Fractionation of elements by particle size and chemical bonding from aerosols followed by ETAAS determination

Abstract

A simple sequential leaching procedure was applied to the determination of the distribution of elements in atmospheric aerosols collected on membrane filters in different particle size ranges. The three-stage sequential leaching procedure divides the elements into environmentally mobile, bound to carbonates and oxides, and bound to silicates and organic matter (environmentally immobile) fractions. Three particle size fractions were collected: fine, d < 1 µm; coarse, d = 1–10 µm; and pre-filter, >10 µm on membrane filters. The sampling site was set up in K-puszta, a Hungarian meteorological background station. The campaign involved the collection of samples every two days weekly for 13 consecutive months. Concentrations of ten elements were determined by electrothermal atomic absorption spectrometry (ETAAS). Dry deposition rates and enrichment factors (EF) were calculated for fine and coarse fractions, as well as the mobile portion of the chemical bonding of the elements. This approach opens a new dimension to understanding the dry deposition of aerosols containing toxic heavy metals. Results were compared with earlier findings from a sampling campaign at K-puszta and other studies at different sampling sites and conditions. Generally, it can be concluded that the pollution load characterizes well the background station and that aerosols are transported from long-range and originate from both anthropogenic and natural sources.

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Introduction

Atmospheric aerosols have important roles in the biogeochemistry and transportation of trace elements in the air. The atmosphere is an important vector of global metal transport between regions, from land to sea, and vice versa. Direct atmospheric deposition makes only a minor contribution to the total metal contents of the lithosphere because of the large reservoir of these metals in soils and rocks. It is well known that environmental effects of aerosol particles depend on their particle size and chemical composition.¹ The aerosol particles influence the solar radiation transfer, cloud–aerosol interactions, and control the optical, electrical and radioactive properties of the atmosphere. Hence, the measurement of these two parameters is of crucial importance.¹

Speciation of trace metals in aerosols is different from speciation in aqueous media. The deposition, transport and inhalation processes are controlled predominantly by the size of the atmospheric aerosols.^2,3 Thus, the primary type of speciation is the fractionation by size. Chemical speciation both in terms of the dissolved/particulate distribution of the metals in precipitation and the inorganic or organic complexes plays an important role in controlling the environmental impact of deposited metals. Chemical speciation of trace elements of aerosols needs clear definition and agreement among scientists on the meaning of operations and experiments. Fractionation is the process of classification of an analyte or a group of analytes in a certain matrix according to physical (size, solubility, morphology) or chemical (bonding, reactivity) properties.⁴ The information on the fractionation indicates the mobility of the elements once the aerosol is mixed directly into natural waters or during scavenging of the aerosol by wet deposition. The distribution of elements can be determined among environmentally mobile, bound to carbonates and oxides (Fe-, Mn-oxides), and bound to organic matter and silicates (environmentally immobile) fractions.⁵ Aerosols sampled within the urban environment exhibit a greater solubility than aerosols with a crustal origin and this is an important finding when the interpretation of results of the sequential leaching is to be carried out. Particular attention has to be paid to distinguishing between environmentally mobile and environmentally immobile fractions because these represent the two extreme modes by which the metals are bound to the solid matrices.⁵ Speciation of metals in the atmosphere was summarized by Spokes and Jickells.⁶ Increased awareness of the importance of the use of ultra-clean techniques at all stages of sample collection and analysis has now enabled studies on atmospheric metal speciation. Samples collected on filters are, in most cases, in extremely low amounts, so sensitive and selective analytical techniques have to be used. ETAAS lends itself as the most suitable method to perform reliable analytical determinations and, therefore, it is widely applied to speciation studies.⁶

The aim of this research was to apply a simple sequential leaching procedure to the determination of the distribution of elements in atmospheric aerosols collected on membrane filters in fine and coarse particle size fractions. A three-stage sequential leaching procedure was applied to establish the distribution of metals among (1) environmentally mobile, (2) bound to carbonates and oxides, and (3) bound to silicates and organic matter (environmentally immobile) fractions. Three particle size fractions were collected (fine: d < 1 µm, coarse: d = 1–10 µm, pre-filter: d > 10 µm), but the third fraction was not analyzed. The classification of metals in the aerosol matrix according to particle size and chemical bonding was carried out as it has been defined as fractionation. For comparison with earlier findings, total elemental composition of aerosols collected at Budapest, Veszprém and Kabhegy was used.

Experimental

Sampling

K-puszta is located 80 km from Budapest south/south-east in central Hungary (h = 130 m asl, λ = 19° 33′ E; φ = 46° 58′ N) (asl = above sea level) and it has a continuous air pollution monitoring station. At K-puszta the sampling height was above the forest canopy at 20 m over the surface. Aerosol samples were collected with a three-stage cascade impactor on 10 cm diameter cellulose filters (pore size 0.45 µm) with a Millipore membrane pump (1.2–1.5 m³ h⁻¹ air). Relative humidity of the air was kept at 40%. Aerosol samples were collected in three particle size fractions (fine d < 1 µm, coarse 1 µm < d < 10 µm, pre-filter d > 10 µm). Each filter was placed in a clean Petri-dish during transport and storage. The sampling period was from June 1995 to June 1996. Samples were collected differently: in some months every week, in some months every two weeks only. In the winter period sampling was not performed every week due to the weather conditions. In certain weeks, when sampling was performed, a two-day period was used and about 60 m³ air was sampled on average. Hence, the samples were merged together monthly, and samples collected during 13 consecutive months were investigated.

