MPB-07240; No of Pages 7 Marine Pollution Bulletin xxx (2015) xxx–xxx
Contents lists available at ScienceDirect
Marine Pollution Bulletin journal homepage: www.elsevier.com/locate/marpolbul
Baseline
Levels and spatial distribution of perfluoroalkyl substances in China Liaodong Bay basin with concentrated fluorine industry parks☆ Hong Chen a,b, Can Zhang a, Jianbo Han a,⁎, Ruijun Sun a, Xiangyun Kong a, Xiaomeng Wang a, Xin He b a Key Laboratory of Coastal Ecology and Environment of State Oceanic Administration, Department of Marine Chemistry, National Marine Environmental Monitoring Center, Linghe Street 42, Dalian 116023, China b Key Laboratory of Industrial Ecology and Environmental Engineering (Ministry of Education), School of Environmental and Biological Science and Technology, Dalian University of Technology, Dalian 116024, China
a r t i c l e
i n f o
Article history: Received 19 August 2015 Received in revised form 9 October 2015 Accepted 11 October 2015 Available online xxxx Keywords: PFASs Level Spatial distribution Mass loading
a b s t r a c t Eighteen different perfluoroalkyl substances (PFASs) were investigated in 35 river water samples and 34 sediment samples collected from rivers in the Liaodong Bay basin containing two fluorine industry parks. Perfluorooctanoate (PFOA) and perfluorooctane sulfonate (PFOS) were the predominant PFASs in freshwater, with median concentrations of 26.5 ng/L and 1.87 ng/L, respectively. However, perfluorobutane sulfonate (PFBS) had the highest maximum concentration (up to 124.1 ng/L, approximately two orders of magnitude higher) in water at a site which is the nearest to the industrial source of PFASs. Total PFASs in water at this site were also the highest. In contrast, PFOA and perfluorooctadecanoate (PFOcDA) were the most abundant PFASs in sediment, with median concentrations of 1.19 ng/g and 0.35 ng/g, respectively. Total PFAS concentrations in sediment from the site near to the industrial park were significantly higher than the other rivers. Mass loading of total PFASs from the rivers flowing into Liaodong Bay was estimated to be 506 kg/year. © 2015 Elsevier Ltd. All rights reserved.
Perfluoroalkyl substances (PFASs) are a large and important class of synthetic chemicals. These PFASs have attractive properties, such as surfactant activity, resistance to acid and high temperatures, and water and oil repellency. They make them useful in commercial products including nonstick coatings, stain-repellant fabrics, paper packaging products and firefighting foams. Consequently, this class of compounds has been produced and used for over 60 years (Lindstrom et al., 2011). This has led to that PFASs are now ubiquitous in various environmental matrices, including river and ocean water, soils, sediments, sludge, biota, and even the polar ice caps (Chen et al., 2012; Lau, 2012; Post et al., 2012; Loi et al., 2013). The structure of PFASs consists of a totally fluorinated carbon chain of varying length and a charged functional group, such as carboxylic or sulfonic acid (Lindstrom et al., 2011). Because of the extreme stability of their carbon–fluorine bonds, PFASs are highly persistent in the environment (Lindstrom et al., 2011). Some of them have been shown to be bioaccumulative and toxic (Wang et al., 2010), including hepatic, developmental, immune and endocrine toxicities in experimental
☆ Capsule: Levels and spatial distribution of 18 PFASs in Liaodong Bay basin containing both of the domestic and industrial sources of PFASs were shown. ⁎ Corresponding author. E-mail addresses: [email protected] (H. Chen), [email protected] (C. Zhang), [email protected], [email protected] (J. Han), [email protected] (R. Sun), [email protected] (X. Kong), [email protected] (X. Wang), [email protected] (X. He).
animals. As compared to the typical POPs, such as PAHs and PCBs, PFASs are highly water-soluble (Post et al., 2012). Perfluorooctanoic acid (PFOA) and perfluorooctane sulfonic acid (PFOS) as the two most-well studied PFASs were previously made and used in the largest amounts. The production of PFOS and its synthetic starting material, perfluorooctyl sulfonyl fluoride (PFOSF), has ceased substantially since their major manufacturer 3M voluntarily phasedout their production in 2002 (U.S. EPA, 2000). Currently the use of PFOS and PFOSF is restricted to certain processes, those who have no viable alternatives under the Stockholm Convention because they were designated as persistent organic pollutants (POPs) by the Convention on POPs in 2009. In 2006, eight major manufacturers participated in a PFOA stewardship program which is an agreement between the United States Environmental Protection Agency and the producers (U.S. EPA, 2006). The goals of this program were to have a complete elimination of PFOA together with precursor chemicals and higher homologue chemicals in emission and products by 2015. However, the environmental concentrations of PFOS and PFOA are not expected to decline. Available information (Jin et al., 2007; Harada et al., 2010) indicated that PFOS in human serum did not significantly decrease in Busan and Seoul (Korea) and in Shenyang (China); PFOA elimination was relatively slow, and in some areas (e.g., in Korea) levels of PFOA and some higher homologue chemicals (e.g., PFNA and PFDA) did not decline but showed an increase (Haug et al., 2009; Kato et al., 2011; Yeung et al., 2013). The above observed continuation of PFAS pollution after the application of PFAS regulations may be explained by the proposed several
http://dx.doi.org/10.1016/j.marpolbul.2015.10.024 0025-326X/© 2015 Elsevier Ltd. All rights reserved.
Please cite this article as: Chen, H., et al., Levels and spatial distribution of perfluoroalkyl substances in China Liaodong Bay basin with concentrated fluorine industry parks, Marine Pollution Bulletin (2015), http://dx.doi.org/10.1016/j.marpolbul.2015.10.024
2
H. Chen et al. / Marine Pollution Bulletin xxx (2015) xxx–xxx
hypotheses including the consumption of stockpiles of PFASs, emission from residuals in commercial products and degradation of the precursors of PFOS and PFOA (Wang et al., 2009). Legislation in some countries such as those in Europe and North America on restriction or elimination of PFOS-contained product resulted in a shift of production to other countries such as China (Wang et al., 2009; UNEP, 2009). Since 2004, Fuxin which is situated in northwest Liaoning Province as shown in Fig. 1 has become a production base of the fluorochemical industry in China. It has been developed into two large fluorochemical industrial parks containing several fluorochemical plants. Currently these industrial parks focused on the production of the potassium salt of perfluorobutane sulfonate (PFBS), ammonium perfluorooctanoate (APFO, the derivatives of PFOA) and fluoropolymers, such as polytetrafluoroethylene (PTFE). The industrial effluent from the parks will be discharged into Daling River and then finally to Liaodong Bay. It may be a significant industrial source of PFAS emissions into the environment. In addition, Liaodong Bay basin in Liaoning Province contains a population of 43 million. The domestic sewage would be another source of PFASs in the environment. Therefore, Liaodong Bay basin in Liaoning Province is a PFASs-polluted typical region in China. Previous reports (Wang et al., 2011a, 2011b, 2012) also indicated that the greatest concentrations of PFASs, such as PFOS and PFOA, were observed in Liaodong Bay basin in Liaoning region. However, PFAS concentrations in this region and their flux into the ocean are still limited. The objectives of this study were to investigate the current extent and the spatial distribution of PFAS contamination in the environment around Liaodong Bay basin in Liaoning Province, and to evaluate the possible sources and flux of PFASs into the ocean. To accomplish these goals, a total of 17 perfluoroalkyl acids and 2 precursors were analyzed in surface water and sediment samples collected from strategic locations in Liaodong Bay basin. These data would be helpful for the evaluation of potential impacts from industries operating in this region and the development of appropriate controls and management solutions for PFAS pollution in this area. Water and sediment samples were collected from Liaodong Bay basin. It contains four major rivers and eight minor rivers flowing into the bay. Waters from these twelve rivers make up most of the total river water discharge into Liaodong Bay. The total area of this basin is approximately 269,570 km2. Because this area is too large, we selected
Liaodong Bay basin in Liaoning Province as the sampled area in this study (the downstream region in this basin, Fig. 1). The study area mainly consists of nine cities including Yingkou, Anshan, Liaoyang, Shenyang, Panjin, Fuxin, Jinzhou, Chaoyang and Huludao. It contains a population of 44 million. One liter samples of river water and 15 g of sediment were collected in Liaodong Bay Basin in November 2013. Samples were collected on days when no rainfall had occurred for two consecutive days. For the major four rivers, most of the sample locations were from the downstream end of the river in each watershed, and these samples should represent the quality of each watershed. In addition, eight river water/ sediment samples were collected from the downstream areas of the eight minor rivers flowing into Liaodong Bay to calculate the mass loadings of PFASs into this bay. River water was collected using a stainless steel bucket and stored in 1-L polypropylene (PP) bottles, while sediment was sampled using a grab sampler and stored in 50-mL PP centrifuge tubes. Water and sediment samples were stored in a cool box and transported to the laboratory for chemical analysis. All the equipment was precleaned by rinsing with methanol and Milli-Q water before sample collection. A map of Liaodong Bay basin and the sampling locations was shown in Fig. 1. The water samples were stored at room temperature for less than a week prior to analysis, and the sediment samples at − 20 °C until analysis. To ensure that PFASs were not introduced as a result of background contamination, twelve field blanks were prepared containing cleaned Milli-Q water, carried to the sampling sites, and brought back to the laboratory with the water samples on each day that samples were collected. Prior to extraction, sediment samples were frozen-dried and ground before being homogenized by a porcelain mortar and pestle. The chemical and internal standards were purchased from Wellington Laboratories (Guelph, ON, Canada). Their abbreviations and purity were listed in Table S-1. Methanol and methyl-tert-butyl ether (MTBE) were purchased from Tedia, USA. Tetrabutylammonium hydrogensulfate (TBAHS, HPLC grade) was purchased from Roe Scientific Inc. Milli-Q water was further treated using Waters Oasis HLB cartridges (200 mg, Milford, MA) to remove the potential PFAS residues. Ammonium acetate was purchased from Sigma-Aldrich Co. (St. Louis, MO). All reagents were used as received.
