Airborne foliar transfer of PM bound heavy metals in Cassia siamea: A less common route of heavy metal accumulation

Airborne foliar transfer of PM bound heavy metals in Cassia siamea: A less common route of heavy metal accumulation

Science of the Total Environment 573 (2016) 123–130 Contents lists available at ScienceDirect Science of the Total Environment journal homepage: www...

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Science of the Total Environment 573 (2016) 123–130

Contents lists available at ScienceDirect

Science of the Total Environment journal homepage: www.elsevier.com/locate/scitotenv

Airborne foliar transfer of PM bound heavy metals in Cassia siamea: A less common route of heavy metal accumulation Triratnesh Gajbhiye a, Sudhir Kumar Pandey a,⁎, Ki-Hyun Kim b,⁎, Jan E. Szulejko b, Satgur Prasad c a b c

Department of Botany, Guru Ghasidas Central University, Bilaspur 495009, CG, India Department of Civil and Environmental Engineering, Hanyang University, Seoul 04763, Republic of Korea Department of Analytical Chemistry, Indian Institute of Toxicological Research, Lucknow 226001, India

H I G H L I G H T S

G R A P H I C A L

A B S T R A C T

• Foliar dust and whole leaves were used to assess the metal pollution. • Metals in soil and road dust were considered for EF and source-apportionment. • Scanning EM helped observe the role of morphology of leaves in accumulation of PM. • Study showed airborne transfer of Pb and Cd in foliar region of Cassia siamea. • Cassia siamea is a suitable plant for phytomonitoring of PM bound metals.

a r t i c l e

i n f o

Article history: Received 25 June 2016 Received in revised form 12 August 2016 Accepted 15 August 2016 Available online xxxx Editor: D. Barcelo Keywords: Dust Metal Phytomonitoring Enrichment Leaf SEM

a b s t r a c t In order to investigate possible foliar transfer of toxic heavy metals, concentrations of Cd, Pb, and Fe were measured in samples of: Cassia siamea leaves (a common tree) Cassia siamea foliar dust, nearby road dust, and soil (Cassia siamea tree roots) at six different sites in/around the Bilaspur industrial area and a control site on the university campus. Bilaspur is located in a subtropical central Indian region. The enrichment factor (EF) values of Pb and Cd, when derived using the crustal and measured soil Fe data as reference, indicated significant anthropogenic contributions to Pb and Cd regional pollution. Based on correlation analysis and scanning electron microscopy (SEM) observations, it was evident that Pb and Cd in foliar part of Cassia siamea were largely from airborne sources. The SEM studies of leaf confirmed that leaf morphology (epidermis, trichome, and stomata) of Cassia siamea helped accumulate the toxic metals from deposited particulate matter (PM). There is a line of evidence that the leaf of Cassia siamea was able to entrap PM in respirable suspended particulate matter (RSPM) range (i.e., both in fine and coarse fractions). The overall results of this study suggest that Cassia siamea can be a potential plant species to control the pollution of PM and PM-bound metals (Pb and Cd) in affected areas. © 2016 Elsevier B.V. All rights reserved.

1. Introduction ⁎ Corresponding authors. E-mail addresses: [email protected] (S.K. Pandey), [email protected] (K.-H. Kim).

http://dx.doi.org/10.1016/j.scitotenv.2016.08.099 0048-9697/© 2016 Elsevier B.V. All rights reserved.

Heavy metals can readily combine with airborne particulate matter (PM) (Charlesworth et al., 2003; Yongming et al., 2006). Consequently,

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deposition of PM-bound heavy metals is responsible for the contamination over considerably large surface areas. Vehicular transportation is responsible for producing and mixing fine dust particles on the street (Zheng et al., 2010). Chronic deposition of heavy metals even at low rates can accumulate in the environment to pose serious environmental and human health hazards (Walkenhorst et al., 1993; Banerjee, 2003; Yongming et al., 2006). These dust bound metals can also be transported over long distances (Tabatabaei et al., 2015). The dusts containing these metals are deposited on surfaces such as plant foliage. Use of soil and dust as indicator to assess the effects of anthropogenic activities is a hot topic (e.g., Simon et al., 2013; Karbassi et al., 2015). Plants may also be used as biological filter, as they are capable of accumulating the PM and associated metals (Nowak et al., 2006; Escobedo et al., 2008). Areas near industrial activities generally consist of higher levels of dust and associated metals (Shparyk and Parpan, 2004; Wilson et al., 2005; Zibret and Sajn, 2008; Simon et al., 2016). The elemental screening of foliar dust and leaves is especially useful to assess the environmental air pollution (Simon et al., 2011). However, the extent can vary due to morphological differences among plants (e.g., cuticular layers with epicuticular-cuticular waxes, raised epidermal cell boundaries, stomatal types, structures, density, and trichomes shape) (Pal et al., 2002; Simon et al., 2014). Roughness and large surface area of leaves may promote the accumulation of PM and toxic metals from environment (Freer-Smith et al., 1997; Meusel et al., 1999; Mulgrew and Williams, 2000; Sæbø et al., 2012). Plant leaf surfaces were also analyzed through scanning electron microscope-energy dispersive X-ray (SEM-EDX). Accordingly, plants leaves were seen to trap small PM (b 2.5 μm) with trace elements (e.g., Pb) (Tomasevic et al., 2004; Tomasevic and Anicic, 2010). Hence, the macro and micro-morphology of a leaf surface can be employed to assess the extent and impact of PM pollution (Lohr and Pearson-Mims, 1996; Freer-Smith et al., 1997; Barber, 2004; Tomasevic et al., 2005; Rico et al., 2011; Schreck et al., 2012a; Sawidis et al., 2012). Sawidis et al. (2012) studied the leaf surfaces of different plants (adaxial and abaxial) to evaluate the leaf morphological characteristics (stomatal and trichome structure and density) which are responsible for trapping PM. In recent years, much research has been conducted on foliar uptake of heavy metal from atmospheric PM deposition (Honour et al., 2009; Uzu et al., 2010; Schreck et al., 2012a, 2012b). The mechanism of tentative pathway for metal uptake from deposited PM has also been postulated. However, the extent of foliar transfer of these toxic metals along with PM can vary across a number of variables such as plant vs. metal type along with environmental conditions. Hence, there is a pressing need to elucidate the uptake process of these PM bound toxic metals by plants growing naturally in affected areas. In this research, we have selected the common plant species Cassia siamea identified as one of the most abundant plant species in the study area. At present, the species “siamea” is categorised under the “Senna” genera (Singh, 2001; Boonkerd et al., 2005; and Acharya et al., 2011). However, we have referred it in its former “Cassia” genera. Cassia siamea is an evergreen tree. Leaves are pinnately compound 10–25 cm long, rachis pubescent. Leaflets are 6–14 pairs ovate-oblong. In Cassia siamea, fewer stomata are present at upper surface in comparison to the lower surface. Trichomes are present along with margin on adaxial (upper) surface and along with midrib, leaflets margin on abaxial (lower) surface. According to Reddy et al. (2012), Cassia siamea is a possible accumulator of Cr in coal mining area. For the phytoremediation of Cd, Cassia siamea was found highly suitable plant (Tripathi et al., 2004; John et al., 2011). Cassia siamea also showed high air pollution tolerance index (APTI) (Jain and Kutty, 2014). This study was thus carried out to investigate the factors and processes controlling the airborne transfer of toxic metals (with an emphasis on Pb and Cd) in Cassia siamea (leaves). For this purpose, their concentrations in foliar dust and leaves of Cassia siamea plant were monitored to evaluate the accumulation of

