A review of sources, multimedia distribution and health risks of perfluoroalkyl acids (PFAAs) in China.

Chemosphere xxx (2014) xxx–xxx

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A review of sources, multimedia distribution and health risks of perfluoroalkyl acids (PFAAs) in China Tieyu Wang a, Pei Wang a,b, Jing Meng a,b, Shijie Liu a,b, Yonglong Lu a,⇑, Jong Seong Khim c, John P. Giesy d a

State Key Lab of Urban and Regional Ecology, Research Center for Eco-Environmental Sciences, Chinese Academy of Sciences, Beijing 100085, China University of Chinese Academy of Sciences, Beijing 100049, China c School of Earth and Environmental Sciences & Research Institute of Oceanography, Seoul National University, Seoul, Republic of Korea d Department of Veterinary Biomedical Sciences and Toxicology Centre, University of Saskatchewan, Saskatoon, Saskatchewan, Canada b

h i g h l i g h t s PFOS and PFOA are the predominant compounds found among environmental media. PFAAs emissions from industrial source are significant higher than domestic source. Higher levels of PFAAs were detected in more industrialized and urbanized areas. Regional urbanization is closely related to emission and contamination of PFAAs. Risk assessment of PFAAs in China has lagged behind those in developed countries.

a r t i c l e

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Article history: Received 2 April 2014 Received in revised form 4 September 2014 Accepted 4 September 2014 Available online xxxx Handling Editor: I. Cousins Keywords: PFOS PFOA Source emission Contamination Risk assessment China

a b s t r a c t Perfluoroalkyl acids (PFAAs) have been recognized as emerging pollutants because of their ubiquitous occurrence in the environment, biota and humans. In order to investigate their sources, fate and environmental effects, a great number of surveys have been carried out over the past several years. In the present review, we summarized the status of sources and emission, concentration, distribution and risks of PFAAs in China. Concentrations of PFAAs, especially perfluorooctane sulfonic acid (PFOS) and perfluorooctanoic acid (PFOA) in various environmental media including water, sediment, soil, rain, snow and organisms, as well as human tissues are summarized based on the available data. Concentrations of PFAAs in aquatic systems are higher in relatively more industrialized and urbanized areas than those from the less populated and remote regions in China, indicating that their emission and distribution are closely related to regional urbanization and industrialization. PFAAs and related products have been widely used over the past several decades, which have brought about high concentrations detected in environmental matrixes, biota and even local residents. Ecological risk assessment of PFAAs is still less developed in China. Most existing studies compared concentrations of PFAAs to guideline values derived for single species to evaluate the risk. In order to reveal the transport, partitioning and degradation of PFAAs in the environment, further studies on their behavior, fate, bioaccumulation and adverse effects in different trophic levels should be conducted. Ó 2014 Elsevier Ltd. All rights reserved.

1. Introduction Perfluoroalkyl acids (PFAAs) are of unique and useful chemical properties including surface activity, thermal and acid resistance, as well as repellency of water and oil, thus they have been used worldwide in various applications such as stain repellents, food packaging and firefighting foams for more than 60 years (Giesy ⇑ Corresponding author. Tel.: +86 10 62849466; fax: +86 10 62918177. E-mail address: [email protected] (Y. Lu).

and Kannan, 2001; Wang et al., 2009; Blaine et al., 2013). Due to high energy of carbon–fluorine bonds, PFAAs are resistant to hydrolysis, photolysis, microbial degradation, and metabolism by vertebrates (Kissa, 2001). PFAAs, especially perfluorooctane sulfonic acid (PFOS) and perfluorooctanoic acid (PFOA), were first reported to be widespread in the environment (Giesy and Kannan, 2001, 2002), and they were subsequently detected in aquatic systems (Fujii et al., 2007; Rayne and Forest, 2009) and wildlife (Kannan et al., 2002; Houde et al., 2006a; Suja et al., 2009). Many PFAAs can be accumulated in aquatic system, leading

http://dx.doi.org/10.1016/j.chemosphere.2014.09.021 0045-6535/Ó 2014 Elsevier Ltd. All rights reserved.

Please cite this article in press as: Wang, T., et al. A review of sources, multimedia distribution and health risks of perfluoroalkyl acids (PFAAs) in China. Chemosphere (2014), http://dx.doi.org/10.1016/j.chemosphere.2014.09.021

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T. Wang et al. / Chemosphere xxx (2014) xxx–xxx

to bioaccumulation and biomagnification through the food chain to wildlife and humans (Loi et al., 2011; Squadrone et al., 2014). There is still long-term public concern over adverse effects of PFAAs on ecosystem and human health as well as secondary release of PFAAs from the environment (Pico et al., 2011; Lu et al., 2013). In recent years, great attention has been paid to PFOS and PFOA, two of the frequently detected predominant PFAAs in the environment. PFOS and related chemicals have recently been added to the list of ‘‘persistent organic pollutants’’ (POPs) under the Stockholm Convention. Inclusion in Annex B of the Stockholm Convention allows limited on-going uses of PFOS, mostly in the semiconductor, photolithography and metal plating industries (Wang et al., 2009). However, in order to meet the growing demand for surfactants and other surface modification applications, China has become one of the largest countries of PFOS and related chemicals production and consumption. From 2001 to 2006, the annual production of PFOS and related chemicals increased from less than 50 t to about 250 t, and kept almost consistently until 2011. The cumulative historical production volume of PFOS and related chemicals were estimated to be 1800 t by 2011 (Xie et al., 2013b). In China, research on PFAAs was initiated much later when compared with more industrially developed countries (Fig. 1). The status of PFAAs in China was first reported by Jin et al. (2003). From 2003 to 2014, studies about PFAAs in China have been focused primarily on four aspects, including detection of PFAAs in different environmental media, risk assessment of PFAAs on environment and ecosystem, toxicology to humans and wildlife, and estimation of sources and emission (Wang et al., 2010; Cai et al., 2012; Wu et al., 2012; Xie et al., 2013a,b; Yeung et al., 2009; Zhang et al., 2011). Previous studies have summarized and compared PFAAs concentration and contamination profiles in aquatic biota and human tissues, emissions and degradation mechanisms, bioaccumulation and biotransformation processes (Houde et al., 2006b, 2011; Lau et al., 2007; Conder et al., 2008; Houde et al., 2011; Liu and Avendano, 2013). However, historically only very

limited information about the profile of PFAAs in China is available in the existing review literature. Zhao et al. (2012a) provided a review of environmental contamination and human exposure for PFOS. In 2009, we published our first review of spatial and temporal distributions of PFOS and PFOA in China, covering the articles published before July 2008 (Chen et al., 2009), in which we summarized production, contamination, and exposure of humans to PFOS and PFOA. However, studies on PFAAs in China and resulting publications have increased dramatically in recent years, especially after PFOS was included in the national priority control list in 2009. We have carried out systematic studies in the coastal regions of the Bohai and Yellow Seas. The studies are focused on correlation between varied degrees of industrialization and PFAAs concentrations in environmental media, classification and identification of PFAAs emissions from major industries, flux of PFAAs emissions in some wastewater treatment plants, and effects of PFAAs emissions from fluoropolymer facilities on the coastal rivers (Chen et al., 2011a,b; Wang et al., 2011b, 2012, 2014b; Xie et al., 2013a). In the present review, we aim to summarize and update information on PFAAs sources, status and trends in concentrations in various media such as water, sediment, soil, sludge, rain and snow, and assessment of risks to humans and wildlife in China. We have also endeavored to assess gaps in knowledge and understanding and suggest possible future research perspectives. 2. Potential sources and emission estimation PFAAs have been manufactured and used in the globe for more than 50 years. At present, although the manufacture and use of PFAAs-based products, especially PFOS and PFOA, have been restricted or eliminated in many developed countries, PFAAs and related substances are still manufactured and used in certain industries including metal plating, photographic, fire-fighting foams, semiconductor and aviation in China (Xie et al., 2013b). China began to produce PFAAs-based products later than

