Science of the Total Environment 490 (2014) 351–359

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

Concentration profiles and spatial distribution of perfluoroalkyl substances in an industrial center with condensed fluorochemical facilities Guoqiang Shan a, Mingcui Wei a, Lingyan Zhu a,⁎, Zhengtao Liu b, Yahui Zhang b a Key Laboratory of Pollution Processes and Environmental Criteria, Ministry of Education, Tianjin Key Laboratory of Environmental Remediation and Pollution Control, College of Environmental Science and Engineering, Nankai University, Tianjin 300071, PR China b Chinese Research Academy of Environmental Sciences, Beijing 100012, PR China

H I G H L I G H T S • • • • •

PFOA and short-chain PFCAs were predominant in all the environmental samples. The center was a potential source of short-chain PFASs in the surrounding environment. The ∑PFASs in suspended particulate matter were much lower than those in dissolved phase. The sediment-derived logKoc of PFOA was one log unit lower than the SPM-derived one. The spatial distribution of PFASs in tree leaves indicated airborne transport from the facilities.

a r t i c l e

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Article history: Received 24 January 2014 Received in revised form 30 April 2014 Accepted 1 May 2014 Available online xxxx Editor: Eddy Y. Zeng Keywords: PFASs PFOA SPM Tree leaf Bark Fluorine chemistry industry center

a b s t r a c t Jiangsu Hi-tech Fluorochemical Industry Park, China, is one of the largest fluorochemical industry centers in Asia and could be a point source of polyfluoroalkyl substances (PFASs) to the surrounding environment. Besides water, sediment and soil samples, tree leaves and bark were also collected to monitor airborne PFASs around the facilities. Perfluorooctanoic acid and short-chain perfluorocarboxylates including perfluorohexanoic acid and perfluoropentanoic acid were found predominantly in all the samples. The target ∑PFASs were distributed in the dissolved phase with a proportion of 96.5 ± 2.9%. High concentrations of ∑PFASs (up to 12,700 ng/L in surface water) were found at sites near and within the wastewater treatment plant and the facilities. The ∑PFASs in the sediment/sludge were in the range of 3.33–324 ng/g dw. For the first time, tree samples were used for bio-monitoring airborne PFASs in the environment. The ∑PFASs in the tree leaf and bark samples were in the range of 10.0–276 and 6.76–120 ng/g dw, respectively. The spatial distribution of ∑PFASs in the tree leaves suggested that airborne PFASs could be transported from the center to the surrounding environment by prevailing wind. © 2014 Elsevier B.V. All rights reserved.

1. Introduction Perfluoroalkyl carboxylic acids (PFCAs), perfluoroalkyl sulfonates (PFSAs), and other related perfluoroalkyl and polyfluoroalkyl substances (PFASs) have been widely used in commercial products such as surfactants, textile, and aqueous fire-fighting foams, for several decades due to their outstanding thermal and chemical stabilities and hydro- and lipophobic properties (Buck et al., 2011). They are also used as additives or processing acids in industrial processes such as polymer manufacturing (EPA Denmark, 2013; Giesy et al., 2010; Post et al., 2012). They are ubiquitous in the environment and are even ⁎ Corresponding author. Tel.: +86 22 23500791; fax: +86 22 23503722. E-mail address: [email protected] (L. Zhu).

http://dx.doi.org/10.1016/j.scitotenv.2014.05.005 0048-9697/© 2014 Elsevier B.V. All rights reserved.

found in remote regions such as the Arctic as a consequence of widespread usage. It is reported that long-chain PFCAs and PFSAs could be bioaccumulated and biomagnified in food chains and may display adverse effects to animals and humans (EPA Denmark, 2013; Giesy et al., 2010). Thus, continuous concerns over their potential impacts on ecosystem and human health have led to regulatory actions against their production and application. For example, perfluorooctane sulfonate (PFOS) was listed as a new persistent organic pollutant by Stockholm Convention in 2009. Although a major PFAS manufacturer in the United States has phased-out the production of PFOS and related chemicals since 2002, the production has gradually increased in China since then (EPA US, 2012a; Xie et al., 2013; Lim et al., 2011). Perfluorooctanoic acid (PFOA) is another PFAS compound which attracts lots of concerns due to its high production volume and high