PM in different size fractions and their possible link to metal pollution in a heavily industrialized area of Bilaspur region, Chhattisgarh, India. 2. Materials and methods 2.1. Study area Bilaspur is located in Chhattisgarh state, central India in a tropical savanna climate zone (Köppen climate classification: Aw). Estimated population is 452,851 in an area of 145.76 km2. The study area (see Fig. 1) was 7.5 km (site 1) to 36 km (site 3) from Bilaspur city center (21°55 N; 82°01 E). Samples were collected (on 15 February 2016) in the dry post-monsoonal winter season. In the two weeks prior to sampling, the weather conditions were (a) the wind direction was variable: SW (19.6%), NNE (12.5%), NE (10.7%), SSW (7.1%), and all other directions (50%) (b) the mean daily temperature during this period ranged between 23.5 and 25.0 °C (min 17 °C, max 36 °C), (c) the relative humidity varied between 25 and 49%, and (d) no rainfall was recorded (http:// www.worldweatheronline.com/). Plants leaf samples were collected from six different industrial locations and one university campus control site (Fig. 1). The sites 1 to 6 represent highly polluted areas due to anthropogenic sources (including road traffic, stone crusher, cement works, iron works, and a waste oil dump). Vehicles are a main source of PM and metals (Furusjö et al., 2007; Ewen et al., 2009). Site 1 was located near an oil processor where pits are used to store waste oil transported by heavy trucks. The sampling site 2 was located near a steel works. Site 3 was located near a stone crusher and site 4 was located near a cement works. Site 5 is located near a sponge iron works and site 6 is located near a coal burning power station. A site with minimal local source pollution, distant from industrial activities and heavy traffic was selected as the control site (Guru Ghasidas Central University, Campus, Koni) (Table 1). 2.2. Sampling of plant material, foliar dust, road dust and soil All samples from each site were collected within one day (i.e., on 15 February 2016) to minimize temporal effects. All plant leaf/foliar dust samples from the Cassia siamea trees were collected at a height of ~ 2 m above ground level at all sites to match the ambient height and minimize road-dust re-suspension during sample collection. Road dust samples were collected from three to four points (spaced a few metres apart) on both roadside shoulders along the paved roads using a fine plastic brush and a cleaned tray for each dust collection. The target soil and Cassia siamea sites were located 2–3 m from the nearest roadside shoulder edge and the road paved width was ~7.5 m. The collected roadside dust samples at each site were then combined and mixed thoroughly to obtain a common 200 g dust sample. All road dust samples are sealed in a polythene bags. Leaf samples were collected in the Ziploc plastic bags. The soil samples near each study site were also collected from topsoil (lower A-Horizon) 30–35 cm depth, around the roots near the Cassia siamea tree trunk by scraper plate method (Faiz et al., 2009). 2.3. Sample preparation for analysis of heavy metals The foliar dust from leaves was removed in a closed glass chamber to avoid loss of fine dust particles during separation. Foliar dust was removed from both surfaces of leaves with the help of fine hair brush. Special care was taken to avoid loss of fine foliar dust particles during the separation of dust from sampled leaves. After removing the heavy foliar dust from leaf surface, leaves were thoroughly washed in 200 ml microdistilled water for 5 min in a ceramic container. The container was hand shaken for 7 min before the leaves are removed. Then, washed leaves were placed in 50 ml of chloroform and rinsed off for 5 min in ceramic container to remove additional PM (those are trapped on epicuticular wax surface on leaf) (Sgrigna et al., 2015). The dust containing

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Fig. 1. Geographical location of study sites (Bilaspur, CG, India). Site 1 - Waste oil dump site, Site 2 - Near a Steel iron industry, Site 3 - Near a Stone crusher, Site 4 - Near a Cement industry, Site 5 - Near a Sponge Iron industry. Site 6 - Near a Coal power plant, and Control site at Guru Ghasidas University Campus, Bilaspur.

suspension was placed into microwave oven to evaporate the liquid content from solution. Then, it was oven dried for 24 h at 30 °C and crushed into fine powder. The road dust and soil samples were airdried in the laboratory at room temperature for 2 days. The handling of these samples was carried out without contact with metals, to prevent cross-contamination. The weight of samples (foliar dust, leaf powder, road dust and soil) was 0.5 g for digestion. All samples (foliar dust, leaf powder, road dust, and soil) were digested separately into 10 ml of aqua-regia solution (HNO3: HCl v/v 3:1). Solution was incubated at room temperature for 24 h. Afterwards, the solutions were heated to reflux for 15 min and set aside to cool. After digestion, solution was filtered through filter paper Whatman no. 1. Distilled H2O was added to Table 1 Characteristics of sampling sites. Sites