Fig. 1. Temporal trends of research progress on PFAAs in China. Reference: (Jin et al., 2003, 2009; So et al., 2006a,b; Ju et al., 2008; Liu et al., 2009a; Shi et al., 2010; Zhang et al., 2010a,b, 2012a,b; Guo et al., 2011; Loi et al., 2011; Xie et al., 2013a,b).



industrially developed countries. From 2003 to 2006, the annual production of PFAAs-related chemicals had been rapidly increased, due to the sharp increase in both domestic and overseas demands in various application fields (Footitt et al., 2004; Liu et al., 2008a,c). As a result, PFAAs-related industrial processes are still the significant sources of PFAAs releases to the environment in China. Nevertheless, studies on PFAAs emissions are still at an initial stage compared with those on environmental exposure and toxic effects. Furthermore, most of the investigations have focused on estimation of PFOA and PFOS emissions. Sources of PFOS and PFOA emitted to the environment in China and other countries are illustrated in Table S1. Historically, global emission of PFOA from both direct and indirect sources was estimated based on the production data (Armitage et al., 2006, 2009; Prevedouros et al., 2006). Wang et al. (2014c) estimated emissions of perfluoroalkyl carboxylic acids (PFCAs) from quantifiable sources from 1951 to 2030. The results showed that 98–100% of historical (1951–2002) PFOA emissions are attributed to direct releases during the life-cycle of products containing PFOA. In order to track the geographical shift of industrial sources, the main countries of manufacturing and using PFOA were divided into two groups: I (Japan, Western Europe and the US); II (Russia, China, India and Poland). According to estimation from PFOA production sites in plausible scenario (i.e. assuming a use rate of 0.3 wt% PFOA-based products in fluoropolymer production), the total annual emissions of PFCAs (C4–C14) in country group II were about 2 t a 1 in 1980, but currently, the emissions exceeded 25 t a 1. PFOA and its derivatives are important additives in the production of fluoropolymers, which is bound to cause pollution of PFOA. It is noted that domestic demand and production PTFE in China rapidly increased from 6.6 kt a 1 in 1999 to about 64 kt a 1 in 2012 (Fang, 2004; Cai, 2009; Kälin et al., 2012; Wang, 2006). Chen et al. (2011b) tried to estimate mass flow of PFOA in some rivers in China, and the results showed that the highest levels of PFOA discharges were found in the Yangtze River (39 200 kg a 1) and the Huangpu River (16 000 kg a 1) in 2005. Information on emission estimation of PFOS can be divided into two categories. The first category is estimation based on rates of PFOS used in specific industries, and emission factors. In this category, PFOS emissions are estimated for both the global and various geographical regions. For example, an estimate of global historical production and environmental releases of PFOS was performed (Paul et al., 2009). In the ‘‘Environment risk evaluation report: PFOS’’ produced by the UK Environment Agency, the PFOS emissions from metal plating, photolithography, photographic industry, aviation, fire-fighting, fabric treatment, paper treatment and coatings in UK and EU were estimated (Brook et al., 2004). In China, estimation of PFOS released from chromium plating and semiconductor manufacturing sites and environmental concentrations of PFOS near the sites was made based on information in the EU technical guidance document and use of the EUSES model. The emission of PFOS from chromium plating site was estimated to be 0.0012 kg d 1 to water and 0.00072 kg d 1 to air, while the emission from semiconductor manufacturing sites was estimated to be 0.007 kg d 1 to water (Liu et al., 2008a,b). Liu et al. (2008c) also estimated the PFOS emission from fabric treatment, metal plating, semiconductor manufacturing, and fire-fighting, using the emission factors in the EU technical guidance document. It was estimated that the total amount of PFOS annual emission in whole China was 1.4 t to water from fabric treatment, 7.9 t to water and 0.005 t to air from metal plating, 0.02 t to water from semiconductor manufacturing, 1.6 t to water and 0.08 t to air from firefighting, respectively. The second category is estimation of production, use and release of PFOS based on concentrations and water flux in rivers or wastewater treatment plants (WWTPs). For instance, Pistocchi and Loos (2009) estimated overall aqueous emission of PFOS from

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the European Continent by using measured concentrations of PFOS in rivers. Similar methods were applied by Kim (2012) to estimate the watershed-based riverine discharge load of PFOS on the Korean peninsula. Total flux of sewage-derived PFOS from Japan was estimated by use of per capita emission factor, which was derived from PFOS loads in the wastewater effluents and the population served by the WWTPs (Murakami et al., 2008). In China, discharges of PFOS from major rivers, effluent and sludge in several WWTPs were estimated by Chen et al. (2011b). Total discharges of PFOS in main rivers from the coastal regions of the North Bohai Sea in 2008 were estimated to be 122 kg a 1, respectively, to which the Daliao River and Daling River made great contributions. PFOS discharges in the Pearl River Delta and Yangtze River in 2005 were found to be 4470 and 807 kg a 1, respectively. As mentioned above, little information is available on PFOS-related industries and emissions of PFOS in China. In order to reflect the overall PFOS emission pattern in China, it is necessary to conduct a systematic assessment of sources and regional emissions of PFOS-related chemicals. PFOS emissions from industrial or domestic sources in China have been undertaken by our group. Emissions of PFOS equivalents from major industrial sources including manufacture and use in textile treatment, metal plating, fire-fighting and semiconductor industry, were estimated at the provincial level (Xie et al., 2013b). Total emission of PFOS equivalents from industrial sources was estimated to be 70 t in 2010 (Fig. 2). This result was about 6 times higher than that conducted by Liu et al. in 2008. There might be two reasons: Firstly, from 2008 to 2010, related industries have developed very fast in China. For example, the production capacity of integrated circuit industry (ICI) has increased 56% from 2008 to 2010, and the increasing capacity on related industries would lead to increasing emission of PFOS; Secondly, Xie et al. provided detailed information on emission methodology but underestimation still exists, while Liu et al. just gave a very general and primary results of PFOS emission, which would lead to more underestimation. Industrial emission of PFOS equivalents in the eastern region of China was remarkably higher than those in central and western regions. For the eastern coastal provinces, the Gross Domestic Products (GDP) is similar to those of upper-middle-income economies (with per capita GDP up to 4136–12745 U.S. dollars) (The World Bank Group, 2014) (Fig. 2). Metal plating contributed the largest portion of emissions of PFOS equivalents in eastern and central China, while fire-fighting services were dominant in the emissions in western China. Estimation of domestic PFOS emission was based on the assumption that both PFOS and precursors released to air, dust and water would enter the municipal wastewater system through cleaning, wiping and washing of the products and indoor environment (Xie et al., 2013a). Total emission from domestic sources in the eastern coastal region of China was 381 kg in 2010. The results showed that emission from domestic sources was much less than that from industrial sources, and the most populous and developed regions, the Pearl River Delta, Bohai coastal region and Yangtze River Delta, were responsible for 74% of the total domestic emission of PFOS in the eastern and coastal regions of China (Fig. 3).