Fig. 1. Map of China and the Liaodong Bay basin marked with the fluorochemical industrial parks and sampling sites.
Please cite this article as: Chen, H., et al., Levels and spatial distribution of perfluoroalkyl substances in China Liaodong Bay basin with concentrated fluorine industry parks, Marine Pollution Bulletin (2015), http://dx.doi.org/10.1016/j.marpolbul.2015.10.024
H. Chen et al. / Marine Pollution Bulletin xxx (2015) xxx–xxx
The water sample extraction method was developed previously by Bao et al. (Bao et al., 2011). Briefly, suspended solids were removed by a glass fiber filter (1.6 μm; Whatman, Florham Park, NJ). The samples were spiked with 2 ng of mass-labeled internal standards before filtration. Solid phase extraction (SPE) was conducted at a flow of 10 mL/min by a Sep-Pak concentrator (Waters Corp., Milford, MA). Waters Oasis HLB cartridges (200 mg, Milford, MA) were first conditioned with 10 mL of methanol and 10 mL of cleaned Milli-Q water. A 1000-mL aliquot of each water sample was then loaded onto the SPE cartridge. After the cartridges were dried completely by purging with high purity nitrogen, the target analytes were eluted from the cartridges with 2 mL of methanol at a flow rate of 1 mL/min. The extract was dried under a gentle stream of nitrogen, and then dissolved in 200 μL of methanol/10 mM ammonium acetate (40:60) to approximate the initial mobile-phase conditions. The extracted sample was passed through a nylon filter (pore diameter 0.2 μm, Millipore, MA) and then transferred to a PP vial for instrumental analysis. Sediment samples were treated and extracted using a method we previously reported (Chen et al., 2012). Briefly, 5 g of sediment was transferred to a 50 mL PP centrifuge tube, and vortexed with 2 mL of cleaned Mill-Q water. 2 mL of 0.25 M Na2CO3 and 1 mL of 0.5 M TBAHS solutions were then added and vortex mixed before duplicate extraction with MTBE. The combined MTBE extracts were brought to dryness under a gentle stream of high purity nitrogen, and reconstituted in a 1 mL mixture of methanol and 10 mM ammonium acetate (2:3, v/v) before final filtration with a 0.22 μm nylon filter. Prior to extraction, mass-labeled ISs (2 ng) were spiked into sediment samples. The PFASs in the extracts of river water and sediment samples were analyzed using an Agilent 1200 high performance liquid chromatography system (Palo Alto, CA) coupled with an Agilent 6410 Triple Quadrupole (QQQ) mass spectrometer (Santa Clara, CA). A 10-μL aliquot of each extract was injected onto an Agilent Eclipse Plus C18 column (2.1 × 100 mm, 3.5 μm, Agilent, Palo Alto, CA) that was maintained at 40 °C. The mobile phase consisted of 10 mM ammonium acetate and methanol at a flow rate of 250 μL/min. The mobile phase gradient began at 40% methanol, was increased to 90% after 3 min, maintained for 17 min, increased to 100% methanol, and then maintained this level for 10 min. Finally, the gradient was returned to its initial conditions. Electrospray negative ionization (ESI) was used in the mass spectrometer source. The gas temperature and capillary voltage were maintained at 350 °C and 4000 V. Ions were monitored with a multiple reaction monitoring (MRM) mode, and analyte-specific mass spectrometer parameters were optimized for each compound. The other
3
parameters for instrument analysis for each measured compound are listed in Table 1. Eleven-point calibration curves using external solvent standards in the range of 0.30 to 500 ng/mL were constructed for quantification of PFASs in the extracts. The correlation coefficients (r2) for each calibration curve were over 0.99. The calibration verification standard was checked every 10 sample injections to monitor the validity of the calibration. Field blank, solvent blank and procedural blank samples were also prepared and analyzed in this experiment. Field blanks were prepared to ensure that PFASs were not introduce as a result of background contamination. Solvent blanks consisting of methanol and 10 mM ammonium acetate (2:3, v/v) were used to run after every ten samples to ensure that the mobile phase materials and analytical instrumentation remained free of contamination during analysis. Procedural blank samples, prepared from 1 L of cleaned Milli-Q water, were also run after every ten samples to ensure that sample processing materials and procedures were free of contamination. No detectable PFAS contamination occurred in the blank samples (b 5 ng/L). Recovery experiments were performed by spiking 2 ng of native and mass-labeled standards into 1000 mL of the river water (n = 3) and 5 g of sediment (n = 6) samples. Matrix recoveries ranged 28.4–116.3% in water, and 43.9–110.9% in sediment samples (listed in Table 1). Only one PFAS (PFBA) had recoveries out of the range between 50 and 120%. It was omitted from this study because of poor accuracy. Duplicate measurements showed good precision. The limit of quantification (LOQ) as the analyte peak required yielding a signal-to noise (S/N) ratio larger than 10:1. The LOQ values for each measured compound for different matrixes are shown in Table 1. During analysis, samples with concentrations below the LOQ were given a value of one-half of the LOQ. In total, thirty-five river water samples were analyzed for 18 PFASs. Concentrations and spatial distributions of PFASs in the rivers of Liaodong Bay basin were shown in Table S2 and Fig. 2, respectively. Of the 35 samples collected, all of them had at least one target compound at concentrations above the LOQ. Six of the target compounds were measured at concentrations exceeding the LOQ in more than half of the samples, with PFHpA in 31 samples (89%), PFOA in 35 (100%), PFNA in 33 (94%), PFDA in 24 (69%), PFHxS in 13 (63%) and PFOS in 33 (94%). It was indicated that PFOA was the most commonly detected compound in the river water samples of Liaodong Bay basin. In terms of levels for each target compound, high concentrations of PFOA (1.4–107.2 ng/L), PFOS (b LOQ ~ 8.5 ng/L) and PFBS (b LOQ ~124.1 ng/L) were found in the collected water samples (Fig. 3). PFOA levels in this study were comparable to those measurements that have
Table 1 Parameters for instrumental analysis, LOQ and extraction process recovery for each measured compound. Measured compound
Internal standard
Precursor ion
PFBA PFPeA PFHxA PFHpA PFOA PFNA PFDA PFUnDA PFDoDA, PFTrDA PFTeDA PFHxDA PFOcDA, PFBS PFHxS PFOS PFDS FOSA FOSAA
[13C4]PFBA [13C2]PFHxA
212.7 262.7 312.7 362.7 412.7 462.7 512.7 562.7 612.7 662.7 712.7 812.7 912.7 298.7 398.7 498.7 598.7 497.7 557.7
[13C4]PFOA [13C5]PFNA [13C2]PFDA [13C2]PFUdA [13C2]PFDoA
[18O2]PFHxS [13C4]PFOS [13C4]PFOS
Product ion
168.7 218.7 268.7 318.7 368.7 418.7 468.7 518.7 568.7 618.7 668.7 768.7 868.7 98.7 98.7 98.7 98.7 77.7 499.7
Fragmentor (V)
Collision (V)
LOQ Water (ng/L)
Sediment (ng/g dw)
Recovery (mean ± SD) Water
Sediment
65 70 65 75 75 80 75 85 95 100 105 120 100 130 170 200 205 155 135
5 5 5 5 5 5 5 5 5 5 5 9 9 29 37 40 40 40 17
0.24 0.16 0.15 0.05 0.05 0.05 0.05 0.05 0.05 0.08 0.08 0.08 0.08 0.12 0.04 0.04 0.04 0.05 0.08
0.50 0.10 0.05 0.05 0.04 0.05 0.05 0.05 0.05 0.10 0.10 0.10 0.10 0.08 0.05 0.04 0.05 0.08 0.10
28.4 ± 11.4 51.6 ± 2.7 83.5 ± 3.2 80.6 ± 2.8 87.5 ± 3.5 72.0 ± 1.3 73.4 ± 0.3 74.8 ± 5.3 108.1 ± 2.8 84.2 ± 3.0 69.8 ± 3.5 60.4 ± 8.2 50.3 ± 8.9 116.3 ± 4.8 98.8 ± 1.0 90.4 ± 5.0 91.7 ± 6.8 74.0 ± 0.6 91.4 ± 9.1
43.9 ± 7.3 68.7 ± 13.7 87.7 ± 5.2 107.6 ± 4.9 109.2 ± 7.4 94.1 ± 4.4 87.0 ± 8.0 98.2 ± 4.7 73.3 ± 7.0 71.9 ± 2.5 63.4 ± 2.5 65.4 ± 1.6 58.3 ± 8.6 110.9 ± 10.1 84.0 ± 10.5 85.0 ± 6.0 78.1 ± 15.4 82.5 ± 8.8 77.7 ± 9.2
Please cite this article as: Chen, H., et al., Levels and spatial distribution of perfluoroalkyl substances in China Liaodong Bay basin with concentrated fluorine industry parks, Marine Pollution Bulletin (2015), http://dx.doi.org/10.1016/j.marpolbul.2015.10.024
4
H. Chen et al. / Marine Pollution Bulletin xxx (2015) xxx–xxx
Fig. 2. Spatial distributions (A) and composition profile of PFASs (B) in the river water samples of Liaodong Bay basin.