Characteristics of sampling site

Control site Site 1 Site 2 Site 3 Site 4 Site 5 Site 6

Guru Ghasidas Central University, remote clean area Waste oil dump site, heavy vehicular traffic (National highway) Near a steel iron industry, medium traffic. Near a stone crusher industry, near a highway Near a cement industry Near a sponge iron industry with moderate vehicular activity Near a coal fired power plant

make up the volume up to 50 ml. and collected into test tube for analysis of heavy metals (US EPA method 3050B, 1996: http://www.epa.gov/ osw/hazard/testmethods/sw846/pdfs/3050b). 2.4. Analysis of metals After sample preparation, heavy metal concentrations were determined using flame atomic absorption spectroscopy (AA 7000, Shimadzu, Japan). Standard solution of target metals were obtained from Inorganic Ventures, 300 technology drive, Christiansburg, VA 24073, USA (Cd: AACd-1, 1000 ± 10 μg/ml, 3% (v/v) HNO3 traceable to NIST, SRM, Pb: AAPb-1, 1000 ± 10 μg/ml, 0.5% HNO3 traceable to NIST-SRM 3128, Fe: AAFe-1, 1000 ± 10 μg/ml, 2% (v/v) HNO3 traceable to NIST-SRM 31269). The standard solutions were diluted to make a five point calibration. From the calibration curve, the concentrations of Fe, Cd and Pb in different samples were quantified after applying a blank correction. All chemical analyses were done in triplicate. The concentration values were expressed as μg g−1. The detection limit (DL: μg g−1) of three target metals were 0.04 (Fe), 0.08 (Pb), and 0.006 (Cd). The precision, when expressed in terms of relative standard deviation (RSD %), was below 5% for all the metal. The average recovery of different metals from randomly selected fortified samples (soil) was as follows: Fe (88.8%), Pb (104.0%), and Cd (98.5%).

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2.5. Scanning electron microscopy (SEM) Small strips of leaf (about 0.5 cm2) were trimmed from area between the margin and mid-rib of leaves. The leaves were kept in 2.5% glutaraldehyde solution and left overnight for fixation (prefixation). Those samples were kept in osmium tetraoxide for post fixation for 1 h. The samples were washed with buffer solution two times for 15 min and then passed through a series of acetone solution (30%, 50%, 70%, 95% and 100%) for dehydration. Drying was done in a critical point drier (CPD) using CO2 as a carrier gas. Scanning electron microscopy (SEM) was carried out using a 30 keV, JEOL JSM-6490LV scanning microscope with standard automated features like auto focus/stigmator, auto gun and auto contrast with multiple live image display. The samples were coated with carbon using a high vacuum system to wet specimens. Both the upper and lower surfaces of the leaves segments were studied and micrographs were taken at various magnifications. 3. Results and discussion 3.1. The distribution of target metals in different sample types The concentrations of heavy metals determined from foliar dust, leaf, road dust, and soil samples are presented in Table 2. Correlation analysis was performed between the concentration of metals in leaf vs. soil, road dust and foliar dust (Table 3). 3.1.1. Fe In this study, Fe was the most abundant element compared to Pb and Cd regardless of site and sample type. It was distributed in a narrow range in the case of its foliar dust concentrations (i.e., 3536 ± 37.1 μg g−1 (site 3) to 3640 ± 81.3 μg g−1 (site 5)) (Table 2(A)). However, if the Fe concentrations were normalized relative to the Control site for each sample type, they ranged: foliar dust (1.81–1.86), leaf (1.67–4.49), road dust (3.57–5.19), and soil (1.49–2.02) (Table 2(B)). This indicates relatively low anthropogenic emissions at all studied industrial sites (1 to 6). The average Fe content in road dust/foliar dust, and soil in this study ranged from 0.33 to 0.38% which was unusually low in comparison to the previous reports (e.g., Tabatabaei et al., 2015). It was also noted that the concentration of Fe in leaves were variable from (1079 ± 46.5 μg g−1) at site 6 to (2911 ± 38.6 μg g−1) at site

Table 3 Correlation analysis between concentration of metals in leaf vs. concentration of metals in foliar dust, road dust, and soil samples. Leaf

Fe

Pb

Cd

Foliar dust Road dust Soil

0.30 −0.28 0.06

0.80a 0.77a 0.20

0.95b 0.80a 0.14

a b

Correlation is significant at 0.05 level. Correlation is significant at 0.001 level.

5. These concentrations were N4 times higher when compared to the value of control site (Table 2(B)). In general, the Fe concentrations in soil, road dust, and foliar dust samples were higher compared to leaf samples. Correlation was not significant between leaf vs. soil/dust/foliar dust.

3.1.2. Pb The presence of Pb in four sample types (foliar dust, road dust, leaf, and soil) at all sites indicates anthropogenic sources. For instance, Pb concentration was found at the order of foliar dust N road dust N leaf N soil. In foliar dust, the concentration of Pb was found in range of 8.67 ± 0.77 μg g− 1 (site 4) to 42.3 ± 0.66 μg g− 1 (site 1). The Pb leaf concentration ranged from 3.34 ± 0.69 μg g−1 (site 4) to 16.0 ± 0.80 μg g−1 (site 1). Our leaf Pb results are comparable to the Pb concentrations (2.79–4.94 μg g−1) found in the leaves of Hypnum cupressiforme, Quercus ilex, and Pinus halepensis trees 2 km from the Barcelona metropolitan area, Spain (Sardans and Penuelas, 2005). The leaf Pb levels (0.70–1.30 μg g−1) in spinach, cabbage, lettuce, carrot and tomato (taken from fertilized farmland and roadside areas in Elaziğ, Turkey (Yaman and Güçer, 1995)) were substantially lower. A tentative correlation was found between leaf vs. foliar dust (r = 0.80, P b 0.05) as well as leaf vs. road dust (r = 0.77, P b 0.05). Hence, in the case of Pb, its partitioning/transfer through airborne route preceded either directly or through road dust re-suspension. The mechanism of airborne Pb transfer has been described (Uzu et al., 2010; Schreck et al., 2012a). Accordingly, Pb (all forms) in leaf deposited PM is converted into PbO2, PbSO4, PbO·PbSO4, and PbCO3 on the leaf surface (Uzu et al., 2010). Consequently, Pb may enter into leaf by two pathways; (1) through the stomatal opening in the form of nanoparticles

Table 2 Metal concentrations of metals at different study sites (μg g−1). (A) Metal

Fe

Pb

Cd

Sample type

Foliar dust Leaf Road dust Soil Foliar dust Leaf Road dust Soil Foliar dust Leaf Road dust Soil

Control site

S1

S2

S3

S4

S5

S6

University campus

Waste oil processor

Steel works

Stone crusher

Cement works

Sponge iron works

Coal power station

1952 ± 35.6 648 ± 11.8 849 ± 17.3 2257 ± 42 –a – – – – – – –

3622 ± 38.4 1969 ± 11.5 3974 ± 71.7 4045 ± 28.8 42.3 ± 0.66 16.0 ± 0.80 23.9 ± 2.27 4.20 ± 0.84 15.6 ± 0.98 12.1 ± 0.7 12.6 ± 0.87 0.74 ± 0.42