3. Concentrations of PFAAs in various environmental media 3.1. PFAAs in freshwater Concentrations of PFAAs have been determined in seven major river systems including main stems and tributaries in China, especially during the last few years. PFAAs are widespread in surface water in China (Fig. 4, original detailed data for histogram were provided in Table S2). PFAAs from the Pearl, Yangtze and Haihe


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Fig. 2. PFOS emissions from industrial sources in China in 2010.

Fig. 3. PFOS emissions from domestic sources in the eastern and coastal regions of China in 2010.

Rivers are the primary focus of monitoring studies. Concentrations of PFOA and PFOS were 0.85–13 and 0.9–99 ng L 1 in surface water from the lower stream of the Pearl River, while those were 2–260 and nd–14 ng L 1 from the lower stream of the Yangtze River in 2004, respectively (So et al., 2007). Concentrations of PFOA and PFOS in water from the lower stream of the Pearl River in 2012 were decreased, especially for PFOS, becoming 0.71–8.7 and 0.52–11 ng L 1, respectively (Zhang et al., 2013e). The decreasing trend may be caused by enforcement of the Stockholm Convention on PFOS and related chemicals. In 2003, concentrations of PFOA and PFOS in surface water from the middle stream of the Yangtze River were 0.2–297.5 and 0.1–37.8 ng L 1, respectively (Jin et al., 2009). There are numerous industrial zones along these major rivers in China. For example, the Yangtze River flows through the highly industrialized and commercial cities such as Chongqing and Wuhan in the middle stream, and Nanjing and Shanghai in the lower stream, where diversified industrial sources and intensive commercial activities could generate high concentrations of PFAAs. In addition, the Haihe River, which is located in Northern

China, also exhibits high concentrations of PFOA and PFOS. Studies of PFAAs in the Haihe River were all concentrated in the vicinity of the city of Tianjin, which is situated in the central zone of the Bohai Economic Rim. The ranges of concentrations were similar for PFOA or PFOS, and relatively higher concentrations were observed in the Binhai New Economic Development Area, where the highest concentrations of PFOA and PFOS were 42.1 and 10.6 ng L 1, respectively (Li et al., 2011; Pan et al., 2011; Wang et al., 2011b). Less data are available for the Songhua, Liaohe, Yellow and Huaihe Rivers. In general, PFAAs in these four rivers were detected with lower concentrations compared with those from the Pearl, Yangtze and Haihe Rivers. In northeast China, mean concentrations of PFOA and PFOS in surface water of the Songhua River were 0.17 and 1.21 ng L 1 (Liu et al., 2007); while those in the Liaohe River were 10.9 and 0.33 ng L 1, respectively (Yang et al., 2011). Even though it is one of the major rivers in China little data is available for the Yellow River. Concentrations of PFOA and PFOS in water of tributaries of the Yellow River were less than 15 ng L 1 (Wang et al., 2012). In central China, data is available for only the Jiangsu reach



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Fig. 4. Mean concentrations of PFOS and PFOA in water and sediment from the major river systems in China.

of the Huaihe River, where the concentrations of PFOA and PFOS were 18 and 4.7 ng L 1, respectively (Yu et al., 2013). Several studies have focused on seriously contaminated tributaries of major rivers, such as the Hanjiang, Huangpu and Daling Rivers (Fig. 4). The Hanjiang River is the largest tributary of the Yangtze River, and flows through the city of Wuhan. The Huangpu River is in the downstream portion of the Yangtze River, which connects Taihu Lake and the East China, and flows through the major city of Shanghai. Pollutants from industrial and domestic wastewater have been discharged into these rivers and cause severe contamination. The mean concentrations of PFOA and PFOS in the Hanjiang River were 81 and 51.8 ng L 1, while in the Huangpu River they were 105 and 5.4 ng L 1, respectively (So et al., 2007; Wang et al., 2013a). Relatively higher mean concentration of PFOA (169.04 ng L 1) was observed in the Daling River, while lower PFOS (0.42 ng L 1) was detected (Bao et al., 2011). The Daling River is located in the Northern Bohai coastal region, which is one of the most developed regions in north China. There are two fluorine chemical parks along the Daling River, which release large amount of wastes directly to the river (Wang et al., 2013c). These two industrial parks were built in 2004 and 2006 respectively, and they currently produce fluoropolymers such as PTFE. In the city of Shenyang, the capital of Liaoning province, concentrations of PFOA and PFOS in the Hunhe River were at moderate levels with mean concentrations of 7.32 and 0.51 ng L 1, compared to those in other rivers in China (Jin et al., 2009; Sun et al., 2011). Studies on PFAAs in surface waters of lakes or reservoirs have been mainly conducted in some great lakes and urban lakes (Fig. 5; original detailed data are provided in Table S2). Concentrations of PFAAs in East and Tangxun Lakes were approximately 100-fold higher than those from other lakes (Chen et al., 2012b; Zhou et al., 2013). Both East and Tangxun Lakes are urban lakes of Wuhan in Hubei province. In Tangxun Lake, it was noted that short chain PFAAs, mainly perfluorobutanesulfonate (PFBS) and