In an earlier study, sampling was performed in a moderately polluted city, Veszprém (h = 260 m asl, λ = 17° 54′ E; φ = 47° 6′ N), and in a regional background sampling site, Kabhegy (h = 600 m asl, λ = 17° 36′ E; φ = 47° 5′ N), on 5 cm diameter Teflon filters (pore size 0.45 µm) with a membrane pump (1.2–1.5 m³ h⁻¹ air). In the two-day periods 55–72 m³ air was sampled. The moderately polluted city of Veszprém is situated in the central part of Transdanubia, 15 km from Lake Balaton. It has about 65 000 inhabitants and there are no industrial activities in the surrounding area. The sampling was carried out on the roof of a university building 20 m above the street level. Kabhegy is part of the Bakony hills and it is located about 20 km north-west from Veszprém. The sampling device was placed at 30 m from the ground.

Leaching procedure

A three-stage sequential leaching procedure was applied to establish the distribution of metals between environmentally mobile, bound to carbonates and oxides, and environmentally immobile (bound to silicates/organic matter) fractions. Details were published elsewhere;⁷ here, only a summary of the procedure is given. 1. Environmentally mobile fraction: filters were folded and placed in centrifuge tubes. Extraction was carried out at room temperature for 15 min with 25 mL of 1 mol L⁻¹ NH₄OAc at pH 7, then the suspensions were centrifuged for 15 min at 3000 rpm. The supernatant phase was separated and stored for analysis. Colloids were not included in the leaching solutions. 2. Bound to carbonate and oxide fraction: residues from stage 1 were extracted at room temperature for 6 h with 25 mL of 1 mol L⁻¹ hydroxylamine hydrochloride + 25% v/v acetic acid and, after leaching, suspensions were centrifuged. The supernatant liquid was stored for analysis. 3. Bound to silicate and organic fraction: the residue was transferred to a PTFE beaker and 10 mL of concentrated HNO₃ were added to each sample. The beakers were placed in a water-bath at a temperature of 95 °C and 2 mL of concentrated HF were added. The beakers were placed in the water-bath and left until all the particles had dissolved. Four millilitres of concentrated HNO₃ were repeatedly added for the evaporation of the HF, and cold solutions were transferred to 50 mL calibrated flasks and made up to volume with 0.1 mol L⁻¹ HNO₃ and stored for analysis.

ETAAS determinations

Elements were determined by electrothermal atomic absorption spectrometry (Perkin-Elmer 5100 PC GEM Software, deuterium background correction, AS-60 autosampler). The operating conditions are summarized in Table 1. The concentrations of Al, As, Cd, Cr, Cu, Fe, Mn, Ni, Pb, and Zn were measured with an RSD of <5% from solutions. For all measurements pyrolytic graphite tubes were used. Standards were prepared from ultra-pure chemicals at 1 g L⁻¹ of each metal and freshly diluted before use. Analytical blanks were performed on filters from the same batch, on centrifuge tubes and on PTFE beakers. The three-stage sequential technique was run on unexposed filters in the same way as the exposed samples. The blank was a reagent + beaker + filter blank and the values were subtracted from the atomic absorption measurements. Before leaching, all glassware and plastic vessels were treated in a solution of 10% m/m HNO₃ for 48 h and washed with doubly distilled water.

Table 1 ETAAS operating parameters, limit of detection (LOD) and graphite furnace program for determination of elements in sequential leaching of aerosols^a

Element	Wavelength/nm	Slit-width/nm	Chemical modifier	Lamp	BC	LOD/µg L^−1	Furnace program^b
							Temp./°C	Hold/s	Ramp/s	Inert gas flow/mL h^−1
a HCL: hollow cathode lamp, BC: background correction. b The temperature of the drying and cleaning steps was 120 and 2600 °C, respectively. c Pre-treatment step. d Atomizing step.
Al	309.3	0.70	Mg(NO3)2	HCL	D2	5	1500^c	30	1	300
							2500^d	5	0	0
As	193.7	0.70	Pd+Mg(NO3)2	HCL	D2	0.5	1300^c	30	1	300
							2300^d	3	0	0
Cd	228.8	0.70	Pd+Mg(NO3)2	HCL	D2	0.1	1700^c	30	1	300
							1800^d	5	0	0
Cr	357.9	0.70	Mg(NO3)2	HCL	—	1	1650^c	30	1	300
							2500^d	5	0	0
Cu	324.8	0.70	Mg(NO3)2	HCL	D2	1	1000^c	30	1	300
							2300^d	5	0	0
Fe	248.3	0.20	Mg(NO3)2	HCL	D2	5	1200^c	30	1	300
							2400^d	5	0	0
Mn	279.5	0.20	Mg(NO3)2	HCL	D2	0.2	1400^c	30	1	300
							2500^d	5	0	0
Ni	232.0	0.20	—	HCL	D2	0.5	1000^d	30	1	300
							2300^d	5	0	0
Pb	283.3	0.70	NH4H2PO4	HCL	D2	0.1	700^c	30	1	300
							1800^d	5	0	0
Zn	213.9	0.70	Mg(NO3)2	HCL	D2	50	700^c	30	1	300
							1800^d	5	0	0