3590 ± 57.3 2765 ± 17 3086 ± 60 3358 ± 35.6 39.2 ± 0.68 13.4 ± 0.66 30.9 ± 1.37 2.48 ± 1.24 8.35 ± 0.28 6.69 ± 0.29 6.82 ± 0.30 0.56 ± 0.14

3536 ± 37.1 1551 ± 42.2 4277 ± 30.8 4553 ± 27.6 36.9 ± 0.98 9.65 ± 0.53 8.22 ± 0.68 1.78 ± 0.56 14.4 ± 0.26 9.57 ± 0.41 11.5 ± 0.79 0.82 ± 0.24

3537 ± 43.5 1708 ± 34.2 4405 ± 45.1 4255 ± 32.7 8.67 ± 0.77 3.34 ± 0.69 6.54 ± 0.33 3.25 ± 0.75 11.5 ± 0.36 8.56 ± 0.62 11.0 ± 0.51 0.52 ± 0.08

3640 ± 81.3 2911 ± 38.6 3384 ± 149 4561 ± 22.3 15.6 ± 0.82 5.29 ± 0.62 8.35 ± 0.78 2.54 ± 0.86 9.42 ± 0.41 6.50 ± 0.45 7.49 ± 0.44 0.78 ± 0.12

3628 ± 51.9 1079 ± 46.5 3030 ± 59.4 3563 ± 45 14.7 ± 0.64 11.5 ± 0.8 9.56 ± 0.38 1.86 ± 1.06 8.82 ± 0.62 7.40 ± 0.5 11.4 ± 0.33 0.84 ± 0.24

(B) Normalized Fe concentrations relative to control for each sample type Study sites Metal

Samples

Control

S1

S2

S3

S4

S5

S6

Fe

Foliar dust Leaf Road dust Soil

1 1 1 1

1.86 3.04 4.68 1.79

1.84 4.27 3.63 1.49

1.81 2.39 5.04 2.02

1.81 2.64 5.19 1.89

1.86 4.49 3.99 2.02

1.86 1.67 3.57 1.58

a

– = not detected.

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in the apoplasm or (2) through hydrophilic pathway by diffusing through the aqueous pore of stomata and cuticle.

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Table 4 Cd and Pb enrichment factors according to sample types at all different sites. (A) Enrichment factors (conventional)a

3.1.3. Cd The concentrations of Cd showed similar order (foliar dust N road dust N leaf N soil) as of Pb in all six sites of 1 to 6. The concentration of Cd in foliar dust samples ranged from 8.35 ± 0.28 μg g− 1 (site 2) to 15.6 ± 0.98 μg g− 1 (site 1). In leaf, concentration of Cd varied from 6.50 ± 0.45 μg g−1 (site 5) to 12.1 ± 0.70 μg g−1 (site 1). Our Cd results are much higher than the Cd concentrations (0.14–0.19 μg g−1) found in the leaves of Hypnum cupressiforme, Quercus ilex, and Pinus halepensis trees 2 km from the Barcelona metropolitan area, Spain; leaf Cd levels in trees located much further away (80–102 km) were an order of magnitude lower (0.019–0.037 μg g−1) (Sardans and Penuelas, 2005). Similarly, the leaf Cd levels (0.15–0.80 μg g−1) in spinach, cabbage, lettuce, carrot and tomato (Yaman and Güçer, 1995) were more than an order of magnitude lower than the present data. Correlation analysis indicated airborne transfer of this toxic metal in the leaves with a significantly higher correlation at 0.001 level: leaf vs. foliar dust (r = 0.95, P b 0.001) and 0.05 level from leaf vs. road dust (r = 0.80, P b 0.05). 3.2. Enrichment factor (EF) EF is one of preferable options to indirectly appraise the dispersal of heavy metals from anthropogenic sources. A number of different EFs can be calculated based on a metal's concentration in a given sample and a reference background sample (Hornung et al., 1989; Lorenzini et al., 2006; Huu et al., 2010). The EF method (Eq. (1)) typically normalises the measured heavy metal content with respect to an abundant reference element (typically Al, Fe, or Si) found in the Earth's crust (Ravichandran et al., 1995; Lorenzini et al., 2006). Generally, Fe has a relatively high abundance in natural soil reservoir. As such, soil Fe is generally not appreciably enriched from anthropogenic sources, it is preferably selected as normalisation basis in the calculation of the EF (Deely and Fergusson, 1994; Niencheski et al., 1994; Neto et al., 2006; Mediolla et al., 2008; Vaezi et al., 2015). Hence, in this study, Fe was taken as a reference element [5 wt% (50,000 μg g−1 in the Earth's crust)]. Thus, the EF of a given metal can be calculated:

EF ¼

 SðEÞ =SðRÞ leaf  CðEÞ =CðRÞ crust

ð1Þ

Metal Sample Control site

• EF b 2 = minimal enrichment • 2 ≤ EF b 5 = moderate enrichment • 5 ≤ EF b 20 = significant enrichment (suggesting low anthropogenic emissions) • 20 ≤ EF b 40 = very high enrichment (suggesting moderate anthropogenic emissions) • EF ≥ 40 = extremely high enrichment (suggesting high anthropogenic emissions) Based on these criteria, the EF results were compared at sites 1–6 (Table 4). For Cd, unusually high enrichment factors (EF: 737–2286) was noted for 3 sample types (foliar dust, leaf, and road dust) at sites 1 to 6 (Table 4). This suggests that Cd is of anthropogenic origin (e.g., car tyres). The Pb EF values were much lower than for Cd: low to

S2

S3

S4

S5

S6

University Waste Steel Stone Cement Sponge Coal campus oil works crusher works iron power dump works station Pb

Cd

Foliar dust Leaf Road dust Soil Foliar dust Leaf Road dust Soil

–b

38.9

36.4

34.8

8.17

14.3

13.5

– 4.87

27.1 20.0

16.2 33.4

20.7 6.41

6.52 4.95

6.06 8.22

35.5 10.5

1.45 –

3.46 1428

2.46 775

1.30 1357

2.55 1084

1.86 863

1.74 810

– 133

2048 1057

807 737

2057 896

1671 832

744 738

2286 1254

26.6

61.0

55.6

60.0

40.7

57.0

78.6

(B) Modified enrichment factors {MEF} ((metal fraction in sample)/(metal fraction in crust)) Metals