perfluorobutanoic acid (PFBA), were detected with extremely high concentrations, with mean concentrations of 4770 and 3660 ng L 1, respectively, while concentrations of PFOA and PFOS were 372 and 357 ng L 1, respectively. The results reveal that C4-PFAAs acting as substitutes of C8-PFAAs have been extensively produced and applied, and monitoring of C4-PFAAs in the environment has become urgent. PFAAs from three great lakes were also reported, including the Poyang, Tai and Chaohu Lakes. These three lakes are all distributed along the Yangtze Watershed and influenced by intensive human activities. Among these three lakes, the Taihu Lake has been the most contaminated water body. Concentrations of PFAAs in the Taihu Lake were similarly reported by Yu et al. (2013) and Yang et al. (2011). Mean concentrations of PFOA and PFOS in surface water were reported to be 56 and 15 ng L 1, respectively (Yu et al., 2013), and 21.7 and 26.5 ng L 1, respectively (Yang et al., 2011). In the Chaohu Lake, only PFOS in water was investigated, and the highest concentration was 400 ng L 1 at a location close to sewage outfall, situated in the industrially intensive district of Chaohu City (Zhang et al., 2012b). The Poyang Lake was slightly contaminated by PFAAs, with mean concentrations of PFOS and PFOA in water of 0.35 and 1.1 ng L 1, respectively (Zhang et al., 2012a). In general, concentrations of PFAAs are higher in lakes in the Yangtze Watershed than those in other lakes. Some lakes in northern and southwestern China have also been studied. Concentrations of PFOA and PFOS in water from the Dianchi Lake, which is located in Yunnan province in southwest China, were 3.4–35.4 and 1.71–40.9 ng L 1, respectively (Zhang et al., 2012c). Similar to those in the Dianchi Lake, moderate concentrations of PFOA (1.71– 43.5 ng L 1) and PFOS (0.11–1.48 ng L 1) were detected in the Baiyangdian Lake, which is the largest natural freshwater body in north China (Shi et al., 2012). Concentrations of PFOA and PFOS in water of the Guanting Reservoir were less than those in other lakes, with levels of 0.55–2.3 ng L 1 and nd–0.52 ng L 1, respectively (Wang et al., 2011a).


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Fig. 5. Mean concentrations of PFOS and PFOA in water and sediment from lakes or reservoirs in China.

Most PFAAs in China are found at the moderate levels compared with those in other countries or regions such as the USA, Canada, Japan and Europe. In general, higher concentrations of PFAAs were detected in mainstreams or tributaries of the Pearl, Yangtze and Haihe Rivers, and the East, Tangxun and Taihu Lakes in China, which were comparable with those in 18 rivers throughout whole Japan (PFOA: 0.76–192 ng L 1; PFOS: nd–191 ng L 1) (Murakami et al., 2008), and the Great Lakes in North America (PFOA: 27– 50 ng L 1; PFOS: 21–70 ng L 1) (Boulanger et al., 2004), while lower concentrations of PFAAs in the studied rivers or lakes in China were similar to those in European rivers, such as the River Elbe in Germany (PFOA: 2.8–9.6 ng L 1; PFOS: 0.5–2.9 ng L 1) (Ahrens et al., 2009), the River Seine in France (PFOA: 1.1– 18.0 ng L 1; PFOS: 9.9–39.7 ng L 1) (Labadie and Chevreuil, 2011), and the River Rhine (PFOA: 0.61–41.4 ng L 1; PFOS: 0.89– 18.6 ng L 1) (Moeller et al., 2010). PFAAs in surface water from some sites contaminated by fluorine chemical plants, mainly including East and Tangxun Lakes in Wuhan (Zhou et al., 2013) and Daling River in Fuxin (Bao et al., 2011), were less than those from a cove into which 3 M plant waste water was directly discharged (PFOA: 3,600 ng L 1; PFOS: 18 200 ng L 1) (Oliaei et al., 2013). 3.2. PFAAs in sediments Data on concentrations of PFAAs in sediments are available mostly for rivers that also have high concentrations in surface water (Fig. 4). The highest concentrations of PFOA and PFOS were detected in the Huangpu River, with 203 and 8.78 ng g 1 dry weight (dw), respectively (Li et al., 2010). Bao et al. (2010) reported that concentrations of PFOA and PFOS in sediments of the Huangpu River were 0.2–0.64 and nd–0.46 ng g 1 dw, which were less than that reported by Li et al. (2010). The difference between these two reports might due to the difference in sample collection. In the

study by Li et al., the highest PFAAs were observed around a fluorine chemical plant for producing PTFE which should be responsible for the higher level of PFOA, while in the study by Bao et al., sediments were collected from the lower reach of the Huangpu River where no point source existed. Concentrations of PFOA in sediments from the Daling River ranged from 0.18 to 48 ng g 1 dw, while PFOS was not detected (Bao et al., 2011). The greater frequency of detection of PFOA than other PFAAs could be due to emissions from the local fluorine chemical parks. Concentrations of PFAAs in sediments from the Liaohe River, adjacent to the Daling River, were also analyzed (Yang et al., 2011), PFOA and PFOS were both detected at concentrations less than 0.48 and 0.18 ng g 1 dw, respectively. Other studies on distribution of PFAAs in sediments of rivers were mostly conducted in the Haihe River (Li et al., 2011; Pan et al., 2011; Wang et al., 2011b), and there was no significant difference among these results. Data on concentrations of PFAAs in sediments of lakes or reservoirs are limited. Mean concentrations of PFOA and PFOS in Tangxun Lake were 2.4 and 74 ng g 1 dw, respectively (Zhou et al., 2013) (Fig. 5). In addition, PFBA and PFBS were detected with extremely high concentrations of 16.3 and 50.8 ng g 1 dw, respectively, which was consistent with the results from the surface water. Concentrations of PFAAs in sediments remained lower levels from other lakes or reservoirs including Guanting Reservoir, Baiyangdian and Taihu Lakes (Wang et al., 2011a; Shi et al., 2012; Yang et al., 2011). It is noted that concentrations of PFOA and PFOS in sediments in most sites from China are comparable to those from other countries or regions, such as Roter Main River, Germany (PFOA: 0.02– 0.07 ng g 1 dw; PFOS: 0.09–0.35 ng g 1 dw) (Becker et al., 2008), San Francisco Bay, USA (PFOA: nd–0.40 ng g 1 dw; PFOS: nd– 3.76 ng g 1 dw) (Higgins et al., 2005), several rivers in Japan (PFOA: nd–3.9 ng g 1 dw; PFOS: nd–11 ng g 1 dw) (Senthilkumar et al., 2007), and coastal areas of Korea (PFOA and PFOS: less than 2.0 ng g 1 dw) (Naile et al., 2010). Relatively higher concentrations