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detection frequency in the environment. It has been considered as a potential human carcinogen by the US Environment Protection Agency (EPA) (EPA US, 2012a). Currently, eight major manufacturers in North America agreed with the US EPA to work toward eliminating emission of PFOA and related chemicals by 2015 (Vierke et al., 2012). In China, due to the booming market and abundant resources of fluorite, the fluorochemical industry is one of the fastest growing industries. Several fluorochemical industry centers were founded in the past years, including Jiangsu Hi-tech Fluorochemical Industry Park. It was founded in 1999 and is one of the largest fluorochemical industrial centers in Asia. It has an area of 15.02 km2 and is located in Jiangsu Province by Yangtze River in southeast China. There are more than twenty facilities manufacturing fluorine-based chemicals, including Arkema, DuPont, Daikin and Solvay in the center. The main products of these facilities are fluoropolymers such as polytetrafluoroethylene (PTFE) and polyvinylidene fluoride (PVDF). Production of these fluoropolymers is the main source of PFOA and perfluorononanoic acid (PFNA) in the environment since they are used as processing acids during manufacturing (Dauchy et al., 2012; Dupont, 2012). It was reported that PFASs with shorter fluorinated chain (CF b 7) have less potential for bioaccumulation and lower toxicities (EPA Denmark, 2013). Therefore, along with the regulatory actions against long-chain PFASs, many manufactures have begun to produce and use PFASs with less fluorinated chain, such as perfluorohexanoic acid (PFHxA), as alternatives or additives of perfluorooctyl-based products (EPA US, 2012b; Dupont, 2013). Perfluorobutane sulfonate (PFBS) is produced as alternatives of PFOS and perfluorohexane sulfonate (PFHxS) as well (Zhou et al., 2013). The large scale manufacturing of fluoropolymers in the center may result in the release of PFASs to the surrounding environment and finally threaten the ecological system (Dupont, 2012; Daikin, 2012). The C8 Health Project aimed to evaluate the adverse effects of PFOA on human health in Ohio and West Virginia communities contaminated by fluoropolymer production facility owned by DuPont (Hoffman et al., 2011; Shin et al., 2011). However, little is known about the impacts of the manufacturing activities in Jiangsu Hi-tech Fluorochemical Industry Park on the surrounding environment. Vegetation can be used to indicate atmospheric contamination of volatile organic pollutants and has been used to identify pollution point sources (Simonich and Hites, 1995). It was reported that employing plant samples such as tree bark is an effective tool for monitoring atmospheric polychlorinated biphenyls (PCBs), polycyclic aromatic hydrocarbons (PAHs) and other semi-volatile pollutants around emission sources (Rodriguez et al., 2012; Meredith and Hites, 1987; de Nicola et al., 2013). For PFASs, most of the studies focused on water, soil, sediment and aquatic organisms, while a few of studies using plants as a tool have been performed considering that they are less volatile. To understand the contamination level of PFASs in the surrounding environment of Jiangsu Hi-tech Fluorochemical Industry Park, water and sediment samples were collected from the canals and ponds in or near the facilities. Besides, tree leaf and bark samples near the industrial facilities were also collected to investigate the possibility of using plants as biomonitoring tool for PFASs in atmosphere. Eleven PFASs were quantified in these samples. The results would provide useful information on the ecological risk assessment of PFASs in the surrounding environment due to the production activity of the industry facilities. 2. Materials and methods 2.1. Reagents and materials The target PFAS analytes include the PFCAs with 4–11 perfluorinated carbons, and C4, C6, and C8-PFSAs. Perfluorobutane sulfonate (PFBS, 98%) and perfluorohexanoic acid (PFHxA, 98%) were obtained from Tokyo Chemical Industry Co., Ltd. (Japan). Perfluoropentanoic acid (PFPeA,

97%), perfluorododecanoic acid (PFDA, 96%) and perfluorododecanoic acid (PFDoA, 97%) were obtained from Acros Organics (Geel, Belgium). Perfluoroheptanoic acid (PFHpA, 98%), perfluorooctanoic acid (PFOA, 95%) and perfluorononanoic acid (PFNA, 97%) were purchased from Shanghai Adamas-Beta Reagent Co., Ltd. (Shanghai, China). Perfluorohexane sulfonate (PFHxS, 98%) was from Sigma-Aldrich (St. Louis, MO, USA). Perfluorododecanoic sulfonate (PFOS, 98%) was purchased from Kasei Kogyo Co., Ltd. (Tokyo, Japan). Perfluorodecanoic acid (PFUnDA, 96%) and the internal standards of perfluoro-1-(1,2,3,413 C4) octanesulfonate (13C4-PFOS) and perfluoro-n-(1,2-13C2) octanoic acid (13C2-PFOA) were purchased from Wellington Laboratories (Guelph, Canada). HPLC-grade methanol (MeOH) was purchased from Concord Chemical Company (Tianjin, China). Mill-Q water was used. 2.2. Sample collection and storage All samples were collected around the fluorochemistry industrial facilities in May, 2012 (Fig. 1). Water samples were collected at sites 1–17, 21, 23, 28 and 29 (totally 21 sites). Among those sites, site 8 was at a pond inside the wastewater treatment plant (WWTP) of the industrial center, sites 1–7 were located near the north part of the center and site 7 was at the drainage canal next to the WWTP, sites 9–13 were in Yangtze River near the south part of the center, sites 14–17 were at Wangyu River flowing into Yangtze River and sites 21, 23, 28 and 29 were along the No. 338 provincial road. Sediment samples were collected at sites 1–6 and site 23, while sediment was not available at other sites. One sludge sample was collected at site 8. Soil, leaves and tree bark samples were collected at sites 1, 2, 4, 7, 15–22 and 25–29 (totally 17 sites). All fluorinated materials such as Teflon coated lab wares were avoided during sample collection, preparation and instrumental analysis to minimize contamination of the samples. Water samples were stored in 1.25 L polypropylene (PP) bottles and surface sediment and sludge samples were stored in PP plastic bags. The bottles and bags were pre-cleaned with MeOH and deionized water. Leaf and bark samples were collected from a typical and widespread local species, camphor, by cutting the leaves at 2 m high above the ground and the bark at height about 120–150 cm above the ground using stainless steel scissors, which were pre-cleaned with MeOH and deionized water. All the leaves contained petiole parts. At the same sampling sites, surface soil samples were collected simultaneously. All the samples were stored in pre-cleaned PP plastic bags. Upon back to laboratory, the leaf and bark samples were freeze-dried directly without washing and grounded. All the samples were then passed through a sieve (250 μm mesh size), and stored in plastic bags until extraction. 2.3. Sample pretreatment and extraction One liter of water was filtered through a glass fiber filter (47 mm, pore diameter 0.7 μm, Whatman, pre-combusted at 450 °C for 5 h) to separate suspended particulate matter (SPM) from water. The dissolved phase was collected in a new PP bottle and extracted with solid phase Cleanert PEP cartridges (500 mg, 6 mL, Agela Technologies, Tianjin, China) as described in previous study (Zhao et al., 2012). The MeOH elute was concentrated to 1 mL under a nitrogen stream and transferred into an auto sampler vial. Five nanograms of internal standards of 13 C4-PFOS and 13C2-PFOA was spiked before instrumental analysis. The SPM samples were extracted with 15 mL of MeOH by sonication for 20 min and then centrifuged (3000 rpm) for 10 min. This process was repeated once and the extracts were combined and diluted to 500 mL to pass through the Cleanert PEP cartridges, following the same procedure as described for the dissolved phase samples. Five grams of dry soil or sediment and 10 mL of MeOH were added in a 50 mL PP tube. The mixture was vortexed to ensure complete mixing. Each tube was sonicated for 30 min and centrifuged at 3000 rpm for 5 min. The supernatant was transferred in a new PP tube. The procedure