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Measurement uncertainty of ETAAS determination

Aerosol samples were collected on membrane filters and the weekly taken samples were merged as a representative sample package for a month. Therefore, no parallel measurements can be performed since the amounts of aerosols are extremely small. Furthermore, there is no CRM on filters available yet for fine and coarse aerosols, so a systematic procedure has to be taken into consideration for estimation of the measurement uncertainty. For the calculation of the result of each measurement three different approaches were considered, namely, (a) calculation of the result using a calibration graph and estimating the confidence limit by the Student's t-distribution, (b) calculation of the combined uncertainty of measurement, and (c) estimation of the sampling errors using the transport and field blanks together with the calculation of the combined uncertainty of measurement. The last gives the most reliable result since it includes all the parameters which have to be considered in the course of sampling, sample preparation and measurement. In this study transport and field blanks were incorporated in the estimation of measurement uncertainty. As a result of the calculation, 3.5 µg L⁻¹ Pb was found in the leaching solution as combined blanks (field, transport and reagent), which proved to be 8.1% of the total Pb concentration. Details of calculations regarding blanks and combined uncertainty of measurements can be found elsewhere.⁸ It was concluded that the precision of measurements from leaching solutions was about 5.2% as RSD.

Calculation of the enrichment factor (EF)

The origin of aerosols can be characterized by the determination of the aerosol–crust enrichment factor, which is based on the analysis of atmospheric aerosol samples and rocks or soils. Enrichment factors were calculated using the equationwhere i is the concentration of the element considered in the aerosol or the crust and Al is the reference element. The concentration of Al in soils is usually used as an average value of different soils in a given area and is taken from the literature. If the EF value is near unity a certain element is considered to be of mineral origin.

Results and discussion

Speciation of trace metals in the atmosphere refers mainly to the mechanism of interaction of the biosphere and the atmosphere and the mechanism of transport in the atmosphere. Metals are primarily transported in the atmosphere on aerosols that can be removed by wet and dry deposition processes. In this study only dry deposition processes are discussed. Fractionation by chemical bonding of aerosols with a three-stage sequential leaching procedure was carried out and particular attention was paid to distinguishing between environmentally mobile and environmentally stable fractions. Geometric means and ranges of the three size fractions of aerosol samples for 13 months are summarized in Table 2.

Table 2 Average, minimum and maximum concentrations of elements in the three particle size fractions of atmospheric aerosols collected at K-puszta, 1995–96

Element	Geometric mean of the elemental content ng m^−3,n = 13 samples (minimum–maximum concentrations)
	Fine	Coarse	Pre-filter
	d < 1 µm	1 µm < d < 10 µm	d > 10 µm
Al	26 (9.3–132)	27 (13–254)	24 (13–67)
Fe	55 (15–204)	75 (28–220)	88 (44–298)
Mn	0.6 (0.2–2.5)	0.7 (0.3–3.4)	32 (3.2–293)
Zn	3.5 (0.6–102)	6.5 (1.5–22)	27 (9.2–115)
Pb	4.7 (0.5–50)	11 (2.8–42)	18 (3.1–58)
As	2.9 (0.7–48)	1.6 (0.5–10)	2.7 (1.2–4.7)
Cd	1.1 (0.4–7.4)	1.1 (0.3–8.9)	0.8 (0.3–6.2)
Cr	0.6 (0.5–4.4)	0.7 (0.6–1.2)	2.6 (0.6–7.8)
Ni	1.9 (0.4–14)	2.0 (0.5–3.4)	20 (5.9–202)
Cu	2.2 (0.5–16)	4.0 (1.2–13)	9.0 (5.0–29)

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As can be seen, Al compounds are distributed almost evenly in the three particle size fractions (24–27 ng m⁻³). Although the range seems large (9–254 ng m⁻³), the average concentration is around the lower part. These values are much less than those reported earlier either from K-puszta (150 ng m⁻³)⁹ or by Pinto et al.¹⁰ measured at Teplice (240 ng m⁻³). In both cases samples were collected on membrane filters and the total elemental concentration was determined. As expected, the pollution in Teplice was high since this city could be exposed to a high air pollution level caused by emissions from local industry, power plants, lignite mines, residential space heating and motor vehicles. The ratio between the fine and coarse fractions is about 1, so both natural and anthropogenic origins can be estimated.