been made in the USA (n.d. ~100 ng/L, Post et al., 2013), but lower than those in other rivers in China (0.2–298 ng/L, Jin et al., 2009; a mean concentration of 3112 ng/L, Wang et al., 2014) and Japan (0.23–3703.6 ng/L, Zushi et al., 2011); PFOS levels in this study were lower than those in Japan (99.4 ng/L, Zushi et al., 2011), USA (n.d. ~ 43 ng/L, Post et al., 2013) and other rivers in China (0.40–12.78 ng/L, Wang et al., 2014). As shown in Fig. 2(B), PFOA and PFOS were dominant compounds in most of the water samples and no obvious spatial variation was
Fig. 3. PFAS concentrations in the river water samples of Liaodong Bay basin.
observed in the patterns of dominant compounds. It was indicated that there are no obvious regional difference in the use and emission of PFASs in the study area. PFOA and PFOS were considered as the two most industrially utilized PFASs (Zushi et al., 2011). It suggested that industrial emission of PFASs was the major source in this area. In fact, as discussed before, there are two fluorochemical industrial parks, containing several fluorochemical plants where PFOA is an important material in the manufacture of various fluoropolymers in the study area. The wastewater from the park discharged into the Xi River and eventually into Daling River, which was included in the Liaodong Bay basin, and this might have contributed to the high levels of PFOA. This result of dominant compounds is consistent with many earlier reports (Liu et al., 2009; Wang et al., 2013), but it differs from the previous report of PFASs in the Tokyo Bay basin of Japan where PFNA was the most prevalent PFAS in the river water (Zushi et al., 2011). Although PFOA was the most commonly detected and dominant PFAS, its maximum concentration was not the highest. The highest maximum concentration was for PFBS at site DLI5 (124.1 ng/L) in Daling River; this is much higher than that determined in surface water from the USA (n.d. ~ 6 ng/L, Post et al., 2013). Higher PFBS level was also found at site DLI4 when compared with those in other rivers investigated in this study. However, the level decreased at other three downstream sites and became comparable to those from other rivers. It appears that the PFBS presented a decreasing trend in concentration as the river proceeds downstream from this area. As shown in Fig. 1, the industrial fluorochemical parks are close to the upstream of Daling River. PFBS may be a significant PFAS composition in effluents and they resulted in PFBS contamination in the upstream of Daling River. Due to its high volatility, contributions of PFBS to the total concentrations of PFASs became less in the downstream of the Daling River and
Please cite this article as: Chen, H., et al., Levels and spatial distribution of perfluoroalkyl substances in China Liaodong Bay basin with concentrated fluorine industry parks, Marine Pollution Bulletin (2015), http://dx.doi.org/10.1016/j.marpolbul.2015.10.024
H. Chen et al. / Marine Pollution Bulletin xxx (2015) xxx–xxx
other rivers in this basin. Our observations are consistent with an earlier study (Bao et al., 2011), in which PFASs in the environment near fluorochemical plants were investigated. This study found that in 2011, PFBS was dominant in some areas near the plants at concentration up to 445 ng/L; this is higher than that measured in this study. The total concentrations of PFASs in river water samples were compared among rivers and it presented obvious spatial trends. Generally, as shown in Fig. 2(B), the total PFAS concentrations at sampling sites in Daling River were the highest, followed by Xiaoling River. PFAS concentrations at sampling sites in other rivers were comparable. This spatial trend in PFAS concentrations (Fig. 2), together with the spatial trend in patterns which discussed in the above paragraph, supports the suggestion that the fluorochemical park made large contributions to the total PFAS pollution in this basin and except this there were no other significant point sources. PFCAs with chain-length longer than 10 and PFDS were rarely detected in river water, with the median concentrations below their corresponding LOQ. The precursors of PFOS (FOSA and FOSAA) were present in low concentrations. All of the FOSAA concentrations were below the LOQ. FOSA concentrations in river water samples were higher (b LOQ ~ 0.17 ng/L) than those of FOSAA, but nearly 2 orders of magnitude lower than those of the transformed compound (PFOS). This indicated that the contribution of the precursors to the PFAS pollution might not be significant in Liaodong Bay basin. However, the concentrations of the precursors in the sediment were comparable when compared to those of PFOS which would be discussed in the next section. Accurate estimation of the precursor contribution to PFASs in the environment needs to be conducted in the future. PFASs in thirty-four sediment samples were analyzed. As shown in Table S3, of the 18 PFASs analyzed, 17 PFASs were found above the
5
LOQ with more than detection frequencies of 50%. Only PFPeA was detected in 12% of the sediment samples at concentration above the LOQ. The PFAS spatial distribution and compositions in sediment were presented in Fig. 4. In sediment samples, the most abundant PFASs or PFAS precursors in all sediment samples were PFOA with a median concentration of 1.19 ng/g (range 0.44–20.8 ng/g), followed by PFOcDA (median 0.35 ng/g, range 0.08–6.22 ng/g), PFHxDA (median 0.33 ng/g, range 0.09–4.59 ng/g) and FOSAA (median 0.32 ng/g, range 0.11– 2.94 ng/g) (Fig. 5). From Figs. 2 and 4, different composition profiles were shown in water and sediment samples from Liaodong Bay basin. On average, the greatest contribution of PFASs in sediment samples was from PFOA (28.5%), FOSAA (8.7%) and PFOcDA (8.3%), while for river water samples the greatest contribution was from PFOA (66.1%) and PFOS (7.6%). In addition, the contributions among PFAS pollutants in sediment samples were more even than those in river water samples as compared by Fig. 2(B) with Fig. 4(B). This is reasonable since PFCAs having more carbon-chain number and FOSAA are relatively hydrophobic, which results in different partition behaviors (Ahrens et al., 2009; Higgins and Luthy, 2006). The PFAS levels and composition profiles in sediment among different rivers were compared. The highest total-PFASs and dominant-PFOA level were found in Xiaoling River, followed by Daling River. This spatial distribution trend is different from that in river water samples. However, it is consistent with that in previous report in this area (Liu et al., 2009). In 2009, geographical distribution of PFASs in human blood from Liaoning Province of China was investigated. It was found that levels and composition of PFASs in human blood in Jinzhou which is located at Xiaoling River basin were comparable with those in Fuxin where the industrial fluorochemical park is located (Liu et al., 2009). Meanwhile, to our knowledge there is no perfluoropolymer
Fig. 4. Spatial distributions (A) and composition profile of PFASs (B) in the sediment samples of Liaodong Bay basin.
Please cite this article as: Chen, H., et al., Levels and spatial distribution of perfluoroalkyl substances in China Liaodong Bay basin with concentrated fluorine industry parks, Marine Pollution Bulletin (2015), http://dx.doi.org/10.1016/j.marpolbul.2015.10.024
6
H. Chen et al. / Marine Pollution Bulletin xxx (2015) xxx–xxx
The fluxes for all PFASs analyzed at concentrations above the LOQ into Liaodong Bay based on the above assumptions were estimated. The estimated total-PFAS loading into Liaodong Bay was 506 kg/year (Table 2). The PFOA and PFOS loadings were calculated to be 238 kg/year and 10 kg/year, respectively. It should be noted that PFBS was the second contributor to PFAS loading, with flux of 216 kg/year. The relatively high mass loading of the C4 acid into Liaodong Bay likely reflects a shift toward lower molecular weight PFASs made by the industrial fluorochemical park in this basin. To our knowledge, this is the first report to estimate mass flux of PFASs into Liaodong Bay where it is considered to be a typical PFAS-polluted area in China. These data will be helpful for source identification and the practical management of PFAS pollution. Acknowledgments
Fig. 5. PFAS concentrations in the sediment samples of Liaodong Bay basin.
manufacture in Xiaoling River basin. It supports the suggestion that high concentration of PFASs in Xiaoling River may be attributed to the industrial discharge or industrial waste distribution from the fluorochemical park in the city of Fuxin. The total-PFAS levels in sediment from Liaodong Bay basin were in the range of 2.6–37.3 ng/g. It is higher than those in other river sediments in China (0.098–1.889 ng/g, Zhu et al., 2014) and Hong Kong (Loi et al., 2013), but lower than those in sediment from Tangxun Lake in China where effluents from wastewater treatment plants serving the industrial zone including a production bases of the fluorochemical industry are directly discharged into (41.8–800 ng/g, Zhou et al., 2013). The mass fluxes of PFASs from the rivers flowing into Liaodong Bay were roughly estimated from the PFAS concentrations in water samples from the downstream of the major rivers and their flow rates. The calculation was based on the four assumptions, which was described previously (Zushi et al., 2011). Briefly, no daily and season variations in PFAS concentrations occurred during fine weather; the PFAS concentrations remained to be consistent despite surface runoff of PFASs during rain events, which was in agreement with previous study (Zushi et al., 2008) where it was found that there was no significant change in concentrations of PFASs despite rain events; and there are no relative data about flow rates of the minor rivers in this study because of their low volume of runoff, we assumed that the flow from the minor rivers was minimal and their PFAS loading could be neglected.