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of PFAAs in sediments have been detected in the surroundings of fluorine chemical plants in the Huangpu River, Daling River and Tangxun Lake. 3.3. PFAAs in soils and sludge Less data is available on concentrations of PFAAs in soils (Table 1). Concentrations of PFAAs in soils from Shanghai were the highest, and PFOA and PFOS in soils from Shanghai were detected at all sampling sites with concentrations in the range of 3.28–47.5 and 8.58–10.4 ng g 1 dw, respectively (Li et al., 2010), which was consistent with the results for concentrations in surface water and sediments. Apart from frequent detections of PFAAs in Shanghai, all other studies conducted in Tianjin, Guanting Reservoir in Beijing and Hebei province, north Bohai Sea and Huaihe Watershed showed slight contamination by PFAAs (Pan et al., 2011; Wang et al., 2011a,b,c; Meng et al., 2013). In these studies, PFAAs were not detected in most samples, and the highest concentration of 2.8 ng g 1 dw for PFOA was observed in Guanting Reservoir, while the highest concentration of 9.4 ng g 1 dw for PFOS was observed in Tianjin Binhai New Economic Development Area. Studies on PFAAs in soils from other countries or regions are also limited. Strynar et al. (2012) collected 60 soil samples from six countries including USA, China, Japan, Norway, Greece and Mexico, and found the concentrations of PFOA and PFOS in the range of nd– 31.7 ng g 1 dw and nd–10.1 ng g 1 dw, respectively. The results showed that relatively higher detections of PFAAs were mostly observed in USA and Japan among the six countries, and the high concentrations were comparable to those in soils from Shanghai. Concentrations of PFAAs have been measured in sludge from wastewater treatment plants (WWTPs), which are mainly situated in some developed cities (Table 1). Sludge from WWTPs can be applied to improve soils. Thus, PFAAs in sludge could affect concentrations in soils to which they are applied. In Shanghai, concentrations of PFOA and PFOS of sludge collected in 2008 from different WWTPs were 9.21–75.5 and 28.1–135 ng g 1 dw, which were higher than those in local soils and sediments. In another study, the highest concentrations of PFOA and PFOS in sludge from 25 WWTPs of Shanghai in 2008 were reported with 298 and 173 ng g 1 dw, respectively (Yan et al., 2012). Concentrations of PFOA and PFOS in sludge from Dalian, Shanghai and Guangzhou were higher than those in soils (Chen et al., 2012a). These results indicated that WWTPs might be important sources of PFAAs if the sludge was used or stacked in soil or other matrices (Li et al., 2010). Concentrations of PFAAs in sludge from Shanghai were highest, followed by those from Guangzhou, and the least concentrations were detected in sludge from Dalian. Big differences were shown in these three studies on PFOA and PFOS concentrations in sludge from Shanghai. Sludge source of WWTPs was considered as

Table 1 Concentrations of PFOS and PFOA (ng g

a

3.4. PFAAs in snow and rain Studies on PFAAs in snow or rain mainly focused on the Bohai Economic Rim. In 2010, Zhao et al. (2012a) made a detailed investigation into PFAAs in precipitation from eastern and central China (18 snow samples and 1 rain samples), where concentrations of PFOA and PFOS were 0.7–88.0 and 0.6–15.6 ng L 1, respectively. Among 19 different sites, concentrations of PFAAs from Weifang, Shandong province were the highest (more than 150 ng L 1), and those from some districts of Tianjin and Changchun were also high (about 80 ng L 1). There might be strong point sources in or around Weifang City. In a study investigating 17 PFAAs levels in the coastal rivers of Shandong Province, large scale of PTFE production was found located in the nearby city Zibo, which put significant influence to the local environment (Wang et al., 2014b). Besides industry sources, domestic applications are also important sources of PFAAs. As reported by D’eon and Mabury (2011), 80% of fluorotelomer-based commercial products were in the polymeric form that applied to carpets and textiles, whereas the other 20% were used in non-polymeric form to produce fluorinated surfactants that applied to personal care products, leveling and wetting agents, and non-stick food packaging. The study about PFAAs in precipitation also reported that, according to detected results in Tianjin, there was no apparent seasonal variation in concentrations of PFAAs, but there was a trend of decreasing concentrations in successive samples during a single precipitation interval. This result indicated that precipitation played a significant role in transportation of PFAAs from air to soil and surface water. In addition, other scattered studies were conducted in Beijing, Shenyang and Dalian. In snow samples from Beijing, concentrations of PFOA and PFOS were not high and in the range of nd–2.96 and 0.04–4.76 ng L 1, respectively. The lower pollution mainly benefited from emigration of polluting enterprises (Wang et al., 2014a). In the city of

dw) in soils and sludge from China.

Media

Location

PFOS

PFOA

Sampling time

Reference

Soil

Tianjin Binhai New Area North Bohai Sea Guanting Reservoir Shanghai Huaihe Watershed Haihe Watershed Shanghai Shanghai Coastal citiesb Tianjin 11 cities

nd–9.4(1.76) nd–0.7(0.58) nd–0.86(0.12) 8.58–10.4(9.54) nd–0.21(0.05) 0.02–2.36(0.19)a 28.1–135(44.25) 27.6–173 0.5–19.8 42–169 0.8–22.5

nd–0.93(0.2) nd–0.47(0.21) nd–2.8(0.4) 3.28–47.5(35.25) nd–0.2(0.08) nd–0.51(0.19)a 9.21–75.5(35.83) 23.2–298 0.5–158 12–68 0.6–6.7

2008 2008 2008 2007 2008 2008 2008 2010 2011 2009 2009

Wang et al. (2011b) Wang et al. (2011c) Wang et al. (2011a) Li et al. (2010) Meng et al. (2013) Pan et al. (2011) Li et al. (2010) Yan et al. (2012) Chen et al. (2012a) Sun et al. (2011) Zhang et al. (2013b)

Sludge

b

1

the most important factor to influence levels of PFAAs. Industrial sludge tends to contain higher PFAAs than domestic sludge (Li et al., 2010; Yan et al., 2012; Chen et al., 2012a). In addition, wastewater treatment technologies and different processing stages also contributed to the variability of PFAAs in sludge. In 2009, sludge was collected from 28 WWTPs in eleven economically developed cities (Zhang et al., 2013b). Concentrations of PFOA and PFOS were 0.6–6.7 and 0.8–22.5 ng g 1 dw, respectively, which were not high compared with the previous studies. Therein, 26 WWTPs just received domestic wastewater, and only 2 WWTPs received both domestic and industrial wastewater, which might be the reason why concentrations of PFAAs in these samples were lower than those from other studies in China. In general, PFAAs in sludge from WWTPs are mainly affected by types of influents, such as different kinds of industrial wastewater or domestic wastewater.

Median value. Including Shanghai, Guangzhou and Dalian.