G. Shan et al. / Science of the Total Environment 490 (2014) 351–359

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A

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Fluorochemical industrial center

28 33

B

8P

rov

inc

ial

27 Ro

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5 25

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Draina

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al

23 19

r

ve

gy

i uR

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Yangtze River

7

9 1011 12 14 13 15 22

16

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17 2.5 km

21

18

Fig. 1. Sampling sites around the fluorochemical industrial center (in orange): (A): Location of the industrial center in Jiangsu Province, China; (B): Sampling sites surrounding the center and in Yangtze River near the center; ☆: The WWTP of the fluorine chemistry industrial center. (For interpretation of the references to color in this figure legend, the reader is referred to the web version of this article.)

was repeated once and the extract was combined. The combined extracts were blown to 5 mL under nitrogen flow. The Cleanert PestiCarb cartridges (500 mg, 6 mL, Agela Technologies) were preconditioned with 5 mL of 0.1% NH4OH in MeOH, 5 mL of water and 5 mL of MeOH. The extracted samples were passed through the cartridges at a rate of 2 mL/min. The cartridges were eluted with 5 mL of MeOH. Both the extract and elute were collected and combined after passing through the cartridge. The collected solution was blown down to 1 mL under a nitrogen stream and transferred into an auto sampler vial. Mass labeled internal standards (5 ng) of 13C4-PFOS and 13C2-PFOA were added before instrumental analysis. Plant samples were extracted by the method described by Yoo et al. with a few minor modifications (Yoo et al., 2011). One gram of dry leaf and bark samples was added in a PP tube. Five milliliters of dichloride methylene was added and the mixture was sonicated for 30 min. Then 5 mL of MeOH was added and the tube was shaken for 60 min and centrifuged for 30 min at 10,000 g. This procedure was repeated once and the supernatants were combined. The combined extracts were dried under a nitrogen stream. Then 2.0 mL of 0.5 M tetrabutylammonium hydrogen sulfate (TBAHS) (adjusted to pH 10) and 4 mL of 0.25 M sodium

carbonate buffer were added and vortexed, 5 mL of methyl tertiary butyl ether (MTBE) was added and the mixture was shaken vigorously for 40 min. The organic layer was separated from the aqueous layer by centrifugation at 3500 rpm for 5 min and then frozen at − 20 °C for 12 h. The MTBE fraction was collected in a new PP tube. The extraction procedure was repeated with another 5 mL of MTBE and the organic layer was combined and evaporated to dry. In order to remove the pigment, the residue was dissolved with 5 mL of methanol. The dissolved solution was cleaned up with Cleanert PestiCarb cartridge and the following cleanup was the same as that described above for soil and sediment samples. 2.4. Instrumental analysis The target PFASs were quantified on an Agilent 1200 liquid chromatography coupled with Agilent 6310 ion trap mass spectrometer in negative electrospray ionization (ESI) mode. Ten microliters of extract was injected on Agilent Eclipse Plus C-18 column (2.1 i.d. × 150 mm). The mobile phase consisted of 2.5 mM ammonium acetate solution and MeOH at a rate of 0.25 mL/min. The gradient was as follows: MeOH

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increased from 10% to 80% at 0.8 min, then to 100% at 12.8 min and ramped to the original condition at 17.8 min, held until 19 min, then followed by equilibrating for 6 min. Selective ion monitoring (SIM) mode was used for quantifying PFASs. The parameters of ion source were as follows: nebulizer 20.0 psi; dry gas 9.0 L/min; dry temperature 350 °C; compound stability 100%; trap dive level 100%. Mass transitions were monitored at m/z: 299 for PFBS, 399 for PFHxS, 499 for PFOS, 219 for PFPeA, 269 for PFHxA, 319 for PFHpA, 369 for PFOA, 419 for PFNA, 469 for PFDA, 519 for PFUnDA, and 569 for PFDoA. 2.5. Quality assurance and statistical analysis Along with each batch of nine samples, one procedure blank (HPLC grade methanol) was run to monitor background contamination. The concentrations of PFOA and PFOS were quantified using internal standards while other PFASs were quantified with external standard method. The limit of detection (LOD) was determined with a signal-to-noise of 3:1, and the limit of quantification (LOQ) was defined with a signalto-noise of 10:1. All the analytical results lower than LODs were reported as bLOQ and zero was assigned for statistical purpose, while those lower than LOQs were reported as bLOQ and the half of LOQ was assigned for statistical purpose. Triplicate matrix spiking experiments were performed by spiking a known amount of PFASs (10 ng/L in water samples, 1 ng/g dw for other samples). The extraction and cleanup of the spiked samples were the same as described in Section 2.3. The recoveries of all target PFASs in the spiked samples were satisfactory within a range of 70–128% for all types of samples, and the relative standard deviation (RSD) for PFOS and PFOA analysis was less than 12%, while it was less than 18% for other PFASs in all type of samples (for details, see Table S1). All PFAS compounds in the blanks were well below the LODs. All other statistical analyses were performed using SPSS for Windows (version 20). 2.6. Calculation of partitioning coefficients The organic carbon normalized partitioning coefficient Koc of PFOA between SPM (or sediment) and dissolved phase was calculated according the following equation: K oc ¼ ðC s =C w Þ  100=f oc