Table 2 Mass loadings of PFASs (kg/year). Compounds
Daliao River
Liao River
Daling River
Xiaoling River
Total flux
FOSA PFPeA PFHxA PFHpA PFOA PFNA PFDA PFUnDA PFDoDA PFBS PFHxS PFOS PFDS PFASs
n.d. n.d. 10.0 n.d. 71.1 n.d. n.d. n.d. n.d. 19.7 n.d. n.d. n.d. 100.8
n.d. n.d. n.d. 3.6 52.3 1.7 n.d. n.d. n.d. n.d. n.d. 2.0 n.d. 59.6
0.2 0.5 11.5 8.3 106.8 1.5 1.0 0.1 n.d. 196.1 0.5 6.7 n.d. 333.2
n.d. n.d. n.d. 0.5 8.1 0.5 0.2 0.6 0.3 n.d. 1.3 1.0 0.4 12.9
0.2 0.5 21.5 12.4 238.3 3.7 1.2 0.7 0.3 215.8 1.8 9.7 0.4 506.5
n.d.: below the LOQ.
We express great appreciation for Chenguang Bao at the National Marine Environmental Monitoring Center, who coordinated the map drawing. We also thank Dr. Xin He at the University of Dalian Technology for her help with the instrumental analysis. This study was financially supported by the National Natural Science Foundation of China (Project Nos. 41006062, 41276098, and 21207014). Appendix A. Supplementary data Supplementary data to this article can be found online at http://dx. doi.org/10.1016/j.marpolbul.2015.10.024. References Ahrens, L., Yamashita, N., Yeung, L.W.Y., Taniyasu, S., Horii, Y., Lam, P.K.S., Ebinghaus, R., 2009. Partitioning behavior of per- and polyfluoroalkyl compounds between pore water and sediment in two sediment cores from Tokyo Bay, Japan. Environ. Sci. Technol. 43, 6969–6975. Bao, J., Liu, W., Liu, L., Jin, Y., Dai, J., Ran, X., Zhang, Z., Tsuda, S., 2011. Perfluorinated compounds in the environment and the blood of residents living near fluorochemical plants in Fuxin, China. Environ. Sci. Technol. 45 (19), 8075–8080. Chen, H., Zhang, C., Han, J., Yu, Y., Zhang, P., 2012. PFOS and PFOA in influents, effluents, and biosolids of Chinese wastewater treatment plants and effluent-receiving marine environments. Environ. Pollut. 170, 26–31. Harada, K.H., Yang, H.R., Moon, C.S., Hung, N.N., Hitomi, T., Inoue, K., Niisoe, T., Watanabe, T., Kamiyama, S., Takenaka, K., Kim, M.Y., Watanabe, K., Takasuga, T., Koizumi, A., 2010. Levels of perfluorooctane sulfonate and perfluorooctanoic acid in female serum samples from Japan in 2008, Korea in 1994–2008 and Vietnam in 2007– 2008. Chemosphere 79, 314–319. Haug, L.S., Thomsen, C., Becher, G., 2009. Time trends and the influence of age and gender on serum concentrations of perfluorinated compounds in archived human samples. Environ. Sci. Technol. 43, 2131–2136. Higgins, C.P., Luthy, R.G., 2006. Sorption of perfluorinated surfactants on sediments. Environ. Sci. Technol. 40, 7251–7256. Jin, Y., Saito, N., Harada, K., Inoue, K., Koizumi, A., 2007. Historical trends in human serum levels of perfluorooctanoate and perfluorooctane sulfonate in Shenyang, China. Tohoku J. Exp. Med. 212 (1), 63–70. Jin, Y., Liu, W., Sato, I., Nakayama, S.F., Sasaki, K., Saito, N., Tsuda, S., 2009. PFOS and PFOA in environmental and tap water in China. Chemosphere 77 (5), 605–611. Kato, K., Wong, L.Y., Jia, L.T., Kuklenyik, Z., Calafat, A.M., 2011. Trends in exposure to polyfluoroalkyl chemicals in the U.S. population: 1999–2008. Environ. Sci. Technol. 45, 8037–8045. Lau, C., 2012. Perfluorinated compounds. EXS 101, 47–86. Lindstrom, A.B., Strynar, M.J., Libelo, E.L., 2011. Polyfluorinated compounds: past, present, and future. Environ. Sci. Technol. 45 (19), 7954–7961. Liu, J., Li, J., Luan, Y., Zhao, Y., Wu, Y., 2009. Geographical distribution of perfluorinated compounds in human blood from Liaoning Province, China. Environ. Sci. Technol. 43, 4044. Loi, E.I.H., Yeung, L.W.Y., Mabury, S.A., Lam, P.K.S., 2013. Detections of commercial fluorosurfactants in Hong Kong marine environment and human blood: a pilot study. Environ. Sci. Technol. 47, 4677–4685. Post, G.B., Cohn, P.D., Cooper, K.R., 2012. Perfluorooctanoic acid (PFOA), an emerging drinking water contaminant: a critical review of recent literature. Environ. Res. 116, 93–117. Post, G.B., Louis, J.B., Lee Lippincott, R., Procopio, N.A., 2013. Occurrence of perfluorinated compounds in raw water from New Jersey public drinking water systems. Environ. Sci. Technol. 47 (23), 13266–13275. U.S. EPA, 2000. 3M: Phase-Out Plan for POSF-Based Products, U.S. EPA Docket ID OPPT2002-0043. U.S. EPA, 2006. Global PFOA Stewardship. http://www.epa.gov/oppt/pfoa.
Please cite this article as: Chen, H., et al., Levels and spatial distribution of perfluoroalkyl substances in China Liaodong Bay basin with concentrated fluorine industry parks, Marine Pollution Bulletin (2015), http://dx.doi.org/10.1016/j.marpolbul.2015.10.024
H. Chen et al. / Marine Pollution Bulletin xxx (2015) xxx–xxx UNEP, 2009. Report of the Conference of the Parties of the Stockholm Convention on Persistent Organic Pollutants on the Work of its Fourth Meeting. http://www.pops. int/documents/meetings/poprc/submissions/submission_PFOS_2008.htm (accessed Feb. 27, 2011). Wang, T., Wang, Y.W., Liao, C.Y., Cai, Y.Q., Jiang, G.B., 2009. Perspectives on the inclusion of perfluorooctane sulfonate into the Stockholm Convention on persistent organic pollutants. Environ. Sci. Technol. 43 (14), 5171–5175. Wang, J.S., Li, Y., Liu, Y., Zhang, H.X., Dai, J.Y., 2010. Disturbance of perfluorooctanoic acid on development and behavior in Drosphila larvae. Environ. Toxicol. Chem. 29, 2117–2122. Wang, T.Y., Lu, Y.L., Chen, C.L., Naile, J.E., Khim, J.S., Park, J., Luo, W., Jiao, W.T., Hu, W.Y., Giesy, J.P., 2011a. Perfluorinated compounds in estuarine and coastal areas of north Bohai Sea, China. Mar. Pollut. Bull. 62 (8), 1905–1914. Wang, T.Y., Chen, C.L., Naile, J.E., Khim, J.S., Giesy, J.P., Lu, Y.L., 2011b. Perfluorinated compounds in water, sediment and soil from Guanting Reservoir, China. Bull. Environ. Contam. Toxicol. 87 (1), 74–79. Wang, T.Y., Khim, J.S., Chen, C.L., Naile, J.E., Lu, Y.L., Kannan, K., Park, J., Luo, W., Jiao, W.T., Hu, W.Y., Giesy, J.P., 2012. Perfluorinated compounds in surface waters from Northern China: comparison to level of industrialization. Environ. Int. 42, 37–46. Wang, P., Wang, T., Giesy, J.P., Lu, Y., 2013. Perfluorinated compounds in soils from Liaodong Bay with concentrated fluorine industry in China. Chemosphere 91, 751–757.
7
Wang, P., Lu, Y., Wang, T., Fu, Y., Zhu, Z., Liu, S., Xie, S., Xiao, Y., Giesy, J.P., 2014. Occurrence and transport of 17 perfluoroalkyl acids in 12 coastal rivers in south Bohai coastal region of China with concentrated fluoropolymer facilities. Environ. Pollut. 190, 115–122. Yeung, L.W.Y., Robinson, S.J., Koschorreck, J., Mabury, S.A., 2013. Part I. A temporal study of PFCAs and their precursors in human plasma from two German cities 1982–2009. Environ. Sci. Technol. 47, 3865–3874. Zhou, Z., Liang, Y., Shi, Y., Xu, L., Cai, Y., 2013. Occurrence and transport of perfluoroalkyl acid (PFAAs), including short-chain PFAAs in Tangxun Lake, China. Environ. Sci. Technol. 47, 9249–9257. Zhu, Z., Wang, T., Wang, P., Lu, Y., Giesy, J.P., 2014. Perfluoroalkyl and polyfluoroalkyl substances in sediments from South Bohai coastal watersheds, China. Mar. Pollut. Bull. 85, 619–627. Zushi, Y., Takeda, T., Masunaga, S., 2008. Existence of nonpoint source of perfluorinated compounds and their loads in the Tsurumi River basin, Japan. Chemosphere 71 (8), 1566–1573. Zushi, Y., Ye, F., Motegi, M., Nojiri, K., Hosono, S., Suzuki, T., Kosugi, Y., Yaguchi, K., Masunaga, S., 2011. Spatially detailed survey on pollution by multiple perfluorinated compounds in the Tokyo Bay basin of Japan. Environ. Sci. Technol. 45 (7), 2887–2893.