A significant difference (Student's t-test, n = 13, t_{n − 1} = 2.179, 95%) was observed in the average concentration of Fe between the fine and coarse fractions. Only 55 (15–204) ng m⁻³ was found in the fine fraction and 75 (28–220) ng m⁻³ in the coarse fraction; an even higher concentration, 88 (44–298) ng m⁻³ was identified in the pre-filter, d > 10 µm, fraction. This proves the natural origin of Fe compounds in aerosols. These values are much less than those measured in other studies in a polluted city (150–260 ng m⁻³ in fine fractions and 440–630 ng m⁻³ in coarse fractions).¹⁰

Fig. 1 shows that Al can mostly be found in the mobile and bound to organic matter/silicate fractions in the fine particle size ranges, and bound to carbonate/metal oxide fraction in the coarse ones. No significant differences were observed in the fractionation of Al either by particle size or chemical bonding. Particles > 1.0 µm are generally produced during mechanical processes and these are primarily the result of low-temperature crustal weathering.¹ The similar total concentration of Al and the comparable distribution between the fine and coarse fractions were attributed to a combination of natural and anthropogenic sources. Iron compounds were considerably concentrated in the bound to carbonate/metal oxide fractions and only 11% was found in the environmentally mobile fraction. These findings differ from our earlier conclusion for the investigation at Veszprém and Kabhegy,¹¹ since the environmentally stable fractions were dominant at a much higher concentration level (433 ng m⁻³ at Veszprém, 85% in the 3rd fraction, and 243 ng m⁻³ at Kabhegy, 89% in the 3rd fraction). In that study fractionation by particle size was not performed, bulk aerosol samples only were collected.

Fig. 1 Distribution of Al and Fe in three chemical fractions in fine and coarse particle size ranges.

Manganese compounds were found in very low concentrations in fine and coarse fractions, 0.6–0.7 ng m⁻³. Earlier, 3.5 ng m⁻³ Mn was reported at K-puszta (Molnár et al.¹²). In addition, much higher concentrations were determined in a study including Veszprém and Kabhegy, 232 and 148 ng m⁻³, respectively.¹³ The reason for the extremely high concentrations lies in the fact that there is an active open-cast mine producing manganese ores at a nearby village, which is situated 20 km north-west from Veszprém and 8 km from Kabhegy. The minor concentration of manganese in the aerosols collected at the background station reflects natural sources, since in a highly polluted city, 8–13 ng m⁻³ Mn was obtained.¹⁰ Data in Fig. 2 show that Mn compounds are highly concentrated in the environmentally immobile fractions and do not have any direct impact on the environment.

Fig. 2 Distribution of Mn and Zn in three chemical fractions in fine and coarse particle size ranges.

Zn compounds were concentrated mostly in coarse fractions, 6.5 (1.5–22) ng m⁻³; 3.5 (0.6–102) ng m⁻³ was found in fine fractions, and 27 (9.2–115) ng m⁻³ in the d > 10 µm particle size fractions. These values suggest a natural origin since there is no industrial emission source around this area. The low concentrations characterize well the Hungarian background concentration level and are one order of magnitude less than those found in the fine fractions of urban and suburban sampling sites at Chinese cities (189–645 ng m⁻³).¹⁴ In a Czech study, 110 ng m⁻³ Zn was determined in the fine fraction and only 17 ng m⁻³ Zn in the coarse fraction in winter, while less Zn was identified in samples collected in summer (69 and 15 ng m⁻³, respectively).¹⁰ Both results reflect anthropogenic sources since Zn is mostly emitted from coal combustion, iron and steel production and waste incineration. It is obvious that Zn was accumulated in the mobile fractions, especially in the coarse particle size range (Fig. 2). Although the concentration of Zn is not high, most of the Zn compounds, on deposition, dissolve in natural waters.

Concerning the trace elements an interesting picture was found. The fine fraction of aerosols contained only 4.7 ng m⁻³ Pb with a range of 0.5–50 ng m⁻³ and this average value was much less than that found in other studies: 16 ng m⁻³ at K-puszta⁹ and 25 ng m⁻³ at Veszprém.¹⁵ It is interesting that a higher concentration of Pb was identified in the coarse fractions, as an average of 11 ng m⁻³ with a range of 2.8–42 ng m⁻³. In a fractionation study by size using an eight-stage Berner-type impactor, lead showed a unimodal distribution at a maximum of <0.71 and 1.4 µm indicating an anthropogenic source.⁷ However, the total concentration of Pb compounds in particles of 11.3, 5.7, 2.8, and 1.4 µm size ranges was higher than that found in 0.71–0.088 µm ranges. In this campaign the aerosol particles were collected at a size of <1.0 µm instead of the common PM 2.5 particle size. Not long ago the lead contribution to the atmosphere was dominated by emissions from vehicle exhausts, although smelting operations also contribute to this atmospheric lead load, emitting both PbO and Pb⁰ compounds.¹⁶ The significance of lead sources in the atmosphere is currently changing as a result of the use of unleaded vehicle fuels. The actual composition of lead compounds in a particular aerosol depends on the other constituents in the atmosphere and the age of the aerosol. The dominant inorganic lead compound in aerosols has been identified as PbSO₄·(NH₄)₂SO₄ by XRD^17,18 and this species arises from transformations of the primary emitted aerosol compounds during atmospheric transport.

Data in Fig. 3 clearly indicate that Pb compounds are concentrated in the environmentally mobile fractions especially in the coarse particle size range. This information gives an indication of the mobility of the elements once the aerosol is mixed directly into natural waters throughout scavenging of the aerosols by wet deposition. During mixing of aerosols with aqueous solutions, anthropogenic metals are preferentially released having, potentially, the most harmful impact on the biological community.

Fig. 3 Distribution of Pb and As in three chemical fractions in fine and coarse particle size ranges.