Samples

Control

S1

S2

S3

S4

S5

S6

Fe

Foliar dust Leaf Road dust Soil Foliar dust Leaf Road dust Soil Foliar dust Leaf Road dust Soil

0.039 0.013 0.017 0.045 –b – 0.08 0.07 – – 2.27 1.20

0.072 0.039 0.079 0.081 2.82 1.07 1.59 0.28 104 80.7 84.0 4.93

0.072 0.055 0.062 0.067 2.61 0.89 2.06 0.17 55.7 44.6 45.5 3.73

0.071 0.031 0.086 0.091 2.46 0.64 0.55 0.12 96.0 63.8 76.7 5.47

0.071 0.034 0.088 0.085 0.58 0.22 0.44 0.22 76.7 57.1 73.3 3.47

0.073 0.058 0.068 0.091 1.04 0.35 0.56 0.17 62.8 43.3 49.9 5.20

0.073 0.022 0.061 0.071 0.98 0.77 0.64 0.12 58.8 49.3 76.0 5.60

Pb

Cd

(C) Simple enrichment factors {SEF} of metals at different study sites (metal fraction in sample)/(metal fraction is soil) at each site Metal

Sample

Control

S1

S2

S3

S4

S5

S6

Fe

Foliar dust Leaf Road dust Soil Foliar dust Leaf Road dust Soil Foliar dust Leaf Road dust Soil

0.86 0.29 0.38 1.00

0.90 0.49 0.98 1.00 10.1 3.81 5.69 1.00 21.1 16.4 17.0 1.00

1.07 0.82 0.92 1.00 15.8 5.40 12.5 1.00 14.9 12.0 12.2 1.00

0.78 0.34 0.94 1.00 20.7 5.42 4.62 1.00 17.6 11.7 14.0 1.00

0.83 0.40 1.04 1.00 2.67 1.03 2.01 1.00 22.1 16.4 21.2 1.00

0.80 0.64 0.74 1.00 6.14 2.08 3.29 1.00 12.1 8.33 9.60 1.00

1.02 0.30 0.85 1.00 7.90 6.18 5.14 1.00 10.5 8.81 13.6 1.00

Pb

Cd

Where S(E) is the concentration of a target metal (E) in the examined environmental sample, S(R) is the concentration of the reference metal in the examined environmental sample, C(E) is the concentration of a target metal (E) in the crust and C(R) is the concentration of the reference metal in crust. The C(R) value for Pb and Cd were 15 and 0.15 ppm (w/w), respectively. Accordingly, five categories of contamination were arbitrarily established as follows (Sutherland, 2000):

S1

a b

b

– 1.27 1.00 – – 1.89 1.00

Fe reference element relative to crust. Not detected.

moderate enrichment factors (EF: 4.95–38.9) were observed for 3 sample types (foliar dust, leaf, and road dust) at sites 1 to 6 This implies that Pb is also mainly the result of anthropogenic activities. However, it should be noted that these significantly high EF values of Pb and Cd are derived due to unusually low (N10 times low) mean concentration values of Fe obtained in this study (i.e., 0.33% (road dust) − 0.38% (soil)) compared to normal earth crust Fe concentration (5 wt% considered for calculation of conventional EF in this study). Moreover, the mean foliar dust/road dust concentration of Fe (~ 0.33%) in this study was considerably low in comparison to previous studies. For instance, it was nearly five times lower to the mean Fe concentration (1.79%) reported in deposited PM at Bushehr, Iran during dust storm events (Tabatabaei et al., 2015). In soil samples, Pb had low to moderate EF values (EF: 1.30–3.46).

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A modified enrichment factors (MEF) was also used to analyze the results (Kim et al., 2016). It is defined as: MEF ¼ ðmetal fraction in sampleÞ=ðmetal fraction in crustÞ

ð2Þ

The Fe MEF values for all samples at all seven sites ranged from 0.013 (Control leaf) to 0.091 (S3 and S5 soil) thus showing severe Fe deficiency compared to the Earth's crust (Table 4(B)). The present work's road dust Fe MEF values ranged from 0.017 (Control) to 0.088 (S4) are in reasonable agreement with PM data obtained in residential areas in 7 USA counties remote from major pollution sources (Fe MEF: 0.08–0.33), and at an 8th residential area site in the arid state of Arizona, the Fe MEF value was 1.11 (Han et al., 2012). In desert regions, Fe MEF is found to be ~ 1. For Pb, the MEF values (0.07 (Control) – 2.82 (S1)) suggest no to little enrichment. In contrast to Pb, Cd MEF values were significantly higher (1.20 (Control) to 104 (S1)) suggesting low to high enrichment in foliar dust N road dust N leaf. However, considering the low concentrations found in soil and significant correlation between leaf and foliar dust, the high enrichment observed in this study appears to be facilitated by aerial route suggesting airborne transfer of Cd. Similarly, other EFs can be defined depending on application. Since no bedrock samples were analyzed, other EF types may be more appropriate for this study, e.g. a simple enrichment factor. A simple enrichment factor (SEF) can be defined as: SEF ¼ ðmetal fraction in sampleÞ=ðmetal fraction in soilÞ

ð3Þ

at each site and is shown in Table 4(C). The present work's leaf Cd SEFs ranged from 8.33 (S5) to 16.4 (S4) and are comparable to the leaf Cd SEFs (9.8–10.6) of pot grown tobacco plants in soil containing 0.81– 9.2 μg g−1 Cd; tobacco plant leaves may be considered as Cd hyper-accumulator (Liu et al., 2016). In previous studies, significant Cd bioaccumulation has been demonstrated by leaves of herbaceous plant species (i.e., Spinacia oleracea; Solanum nigrum) grown in contaminated/spiked soil (Wei et al., 2005; Pathak et al., 2013). 4. Scanning electron microscopy (SEM) based observations: size distribution and foliar uptake of PM In order to study the accumulation patterns of PM inside the leaf, samples of matured and healthy leaves were taken from control and polluted site (site 1) for SEM analysis. At control site, leaf from adaxial surface showed almost negligible dust deposition on epidermal surface and trichome (Fig. 2a). At adaxial surface of polluted leaf, there were significant aggregate patches of deposited dust on epidermis. It has also been reported that rough surface of leaf facilitates accumulation of PM bound heavy metals (Sawidis et al., 1995; Sawidis et al., 2001; Simon et al., 2014). Due to roughness of wall surface of trichome, leaves of Cassia siamea were probably able to trap and accumulate large amount of PM (Fig. 2b). As shown in Fig. 2c, PM were heavily deposited on epidermal surface to cover the stomata and their openings on adaxial surface of leaf. Begum et al. (2014) reported that the stomata in Cassia siamea generally consist of 3 types, i.e., (1) in epidermal region, it was polygonal type, (2) in anticlinal wall, it was straight, and (3) at abaxial surface, it was anisocytic or paracytic type. The presence of three types of stomata and rough surface of trichomes in Cassia siamea is highly supportive to hold PM. At abaxial surface, PM covered the entire thinner portion of stomata around guard cell and accumulated in lower epidermis and trichome region (Fig. 3a). In this study, fine PM (1 to 8 μm) blocked the stomatal opening with some particles larger than stomatal openings (Fig. 3a and b). As shown in Figs. 2b/c and 3, there were aggregations of fine particles (b1–2 μm). Stomatal opening and guard cells are fully covered by fine PM (b2 μm). It has been reported that fine pores were present inside stomata and leaf cuticles through which the PM bound metal could penetrate inside the leaf tissue (e.g., Betula pubescens); (Kozlov et al., 2000; Eichert et al.,