ð1Þ

the concentrations of PFOA in SPM and sediment (ng·g−1 dw). the concentration of PFOA in water. the percentage of organic carbon in SPM or sediment, which was determined using the wet chemistry technique

Cs Cw foc

(Potassium Dichromate Oxidation-Ferrous Sulfate Titrimetry) suggested by the US EPA (Zhao et al., 2012).

3. Results and discussion 3.1. The levels and distribution of PFASs in water (dissolved phase, SPM) The concentrations of individual PFASs in the dissolved phase and SPM are summarized in Table 1. In the dissolved phase, PFPeA, PFHxA, PFHpA and PFOA, were the predominant PFASs with detection frequency of 100%, and they contributed 97% to the ∑ PFASs. They were followed by PFNA and PFUnDA with detection frequency of 95% and 71%, respectively. PFDA, PFDoA and three PFSAs were detected in only 24–52% of the samples. The mean (median) concentration of PFOA in the dissolved phase was 294 ng/L (249 ng/L) within a range of 33.5–1620 ng/L, and the mean (median) concentration of PFHxA was 827 ng/L (26.2 ng/L) within a range of 4.40–10,300 ng/L, respectively. The mean (median) concentration of PFPeA was 29.2 ng/L (27.5 ng/L) within a range of 8.89–62.4 ng/L, and that of PFHpA was 25.7 ng/L (14.5 ng/L) within a range of 4.50–117 ng/L, respectively. The mean (median) concentrations of PFNA, PFDA, PFUnDA, and PFDoA with longer chain length (CF N 7) were in the range of 1.91– 12.5 ng/L (b 0.30–3.41 ng/L), much lower than those of the PFCAs with short chain length (CF ≤ 7). The mean levels of the PFSAs, including PFBS, PFHxS and PFOS, were also relatively low and in the range of 1.24–2.92 ng/L. In the SPM, only PFOA was detected with 100% detection frequency, which was followed by PFOS, PFPeA and PFHpA with detection frequency of 71, 62 and 62%, respectively. The mean (median) concentrations of these four PFASs, including PFOA, PFOS, PFPeA and PFHpA, were 15.9 (3.47), 0.66 (b 0.60), 5.47 (1.95) and 1.25 (b 0.80) ng/L, respectively. Other PFASs were detected in 29–52% of the SPM samples. The mean level of PFHxA was 10.0 ng/L and other PFASs were in the range of 0.41–3.80 ng/L. The ∑ PFASs in the SPM at each site were in the range of 0.35–343 ng/L and approximately one to two orders of magnitude lower than those in the dissolved phase (range 73.8–12,400 ng/L) and 96.5 ± 2.9% of the ∑ PFASs in water were distributed in the dissolved phase (Fig. 2). This is in agreement with the investigation in the basin of Tokyo Bay, Japan (Zushi et al., 2012), where around 96.5% of the ∑ PFASs was partitioned in the dissolved phase. It was also reported that 93% of the ∑ PFASs were in the dissolved fraction in the river Elbe and North Sea, Germany (Ahrens et al., 2009). The results are significantly different from those of typical hydrophobic pollutants such as polybrominated diphenyl ethers. It was reported that up to 70% of the total tri-, tetra- and penta-brominated diphenyl ether congeners were partitioned in the SPM (Wurl et al., 2006).

Table 1 Concentrations (ng/L) of eleven PFASs in water (dissolved phase and SPM) samples.a Dissolved phase

PFPeA PFHxA PFHpA PFOA PFNA PFDA PFUnDA PFDoA PFBS PFHxS PFOS ∑PFASs a

SPM

Mean

Median

Range

DF (%)

Mean

Median

Range

DF (%)

29.2 827 25.7 294 12.5 11.5 3.38 1.91 2.00 1.24 2.92 1210

27.5 26.2 14.5 249 3.41 b0.40 2.43 b0.40 1.28 b0.20 b0.20 404

8.89–62.3 4.40–10,300 4.50–117 33.5–1620 b0.50–132 b0.40–116 b0.30–12.1 b0.40–28.0 b0.50–6.21 b0.20–20.8 b0.20–18.9 73.8–12,400

100 100 100 100 95 48 71 33 52 24 43

5.47 10.0 1.25 15.9 1.72 1.53 1.63 1.48 3.80 0.41 0.66 43.8

1.95 b0.20 b0.80 3.47 b0.80 b0.50 b0.20 b0.10 b0.60 b0.10 b0.60 12.2

b0.60–57.4 b0.20–15.4 b0.40–5.14 b0.70–197 b0.50–13.4 b0.50–25.2 b0.20–9.80 b0.10–25.5 b0.60–29.7 b0.10–4.95 b0.20–3.69 0.35–343

62 43 62 100 52 29 48 33 48 33 71

DF = detection frequency.