8


Shenyang with heavy industry, concentrations of PFOA and PFOS in snow in 2006 were 1.6–22.4 and 0.4–46.2 ng L 1, respectively (Liu et al., 2007), while those collected in 2007 were 0.82–13 and nd– 51 ng L 1, respectively (Liu et al., 2009b). Results from the two studies did not show big difference. Another study reported the relatively higher concentrations of PFOA (8.08–65.8 ng L 1) and PFOS (26.9–545 ng L 1) of precipitation from Dalian in 2006 (Liu et al., 2009c). This result is unexpected since Dalian is a less industrialized city near the coast. Apart from local application of PFAAs, the coastal climate is considered to be a main reason for the relatively high concentrations, including the evolution of regional seabreeze circulation and exchange of PFAAs at air–sea interface (Tong et al., 2005; Ju et al., 2008). Sea-land breeze may cause PFAAs unable to disperse, therefore accumulate within the city, and many PFAAs acting as surfactants may also accumulate on the air/water interface. Some factors can promote sea-to-air exchange of PFAAs with low volatility, such as bursting process of the bubbles generated by breaking waves (Saint-Louis and Pelletier, 2004). And significantly higher concentrations of PFOA in aerosols compared with those in the water illustrated the transportation from water into atmosphere (Mcmurdo et al., 2008). More studies need to be conducted to compare concentrations in air from inland and coastal regions. 4. Risk assessment 4.1. Risk of PFAAs on biota 4.1.1. Occurence of PFAAs in biota PFOS is the predominant PFAA in zooplankton and blood serum of fish from a lake receiving wastewater with mean concentrations ranging from 5.7 to 64.2 ng mL 1, and biomagnification of PFOS in the food chain was also observed (Li et al., 2008). For large wild fish, such as the Chinese Sturgeon, PFOS was the dominant PFAA in eggs, while longer chain C11 to C14 and C16 PFCAs were more accumulated in adult than PFOS (Peng et al., 2010). Age and gender effects were found for certain PFAAs. For example, concentrations of PFOS and Perfluoro-undecanoic acid (PFUdA) were significantly higher in male alligators than those in females (Wang et al., 2013b). Concentrations of PFOS were higher in piscivorous waterbirds than those in the fishes in their diet. Concentrations in eggs of waterbirds were in the range of 14.4–343 ng g 1 dw (Wang et al., 2008). Even in the remote region of the Qinghai-Tibetan Plateau, which is one of the least urbanized and industrialized regions in China, PFOS was detected as the dominant PFAA with mean concentrations ranging from 0.2 to 5.2 ng g 1 dw in fish muscle (Shi et al., 2010). These results indicated widespread distribution of PFAAs from direct emission and through long range transport of volatile precursors with subsequent degradation (Stock et al., 2007). 4.1.2. Risk criteria for protection of aquatic ecosystems Recently, criteria for protection of aquatic organisms in China have been derived for PFOS and PFOA. Data were obtained from toxicity tests, including chronic toxicity tests using three species and acute toxicity tests using nine aquatic species, along with the toxicity information from other studies in China. The criteria maximum concentration (CMC) and criteria continuous concentration (CCC) for protection of aquatic organisms, were calculated to be 3.78 and 0.25 mg L 1 for PFOS, respectively. While for PFOA, the CMC value was 45.54 mg L 1 and the CCC value was 3.52 mg L 1 (Yang et al., 2014). The CMC and CCC values for both PFOS and PFOA in China were significantly higher than values derived in North America, especially for PFOS, where CMC and

CCC values were 21 and 5.1 lg L 1 for PFOS, 25 and 2.9 mg L 1 for PFOA, respectively (Giesy et al., 2010). Given the method used in both studies was the same, the main reason for such significant differences might come from the different native species, which showed different toxic responses to PFOS and PFOA. Although the CMC and CCC values derived in China for PFOS and PFOA are higher than those derived in North America, the actual levels are opposite in the aquatic environments. Historically, discharge from 3 M plants led to high level of PFOS in the aquatic environment. For example, a cave received 3 M plant wastewater contained PFOS up to 18.2 lg L 1 which was higher than the CCC value and even close to the CMC value derived for North America (Oliaei et al., 2013). With strict control over PFOS production and emission, the decreasing trend has already been observed in the environment of the developed countries (Myers et al., 2012; Olsen et al., 2012). While in China, increasing production capacity of fluorinated surfactants and polymers has brought a great amount of PFAAs into the environment (Bao et al., 2011; Wang et al., 2014b; Zhou et al., 2013). Even though none of the reported data has failed the criteria derived for native Chinese species, there is still much need for monitoring the fluorinated polymer industry to make sure that the emission will not pose adverse effect on local aquatic ecosystem in the future. 4.2. Risks of exposure to PFAAs on health of humans 4.2.1. Exposure of humans to PFOS and PFOA It has been found that many PFAAs, like PFOS and PFOA, mainly bind to proteins other than lipids (Li et al., 2013). Until now, blood, usually in whole blood or serum, has been used in biomonitoring of human exposure to PFAAs. In order to make a comparison, concentrations in whole blood and plasma are converted to those in serum by multiplying a factor of 1 and 2, respectively (Ehresman et al., 2007). PFAAs were widely detected in the general population of Chinese people (Zhao et al., 2012b). The spatial distribution of PFAAs concentrations in blood indicated that people living in eastern cities exhibited much higher concentrations and frequency of detection of PFAAs than people living in western cities (Liu et al., 2012). As one of the most developed coastal areas in China, the Bohai-Rim Economic Circle (BREC) is the most frequently studied area, with concentrated fluoropolymers production, textile, paper making and electroplating industries. Generally, PFOS was the dominant PFAA with the most detection frequency, the highest geometric mean (GM) concentrations and the highest maximum concentrations (Liu et al., 2009a; Guo et al., 2011. Shenyang City and Shijiazhuang City exhibited significantly higher levels of PFOS than other cities investigated with GM concentrations of 56.3 and 34.0 ng mL 1, respectively, due to higher emissions (Yeung et al., 2008; Pan et al., 2010). This observation is consistent with the fact that there is still a large amount of PFOS production and emission in BREC, mostly from metal plating industry (Xie et al., 2013b). Concentrations of PFOA were less than those of PFOS in most cities studied, such as Beijing, Jinzhou, Shenyang, Yingkou, Dalian, Huludao, Shijiazhuang, Qingdao, Tangshan and Weihai City, except for Fuxin City and Zouping City. The mean concentration of PFOA in serum of residents in Fuxin was 7.6 ng mL 1, with a maximum of 15 ng mL 1 (Bao et al., 2011). Two fluorine industry parks in the city of Fuxin have been built recently and emission of PFOA has posed a significant impact to local residents and the environment (Bao et al., 2011; Wang et al., 2013c). Concentrations of PFAAs in blood of humans varied among gender and age (Fu et al., 2014). Positive correlations were often observed between age and concentrations of PFOS and PFOA in serum. Generally PFAAs concentrations in blood of males were higher than those of females, the difference was small for most PFAAs except for PFOA, and females exhibited higher concentrations on PFOA serum than males in


9


children less than 15 years of age (Liu et al., 2009a; Bao et al., 2011; Guo et al., 2011; Zhang et al., 2010b). However, inconsistency with the influence of age and gender did exist in studies by different researchers. For example, PFAAs levels in blood except for PFOA in senior people were significantly higher than those in low age people in the population of Liaoning Province, China; whereas in the National Health and Nutrition Examination Survey of the United States (U.S. NHANES), no age effects were observed (Liu et al., 2009a). The reason for the inconsistency might come from multiple factors, including lifestyle, occupation, diet and place of habitation (Liu et al., 2009a). Isomers of PFOS and PFOA were also identified in human serum in Shijiazhuang City and Handan City of north China. A significant difference was observed between profiles of individual isomers of PFOS and PFOA. On average, linear PFOS accounted for 48% of the total PFOS, while linear PFOA accounted for as much as 96% of the total PFOA (Zhang et al., 2013c). The two main processes used to create compounds containing perfluoroalkyl chains produce different isomers. Telomerization produces 100% linear isomer, while electrochemical fluorination produces approximately 70–80% linear and 20–30% branched isomers (Paul et al., 2009). The less proportion of linear PFOS in serum might be used as a biomaker of exposure to PFOS-precursors (Martin et al., 2010). The temporal trend of PFAAs concentrations in three cities of China implied that concentrations of PFOS in serum have decreased, while concentrations of PFOA have increased in recent years. Mean concentration of PFOS in blood of residents in Shenyang City decreased from 112.6 in 2004 to 14.38 ng mL 1 in 2007, while the mean concentration of PFOA increased from nd to 1.96 ng mL 1 (Liu et al., 2009a; Yeung et al., 2008). Similar trends were also observed in residents of Guiyang City and Shijiazhuang City (Pan et al., 2010; Zhang et al., 2013c). The decreasing trend of PFOS observed in China was consistent with those in developed countries, which was due to the phase-out of PFOS and related chemicals since 2000. However, there might be time-lag on the trend in China as production shift from developed countries to China might take years. For PFOA, the increasing trend might be due to the increasing demand for fluoropolymers with PFOA as processing addictive, which has happened in BREC (Wang et al., 2014b; Xie et al., 2013b). In recent years, researchers attempted to explore non-invasive sampling for biomonitoring of PFAAs in humans due to the disadvantages of collecting blood, such as its invasiveness to participants, difficulty in storage and relatively high cost. Almost all PFAAs except for PFOA had similar fingernail/toenail levels or higher fingernail levels than toenail levels. Concentration of PFOS