Arsenic compounds were concentrated in the fine fractions, 2.9 (0.4–14) ng m⁻³, and only 1.6 (0.5–10) ng m⁻³ was identified in the coarse fractions. This unambiguously points to an anthropogenic origin, although the concentrations are small. Arsenic has been shown to occur as both As(III) and As(V) in natural samples.¹⁹ Emissions of arsenic from smelters and coal-fired power plants are primarily as the reduced oxide, As₂O_3, and are associated with sub-µm aerosols. If coal and smelting sources dominate in the aerosol, the resulting As(III) released must be relatively rapidly oxidized to As(V). In a Chinese study, 17–30 ng m⁻³ As was determined in the fine fractions, and 2.6–6.6 ng m⁻³ As in the coarse fractions at eight sampling sites in four cities.¹⁴ These findings indicate that urban pollution sources have a much larger contribution to fine particles.

In our earlier study a monitoring system was developed around Lake Balaton and a long-term campaign was performed at the northern part. Results showed that 7.3 ng m⁻³ As, as an average value of 82 measurements, was found.⁷ This concentration is higher compared with the other monitoring stations in different parts of Hungary (1.9–2.4 ng m⁻³) (Molnár et al.¹²). However, similarly to this conclusion, in the 13 samples collected at K-puszta, 77% of the total amount of As was associated with the bound to organic matter/silicates fraction in the fine particles, and only 13% to the environmentally mobile fraction. Furthermore, if fractionation by particle size was considered, As compounds were connected to coarse particles having a maximum value of 2.8 µm in aerosols collected at Veszprém.⁷ The present finding shows that As is predominantly produced from anthropogenic sources during high temperature volatilization processes and subsequently becomes bound to particle surfaces. In the coarse fractions the chemical distribution of As ranges among the three fractions from 25 to 40%.

Cadmium and chromium compounds were determined in small concentrations both in fine and coarse fractions, 1.1 and 0.6 ng m⁻³, respectively. Along with the average concentration of samples collected over 13 months the ranges proved to be narrow (0.3–9 ng m⁻³ for Cd, and 0.5–4.4 ng m⁻³ for Cr), which means that the pollution originating from these two compounds was nearly constant. It was surprising that in this study the Cd concentration was higher than that of Cr compounds, since only 0.73 ng m⁻³ Cd was reported at K-puszta.¹² In the former campaign atmospheric aerosol particles were collected without size separation and the total concentration was measured.

The concentration of Cd in aerosols depends considerably on the location, pollution sources, time, as well as meteorological conditions, and ranges between 1 and 300 ng m⁻³ in major cities.²⁰ In an earlier study much less Cd was found at our sampling sites, 1.35 ng m⁻³ (Kabhegy) and 2.9 ng m⁻³(Veszprém) on average.¹¹ Different samples of aerosol particles over the North Atlantic, South Atlantic, an area near the equator influenced by Saharan mineral dust, and the Antarctic Ocean were collected by a six-stage cascade impactor and analyzed for Cd, Pb, Tl, Ni, Cr, and Fe.²¹ In all samples the heavy metal content was mostly below 1 ng m⁻³. The more anthropogenically influenced aerosol from an industrial area showed higher mean metal concentrations except for iron and chromium. For these two elements, typically of crustal origin, especially high concentrations were found in samples influenced by Saharan mineral dust (0.2–0.95 ng m⁻³ for Cr and 34–524 ng m⁻³ for Fe).

The chemical bonding of cadmium resulted in its association with the environmentally mobile fractions (55% in fine and 68% in coarse fractions), and a smaller amount was found in the fractions as bound to carbonate/oxides (30 and 20%, respectively) and bound to silicate/organic matter (15 and 12%, respectively). In an earlier study it was reported that Cd in urban aerosol was found to be almost completely in exchangeable form.²² Cadmium, liberated during combustion processes, has been shown to occur in elemental and oxide forms whereas emissions from refuse incineration were predominantly as CdCl₂.²²

Chromium compounds are generally emitted into the atmosphere from waste incineration, coal combustion, smelting furnaces, kilns, and metal industries. Fly ash emitted from coal-fired power plants may contain 10–600 mg kg⁻¹ Cr.²³ Concentration ranges (ng m⁻³) in urban regions are: Canada 4–26, USA 2.2–124, and Europe 3.7–277.²⁴ In our earlier campaign a unimodal distribution was obtained for this element, at a maximum of 0.35–0.71 µm particle size ranges with 0.43 ng m⁻³ Cr on average.⁷ This phenomenon can be explained as a combination of anthropogenic sources and dynamic processes. Chromium was partitioned evenly into the three fractions in both particle size ranges (Fig. 4). High proportions of Cr were identified in stable fractions in urban particulate matter analyzed by a sequential extraction procedure.²²

Fig. 4 Distribution of Cd and Cr in three chemical fractions in fine and coarse particle size ranges.