Fig. 2. Scanning electron micrographs of Cassia siamea showing (a) healthy trichomes and epidermis cells on adaxial surface of leaf at control site. (b) PM trapped on trichomes and epidermis cells on adaxial surface of leaf at polluted site (site 1). (c) Deposition of PM over stomata and epidermal cells of adaxial surface of leaves (site 1). Size ranges of PM are shown with arrows.

2008; Fernandez and Eichert, 2009; Schreck et al., 2012a). On the basis of significant concentration of metals in leaf of Cassia siamea, there is strong evidence that PM was holding large amount of toxic metals, especially Pb and Cd. These metals appeared to be mainly transferred/accumulated from airborne sources. Hence, observations based on SEM also provided a line of evidence that the PM and PM bound toxic metals (Pb and Cd) can be transferred in foliar region of plants in RSPM range (b10 μm) which is significantly important from human health point of view.

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environmental variables. Moreover, the negative impact caused on the plant may also have to be taken in to the account. Acknowledgments The first author is thankful to UGC, New Delhi for Rajiv Gandhi National Fellowship (RGNF). The corresponding author gratefully acknowledges the financial support from a UGC start-up grant, New Delhi, India. (No. F. 20-1/2012(BSR)/20-2(3)/2012(BSR)) and UGCMRP grant (F. No.-43-311/2014 (SR)). The third and fourth authors are thankful to a grant from the Korea Research Foundation (KRF2006-341-C00026) funded by the Korean government (MOEHRD). We also acknowledge the USIC, BBAU Lucknow for the SEM study and the CSIR-Indian Institute of Toxicology Research (IITR), Lucknow for conducting the metal analyses. We gratefully acknowledge the constructive criticism provided by two anonymous reviewers which immensely helped in improving the final version of the manuscript. References

Fig. 3. Scanning electron micrographs of Cassia siamea showing (a) airborne PM deposited on guard cells and inner and outer portion of stomata at abaxial surface (site 1). (b) Destructions of trichomes, stomata, and epidermal cell at abaxial surface of leaf (site 1). (c) Heavy deposition of PM on epidermal layers in abaxial surface of leaf (site 1). (Size ranges of PM are shown with arrows.)

5. Conclusions The results of this study revealed that Cd and Pb were partitioned/ transferred to the leaves of Cassia siamea from airborne sources. The high Cd and moderate Pb EF values which suggests anthropogenic origins (such as traffic and industrial activities). It is observed that the leaves of Cassia siamea may be suitable to accumulate Pb and Cd in polluted environments. The results of our study suggest that Cassia siamea should be a suitable plant species for control of PM (especially in RSPM range) and associated heavy metals (such as Pb and Cd). However, it needs thorough investigation to explicitly explain the significance of such processes in relation to diverse locations, seasonality, and other

Acharya, L., Mukherjee, A.K., Panda, P.C., 2011. Separation of the genera in the subtribe Cassiinae (Leguminosae: Caesalpinioidae) using molecular markers. Acta Bot. Bras. 25, 223–233. Banerjee, A.D.K., 2003. Heavy metal levels and solid phase speciation in street dusts of Delhi, India. Environ. Pollut. 123, 95–105. Barber, J., 2004. Current issues and uncertainties in the measurement and modelling of air–vegetation exchange and within-plant processing of POPs. Environ. Pollut. 128, 99–138. Begum, A., Rahman, M.O., Begum, M., 2014. Stomatal and trichome diversity in Senna mill. from Bangladesh. Bangladesh J. Plant Taxon. 21, 43–51. Boonkerd, T., Pechsri, S., Baum, B.R., 2005. A phenetic study of Cassia s.l. (LeguminosaeCaesalpinioideae: Cassieae: Cassiinae) in Thailand. Plant Syst. Evol. 252, 153–165. Charlesworth, S., Everett, M., McCarthy, R., Ordonez, A., Miguel, E., 2003. A comparative study of heavy metal concentration and distribution in deposited street dusts in a large and a small urban area: Birmingham and Coventry, West Midlands, UK. Environ. Int. 29, 563–573. Deely, J.M., Fergusson, J.E., 1994. Heavy metal and organic matter concentration and distributions in dated sediments of a small estuary adjacent to a small urban area. Sci. Total Environ. 153, 97–111. Eichert, T., Kurtz, A., Steiner, U., Goldbach, H.E., 2008. Size exclusion limits and lateral heterogeneity of the stomatal foliar uptake pathway for aqueous solutes and water suspended nanoparticles. Physiol. Plant. 134, 151–160. Escobedo, F.J., Wagner, J.E., Nowak, D.J., De la Maza, C.L., Rodriguez, M., Crane, D.E., 2008. Analyzing the cost effectiveness of Santiago, Chile's policy of using urban forests to improve air quality. J. Environ. Manag. 86, 148–157. Ewen, C., Anagnostopoulou, M.A., Ward, N.I., 2009. Monitoring of heavy metal levels in roadside dusts of Thessaloniki, Greece in relation to motor vehicle traffic density and flow. Environ. Monit. Assess. 157, 483–498. Faiz, Y., Tufail, M., Tayyeb, J.M., Chaudhry, M.M., Siddique, N., 2009. Road dust pollution of Cd, Cu, Ni, Pb and Zn along Islamabad Expressway, Pakistan. Microchem. J. 92, 186–192. Fernandez, V., Eichert, T., 2009. Uptake of hydrophilic solutes through plant leaves: current state of knowledge and perspectives of foliar fertilization. Crit. Rev. Plant Sci. 28, 36–68. Freer-Smith, P.H., Hollway, S., Goodman, A., 1997. The uptake of particulates by urban woodland: site description and particulate composition. Environ. Pollut. 95, 27–35. Furusjö, E., Sternbeck, J., Cousins, A.P., 2007. PM10 source characterization at urban and highway roadside locations. Sci. Total Environ. 387, 206–219. Han, I., Mihalic, J.N., Ramos-Bonilla, J.P., Rule, A.M., Polyak, L.M., Peng, R.D., Breysse, P.N., 2012. Assessment of heterogeneity of metal composition of fine particulate matter collected from eight US counties using principal component analysis. J. Air Waste Manage. Assoc. 62, 773–782. Honour, S.L., Bell, J.N.B., Ashenden, T.W.A., Cape, J.N., Power, S.A., 2009. Responses of herbaceous plants to urban air pollution: effects on growth, phenology and leaf surface characteristics. Environ. Pollut. 157, 1279–1286. Hornung, H., Karm, M.D., Cohen, Y., 1989. Trace metal distribution on sediments and benthic fauna of Haifa Bay, Israel. Estuar. Coast. Shelf Sci. 29, 43–56. Huu, H.H., Rudy, S., Van Damme, A., 2010. Distribution and contamination status of heavy metals in estuarine sediments near CauOng harbor, Ha Long Bay, Vietnam. Geol. Belg. 13, 37–47. Jain, A., Kutty, C.S., 2014. Biomonitoring of dust pollution of road side of Harda using Air Pollution Tolerance Index (APTI). Int. J. Pure App. Biosci. 2, 233–238. John, A.C., Ibironke, L.,.O., Adedeji, Victor, Oladunni, O., 2011. Equilibrium and kinetic studies of the biosorption of heavy metal (cadmium) on Cassia siamea bark. Am. Eurasian J. Sci. Res. 6, 123–130. Karbassi, A.R., Tajziehchi, S., Afshar, S., 2015. An investigation on heavy metals in soils around oil field area. Glob. J. Environ. Sci. Manag. 4, 275–282. Kim, K.H., Hong, Y.J., Szulejko, J.E., Kang, C.H., Chambers, S., Feng, X., Kim, Y.H., 2016. Airborne iron across major urban centers in South Korea between 1991 and 2012. Sci. Total Environ. 550, 309–320.