in fingernail was significantly correlated with concentration of PFOS in serum (Liu et al., 2011b). Another study showed that concentrations of PFOS in nail, hair and urine were significantly correlated with those in serum. A similar relationship was observed between concentrations of PFOA in nail and serum (Li et al., 2013). The relative proportions of PFOS and PFOA levels were also similar between nail (21.2% of PFOA and 78.8% of PFOS) and serum (13.4% PFOA and 86.6% of PFOS). Therefore, nail was suggested as the preferred non-invasive biomarker for monitoring human body loadings of PFAAs (Li et al., 2013). Urine was used for estimation of renal clearance and considered as the major elimination route for short-chain PFAAs (Zhang et al., 2013c) (Fig. 6). 4.2.2. Exposure pathway and risk assessment For the general Chinese population, human exposure to PFAAs is evaluated based on food consumption and drinking water. Levels of PFAAs in various foodstuffs were analyzed, including poultry, livestock and seafood (Wang et al., 2010a; Lu et al., 2011; Chen et al., 2011c; Zhang et al., 2010a). Mean concentrations of PFOS and PFOA in foodstuffs collected from 17 cities of 15 provinces in China were 0.05–1.99 and 0.06–12.5 ng g 1 fresh weight (fwt), respectively (Zhang et al., 2010a). Fish usually contained higher concentrations of PFOS and lower concentrations of PFOA compared with the poultry (Lu et al., 2011), which is probably due to the difference in bioconcentration in aquatic environment or different emission profiles. In different edible parts of poultry and livestock, the total PFAAs accumulated in a decreasing order as follows: liver > kidney > meat (Wang et al., 2010a). Concentrations of PFAAs increased with tropic level, especially in seafood (Chen et al., 2011c). For Chinese adults, the mean total daily intake (TDI) of PFOS and PFOA were calculated to be 1.19 and 9.83 ng kg 1 bw d 1, respectively (Zhang et al., 2010a). The TDI was calculated through multiplication of mean PFOS or PFOA concentrations in certain foodstuff, drinking water and dust inhalation. The foodstuff included meat, animal liver, animal blood, eggs, fish and seafood, which was based on the dietary pattern and nutrition survey in China. Seafood was the main dietary source of PFOS, accounting for 78.9% of the TDI, while meat was the main contributor to dietary exposure of PFOA, accounting for 93.2% of the TDI (Zhang et al., 2010a). Drinking water accounted for as much as 13% of the TDI of PFOA for residents living near fluorochemical plants (Bao et al., 2011). Although exposure from indoor dust might contribute a small portion to the TDI of PFOS and PFOA for the nonoccupational population (Zhang et al., 2010a), it might be a significant source of exposure to PFOS and PFOA or other PFAAs

Fig. 6. Major exposure pathway and biomonitoring of PFAAs for humans in China.

1, 2, 4

Zhang et al., 2010a;

3

Bao et al., 2010.


10


under some extreme conditions, for example, for the workers in fluoropolymer facilities (Wang et al., 2010b). For fetuses and infants, two exposure pathways of PFAAs, including maternal-fetal transmission and breast feeding transmission, have attracted intensive attention in recent years (So et al., 2006b; Tao et al., 2008; Liu et al., 2010; Glynn et al., 2012; Zhang et al., 2013a,d; Nøst et al., 2014). For fetuses, significant positive correlations have been found between concentrations of PFAAs in paired samples of maternal whole blood and placenta, and between placenta and cord blood (Zhang et al., 2013a) (Fig. 6). Effects of carbon chain length on efficiency of transfer of PFAAs from the female to fetus were also evaluated, and a U-shaped trend with increasing carbon chain length was found for C7–C12 PFCAs. The trend could be explained by the opposite ratios between the transfer efficiencies of individual PFAA (C6–C12) from maternal blood to placenta and those from placenta to cord blood (Zhang et al., 2013a). For newborns, a significant correlation was observed among concentrations of PFAAs in matched maternal serum, cord serum and breast milk. Mean concentrations of PFOS and PFOA were 3.2 and 1.66 ng mL 1 in maternal serum, 1.79 and 1.5 ng mL 1 in cord serum, 0.06 and 0.18 ng mL 1 in breast milk, respectively. So the efficiency of transport through both placental barrier and lactation was relatively higher for PFOA than PFOS (Liu et al., 2011a). Concentration of PFOA in breast milk was usually observed at least 10-fold less than that in matched maternal serum (Liu et al., 2010, 2011a). The measured TDI values of PFOS and PFOA for adult or infant were generally far below the criteria values derived by several governmental agencies (So et al., 2006a; Alexander et al., 2008; Roos et al., 2008; Fromme et al., 2009) (Table S3). However, the highest TDI of PFOS (30 ng kg 1 d 1) observed in Zhoushan exceeded the reference dose of 25 ng kg 1 d 1 derived by the environmental working group of U.S. (So et al., 2006b), and the highest TDI of PFOA in Shanghai (88.4 ng kg 1 d 1) was close to the criteria value of 100 ng kg 1 d 1 derived by the German Federal Institute for Risk Assessment and the Drinking Water Commission (Liu et al., 2010). These results suggested that there might be a potential risk of PFOS or PFOA for infants via breast-breeding and also adults via various pathways like diet or drinking water in both cities. TDIs of PFOS for infant (usually used as EDIs, which stand for estimated dietary intakes for infant) in China were less than or comparable to those from other Asian countries (Tao et al., 2008), but higher than certain western countries like Canada, Germany and Spain (Liu et al., 2010). Mean TDIs of PFOA for both adult and infant in China were higher than those from other countries (Tao et al., 2008; Liu et al., 2010), which implied that Chinese population might be exposed to higher PFOA concentrations through daily intake (Table S3).