Copper compounds were found to be 2.2 (0.5–16) ng m⁻³ in fine fractions and 4.0 (1.2–13) ng m⁻³ in coarse fractions, respectively, while a considerably higher concentration was achieved in the pre-filter fraction, viz., 9.0 (5.0–29) ng m⁻³. This means that copper originates mostly from natural sources, soil erosion and wind-blown dust. In Teplice, 12 ng m⁻³ in fine fractions and 6.7 ng m⁻³ in coarse fractions were determined in winter, and 17 ng m⁻³ in fine fractions and 12 ng m⁻³ in coarse fractions were identified in summer, respectively. The sampling site was set up in an industrial city in northern Bohemia, so the pollution can be attributed to anthropogenic sources.¹⁰ In different Chinese sampling sites the concentration of Cu was dependent on the urban and suburban locations and ranged from 8.4 to 25.2 ng m⁻³ in urban areas and from 2.3 to 12.2 ng m⁻³ in less polluted areas, respectively.¹⁴ Our findings agree well with those of others, since K-puszta is a Hungarian background monitoring station and aerosols are transported from long distance with natural origin.

A similar tendency was observed for Ni compounds. In the fine fractions, 1.9 (0.4–14) ng m⁻³ was found, while for the 1–10 µm particle size range 2.0 (0.5–3.4) ng m⁻³ was found, and in the d > 10 µm fraction, an order of magnitude higher value, 20 (5.9–202) ng m⁻³, was obtained. The last concentration was considerably above that found in an industrialized city, 0.4–2.5 ng m⁻³ in Teplice.¹⁰ In our earlier long-term study at Lake Balaton, 4.5 ng m⁻³ Ni was measured as an average of 82 measurements.⁷ Unusually high concentrations of Ni were found in a 14 week sampling campaign set up in Veszprém and Kabhegy (regional background station): 44 and 52.7 ng m⁻³, respectively, probably due to local emission sources.¹¹

Fig. 5 shows that Cu is mostly accumulated to the bound to carbonate/oxide fractions (63% in the fine and 44% in the coarse particle size ranges). The mobile portion is not significant considering that the concentrations are low. Nickel compounds behaved differently; the environmentally mobile fraction was 32% in the fine and 48% in the coarse particle size ranges. It can also be concluded, however, that at the background station the Ni pollution is minor. Only half of the Ni compounds can be dissolved in natural waters after dry deposition.

Fig. 5 Distribution of Ni and Cu in three chemical fractions in fine and coarse particle size ranges.

In an earlier study, the dry deposition rates were calculated to estimate the atmospheric aerosol budget of Hungary.¹¹ Samples were collected at different sampling sites such as K-puszta (Hungarian background station with continuous air pollution monitoring system), Kabhegy (local background station in the western part of Hungary), and Veszprém (non-industrialized medium-size city). The dry deposition rates were determined from the results of elemental contents of atmospheric aerosols. The deposition velocity of the aerosol particles was determined with the following equation:⁹

where v is the dry deposition velocity (cm s⁻¹), c_i the concentration of the ith element (ng m⁻³), and v_i the deposition velocity of the ith particle.

Molnár et al.⁹ estimated the deposition velocity using three independent calculations by taking into account the concentrations, the geometric mean aerodynamic diameter of each size range as well as the published dry depositions. Values agreed well with others estimated for different conditions.⁹

The dry deposition (D_d, mg m⁻² yr⁻¹) was determined for coarse, fine and mobile fractions with the following expression:⁹

where D_d is the dry deposition (mg m⁻² yr⁻¹), c_i the concentration of the ith element (ng m⁻³), v the dry deposition velocity (cm s⁻¹), and 0.315 is a calculation factor (for combination of different units).

The dry deposition rates were calculated for different fractions by means of the dry deposition velocities and concentration of elements in particle size and chemical fractions (Table 3). Dry deposition velocities were calculated by Molnár et al.⁹ using the size distributions of aerosols sampled by a cascade impactor. For the environmentally mobile fractions the dry deposition velocities were also calculated. It is assumed that the size distributions measured at one location are, more or less, valid for the background air over the whole country. This is, of course, not completely true, so the data calculated can only be used as an approximate estimation and compared with each other with some caution.⁹ Results are summarized in Table 3.

Table 3 Dry deposition velocities (v/cm s⁻¹), dry deposition rates, D_dry (mg m⁻² yr⁻¹) and those of the environmentally mobile fractions. Samples were collected at K-puszta (national background station, Nat.), Kabhegy (regional background station in the western part of Hungary, Reg.),¹³ and Veszprém (non-industrialized medium-size city, City)¹³

Element	v coarse/cm s^−1	v fine/cm s^−1	Nat. Ddcoarse	Nat. Ddfine	Nat. Ddmob	Reg. Ddtot ^13	Reg. Ddmob ^13	City Ddtot ^13	City Ddmob ^13
Al	0.51	0.026	4.3	0.21	1.7	70	10.6	77.5	3.7
Mn	0.52	0.024	0.11	0.004	0.012	31.7	8.2	49.8	7.8
As	0.33	0.026	0.17	0.024	0.07	0.14	0.06	0.17	0.01
Cd	0.45	0.027	0.16	0.009	0.07	0.04	0.01	0.09	0.007
Cr	0.43	0.027	0.095	0.005	0.019	0.06	0.02	0.33	0.13
Cu	0.48	0.026	0.60	0.018	0.105	0.13	0.01	0.22	0.03
Ni	0.71	0.029	0.58	0.017	0.103	3.6	0.42	3.1	0.5
Pb	0.38	0.025	1.3	0.04	0.6	0.83	0.30	1.51	0.71