G. Shan et al. / Science of the Total Environment 490 (2014) 351–359

Concentration of total PFASs (ng/L)

13000

SPM dissolved phase 12000

4000

2000

0 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 21 23 28 29

Sampling sites Fig. 2. The spatial distribution of ∑PFASs in dissolved phase and SPM at the 21 sample sites.

The results imply that PFASs are more subjected for transport via aqueous phase than the hydrophobic pollutants. The dissolved phase and SPM samples displayed different PFAS profiles due to the different partition behaviors of individual PFASs. Sites 5, 7, 8 and 9 were in the outlets of facilities or in/near the WWTP and unexpected high level of PFHxA (911–10,300 ng/L) was detected in the dissolved phase at these sites. The reason for this is unclear. One possible reason is that PFHxA is being produced or used in large amount as an alternative of PFOA by some facilities in the center. Due to the unexpected high level of PFHxA in these four samples, they were not included in the calculation of PFAS profiles in dissolved phase and SPM. As shown in Fig. 3, the four PFCAs with short fluorinated chain ≤7 (including PFPeA, PFHxA, PFHpA and PFOA), the four PFCAs with long fluorinated chain N 7 (including PFNA, PFDA, PFUnDA and PFDoA), and the three PFSAs (including PFBS, PFHxS and PFOS) contributed 95.4, 3.3, and 1.3% to the ∑PFASs in the dissolved phase, respectively, and their contributions were 54.1, 33.2, and 12.7% in the SPM, respectively. The major PFASs in the dissolved phase were PFOA and PFHxA, with a mean contribution of 55.6 and 27.0% to the ∑PFASs, respectively. However, their contribution in the SPM decreased to 25.3 and 5.9%, respectively, due to increased contribution of the PFCAs with longer chain length (CF N 7) and PFSAs in SPM. The summed contribution of PFNA, PFDA, PFUnDA and PFDoA in the SPM was 33.2%, distinctively higher than that in the dissolved phase (only 3.3%). The contribution of the three PFSAs (∑PFSAs) was also significantly enhanced in the SPM (12.7%) as compared to the dissolved phase (1.3%). These suggest that the PFCAs with longer chain length (CF N 7) and the PFSAs are more inclined to partition to SPM, which could be attributed to their higher sorption affinity to the

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particles due to their stronger hydrophobicity (Higgins and Luthy, 2006). The spatial distribution of ∑PFASs in the dissolved phase and SPM of the 21 sites is shown in Fig. 2. The highest level of ∑PFASs in the dissolved phase was detected at site 8 (12,400 ng/L), the pond inside the WWTP of the center, which was followed by site 9 (4040 ng/L), which is at a discharge outlet of the center, and then site 7 (1730 ng/L), which is at the downstream next to the WWTP. The result implies that direct discharge from the facilities and the effluent from the WWTP could be potential contamination sources for the surrounding surface water. Conventional WWTPs might be ineffective to remove PFASs (Xiao et al., 2012a; Weinberg et al., 2011; Shivakoti et al., 2010). It was reported that the concentrations of some PFASs actually increased in the WWTP effluent as compared to those in the influent due to the biotransformation of PFAS precursors in the process of WWTPs (Xiao et al., 2012a). The ∑PFASs in the SPM followed a similar spatial trend in the dissolved phase, and the highest concentration of ∑PFASs in the SPM was also found at site 8. Compared to other districts with PFAS-related manufacturing activities in China, the PFOA contamination in water near Jiangsu Hi-tech Industrial Park was at a moderate level and that of PFOS was at a very low level. The mean PFOA and PFOS levels in water (241 and 3.75 ng/L, the summed concentrations in the dissolved and SPM phases, not including site 8 at the WWTP pond) were much higher than those (mean 10.0, 14.1 ng/L, respectively) in the ponds near a PFAS-manufacturing factory in Wuhan, China. However, the levels of PFOA and PFOS in the effluent of a WWTP near the factory in Wuhan, China, were 153 and 1021 μg/L, respectively (Wang et al., 2010), much higher than the levels of PFOA (1698 ng/L) and PFOS (0.25 ng/L) at site 8, the pond inside the WWTP in the current study. The mean levels of PFOA, PFHxA and PFOS in the dissolved phase in the present study were 228, 352 and 3.05 ng/L, respectively, higher than those in a river near the fluorochemical facilities in Fuxin, China, which were 169, 0.27 and 0.36 ng/L, respectively (Bao et al., 2011). The differences may be explained by the fact that different fluorochemicals are produced in different industrial centers. PTFE, PVDF and other fluoropolymers are manufactured by the factories in Jiangsu Hi-tech Park and PFCAs such as PFOA may be used as a processing acid to facilitate the solubilization of the fluorotelomers and their aqueous polymerization (Dauchy et al., 2012), which could lead to the release of PFCAs to the surrounding environment. Even though PFOA was predominant in the water samples, its level was distinctly lower than that in the outfalls at DuPont Washington facility during 2000–2003, which reached 239–915 μg/L (Ritchey, 2006). It was unexpected that high level of PFHxA (10,300 ng/L) was detected in the wastewater of the WWTP pond (site 8), which was followed by site 9 (3410 ng/L) and site 5 (1180 ng/L). It was much higher than 1136 ng/L in the water samples obtained in a fluoropolymer manufacturing plant in France and 761–1767 ng/L in WWTPs in PFOS PFHxS PFBS PFDoA PFUnDA PFDA PFNA PFOA PFHpA PFHxA PFPeA

SPM

dissovled phase

0

20

40

60

80

100

% of total PFASs Fig. 3. Contribution of individual PFASs to the ∑PFASs in SPM and dissolved phase at the 17 sites. Sites 5, 7, 8 and 9 are not included because unexpected high levels of PFHxA (911–10,300 ng/L) were detected in the dissolved phase at these sites, which are in the outlet of the facilities or in/near the WWTP.