Please cite this article as: Chen, H., et al., Levels and spatial distribution of perfluoroalkyl substances in China Liaodong Bay basin with concentrated fluorine industry parks, Marine Pollution Bulletin (2015), http://dx.doi.org/10.1016/j.marpolbul.2015.10.024
Contents lists available at ScienceDirect
Marine Pollution Bulletin journal homepage: www.elsevier.com/locate/marpolbul
Baseline
Levels and spatial distribution of perfluoroalkyl substances in China Liaodong Bay basin with concentrated fluorine industry parks☆ Hong Chen a,b, Can Zhang a, Jianbo Han a,⁎, Ruijun Sun a, Xiangyun Kong a, Xiaomeng Wang a, Xin He b a Key Laboratory of Coastal Ecology and Environment of State Oceanic Administration, Department of Marine Chemistry, National Marine Environmental Monitoring Center, Linghe Street 42, Dalian 116023, China b Key Laboratory of Industrial Ecology and Environmental Engineering (Ministry of Education), School of Environmental and Biological Science and Technology, Dalian University of Technology, Dalian 116024, China
a r t i c l e
i n f o
Article history: Received 19 August 2015 Received in revised form 9 October 2015 Accepted 11 October 2015 Available online xxxx Keywords: PFASs Level Spatial distribution Mass loading
a b s t r a c t Eighteen different perfluoroalkyl substances (PFASs) were investigated in 35 river water samples and 34 sediment samples collected from rivers in the Liaodong Bay basin containing two fluorine industry parks. Perfluorooctanoate (PFOA) and perfluorooctane sulfonate (PFOS) were the predominant PFASs in freshwater, with median concentrations of 26.5 ng/L and 1.87 ng/L, respectively. However, perfluorobutane sulfonate (PFBS) had the highest maximum concentration (up to 124.1 ng/L, approximately two orders of magnitude higher) in water at a site which is the nearest to the industrial source of PFASs. Total PFASs in water at this site were also the highest. In contrast, PFOA and perfluorooctadecanoate (PFOcDA) were the most abundant PFASs in sediment, with median concentrations of 1.19 ng/g and 0.35 ng/g, respectively. Total PFAS concentrations in sediment from the site near to the industrial park were significantly higher than the other rivers. Mass loading of total PFASs from the rivers flowing into Liaodong Bay was estimated to be 506 kg/year. © 2015 Elsevier Ltd. All rights reserved.
Perfluoroalkyl substances (PFASs) are a large and important class of synthetic chemicals. These PFASs have attractive properties, such as surfactant activity, resistance to acid and high temperatures, and water and oil repellency. They make them useful in commercial products including nonstick coatings, stain-repellant fabrics, paper packaging products and firefighting foams. Consequently, this class of compounds has been produced and used for over 60 years (Lindstrom et al., 2011). This has led to that PFASs are now ubiquitous in various environmental matrices, including river and ocean water, soils, sediments, sludge, biota, and even the polar ice caps (Chen et al., 2012; Lau, 2012; Post et al., 2012; Loi et al., 2013). The structure of PFASs consists of a totally fluorinated carbon chain of varying length and a charged functional group, such as carboxylic or sulfonic acid (Lindstrom et al., 2011). Because of the extreme stability of their carbon–fluorine bonds, PFASs are highly persistent in the environment (Lindstrom et al., 2011). Some of them have been shown to be bioaccumulative and toxic (Wang et al., 2010), including hepatic, developmental, immune and endocrine toxicities in experimental
☆ Capsule: Levels and spatial distribution of 18 PFASs in Liaodong Bay basin containing both of the domestic and industrial sources of PFASs were shown. ⁎ Corresponding author. E-mail addresses: [email protected] (H. Chen), [email protected] (C. Zhang), [email protected], [email protected] (J. Han), [email protected] (R. Sun), [email protected] (X. Kong), [email protected] (X. Wang), [email protected] (X. He).
animals. As compared to the typical POPs, such as PAHs and PCBs, PFASs are highly water-soluble (Post et al., 2012). Perfluorooctanoic acid (PFOA) and perfluorooctane sulfonic acid (PFOS) as the two most-well studied PFASs were previously made and used in the largest amounts. The production of PFOS and its synthetic starting material, perfluorooctyl sulfonyl fluoride (PFOSF), has ceased substantially since their major manufacturer 3M voluntarily phasedout their production in 2002 (U.S. EPA, 2000). Currently the use of PFOS and PFOSF is restricted to certain processes, those who have no viable alternatives under the Stockholm Convention because they were designated as persistent organic pollutants (POPs) by the Convention on POPs in 2009. In 2006, eight major manufacturers participated in a PFOA stewardship program which is an agreement between the United States Environmental Protection Agency and the producers (U.S. EPA, 2006). The goals of this program were to have a complete elimination of PFOA together with precursor chemicals and higher homologue chemicals in emission and products by 2015. However, the environmental concentrations of PFOS and PFOA are not expected to decline. Available information (Jin et al., 2007; Harada et al., 2010) indicated that PFOS in human serum did not significantly decrease in Busan and Seoul (Korea) and in Shenyang (China); PFOA elimination was relatively slow, and in some areas (e.g., in Korea) levels of PFOA and some higher homologue chemicals (e.g., PFNA and PFDA) did not decline but showed an increase (Haug et al., 2009; Kato et al., 2011; Yeung et al., 2013). The above observed continuation of PFAS pollution after the application of PFAS regulations may be explained by the proposed several
http://dx.doi.org/10.1016/j.marpolbul.2015.10.024 0025-326X/© 2015 Elsevier Ltd. All rights reserved.
Please cite this article as: Chen, H., et al., Levels and spatial distribution of perfluoroalkyl substances in China Liaodong Bay basin with concentrated fluorine industry parks, Marine Pollution Bulletin (2015), http://dx.doi.org/10.1016/j.marpolbul.2015.10.024
2
H. Chen et al. / Marine Pollution Bulletin xxx (2015) xxx–xxx
hypotheses including the consumption of stockpiles of PFASs, emission from residuals in commercial products and degradation of the precursors of PFOS and PFOA (Wang et al., 2009). Legislation in some countries such as those in Europe and North America on restriction or elimination of PFOS-contained product resulted in a shift of production to other countries such as China (Wang et al., 2009; UNEP, 2009). Since 2004, Fuxin which is situated in northwest Liaoning Province as shown in Fig. 1 has become a production base of the fluorochemical industry in China. It has been developed into two large fluorochemical industrial parks containing several fluorochemical plants. Currently these industrial parks focused on the production of the potassium salt of perfluorobutane sulfonate (PFBS), ammonium perfluorooctanoate (APFO, the derivatives of PFOA) and fluoropolymers, such as polytetrafluoroethylene (PTFE). The industrial effluent from the parks will be discharged into Daling River and then finally to Liaodong Bay. It may be a significant industrial source of PFAS emissions into the environment. In addition, Liaodong Bay basin in Liaoning Province contains a population of 43 million. The domestic sewage would be another source of PFASs in the environment. Therefore, Liaodong Bay basin in Liaoning Province is a PFASs-polluted typical region in China. Previous reports (Wang et al., 2011a, 2011b, 2012) also indicated that the greatest concentrations of PFASs, such as PFOS and PFOA, were observed in Liaodong Bay basin in Liaoning region. However, PFAS concentrations in this region and their flux into the ocean are still limited. The objectives of this study were to investigate the current extent and the spatial distribution of PFAS contamination in the environment around Liaodong Bay basin in Liaoning Province, and to evaluate the possible sources and flux of PFASs into the ocean. To accomplish these goals, a total of 17 perfluoroalkyl acids and 2 precursors were analyzed in surface water and sediment samples collected from strategic locations in Liaodong Bay basin. These data would be helpful for the evaluation of potential impacts from industries operating in this region and the development of appropriate controls and management solutions for PFAS pollution in this area. Water and sediment samples were collected from Liaodong Bay basin. It contains four major rivers and eight minor rivers flowing into the bay. Waters from these twelve rivers make up most of the total river water discharge into Liaodong Bay. The total area of this basin is approximately 269,570 km2. Because this area is too large, we selected
Liaodong Bay basin in Liaoning Province as the sampled area in this study (the downstream region in this basin, Fig. 1). The study area mainly consists of nine cities including Yingkou, Anshan, Liaoyang, Shenyang, Panjin, Fuxin, Jinzhou, Chaoyang and Huludao. It contains a population of 44 million. One liter samples of river water and 15 g of sediment were collected in Liaodong Bay Basin in November 2013. Samples were collected on days when no rainfall had occurred for two consecutive days. For the major four rivers, most of the sample locations were from the downstream end of the river in each watershed, and these samples should represent the quality of each watershed. In addition, eight river water/ sediment samples were collected from the downstream areas of the eight minor rivers flowing into Liaodong Bay to calculate the mass loadings of PFASs into this bay. River water was collected using a stainless steel bucket and stored in 1-L polypropylene (PP) bottles, while sediment was sampled using a grab sampler and stored in 50-mL PP centrifuge tubes. Water and sediment samples were stored in a cool box and transported to the laboratory for chemical analysis. All the equipment was precleaned by rinsing with methanol and Milli-Q water before sample collection. A map of Liaodong Bay basin and the sampling locations was shown in Fig. 1. The water samples were stored at room temperature for less than a week prior to analysis, and the sediment samples at − 20 °C until analysis. To ensure that PFASs were not introduced as a result of background contamination, twelve field blanks were prepared containing cleaned Milli-Q water, carried to the sampling sites, and brought back to the laboratory with the water samples on each day that samples were collected. Prior to extraction, sediment samples were frozen-dried and ground before being homogenized by a porcelain mortar and pestle. The chemical and internal standards were purchased from Wellington Laboratories (Guelph, ON, Canada). Their abbreviations and purity were listed in Table S-1. Methanol and methyl-tert-butyl ether (MTBE) were purchased from Tedia, USA. Tetrabutylammonium hydrogensulfate (TBAHS, HPLC grade) was purchased from Roe Scientific Inc. Milli-Q water was further treated using Waters Oasis HLB cartridges (200 mg, Milford, MA) to remove the potential PFAS residues. Ammonium acetate was purchased from Sigma-Aldrich Co. (St. Louis, MO). All reagents were used as received.