130

T. Gajbhiye et al. / Science of the Total Environment 573 (2016) 123–130

Kozlov, M., Haukioja, E., Bakhtiarov, A., Stroganov, D., Zimina, S., 2000. Root versus canopy uptake of heavy metals by birch in an industrially polluted area: contrasting behaviour of nickel and copper. Environ. Pollut. 107, 413–420. Liu, H., Wang, H., Ma, Y., Wang, H., Shi, Y., 2016. Role of transpiration and metabolism in translocation and accumulation of cadmium in tobacco plants (Nicotiana tabacum L.). Chemosphere 144, 1960–1965. Lohr, V.I., Pearson-Mims, C.H., 1996. Particulate matter accumulation on horizontal surfaces in interiors: influence of foliage plants. Atmos. Environ. 30, 2565–2568. Lorenzini, G., Grassi, C., Nali, C., Petiti, A., Loppi, S., Tognotti, L., 2006. Leaves of Pittosporum tobira as indicator of airborne trace element and PM10 distributions in central Italy. Atmos. Environ. 40, 4025–4039. Mediolla, L.L., Domingues, M.C.D., Sandoval, M.R.G., 2008. Environmental assessment of an active tailings pile in the State of Mexico (Central Mexico). Res. J. Environ. Sci. 2, 197–208. Meusel, I., Neinhuis, C., Markstadter, C., Barthlott, W., 1999. Ultra structure, chemical composition, and recrystallization of epicuticular waxes: transversely ridged rodlets. Can. J. Bot. 77, 706–720. Mulgrew, A., Williams, P., 2000. Biomonitoring of Air Quality Using Plants. WHO. Neto, J.A.B., Gingele, F.X., Leipe, T., Brehme, I., 2006. Spatial distribution of heavy metals in surficial sediments from Guanabara Bay: Rio de Janeiro, Brazil. Environ. Geol. 49, 1051–1063. Niencheski, L.F., Windom, H.L., Smith, R., 1994. Distribution of particulate trace metal in Patos Lagoon Estuary (Brazil). Mar. Pollut. Bull. 28, 96–102. Nowak, D.J., Crane, D.E., Stevens, J.C., 2006. Air pollution removal by urban trees and shrubs in the United States. Urban For. Urban Green. 4, 115–123. Pal, A., Kulshreshtha, K., Ahmad, K.J., Behl, H.M., 2002. Do leaf surface characters play a role in plant resistance to auto exhaust pollution. Flora 197, 47–55. Pathak, C., Chopra, A.K., Srivastava, S., 2013. Accumulation of heavy metals in Spinacia oleracea irrigated with paper mill effluent and sewage. Environ. Monit. Assess. 185, 7343–7352. Ravichandran, M., Baskaran, M., Santschi, P.H., Bianchi, T., 1995. History of trace metal pollution in Sabine-Neches Estuary, Beaumont, Texas. Environ. Sci. Technol. 29, 1495–1503. Reddy, L.C.S., Reddy, K.V.R., Humane, S.K., Damodaram, B., 2012. Accumulation of chromium in certain plant species growing on mine dump from Byrapur, Karnataka, India. Res. J. Chem. Sci. 2, 17–20. Rico, C.M., Majumdar, S., Duarte-Gardea, M., Peralta-Videa, J.R., Gardea-Torresdey, J.L., 2011. Interaction of nanoparticles with edible plants and their possible implications in the food chain. J. Agric. Food Chem. 59, 3485–3498. Sæbø, A., Popek, R., Nawrot, B., Hanslin, H.M., Gawronska, H., Gawronski, S.W., 2012. Plant species differences in particulate matter accumulation on leaf surfaces. Sci. Total Environ. 427–428, 347–354. Sardans, J., Penuelas, J., 2005. Trace element accumulation in the moss Hypnum cupressiforme Hedw. and the trees Quercus ilex L. and Pinus halepensis mill. in Catalonia. Chemosphere 60, 1293–1307. Sawidis, T., Marnasidis, A., Zachariadis, G.A., Stratis, J.A., 1995. A study o fair pollution with heavy metals in Thessaloniki city (Greece) using trees as biological indicators. Arch. Environ. Contam. Toxicol. 28, 118–124. Sawidis, T., Chettri, M.K., Papaioannou, A., Zachariadis, G.A., Stratis, J., 2001. A study of metal distribution from lignite fuels using trees as biological monitors. Ecotoxicol. Environ. Saf. 48, 27–35. Sawidis, T., Krystallidis, P., Veros, D., Chettri, M., 2012. A study of air pollution with heavy metals in Athens city and Attica basin using evergreen trees as biological indicators. Biol. Trace Elem. Res. 148, 396–408. Schreck, E., Foucault, Y., Sarret, G., Sobanska, S., Cecillon, L., Castrec-Rouelle, M., Uzu, G., Dumat, C., 2012a. Metal and metalloid foliar uptake by various plant species exposed to atmospheric industrial fallout: mechanisms involved for lead. Sci. Total Environ. 427, 253–262. Schreck, E., Bonnard, R., Laplanche, C., Leveque, T., Foucault, Y., Dumat, C., 2012b. DECA: a new model for assessing the foliar uptake of atmospheric lead by vegetation, using Lactuca sativa as an example. J. Environ. Manag. 112, 233–239.