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Thailand (Dauchy et al., 2012; Shivakoti et al., 2010). This implies that the PFCAs with shorter chain (CF b 7) such as PFHxA or their precursors might have been used as alternatives of PFOA (Xiao et al., 2012a; Cui et al., 2013).

of PFAS-based products from wind or rain runoff within the facilities (Xiao et al., 2012b, 2013). On the other hand, freshwater inputs of PFASs are transferred from water to sediment and the sediment could be under-equilibrium with the water, leading to the relatively lower level of PFASs in the sediment (Sakurai et al., 2010).

3.2. The levels of PFASs in sediment and sludge 3.3. PFASs in soil, leaf and bark samples The levels of PFASs in the seven sediment samples and one sludge sample in the WWTP are shown in Fig. 4. PFOA and PFPeA predominated in the sediment and sludge samples with 100% detection frequency, and their concentrations were in the range of b0.60–78.0 ng/g and b0.80–51.7 ng/g dw, respectively. PFDoA, PFHxA, PFHpA and PFOS were also detected with high frequency at 87–100%. However, they were lower than LOQ in more than half of the sediment samples. The ∑ PFASs in the sediment or sludge were in the range of 3.33–324 ng/g dw, significantly lower than that in the SPM, which was 322–21,500 ng/g dw. It is not surprising that the ∑PFASs in the sludge of WWTP (site 8) were much higher than those in the river sediment at other sites. The predominant PFASs at site 8 were also PFOA, PFPeA and PFHxA, with concentrations of 78.0, 51.7 and 49.7 ng/g dw, respectively, and the four PFCAs (CF N 7) were in the range of 24.1–43.2 ng/g dw. The three PFSAs displayed a low level in the range of b 0.50–2.34 ng/g dw. Previous studies demonstrated that the distribution of PFASs between the dissolved phase and sediment (or SPM) is influenced by organic carbon fraction (Ahrens et al., 2010; Zhao et al., 2012). In the present study, the particular organic carbon (POC) content (foc) of the SPM (4.79–46.6%) was much higher than that of the sediment (0.80–7.76%). The log(Koc/[cm3/g]) values of PFOA in the sediment and SPM were calculated to be 2.5 ± 0.2, 4.0 ± 0.3, respectively. The partitioning coefficients for other PFASs were not calculated due to their low detection frequency in dissolved phase, SPM or sediment samples. The sediment-derived logKoc was much lower than the SPMderived one, which is in agreement with those reported in other studies (Ahrens et al., 2010; Zhang et al., 2012), suggesting that SPM has higher sorption capacity to PFASs than sediment. Usually, SPM has smaller particle size than sediment. Previous study demonstrated that the fractionated sediment with smaller particle size displayed stronger affinity to PFASs than the fractions with large size (Zhao et al., 2012). Thus, the large surface area associated with fine particle size of SPM might lead to the high sorption logKoc for PFASs in SPM than sediment. The types and compositions of the organic carbon or POC might also make contribution to PFAS sorption difference between SPM and sediment (Zhao et al., 2012; Xiao, 2013; Zareitalabad et al., 2013). The large sorption in SPM could also influenced by the mass flux of SPM containing the debris

Concentration of total PFASs (ng/g dw)

105 SPM sediment or sludge

104

103

102

101

100 1

2

3

4

5

6

8

23

Sampling sites Fig. 4. The spatial distribution of ∑PFASs in the SPM and sediment (or sludge) samples at the 8 sampling sites.