Fig. 1. Map of China and the Liaodong Bay basin marked with the fluorochemical industrial parks and sampling sites.
Please cite this article as: Chen, H., et al., Levels and spatial distribution of perfluoroalkyl substances in China Liaodong Bay basin with concentrated fluorine industry parks, Marine Pollution Bulletin (2015), http://dx.doi.org/10.1016/j.marpolbul.2015.10.024
H. Chen et al. / Marine Pollution Bulletin xxx (2015) xxx–xxx
The water sample extraction method was developed previously by Bao et al. (Bao et al., 2011). Briefly, suspended solids were removed by a glass fiber filter (1.6 μm; Whatman, Florham Park, NJ). The samples were spiked with 2 ng of mass-labeled internal standards before filtration. Solid phase extraction (SPE) was conducted at a flow of 10 mL/min by a Sep-Pak concentrator (Waters Corp., Milford, MA). Waters Oasis HLB cartridges (200 mg, Milford, MA) were first conditioned with 10 mL of methanol and 10 mL of cleaned Milli-Q water. A 1000-mL aliquot of each water sample was then loaded onto the SPE cartridge. After the cartridges were dried completely by purging with high purity nitrogen, the target analytes were eluted from the cartridges with 2 mL of methanol at a flow rate of 1 mL/min. The extract was dried under a gentle stream of nitrogen, and then dissolved in 200 μL of methanol/10 mM ammonium acetate (40:60) to approximate the initial mobile-phase conditions. The extracted sample was passed through a nylon filter (pore diameter 0.2 μm, Millipore, MA) and then transferred to a PP vial for instrumental analysis. Sediment samples were treated and extracted using a method we previously reported (Chen et al., 2012). Briefly, 5 g of sediment was transferred to a 50 mL PP centrifuge tube, and vortexed with 2 mL of cleaned Mill-Q water. 2 mL of 0.25 M Na2CO3 and 1 mL of 0.5 M TBAHS solutions were then added and vortex mixed before duplicate extraction with MTBE. The combined MTBE extracts were brought to dryness under a gentle stream of high purity nitrogen, and reconstituted in a 1 mL mixture of methanol and 10 mM ammonium acetate (2:3, v/v) before final filtration with a 0.22 μm nylon filter. Prior to extraction, mass-labeled ISs (2 ng) were spiked into sediment samples. The PFASs in the extracts of river water and sediment samples were analyzed using an Agilent 1200 high performance liquid chromatography system (Palo Alto, CA) coupled with an Agilent 6410 Triple Quadrupole (QQQ) mass spectrometer (Santa Clara, CA). A 10-μL aliquot of each extract was injected onto an Agilent Eclipse Plus C18 column (2.1 × 100 mm, 3.5 μm, Agilent, Palo Alto, CA) that was maintained at 40 °C. The mobile phase consisted of 10 mM ammonium acetate and methanol at a flow rate of 250 μL/min. The mobile phase gradient began at 40% methanol, was increased to 90% after 3 min, maintained for 17 min, increased to 100% methanol, and then maintained this level for 10 min. Finally, the gradient was returned to its initial conditions. Electrospray negative ionization (ESI) was used in the mass spectrometer source. The gas temperature and capillary voltage were maintained at 350 °C and 4000 V. Ions were monitored with a multiple reaction monitoring (MRM) mode, and analyte-specific mass spectrometer parameters were optimized for each compound. The other
3
parameters for instrument analysis for each measured compound are listed in Table 1. Eleven-point calibration curves using external solvent standards in the range of 0.30 to 500 ng/mL were constructed for quantification of PFASs in the extracts. The correlation coefficients (r2) for each calibration curve were over 0.99. The calibration verification standard was checked every 10 sample injections to monitor the validity of the calibration. Field blank, solvent blank and procedural blank samples were also prepared and analyzed in this experiment. Field blanks were prepared to ensure that PFASs were not introduce as a result of background contamination. Solvent blanks consisting of methanol and 10 mM ammonium acetate (2:3, v/v) were used to run after every ten samples to ensure that the mobile phase materials and analytical instrumentation remained free of contamination during analysis. Procedural blank samples, prepared from 1 L of cleaned Milli-Q water, were also run after every ten samples to ensure that sample processing materials and procedures were free of contamination. No detectable PFAS contamination occurred in the blank samples (b 5 ng/L). Recovery experiments were performed by spiking 2 ng of native and mass-labeled standards into 1000 mL of the river water (n = 3) and 5 g of sediment (n = 6) samples. Matrix recoveries ranged 28.4–116.3% in water, and 43.9–110.9% in sediment samples (listed in Table 1). Only one PFAS (PFBA) had recoveries out of the range between 50 and 120%. It was omitted from this study because of poor accuracy. Duplicate measurements showed good precision. The limit of quantification (LOQ) as the analyte peak required yielding a signal-to noise (S/N) ratio larger than 10:1. The LOQ values for each measured compound for different matrixes are shown in Table 1. During analysis, samples with concentrations below the LOQ were given a value of one-half of the LOQ. In total, thirty-five river water samples were analyzed for 18 PFASs. Concentrations and spatial distributions of PFASs in the rivers of Liaodong Bay basin were shown in Table S2 and Fig. 2, respectively. Of the 35 samples collected, all of them had at least one target compound at concentrations above the LOQ. Six of the target compounds were measured at concentrations exceeding the LOQ in more than half of the samples, with PFHpA in 31 samples (89%), PFOA in 35 (100%), PFNA in 33 (94%), PFDA in 24 (69%), PFHxS in 13 (63%) and PFOS in 33 (94%). It was indicated that PFOA was the most commonly detected compound in the river water samples of Liaodong Bay basin. In terms of levels for each target compound, high concentrations of PFOA (1.4–107.2 ng/L), PFOS (b LOQ ~ 8.5 ng/L) and PFBS (b LOQ ~124.1 ng/L) were found in the collected water samples (Fig. 3). PFOA levels in this study were comparable to those measurements that have
Table 1 Parameters for instrumental analysis, LOQ and extraction process recovery for each measured compound. Measured compound
Internal standard
Precursor ion
PFBA PFPeA PFHxA PFHpA PFOA PFNA PFDA PFUnDA PFDoDA, PFTrDA PFTeDA PFHxDA PFOcDA, PFBS PFHxS PFOS PFDS FOSA FOSAA
[13C4]PFBA [13C2]PFHxA
212.7 262.7 312.7 362.7 412.7 462.7 512.7 562.7 612.7 662.7 712.7 812.7 912.7 298.7 398.7 498.7 598.7 497.7 557.7
[13C4]PFOA [13C5]PFNA [13C2]PFDA [13C2]PFUdA [13C2]PFDoA
[18O2]PFHxS [13C4]PFOS [13C4]PFOS
Product ion
168.7 218.7 268.7 318.7 368.7 418.7 468.7 518.7 568.7 618.7 668.7 768.7 868.7 98.7 98.7 98.7 98.7 77.7 499.7
Fragmentor (V)
Collision (V)
LOQ Water (ng/L)
Sediment (ng/g dw)
Recovery (mean ± SD) Water
Sediment
65 70 65 75 75 80 75 85 95 100 105 120 100 130 170 200 205 155 135
5 5 5 5 5 5 5 5 5 5 5 9 9 29 37 40 40 40 17
0.24 0.16 0.15 0.05 0.05 0.05 0.05 0.05 0.05 0.08 0.08 0.08 0.08 0.12 0.04 0.04 0.04 0.05 0.08
0.50 0.10 0.05 0.05 0.04 0.05 0.05 0.05 0.05 0.10 0.10 0.10 0.10 0.08 0.05 0.04 0.05 0.08 0.10
28.4 ± 11.4 51.6 ± 2.7 83.5 ± 3.2 80.6 ± 2.8 87.5 ± 3.5 72.0 ± 1.3 73.4 ± 0.3 74.8 ± 5.3 108.1 ± 2.8 84.2 ± 3.0 69.8 ± 3.5 60.4 ± 8.2 50.3 ± 8.9 116.3 ± 4.8 98.8 ± 1.0 90.4 ± 5.0 91.7 ± 6.8 74.0 ± 0.6 91.4 ± 9.1
43.9 ± 7.3 68.7 ± 13.7 87.7 ± 5.2 107.6 ± 4.9 109.2 ± 7.4 94.1 ± 4.4 87.0 ± 8.0 98.2 ± 4.7 73.3 ± 7.0 71.9 ± 2.5 63.4 ± 2.5 65.4 ± 1.6 58.3 ± 8.6 110.9 ± 10.1 84.0 ± 10.5 85.0 ± 6.0 78.1 ± 15.4 82.5 ± 8.8 77.7 ± 9.2
Please cite this article as: Chen, H., et al., Levels and spatial distribution of perfluoroalkyl substances in China Liaodong Bay basin with concentrated fluorine industry parks, Marine Pollution Bulletin (2015), http://dx.doi.org/10.1016/j.marpolbul.2015.10.024
4
H. Chen et al. / Marine Pollution Bulletin xxx (2015) xxx–xxx
Fig. 2. Spatial distributions (A) and composition profile of PFASs (B) in the river water samples of Liaodong Bay basin.