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Data in Table 3 indicate that dry deposition rates range between 0.095 and 4.34 mg m⁻² yr⁻¹ for coarse fractions, and between 0.004 and 0.21 mg m⁻² yr⁻¹ for fine fractions. It can also be observed that the dry deposition rates in the coarse fractions are always much higher (in some cases one order of magnitude), than those in the fine fractions in samples collected at K-puszta. This means that the pollution originates mostly from natural sources. If the three sampling sites are compared an interesting picture can be seen. For Al, Mn, and Ni the dry deposition rates are smaller than those in samples collected at either Kabhegy or Veszprém. The largest differences among the values were observed for Ni: in K-puszta 0.5, in Kabhegy 3.6 and in Veszprém 3.1 mg m⁻² yr⁻¹ was found, respectively. However, for other elements higher dry deposition rates were found at the national background station than at the other two sites. The dry deposition rates of chromium were the lowest both in the fine and coarse fractions (0.005 and 0.095 mg m⁻² yr⁻¹, respectively). Arsenic has a dry deposition rate of 0.17 mg m⁻² yr⁻¹ in the coarse fractions at K-puszta, while values of 0.14 and 0.17 mg m⁻² yr⁻¹, respectively, were found at the regional background station and the city. Dry deposition rates of Cd and Cu at K-puszta proved to be four times higher than those found at Kabhegy and this value was even higher than we obtained in Veszprém. This finding indicates that, for some elements, the western part of the country is less polluted and there are no local pollution sources of As, Cd, Cr, Cu and Pb at Kabhegy and Veszprém.

If the dry deposition rates obtained for these two sampling sites are compared, it can be seen that, except for Ni, lower values were always found at Kabhegy (local background station) than at Veszprém (see Table 3). Along with the total dry deposition rates, those of the environmentally mobile fractions are also tabulated. As was pointed out in another study, the deposition of Mn is extremely high in this area, both at Kabhegy and at Veszprém, compared with other parts of Hungary, due to the nearby mining activities.¹³ On comparing our findings with those of other studies, it was observed that the deposition of Pb, Cd, Cu, and Cr does not differ considerably from other published values^25,26 based on different approaches. However, taking into account the data derived for the environmentally mobile fractions, 0.71 mg m⁻² yr⁻¹ of Pb, a completely non-crustal element, is directly available to the biosphere. For Veszprém this is equivalent to a 13.6 kg dry Pb deposition annually; half of this Pb pollution is in an environmentally mobile form and can be harmful to the direct environment.¹³

A frequently used method for relating an element in atmospheric aerosols to its source is to calculate enrichment factors (EF) by employing an indicator element. For crustal aerosols, Al is normally used as the indicator element. The enrichment factor was calculated using the equation

The concentration of Al in soils is usually used as an average value of different soils in a given area and is taken from the literature.⁹ Those elements that have EF values between 1 and 10 are usually termed as crustal, while elements with EF values in the range 10–5000 are generally emitted by anthropogenic sources.⁵ Results are summarized in Table 4. Along with EF values calculated for the total concentration (sum of fine and coarse fractions) of each element, EF values of fine and coarse fractions alone, as well as environmentally mobile fractions, are also tabulated.

Table 4 Enrichment factors (EF) of aerosol particles for total concentration of elements and those for fine and coarse fractions, as well as for environmentally mobile fractions in samples collected at K-puszta in 1996. Other data for Veszprém (Vp),¹³ Kabhegy (Kab.),¹³ K-puszta (1993)¹² and Budapest¹²

Element	EF K-puszta	EFfine K-puszta	EFfine, mobileK-puszta	EFcoarse K-puszta	EFcoarse, mobileK-puszta	EF K-puszta1993	EF Vp. 1994–95	EFVpmob	EF Kab.1994–95	EFKabmob	EF Buda-pest, 1993
Al	1	1	1	1	1	1	1	1	1	1	1
Fe	4.7	4.1	1.1	5.4	0.6	2.9	3.7	0.06	2.5	0.09	5.1
Mn	3	2.8	0.5	3.1	1.2	3.1	108	9.8	76	11.3	5.9
Pb	3086	1883	2780	4244	13517	824	1685	1206	1028	578	7710
Cd	17 296	17 628	23 085	16 975	42 753	—	4665	241	2404	356	—
Cr	39	37	24	42	46	59.4	24.2	12.4	20	8.9	57.7
Cu	235	170	57	298	243	68	59.8	10.0	38.7	4.3	158
Ni	135	134	102	136	460	26.1	312	84	414	78.3	40.9
As	6847	8995	2784	4779	4425	2265	809	57.2	655	57.8	6443