Sgrigna, G., Sæbø, A., Gawronski, S., Popek, R., Calfapietra, C., 2015. Particulate matter deposition on Quercus ilex leaves in an industrial city of central Italy. Environ. Pollut. 197, 187–194. Shparyk, Y.S., Parpan, V.I., 2004. Heavy metal pollution and forest health in the Ukrainian Carpathians. Environ. Pollut. 130, 55–63. Simon, E., Braun, M., Vidic, A., Bogyó, D., Fábián, I., Tóthmérész, B., 2011. Air pollution assessment based on elemental concentration of leaves tissue and foliage dust along an urbanization gradient in Vienna. Environ. Pollut. 159, 1229–1233. Simon, E., Vidic, A., Braun, M., Bogyó, D., Fábián, I., Tóthmérész, B., 2013. Trace element concentrations in soils along urbanization gradients in the city of Wien, Austria. Environ. Sci. Pollut. Res. 20, 917–924. Simon, E., Baranyai, E., Braun, M., Cserháti, C., Fábián, I., Tóthmérész, B., 2014. Elemental concentrations in deposited dust on leaves along an urbanization gradient. Sci. Total Environ. 490, 514–520. Simon, E., Harangi, S., Baranyai, E., Fabian, I., Tothmeresz, B., 2016. Influence of past industry and urbanization on elemental concentrations in deposited dust and tree leaf tissue. Urban For. Urban Green. 20, 12–19. Singh, V., 2001. Monograph on Indian Subtribe Cassiinae (Caesalpiniaceae). Scientific Editions, Jodhpur, India. Sutherland, R.A., 2000. Bed sediment-associated trace metals in an urban stream, Oahu, Hawaii. Environ. Geol. 39, 611–637. Tabatabaei, T., Karbassi, A.R., Moatar, F., Monavari, S.M., 2015. Geospatial patterns and background levels of heavy metal in deposited particulate matter in Bushehr, Iran. Arab. J. Geosci. 8, 2081–2093. Tomasevic, M., Anicic, M., 2010. Trace element content in urban tree leaves and SEMEDAX characterization of deposited particles. FU Phys. Chem. Technol. 8, 1–13. Tomasevic, M., Rajsic, S., Dordevic, D., Taic, M., Krstic, J., Novakovic, V., 2004. Heavy metals accumulation in tree leaves from urban areas. Environ. Chem. Lett. 2, 151–154. Tomasevic, M., Vukmirovic, Z., Rajsic, S., Tasic, M., Stevanovic, B., 2005. Characterization of trace metal particles deposited on some deciduous tree leaves in an urban area. Chemosphere 6, 753–760. Tripathi, R.D., Vajpayee, P., Singh, N., Rai, U.N., Kumar, A., Ali, M.B., Kumar, B., Yunus, M., 2004. Efficacy of various amendments for amelioration of fly-ash toxicity: growth performance and metal composition of Cassia siamea Lamk. Chemosphere 54, 1581–1588. United states Environmental Protection Agency (US EPA), 1996. (Available at:) http:// www.epa.gov/osw/hazard/testmethods/sw846/pdfs/3050b.pdf. Uzu, G., Sobanska, S., Sarret, G., Munoz, M., Dumat, C., 2010. Foliar lead uptake by lettuce exposed to atmospheric fallouts. Environ. Sci. Technol. 44, 1036–1042. Vaezi, A.R., Karbassi, A.R., Valavi, S., Ganjali, M.R., 2015. Ecological risk assessment of metals contamination in the sediment of the Bamdezh wetland, Iran. Int. J. Environ. Sci. Technol. 12, 951–958. Walkenhorst, A., Hagemeyer, J., Breckle, W.S., 1993. Passive monitoring of airborne pollutant, particularly trace metals, with tree bark. In: Markert, B. (Ed.), Plants as Biomonitors. Indicator of Heavy Metals in the Terrestrial Environment. VCH, Weinheim, pp. 3–27. Wei, S.H., Zhou, Q.X., Wang, X., Zhang, K.S., Guo, G.L., Ma, Q.Y.L., 2005. A newly-discovered Cd-hyperaccumulator Solanum nigrum L. Chin. Sci. Bull. 50, 33–38. Wilson, B., Lang, B., Pyatt, F.B., 2005. The dispersion of heavy metals in the vicinity of Britannia Mine, British Columbia, Canada. Ecotoxicol. Environ. Saf. 60, 269–276. Yaman, M., Güçer, S., 1995. Determination of cadmium and lead in vegetables after activated-carbon enrichment by atomic-absorption spectrometry. Analyst 120, 101–105. Yongming, H., Peixuan, D., Junji, C., Posmentier, E., 2006. Multivariate analysis of heavy metal contamination in urban dusts in Xian, Central China. Sci. Total Environ. 355, 176–186. Zheng, N., Liu, J., Wang, Q., Liang, Z., 2010. Health risk assessment of heavy metal exposure to street dust in the zinc smelting district, Northeast of China. Sci. Total Environ. 408, 726–733. Zibret, G., Sajn, R., 2008. Modelling of atmospheric dispersion of heavy metals in the Celje area, Slovenia. J. Geochem. Explor. 97, 29–41.