been made in the USA (n.d. ~100 ng/L, Post et al., 2013), but lower than those in other rivers in China (0.2–298 ng/L, Jin et al., 2009; a mean concentration of 3112 ng/L, Wang et al., 2014) and Japan (0.23–3703.6 ng/L, Zushi et al., 2011); PFOS levels in this study were lower than those in Japan (99.4 ng/L, Zushi et al., 2011), USA (n.d. ~ 43 ng/L, Post et al., 2013) and other rivers in China (0.40–12.78 ng/L, Wang et al., 2014). As shown in Fig. 2(B), PFOA and PFOS were dominant compounds in most of the water samples and no obvious spatial variation was
Fig. 3. PFAS concentrations in the river water samples of Liaodong Bay basin.
observed in the patterns of dominant compounds. It was indicated that there are no obvious regional difference in the use and emission of PFASs in the study area. PFOA and PFOS were considered as the two most industrially utilized PFASs (Zushi et al., 2011). It suggested that industrial emission of PFASs was the major source in this area. In fact, as discussed before, there are two fluorochemical industrial parks, containing several fluorochemical plants where PFOA is an important material in the manufacture of various fluoropolymers in the study area. The wastewater from the park discharged into the Xi River and eventually into Daling River, which was included in the Liaodong Bay basin, and this might have contributed to the high levels of PFOA. This result of dominant compounds is consistent with many earlier reports (Liu et al., 2009; Wang et al., 2013), but it differs from the previous report of PFASs in the Tokyo Bay basin of Japan where PFNA was the most prevalent PFAS in the river water (Zushi et al., 2011). Although PFOA was the most commonly detected and dominant PFAS, its maximum concentration was not the highest. The highest maximum concentration was for PFBS at site DLI5 (124.1 ng/L) in Daling River; this is much higher than that determined in surface water from the USA (n.d. ~ 6 ng/L, Post et al., 2013). Higher PFBS level was also found at site DLI4 when compared with those in other rivers investigated in this study. However, the level decreased at other three downstream sites and became comparable to those from other rivers. It appears that the PFBS presented a decreasing trend in concentration as the river proceeds downstream from this area. As shown in Fig. 1, the industrial fluorochemical parks are close to the upstream of Daling River. PFBS may be a significant PFAS composition in effluents and they resulted in PFBS contamination in the upstream of Daling River. Due to its high volatility, contributions of PFBS to the total concentrations of PFASs became less in the downstream of the Daling River and
Please cite this article as: Chen, H., et al., Levels and spatial distribution of perfluoroalkyl substances in China Liaodong Bay basin with concentrated fluorine industry parks, Marine Pollution Bulletin (2015), http://dx.doi.org/10.1016/j.marpolbul.2015.10.024
H. Chen et al. / Marine Pollution Bulletin xxx (2015) xxx–xxx
other rivers in this basin. Our observations are consistent with an earlier study (Bao et al., 2011), in which PFASs in the environment near fluorochemical plants were investigated. This study found that in 2011, PFBS was dominant in some areas near the plants at concentration up to 445 ng/L; this is higher than that measured in this study. The total concentrations of PFASs in river water samples were compared among rivers and it presented obvious spatial trends. Generally, as shown in Fig. 2(B), the total PFAS concentrations at sampling sites in Daling River were the highest, followed by Xiaoling River. PFAS concentrations at sampling sites in other rivers were comparable. This spatial trend in PFAS concentrations (Fig. 2), together with the spatial trend in patterns which discussed in the above paragraph, supports the suggestion that the fluorochemical park made large contributions to the total PFAS pollution in this basin and except this there were no other significant point sources. PFCAs with chain-length longer than 10 and PFDS were rarely detected in river water, with the median concentrations below their corresponding LOQ. The precursors of PFOS (FOSA and FOSAA) were present in low concentrations. All of the FOSAA concentrations were below the LOQ. FOSA concentrations in river water samples were higher (b LOQ ~ 0.17 ng/L) than those of FOSAA, but nearly 2 orders of magnitude lower than those of the transformed compound (PFOS). This indicated that the contribution of the precursors to the PFAS pollution might not be significant in Liaodong Bay basin. However, the concentrations of the precursors in the sediment were comparable when compared to those of PFOS which would be discussed in the next section. Accurate estimation of the precursor contribution to PFASs in the environment needs to be conducted in the future. PFASs in thirty-four sediment samples were analyzed. As shown in Table S3, of the 18 PFASs analyzed, 17 PFASs were found above the
5
LOQ with more than detection frequencies of 50%. Only PFPeA was detected in 12% of the sediment samples at concentration above the LOQ. The PFAS spatial distribution and compositions in sediment were presented in Fig. 4. In sediment samples, the most abundant PFASs or PFAS precursors in all sediment samples were PFOA with a median concentration of 1.19 ng/g (range 0.44–20.8 ng/g), followed by PFOcDA (median 0.35 ng/g, range 0.08–6.22 ng/g), PFHxDA (median 0.33 ng/g, range 0.09–4.59 ng/g) and FOSAA (median 0.32 ng/g, range 0.11– 2.94 ng/g) (Fig. 5). From Figs. 2 and 4, different composition profiles were shown in water and sediment samples from Liaodong Bay basin. On average, the greatest contribution of PFASs in sediment samples was from PFOA (28.5%), FOSAA (8.7%) and PFOcDA (8.3%), while for river water samples the greatest contribution was from PFOA (66.1%) and PFOS (7.6%). In addition, the contributions among PFAS pollutants in sediment samples were more even than those in river water samples as compared by Fig. 2(B) with Fig. 4(B). This is reasonable since PFCAs having more carbon-chain number and FOSAA are relatively hydrophobic, which results in different partition behaviors (Ahrens et al., 2009; Higgins and Luthy, 2006). The PFAS levels and composition profiles in sediment among different rivers were compared. The highest total-PFASs and dominant-PFOA level were found in Xiaoling River, followed by Daling River. This spatial distribution trend is different from that in river water samples. However, it is consistent with that in previous report in this area (Liu et al., 2009). In 2009, geographical distribution of PFASs in human blood from Liaoning Province of China was investigated. It was found that levels and composition of PFASs in human blood in Jinzhou which is located at Xiaoling River basin were comparable with those in Fuxin where the industrial fluorochemical park is located (Liu et al., 2009). Meanwhile, to our knowledge there is no perfluoropolymer
Fig. 4. Spatial distributions (A) and composition profile of PFASs (B) in the sediment samples of Liaodong Bay basin.
Please cite this article as: Chen, H., et al., Levels and spatial distribution of perfluoroalkyl substances in China Liaodong Bay basin with concentrated fluorine industry parks, Marine Pollution Bulletin (2015), http://dx.doi.org/10.1016/j.marpolbul.2015.10.024
6
H. Chen et al. / Marine Pollution Bulletin xxx (2015) xxx–xxx
The fluxes for all PFASs analyzed at concentrations above the LOQ into Liaodong Bay based on the above assumptions were estimated. The estimated total-PFAS loading into Liaodong Bay was 506 kg/year (Table 2). The PFOA and PFOS loadings were calculated to be 238 kg/year and 10 kg/year, respectively. It should be noted that PFBS was the second contributor to PFAS loading, with flux of 216 kg/year. The relatively high mass loading of the C4 acid into Liaodong Bay likely reflects a shift toward lower molecular weight PFASs made by the industrial fluorochemical park in this basin. To our knowledge, this is the first report to estimate mass flux of PFASs into Liaodong Bay where it is considered to be a typical PFAS-polluted area in China. These data will be helpful for source identification and the practical management of PFAS pollution. Acknowledgments
Fig. 5. PFAS concentrations in the sediment samples of Liaodong Bay basin.
manufacture in Xiaoling River basin. It supports the suggestion that high concentration of PFASs in Xiaoling River may be attributed to the industrial discharge or industrial waste distribution from the fluorochemical park in the city of Fuxin. The total-PFAS levels in sediment from Liaodong Bay basin were in the range of 2.6–37.3 ng/g. It is higher than those in other river sediments in China (0.098–1.889 ng/g, Zhu et al., 2014) and Hong Kong (Loi et al., 2013), but lower than those in sediment from Tangxun Lake in China where effluents from wastewater treatment plants serving the industrial zone including a production bases of the fluorochemical industry are directly discharged into (41.8–800 ng/g, Zhou et al., 2013). The mass fluxes of PFASs from the rivers flowing into Liaodong Bay were roughly estimated from the PFAS concentrations in water samples from the downstream of the major rivers and their flow rates. The calculation was based on the four assumptions, which was described previously (Zushi et al., 2011). Briefly, no daily and season variations in PFAS concentrations occurred during fine weather; the PFAS concentrations remained to be consistent despite surface runoff of PFASs during rain events, which was in agreement with previous study (Zushi et al., 2008) where it was found that there was no significant change in concentrations of PFASs despite rain events; and there are no relative data about flow rates of the minor rivers in this study because of their low volume of runoff, we assumed that the flow from the minor rivers was minimal and their PFAS loading could be neglected.