The concentrations of PFASs in the soil, leaf and bark samples are summarized in Table 2 and shown in Fig. 5. As similar to the compositions of PFASs in dissolved phase and SPM, the predominant PFASs in soil, leaves and bark were also PFOA, PFPeA and PFHxA with detection frequency of 65–100%. PFOA predominated in soil, leaf and bark samples, and contributed 31, 50 and 44% to the ∑ PFASs, respectively. It was followed by PFPeA with mean contribution of 19, 12 and 16% and PFHxA with a mean contribution of 4, 16 and 15%, respectively. PFNA and PFUnDA were detected in 64–94% of the leaf and bark samples, but less detected in the soil samples (29 and 59% respectively) (Fig. 5B). The ∑ PFASs in the soil, leaf and bark samples were in the range of 0.75–28.8, 10.0–276 and 6.76–120 ng/g dw, respectively. The levels of ∑ PFASs in the leaf (mean 85.2 ng/g dw) were higher than those in the bark (mean 37.9 ng/g dw), and much higher than that in the soil samples (mean 7.17 ng/g dw). As the predominant PFAS, the mean concentration of PFOA in the soil (mean 2.24 ng/g dw, range b0.20–9.03 ng/g dw), was significantly lower than that in the leaf (mean 42.5 ng/g dw, range b0.05–194 ng/g dw) and bark samples (mean 16.5 ng/g dw, range b 0.07–61.3 ng/g dw) (Table 2). It is well known that plants take up organic pollutants via root uptake from soil followed by transportation along with the transpiration stream or by adsorption and absorption after dry or wet deposition in gas or particulate-bound phase (Collins et al., 2006; Zhao et al., 2014). Suggesting that trees accumulate PFASs exclusively from soil, a biotasoil-accumulation-factor (BSAF) could be estimated by the equation: BSAF = Cleaf / Csoil, where Cleaf and Csoil are the PFAS concentrations in tree leaves and soil, respectively. The mean BSAF of PFOA in the present study was calculated to be 13. However, Zhao et al. investigated the accumulation of PFASs in wheat from soil and reported that PFASs could be accumulated in the root of wheat, but it was very difficult for them to translocate to the shoot, and the BSAF of PFOA based on the whole wheat (including root and shoot) was only 0.06–0.16 (Zhao et al., 2014). Yoo et al. determined the above-ground grass/soil accumulation factors of PFOA and other perfluorinated chemicals, and reported that the mean accumulation factor was 0.25 for PFOA (Yoo et al., 2011). The wheat straw/soil accumulation factor of PFOA was 3.99 based on the data given by Yoo et al. (2011), Stahl et al. (2008) and Lechner and Knapp (2011). In addition, prior studies demonstrated that compounds with logKow N 3.5 have low translocation capacity after absorbed from soil (de Kok and Hawkesford, 2011). The estimated logKow values of PFASs excluding PFPeA in the present study are in the range of 3.40–7.77 (Wang et al., 2011). Thus, it is speculated that it is hard for these PFASs to be transported from root to leaves and bark. On the other hand, there was evidence that PFOA was directly released to the atmosphere at fluoropolymer manufacturing facilities (Shin et al., 2011; Barton et al., 2006; Niisoe et al., 2010; Davis et al., 2007). Due to the low volatility, airborne PFASs are mostly bounded with particulate matters (Barton et al., 2006; Davis et al., 2007). The PFASs in the surrounding environmental media could be a result of the deposition of particulate-phase PFASs (Davis et al., 2007). Thus, it is reasonable to deduce that the high level of PFASs in the tree leaves or bark was mainly from airborne PFASs and their precursors rather than from soil. In general, the ∑PFASs in bark were lower than those in the leaves (Fig. 5A). This may result from the differences in structure characteristics between them. The waxy cuticle of the leaves could effectively trap the particulate-phase PFASs, which is followed by absorption and transfer through the stomata at the surface. Bark does not have the waxy layer and usually has solid and condense suberin surface,

G. Shan et al. / Science of the Total Environment 490 (2014) 351–359

357

Table 2 Concentrations (ng/g dw) of eleven PFASs in soil, tree leaf and bark samples.a Soil

PFPeA PFHxA PFHpA PFOA PFNA PFDA PFUnDA PFDoA PFBS PFHxS PFOS ∑PFASs a

Leaves

Bark

Mean

Range

DFa

Mean

Range

DFa

Mean

Range

DFa

1.40 0.28 0.59 2.24 0.46 0.46 0.48 0.21 0.21 0.80 0.06 7.17

b0.50–4.09 b0.40–1.42 b0.70–5.86 b0.20–9.03 b0.20–4.60 b0.70–4.06 b0.15–3.90 b0.50–2.34 b0.50–3.49 b0.20–13.7 b0.10–0.34 0.75–28.8

65 59 71 100 29 29 59 59 6 6 29

9.87 13.7 2.97 42.5 4.55 3.76 4.28 2.66 0.42 0.23 0.22 85.2

b0.12–40.9 b0.10–59.8 b0.09–11.5 b0.05–194 b0.10–7.78 b0.09–12.3 b0.08–8.61 b0.10–12.8 b0.20–2.43 b0.10–1.65 b0.08–1.12 10.0–276

88 82 59 88 94 53 65 35 24 29 35

6.19 5.83 1.01 16.5 3.54 0.21 3.52 0.93 0.07 0.06 0.04 37.9

b0.12–23.6 b0.10–28.6 b0.10–6.66 b0.07–61.3 b0.10–14.1 b0.10–3.51 b0.10–9.12 b0.06–5.04 b0.20–0.86 b0.10–0.81 b0.07–0.41 6.76–120

88 94 41 88 76 6 65 24 12 12 18

DF = detection frequency.

which may prevent the pollutants from entering the inner part (Meredith and Hites, 1987). The highest level of ∑PFASs (276 ng/g dw) in the tree leaves was found at site 7 in front of the WWTP (Fig. 5A). PFASs in the WWTP wastewater might be released to atmosphere through volatilization or forming aerosols at the air–water interface, leading to increased PFASs in the air near WWTP (Ahrens et al., 2011). High level of ∑PFASs was also observed in the tree leaves at site 2 (143 ng/g dw) and site 4 (101 ng/g dw), which are located in the center. The results suggest that the WWTP and the industrial facilities are potential sources for airborne PFASs and their precursors in the surrounding environment.

A soil leaf bark

Concentration of total PFASs (ng/g dw)

250

125 100 75 50 25

It was reported that airborne PFASs or their precursors could be carried by prevailing wind to the downwind districts (Barton et al., 2006). In the studied district, the main wind direction in June, 2011–May, 2012 was in northeast and southeast. As for the sampling sites outside the center, sites 16, 19, 20, 25 and 26, located in the southwest within 3 km from the facilities, displayed higher level of ∑PFASs (147, 93.4, 129, 129, and 99.3 ng/g dw, respectively) than other sites, which were from the facilities with long distance N 3 km, such as sites 17 (51.1 ng/g dw) and 18 (63.1 ng/g dw), or were not in the main wind direction, such as sites 15, 21 and 22 (31.7, 46.7, and 10.0 ng/g dw, respectively). Sites 27, 28 and 29, which are in the northwest but N 3 km away from the facilities, also showed low level of ∑ PFASs (10.5, 27.9, and 40.7 ng/g dw, respectively) (Fig. S1). Fig. 6 illustrates a simple relationship between the PFAS concentrations in the tree leaves and the distance from the facilities (site O was the central point of sites 2, 4 and 7, which displayed the highest level of ∑PFASs in the tree leaves). Since sites 1, 15, 21 and 22 are not in the main wind direction, they were not included in the figure. The PFAS concentrations in the tree leaves decreased as the distance from site O increased. This suggests that the spatial distribution of PFASs in the leaf samples was distinctly influenced by the prevailing wind flow. David et al. surveyed the air dispersion of PFOA in West Virginia and Ohio (Davis et al., 2007). They reported that airborne PFOA was mostly bounded with particulate matters due to its low volatility, and the particulate-phase PFOA was transported by airflow and then deposited to the surrounding environmental media within several kilometers from the facilities (Davis et al., 2007), leading to higher level of PFOA in the water samples near the facilities.