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Iron is a crust-dominated element with EF < 10. Manganese shows small values at K-puszta and Budapest, and, as was found earlier, a considerable anthropogenic source was identified near Kabhegy and Veszprém.¹³ This is due to an active open-cast mine producing manganese ores at a nearby village. EF values show that Pb, Cd, Cr, Cu, Ni, and As are emitted from anthropogenic sources. Comparing the EF values of fine and coarse fractions it can be seen that there is no straightforward tendency for the enrichment, since, depending on the elements, either EF_fine or EF_coarse is larger. These values are largely determined by the Al content of different fractions. As is obvious from Table 2, the Al content of aerosols collected at K-puszta is considerably less than that found in other parts of Hungary.^12,13 Therefore, as a consequence, the EF values are much higher than those found in studies performed in 1993–95. In some cases the environmentally mobile fractions show extremely high EF values, providing clear evidence of the anthropogenic origin. If the EF values of two studies at K-puszta are compared it can be concluded that, except for Cr, much higher EF values were measured during the campaign in 1996. Nevertheless, the Al content was found to be 131 ng m⁻³ at that time compared with 53 ng m⁻³ in the later experiments (fine + coarse fractions). Concerning the other elements, there were no considerable differences in the concentrations.¹² Extremely high EF values were calculated for Cd in all fractions and particle size ranges in samples collected at K-puszta. Although the total amount of Cd is emitted from anthropogenic sources, EF ranges between 4665 (Veszprém) and 2404 (Kabhegy).¹³ The cadmium species in aerosols are probably Cd, CdS, CdO, Cd(OH)₂ and mixed oxides with copper and zinc.²⁷

EF values of Cr, Ni and Cu indicate that these elements are also emitted by anthropogenic sources; however, the factors of Cr are lower than those calculated for Budapest and K-puszta in 1993.¹² The range is narrow, independent of the particle size fractions and the concentration of total, as well as mobile fractions; EF values between 9 and 60 were found. On the other hand, EF values of Cu and Ni are higher than those calculated for aerosols collected at other sampling sites. Considerable portions of Ni were identified in mobile forms (32% in the fine and 48% in the coarse fractions) and these Ni compounds could be dissolved, on deposition, in the receiving media.

EF values of As are much higher than those of the other sites, except at Budapest, indicating the anthropogenic sources in this area. About the same concentrations of As were found in the aerosol samples collected at K-puszta compared with those found in Kabhegy and Veszprém.¹¹ A considerable portion of As compounds was identified in the stable fraction (Fig. 3). This is entirely consistent with the long-range transport of As from its industrial sources, since it is released during high temperature volatilization processes and subsequently becomes bound to particle surfaces or is incorporated in individual compounds. The major chemical species of arsenic are As₂O₃, As₂S₃ and organoarsenic compounds from combustion and metallurgical technologies.²⁷

In this work a further step towards the estimation of the origin of aerosols was developed and applied, namely, the calculation of EF values, along with the total concentration, for fine and coarse fractions, as well as the environmentally mobile portions. A comparison of our findings with others measured in other parts of the world can be accomplished by use of EF values of total aerosol concentration, since these are available in the literature. However, a more serious analysis can be performed by using the detailed data on physical (particle size) and chemical (bound to mineralogical phases) fractionation. Nevertheless, from the results of new findings and the characterization of the state of the environment, some other factors should also be taken into account. The reliability of results of analytical measurements greatly depends on the sampling (pollution from sampling head, sampling device, or surroundings of samplers), sample preparation and measurement uncertainty.

Conclusion

Metals can be deposited from the atmosphere by dry and wet deposition. In this study only dry deposition was thoroughly investigated by a monitoring system over a 13 month period. Detailed information on the origin, amount and bioavailability of elements can be collected by the fractionation of aerosols according to particle size and chemical bonding. In a two year sampling campaign set up at K-puszta, a Hungarian background meteorological station, aerosols were collected by a cascade impactor in three size fractions, viz., <1 µm (fine), 1–10 µm (coarse) and <10 µm (pre-filter). Only the fine and coarse fractions were studied further. With a simple sequential leaching procedure elements were classified into three fractions: (i) environmentally mobile, (ii) bound to carbonate/oxide, and (iii) bound to silicate/organic matter. Since the concentrations of elements were low, all analyses were performed by ETAAS. The mobility order of elements in fine and coarse fractions is shown in Fig. 6 and 7, respectively.

Fig. 6 Mobility order of elements in fine particles.

Fig. 7 Mobility order of elements in coarse particles.

It is obvious that Pb, Cd and Zn can be identified, in the greatest amount, in mobile fractions. Furthermore, in coarse fractions, the mobile portion proved to be even higher (Fig. 7) than in fine fractions, which means that the natural pollutants have considerable amounts of elements available directly to the biosphere and hydrosphere. On the basis of EF values it can be concluded that the trace elements originate from anthropogenic sources even in the present case, where the sampling station is located more than 50 km from industrial activities. In the Hungarian investigations, aerosol sampling was always carried out under national background conditions for studying the long-range transport of pollution in aerosol form.

On comparing the data of the present campaign with those of 1993 it was found that the elemental composition of aerosols had not changed considerably,¹² the concentration of some elements (Al, Fe, Cr, Mn and Zn) had decreased insignificantly, and others had slightly increased. Differences among EF values were much greater due mostly to differences of Al concentrations in the fine and coarse fractions, as well as the mobile portion from sequential leaching. These factors are appropriate indicators of the environmentally and biologically available fractions of elemental pollution in aerosols.

acknowledgement

This research was supported by the Hungarian National Science Foundation (OTKA) T 029250 and FKFP 0084/1999. The financial support is greatly acknowledged. The authors thank Dr Agnes Molnar for helpful discussions.

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