefficients between normalized concentrations of target compounds and meteorological factors Size fraction (mm)

T ( C)

RH (%)

WS (m/s)

PM PM

0.95–1.5 3.07.5

0.67**b 0.23

0.52 0.22

0.45 0.71**

Npa Np Ph Fl Fl Py B[a]An B[a]An B[k]Fl C18 C18 C18 C18 C18 C19 C19 C24 C24 C24 C28 Phytane Phytane Phytane Phytane Aldrin Endrin

1.5–3.0 >7.5 0.95–1.5 0.95–1.5 3.07.5 3.07.5 0.95–1.5 1.5–3.0 >7.5 <0.95 0.95–1.5 1.5–3.0 3.07.5 >7.5 <0.95 0.95–1.5 0.95–1.5 1.5–3.0 >7.5 1.5–3.0 <0.95 0.95–1.5 3.07.5 >7.5 1.5–3.0 3.07.5

0.71** 0.73** 0.69** 0.71** 0.35 0.21 0.73** 0.52 0.78** 0.86** 0.77** 0.79** 0.84** 0.73** 0.73** 0.76** 0.49 0.66** 0.66** 0.69** 0.67** 0.66** 0.91** 0.70** 0.56 0.01

0.45 0.54* 0.42 0.37 0.15 0.10 0.50 0.44 0.53 0.64* 0.51 0.52 0.56* 0.41 0.33 0.53 0.39 0.46 0.46 0.46 0.42 0.50 0.69** 0.49 0.75** 0.95**

0.13 0.01 0.12 0.23 0.73** 0.75** 0.47 0.67** 0.17 0.29 0.38 0.18 0.39 0.04 0.03 0.49 0.69** 0.45 0.45 0.34 0.32 0.06 0.12 0.05 0.20 0.47

*Significant at the 0.05 level (2-tailed). **Significant at the 0.01 level (2-tailed). a PAHs abbreviations are given in text.

fraction 1.5–3.0 mm was positively correlated with WS reflecting potential transport from neighboring sources. 4. Conclusion The size distribution of PM and associated organic compounds (PAHs, AHs, OCPs and PCBs) was investigated at an urban site during the cold and warm seasons. A bimodal size distribution was found for PM with peaks at the fine size range (0.95–1.5 mm in winter shifted to <0.95 mm in summer) and the coarse size range (3.0–7.5 mm), attributable to traffic emissions and traffic-induced road dust resuspension, respectively. The MMD for total PM was around 1.2 mm, without significant seasonal difference, verifying that particulate pollution in this area may pose adverse health risks. Ambient temperature and wind speed were found to significantly affect the size distribution of particles. Most organic compounds exhibited highest proportions in particle size fraction <0.95 mm in both seasons, with some species presenting a shift to larger particles in summer. Their sizenormalized distribution was nearly unimodal but peaked at 0.95– 1.5 mm size range due to the low fractionation resolution of the particle fraction <0.95 mm. A second minor mode located at larger particles (3.0–7.5 mm) was observed for some compounds preferably in the cold (aldrin, endrin, g-HCH) or the warm season (C19, aendosulfan, p-p0 -DDD) suggesting redistribution due to volatilization and condensation. Acknowledgements The authors would like to thank Dr. M.J. Petrakakis and Dr. A.G. Kelesis from the Environmental Department, of Municipality of Thessaloniki, for providing the meteorological data.

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