0 1 2 4 7 15 16 17 18 19 20 21 22 25 26 27 28 29

Sampling sites

100

PFOS PFHxS PFBS PFDoA PFUnDA PFDA PFNA PFOA PFHpA PFHxA PFPeA

% of total PFASs

80

60

40

Concentration of total PFASs (ng/g dw)

B

O 160

16 20

25 120

26

19

80

18 17

29

40

28 27

0

20

0

2.5

5.0

7.5

10

Distance (km) soil

leaf

bark

Fig. 5. The spatial distribution of ∑PFASs in the soil, leaf and bark samples (A); and the mean contribution of each PFAS to the ∑PFASs in the soil, leaf and bark samples (B).

Fig. 6. The relation between the concentration of ∑PFASs in leaf samples and the distance from the center. The level of ∑PFASs at site O, the central point of sites 2, 4 and 7 in the center, is the mean concentration of the PFASs at the three sites.

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4. Conclusions Although PFAS contamination widely existed around Jiangsu fluorochemical center, their contamination was moderate compared to similar areas around the world. PFOA and other PFCAs with shorter chain length such as PFHxA, were the main PFASs. The ∑PFASs in the SPM was one to two orders of magnitude lower than those in the dissolved phase, suggesting that PFASs are more subjected for transport via aqueous phase. The sediment-derived logKoc of PFOA was one log unit lower than the SPM-derived, indicating the difference in partitioning behavior of PFASs between SPM- and sediment-dissolved phase systems. The spatial distribution of ∑PFASs in the water samples indicated that discharge of effluent from the WWTP and direct discharge from the facilities could be potential sources for the surrounding water body. Local tree leaves were used to monitor the airborne PFASs and the ∑ PFASs in the tree leaves decreased significantly as the distance from the facilities increased, suggesting that PFASs could be carried by prevailing wind to downwind district. More monitoring research is necessary in the future to assess the ecological risk of PFASs in this district. Declaration We declared that there is no any actual or potential conflict of interest including any financial, personal or other relationships with other people or organizations within three years of beginning the submitted work that could inappropriately influence, or be perceived to influence in this work. Acknowledgments The authors gratefully acknowledge the financial support of the Chinese National Natural Science Foundation Grants (21325730, 21277077), the Ministry of Education (20130031130005), the Ministry of Environmental Protection (201009026), the Ministry of Science and Technology (2012ZX07529-003), and the Ministry of Education innovation team IRT 13024. Appendix A. Supplementary data Supplementary data to this article can be found online at http://dx. doi.org/10.1016/j.scitotenv.2014.05.005. References Ahrens L, Plassmann M, Xie Z, Ebinghaus R. Determination of polyfluoroalkyl compounds in water and suspended particulate matter in the river Elbe and North Sea, Germany. Front Environ Sci Eng China 2009;3:152–70. Ahrens L, Taniyasu S, Yeung LW, Yamashita N, Lam PK, Ebinghaus R. Distribution of polyfluoroalkyl compounds in water, suspended particulate matter and sediment from Tokyo Bay, Japan. Chemosphere 2010;79:266–72. Ahrens L, Shoeib M, Harner T, Lee SC, Guo R, Reiner EJ. Wastewater treatment plant and landfills as sources of polyfluoroalkyl compounds to the atmosphere. Environ Sci Technol 2011;45:8098–105. Bao J, Liu W, Liu L, Jin Y, Dai J, Ran X, et al. Perfluorinated compounds in the environment and the blood of residents living near fluorochemical plants in Fuxin, China. Environ Sci Technol 2011;45:8075–80. Barton CA, Butler LE, Zarzecki CJ, Flaherty J, Kaiser M. Characterizing perfluorooctanoate in ambient air near the fence line of a manufacturing facility: comparing modeled and monitored values. J Air Waste Manag Assoc 2006;56:48–55. Buck RC, Franklin J, Berger U, Conder JM, Cousins IT, de Voogt P, et al. Perfluoroalkyl and polyfluoroalkyl substances in the environment: terminology, classification, and origins. Integr Environ Assess Manag 2011;7:513–41. Collins C, Fryer M, Grosso A. Plant uptake of non ionic organic chemicals. Environ Sci Technol 2006;40:45–52. Cui R, Zhang Y, Wang J, Dai J. Levels and composition distribution of perfluoroalkyl substances in water and biological samples from Jiangsu Hi-tech Fluorochemical Industry Park in Changshu, China. Environ Chem 2013;32:1318–27. [in Chinese]. Daikin. Progress in removing PFOA in fluorochemical products. http://www.daikin.com/ press/2012/121221/index.htm, 2012. [accessed Dec 12, 2013]. Dauchy X, Boiteux V, Rosin C, Munoz JF. Relationship between industrial discharges and contamination of raw water resources by perfluorinated compounds. Part I: case

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