5. Conclusions and perspectives PFAAs are released to the environment from a number of point sources associated with industrial processes in manufacturing or using PFAAs-related chemicals, or via diffuse widespread consumer use and disposal of PFAAs-containing products. Due to their significant lower vapor pressure and higher water-solubility, PFAAs accumulate and disperse in aquatic environment, so studies on their adverse effects and risks mostly focus on aquatic ecosystems. In general, concentrations of PFAAs especially PFOS and PFOA in aquatic organisms collected near cities in most coastal regions are much higher, due to rapid industrialization and urbanization in recent years, than those in more remote areas in China. This observation indicated that the emission and pollution of PFAAs as well as their distribution are closely related to regional urbanization processes. PFOS, PFOA and related products have been

widely used in industrial and commercial areas over the past several decades, with relatively high concentrations detected in environmental matrixes, biota and even local residents. Studies on PFAAs in China have lagged behind those in more developed countries. Therefore, the foundation for PFAAs studies is frail, especially in risk assessment. Risk assessment of PFAAs is still less developed in China, and most studies have just compared concentrations of PFAAs with the guideline values derived for certain species or receptors to evaluate the risks. Although there seems to be little risk of adverse effects of PFAAs on aquatic environments, risks to organisms at high trophic levels such as piscivorous birds and mammals may be great due to trophic biomagnification. However, studies on biomagnification of PFAAs through food chain is still limited in China. PFOS and its salts have been included in the Stockholm Convention, but exemption allows its continuous production and use in China. China has become one of the largest production and consumption countries to meet the growing demand for surfactants and other surface modification applications. Excitingly, the government and industries have devoted to developing alternatives to replace PFOS, PFOA and related salts, including PFBS and PFBA (Holt, 2011). Based on the previous studies on PFAAs in China, more strategies and legislation for management of PFAAs need to be further developed and strengthened. A specific agency should be designated to investigate these pollutant residues in environment matrixes and different trophic levels on a regular basis. With rapid industrialization and urbanization in China, emission, transport and fate of PFAAs in environment will still be a great concern for ecosystem and human health. However, data gaps on the identification of various sources, industrial and domestic emission, spatial and temporal distribution of typical PFAAs from sources to the environment, toxic effect of certain PFAAs to typical aquatic organisms still exist, which should be taken as priority in the future plan for national actions. Acknowledgements This work supported by the National Natural Science Foundation of China (Nos. 41171394 and 41371488), the Key Research Program of the Chinese Academy of Sciences (No. KZZD-EW-TZ12), and the National Fundamental Field Study Program (No. 2013FY11110). Prof. Giesy was supported by the Canada Research Chair Program and the Einstein Professor Program of the Chinese Academy of Sciences. Appendix A. Supplementary material Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.chemosphere. 2014.09.021. References Ahrens, L., Felizeter, S., Sturm, R., Xie, Z.Y., Ebinghaus, R., 2009. Polyfluorinated compounds in waste water treatment plant effluents and surface waters along the River Elbe, Germany. Mar. Pollut. Bull. 58 (9), 1326–1333. Alexander, J., Auðunsson, G.A., Benford, D., Cockburn, A., Cravedi, J.P., Dogliotti, E., Domenico, A.D., Fernández-Cruz, M.L., Fink-Gremmels, J., Fürst, P., Galli, C., Grandjean, P., Gzyl, J., Heinemeyer, G., Johansson, N., Mutti, A., Schlatter, J., Leeuwen, R.V., Peteghem, C.V., Verger, P., 2008. Perfluorooctane sulfonate (PFOS), perfluorooctanoic acid (PFOA) and their salts. EFSA J. 653, 1–131. Armitage, J., Cousins, I.T., Buck, R.C., Prevedouros, K., Russell, M.H., MacLeod, M., Korzeniowski, S.H., 2006. Modeling global-scale fate and transport of perfluorooctanoate emitted from direct sources. Environ. Sci. Technol. 40 (22), 6969–6975. Armitage, J.M., MacLeod, M., Cousins, I.T., 2009. Modeling the global fate and transport of perfluorooctanoic acid (PFOA) and perfluorooctanoate (PFO) emitted from direct sources using a multispecies mass balance model. Environ. Sci. Technol. 43 (4), 1134–1140.


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Occurrence, Distribution, and Risk Assessment of Perfluoroalkyl Acids (PFAAs) in Muscle and Liver of Cattle in Xinjiang, China.

Spatial distribution and partition of perfluoroalkyl acids (PFAAs) in rivers of the Pearl River Delta, southern China.

Long-chain perfluoroalkyl acids (PFAAs) affect the bioconcentration and tissue distribution of short-chain PFAAs in zebrafish (Danio rerio).

Temporal-spatial distributions and ecological risks of perfluoroalkyl acids (PFAAs) in the surface water from the fifth-largest freshwater lake in China (Lake Chaohu).

Survey of perfluoroalkyl acids (PFAAs) and their precursors present in Japanese consumer products.

Bioaccumulation characteristics of perfluoroalkyl acids (PFAAs) in coastal organisms from the west coast of South Korea.

Dietary exposure to selected perfluoroalkyl acids (PFAAs) in four European regions.

Arsenic contamination of groundwater: a review of sources, prevalence, health risks, and strategies for mitigation.

Spatial Distribution, Sources Apportionment and Health Risk of Metals in Topsoil in Beijing, China.

Accumulation potentials of perfluoroalkyl carboxylic acids (PFCAs) and perfluoroalkyl sulfonic acids (PFSAs) in maize (Zea mays).

Distribution, sources, and ecological risks of organochlorine pesticides in surface sediments from the Yellow River Estuary, China.

Distribution, source characterization and inventory of perfluoroalkyl substances in Taihu Lake, China.

Trihalomethanes in marine mammal aquaria: occurrences, sources, and health risks.

Spatial distribution and fate of perfluoroalkyl substances in sediments from the Pearl River Estuary, South China.

Assessment of Industry-Induced Urban Human Health Risks Related to Benzo[a]pyrenebased on a Multimedia Fugacity Model: Case Study of Nanjing, China.

Hazard assessment of fluorinated alternatives to long-chain perfluoroalkyl acids (PFAAs) and their precursors: status quo, ongoing challenges and possible solutions.

Food adulteration: Sources, health risks, and detection methods.

Field study on the uptake and translocation of perfluoroalkyl acids (PFAAs) by wheat (Triticum aestivum L.) grown in biosolids-amended soils.

Characterizing direct emissions of perfluoroalkyl substances from ongoing fluoropolymer production sources: A spatial trend study of Xiaoqing River, China.

Distribution, possible sources, and health risk assessment of SVOC pollution in small streams in Pearl River Delta, China.

Perfluoroalkyl acids and the isomers of perfluorooctanesulfonate and perfluorooctanoate in the sera of 50 new couples in Tianjin, China.

Level and temporal trend of perfluoroalkyl acids in Greenlandic Inuit.

Occurrence of perfluoroalkyl acids in environmental waters in Vietnam.

Analytical method for biomonitoring of perfluoroalkyl acids in human urine.