Table 2 Mass loadings of PFASs (kg/year). Compounds
Daliao River
Liao River
Daling River
Xiaoling River
Total flux
FOSA PFPeA PFHxA PFHpA PFOA PFNA PFDA PFUnDA PFDoDA PFBS PFHxS PFOS PFDS PFASs
n.d. n.d. 10.0 n.d. 71.1 n.d. n.d. n.d. n.d. 19.7 n.d. n.d. n.d. 100.8
n.d. n.d. n.d. 3.6 52.3 1.7 n.d. n.d. n.d. n.d. n.d. 2.0 n.d. 59.6
0.2 0.5 11.5 8.3 106.8 1.5 1.0 0.1 n.d. 196.1 0.5 6.7 n.d. 333.2
n.d. n.d. n.d. 0.5 8.1 0.5 0.2 0.6 0.3 n.d. 1.3 1.0 0.4 12.9
0.2 0.5 21.5 12.4 238.3 3.7 1.2 0.7 0.3 215.8 1.8 9.7 0.4 506.5
n.d.: below the LOQ.
We express great appreciation for Chenguang Bao at the National Marine Environmental Monitoring Center, who coordinated the map drawing. We also thank Dr. Xin He at the University of Dalian Technology for her help with the instrumental analysis. This study was financially supported by the National Natural Science Foundation of China (Project Nos. 41006062, 41276098, and 21207014). Appendix A. Supplementary data Supplementary data to this article can be found online at http://dx. doi.org/10.1016/j.marpolbul.2015.10.024. References Ahrens, L., Yamashita, N., Yeung, L.W.Y., Taniyasu, S., Horii, Y., Lam, P.K.S., Ebinghaus, R., 2009. Partitioning behavior of per- and polyfluoroalkyl compounds between pore water and sediment in two sediment cores from Tokyo Bay, Japan. Environ. Sci. Technol. 43, 6969–6975. Bao, J., Liu, W., Liu, L., Jin, Y., Dai, J., Ran, X., Zhang, Z., Tsuda, S., 2011. Perfluorinated compounds in the environment and the blood of residents living near fluorochemical plants in Fuxin, China. Environ. Sci. Technol. 45 (19), 8075–8080. Chen, H., Zhang, C., Han, J., Yu, Y., Zhang, P., 2012. PFOS and PFOA in influents, effluents, and biosolids of Chinese wastewater treatment plants and effluent-receiving marine environments. Environ. Pollut. 170, 26–31. Harada, K.H., Yang, H.R., Moon, C.S., Hung, N.N., Hitomi, T., Inoue, K., Niisoe, T., Watanabe, T., Kamiyama, S., Takenaka, K., Kim, M.Y., Watanabe, K., Takasuga, T., Koizumi, A., 2010. Levels of perfluorooctane sulfonate and perfluorooctanoic acid in female serum samples from Japan in 2008, Korea in 1994–2008 and Vietnam in 2007– 2008. Chemosphere 79, 314–319. Haug, L.S., Thomsen, C., Becher, G., 2009. Time trends and the influence of age and gender on serum concentrations of perfluorinated compounds in archived human samples. Environ. Sci. Technol. 43, 2131–2136. Higgins, C.P., Luthy, R.G., 2006. Sorption of perfluorinated surfactants on sediments. Environ. Sci. Technol. 40, 7251–7256. Jin, Y., Saito, N., Harada, K., Inoue, K., Koizumi, A., 2007. Historical trends in human serum levels of perfluorooctanoate and perfluorooctane sulfonate in Shenyang, China. Tohoku J. Exp. Med. 212 (1), 63–70. Jin, Y., Liu, W., Sato, I., Nakayama, S.F., Sasaki, K., Saito, N., Tsuda, S., 2009. PFOS and PFOA in environmental and tap water in China. Chemosphere 77 (5), 605–611. Kato, K., Wong, L.Y., Jia, L.T., Kuklenyik, Z., Calafat, A.M., 2011. Trends in exposure to polyfluoroalkyl chemicals in the U.S. population: 1999–2008. Environ. Sci. Technol. 45, 8037–8045. Lau, C., 2012. Perfluorinated compounds. EXS 101, 47–86. Lindstrom, A.B., Strynar, M.J., Libelo, E.L., 2011. Polyfluorinated compounds: past, present, and future. Environ. Sci. Technol. 45 (19), 7954–7961. Liu, J., Li, J., Luan, Y., Zhao, Y., Wu, Y., 2009. Geographical distribution of perfluorinated compounds in human blood from Liaoning Province, China. Environ. Sci. Technol. 43, 4044. Loi, E.I.H., Yeung, L.W.Y., Mabury, S.A., Lam, P.K.S., 2013. Detections of commercial fluorosurfactants in Hong Kong marine environment and human blood: a pilot study. Environ. Sci. Technol. 47, 4677–4685. Post, G.B., Cohn, P.D., Cooper, K.R., 2012. Perfluorooctanoic acid (PFOA), an emerging drinking water contaminant: a critical review of recent literature. Environ. Res. 116, 93–117. Post, G.B., Louis, J.B., Lee Lippincott, R., Procopio, N.A., 2013. Occurrence of perfluorinated compounds in raw water from New Jersey public drinking water systems. Environ. Sci. Technol. 47 (23), 13266–13275. U.S. EPA, 2000. 3M: Phase-Out Plan for POSF-Based Products, U.S. EPA Docket ID OPPT2002-0043. U.S. EPA, 2006. Global PFOA Stewardship. http://www.epa.gov/oppt/pfoa.
Please cite this article as: Chen, H., et al., Levels and spatial distribution of perfluoroalkyl substances in China Liaodong Bay basin with concentrated fluorine industry parks, Marine Pollution Bulletin (2015), http://dx.doi.org/10.1016/j.marpolbul.2015.10.024
H. Chen et al. / Marine Pollution Bulletin xxx (2015) xxx–xxx UNEP, 2009. Report of the Conference of the Parties of the Stockholm Convention on Persistent Organic Pollutants on the Work of its Fourth Meeting. http://www.pops. int/documents/meetings/poprc/submissions/submission_PFOS_2008.htm (accessed Feb. 27, 2011). Wang, T., Wang, Y.W., Liao, C.Y., Cai, Y.Q., Jiang, G.B., 2009. Perspectives on the inclusion of perfluorooctane sulfonate into the Stockholm Convention on persistent organic pollutants. Environ. Sci. Technol. 43 (14), 5171–5175. Wang, J.S., Li, Y., Liu, Y., Zhang, H.X., Dai, J.Y., 2010. Disturbance of perfluorooctanoic acid on development and behavior in Drosphila larvae. Environ. Toxicol. Chem. 29, 2117–2122. Wang, T.Y., Lu, Y.L., Chen, C.L., Naile, J.E., Khim, J.S., Park, J., Luo, W., Jiao, W.T., Hu, W.Y., Giesy, J.P., 2011a. Perfluorinated compounds in estuarine and coastal areas of north Bohai Sea, China. Mar. Pollut. Bull. 62 (8), 1905–1914. Wang, T.Y., Chen, C.L., Naile, J.E., Khim, J.S., Giesy, J.P., Lu, Y.L., 2011b. Perfluorinated compounds in water, sediment and soil from Guanting Reservoir, China. Bull. Environ. Contam. Toxicol. 87 (1), 74–79. Wang, T.Y., Khim, J.S., Chen, C.L., Naile, J.E., Lu, Y.L., Kannan, K., Park, J., Luo, W., Jiao, W.T., Hu, W.Y., Giesy, J.P., 2012. Perfluorinated compounds in surface waters from Northern China: comparison to level of industrialization. Environ. Int. 42, 37–46. Wang, P., Wang, T., Giesy, J.P., Lu, Y., 2013. Perfluorinated compounds in soils from Liaodong Bay with concentrated fluorine industry in China. Chemosphere 91, 751–757.
7
Wang, P., Lu, Y., Wang, T., Fu, Y., Zhu, Z., Liu, S., Xie, S., Xiao, Y., Giesy, J.P., 2014. Occurrence and transport of 17 perfluoroalkyl acids in 12 coastal rivers in south Bohai coastal region of China with concentrated fluoropolymer facilities. Environ. Pollut. 190, 115–122. Yeung, L.W.Y., Robinson, S.J., Koschorreck, J., Mabury, S.A., 2013. Part I. A temporal study of PFCAs and their precursors in human plasma from two German cities 1982–2009. Environ. Sci. Technol. 47, 3865–3874. Zhou, Z., Liang, Y., Shi, Y., Xu, L., Cai, Y., 2013. Occurrence and transport of perfluoroalkyl acid (PFAAs), including short-chain PFAAs in Tangxun Lake, China. Environ. Sci. Technol. 47, 9249–9257. Zhu, Z., Wang, T., Wang, P., Lu, Y., Giesy, J.P., 2014. Perfluoroalkyl and polyfluoroalkyl substances in sediments from South Bohai coastal watersheds, China. Mar. Pollut. Bull. 85, 619–627. Zushi, Y., Takeda, T., Masunaga, S., 2008. Existence of nonpoint source of perfluorinated compounds and their loads in the Tsurumi River basin, Japan. Chemosphere 71 (8), 1566–1573. Zushi, Y., Ye, F., Motegi, M., Nojiri, K., Hosono, S., Suzuki, T., Kosugi, Y., Yaguchi, K., Masunaga, S., 2011. Spatially detailed survey on pollution by multiple perfluorinated compounds in the Tokyo Bay basin of Japan. Environ. Sci. Technol. 45 (7), 2887–2893.
Please cite this article as: Chen, H., et al., Levels and spatial distribution of perfluoroalkyl substances in China Liaodong Bay basin with concentrated fluorine industry parks, Marine Pollution Bulletin (2015), http://dx.doi.org/10.1016/j.marpolbul.2015.10.024
