State of polybrominated diphenyl ethers in China: An overview

Chemosphere 88 (2012) 769–778

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Review

State of polybrominated diphenyl ethers in China: An overview Jin Ma a,⇑, Xinghua Qiu b,⇑, Jinliang Zhang a, Xiaoli Duan a, Tong Zhu b a b

State Key Laboratory of Environmental Criteria and Risk Assessment, Chinese Research Academy of Environmental Sciences, Beijing 100012, China State Key Joint Laboratory for Environmental Simulation and Pollution Control, College of Environmental Sciences and Engineering, Peking University, Beijing 100871, China

a r t i c l e

i n f o

Article history: Received 1 September 2011 Received in revised form 7 March 2012 Accepted 31 March 2012 Available online 28 April 2012 Keywords: Polybrominated diphenyl ethers Persistent organic pollutants E-waste Exposure China

a b s t r a c t Polybrominated diphenyl ethers (PBDEs), extensively used as flame retardants, are ubiquitous environmental contaminants that are found in both abiotic and biotic environmental samples. Sufficient evidence shows that PBDEs have been rapidly accumulating in the environment of China, especially in the Southeast regions. This paper summarizes and critically reviews the published scientific data on PBDEs in China, including the levels of PBDEs in the air, soil, water, sediment, the terrestrial and marine organisms, and human samples in China. The data preliminarily reveal the state of PBDEs in China. Ó 2012 Elsevier Ltd. All rights reserved.

Contents 1. 2.

3.

4. 5.

6.

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sources of PBDEs in China . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.1. Non-point sources . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.2. E-waste recycling . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.3. Production . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . PBDEs in the environment materials of China . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.1. Air. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.2. Soil . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.3. Water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.4. Sediment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . PBDEs in biological samples of China . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . PBDEs in human samples of China . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.1. Breast milk. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.2. Blood . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.3. Hair . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.4. Tissues . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.5. Semen . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusions. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Acknowledgment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

⇑ Corresponding authors. Tel./fax: +86 10 84916422 (J. Ma), tel.: +86 10 62753184; fax: +86 10 62760755 (X. Qiu). E-mail addresses: [email protected] (J. Ma), [email protected] (X. Qiu). 0045-6535/$ - see front matter Ó 2012 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.chemosphere.2012.03.093

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1. Introduction Polybrominated diphenyl ethers (PBDEs) are brominated flame retardants used in polymeric materials for fire safety of furniture, textiles and electronics (Hites, 2004). There are three major commercial formulations produced and used in the market: pentabromodiphenyl ethers (Penta-BDEs), octabromodiphenyl ethers (OctaBDEs), and decabromodiphenyl ether (Deca-BDEs) (Darnerud et al., 2001). The global demand for PBDEs has increased rapidly since the 1970s. In 1992, the global production of PBDEs was about 40 000 tons, but in 2001 this had increased to approximately 67 000 tons (BSEF, 2006). Worldwide PBDE production was dominated previously by Deca-BDE technical mixture, with an estimated global demand of 56 100 tons in 2001 (BSEF, 2004). Due to their high production volume, widespread usage, and environmental persistence, PBDEs have become ubiquitous contaminants that are found in both abiotic and biotic environmental samples (Wang et al., 2007a,b). PBDEs aroused extensively concern for their persistence, bioaccumulation, and potential adverse effects on wildlife and humans (Hites, 2004). PBDEs are structurally similar to polychlorinated biphenyls (PCBs) and so have similar properties including toxicities (Rahman et al., 2001). Several studies have recently found associations between human body burdens of PBDEs (primarily Penta-BDE) and health effects such as thyroid hormone and androgen abnormalities, cryptorchidism, and low birth weights (Main et al., 2007; Turyk et al., 2008; Meeker et al., 2009). In order to balance potential adverse effects from PBDEs, two of the three commercial PBDE mixtures (Penta-BDE and Octa-BDE) were banned or voluntarily phased out from use beginning in 2002 in many countries or regions of the world (Stapleton et al., 2009), China officially banned the usage of Penta-BDE and Octa-BDE in 2006 (Wang et al., 2007a,b). Deca-BDE is still used around the world, however, the EU has prohibited the use of Deca-BDE in electrical and electronic equipment, and several US States (Vermont, Maine, Washington and Oregon) have prohibited the use of Deca-BDE in consumer products, Deca-mixture production will be phased out in the US by the end of 2012 (BSEF, 2010). In 2009, PBDEs were added to the list of banned chemicals included in the Stockholm Convention on Persistent Organic Pollutants. However, the potential environmental hazard posed by these BDE compounds will not disappear immediately because of new productions with recycled PBDE-containing materials, use of PBDE-containing equipment, and disposal of e-waste (Watanabe and Sakai, 2003). This paper is a critical review of the current status of PBDEs in China, including the sources of PBDEs, PBDEs in environmental materials, biological samples and human samples from the published studies. To our best knowledge, this is the first review on PBDEs in China.

2. Sources of PBDEs in China 2.1. Non-point sources PBDEs could enter the environment through a number of pathways including atmospheric emissions during manufacture, recycling of wastes containing PBDEs, volatilization from consumer products, and leaching from disposal sites (Watanabe and Sakai, 2003). As mentioned above, PBDEs were used as flame retardants in various polymers. In fact, they could migrate from the treated products over their entire lifetimes, owing to there are no chemical bonds between the PBDEs and the polymers (Strandberg et al., 2001). This was a common non-point source of PBDEs in the environment of the world, which was also true for China. For example, old electronic/electrical appliances, especially computers were

found to be the primary sources of PBDEs in indoor air of Guangzhou, China (Chen et al., 2008). Except for the common non-point source, there are still two particular point sources of PBDEs in China, namely e-waste recycling (Leung et al., 2007) and PBDE production (Jin et al., 2011). 2.2. E-waste recycling With newer generations of technology and constant upgrade of the products, the quantities of discarded electronic equipment (socalled e-waste) increase rapidly in the world, especially in developed countries (Robinson, 2009). It was estimated that there were about 50 million tons of e-waste produced annually in the world (Leung et al., 2007). E-waste is chemically and physically distinct from other forms of municipal or industrial waste, it contains both valuable and hazardous materials that require special handling and recycling methods to avoid environmental contamination and detrimental effects on human health. Recycling can recover reusable components and base materials, especially Cu and precious metals. However, due to lack of facilities, high labor costs and tough environmental regulations, rich countries tend not to recycle e-waste. Instead, it is either landfilled, or exported from rich countries to poor countries, where it may be recycled using primitive techniques and little regard for worker safety of environmental protection (Cobbing, 2008). It was reported that e-wastes were transported in massive quantities around the world to developing countries. About 80% of computer e-wastes were exported to Asia, and 90% of these exports were sent to China through illegal imports, for recycling (UNEP, 2005). It was estimated that up to 261 000 tons of PBDEs were imported into Guangdong Province in 2002 in scrap electronic devices (Martin et al., 2004). For example, Guiyu with its surrounding towns in the Guangdong Province of China has become the largest e-waste recycling site around the world. Nearly 80% of families have members who have engaged in e-waste recycling operations (Li et al., 2008). Recycling has been occurring since 1995 in the region (Wong et al., 2007) with crude techniques, including the heating and manual removal of components from printed circuit boards, open burning to reduce volumes and recover metals, and open acid digestion of e-waste to recover precious metals. During the crude recycling of e-wastes, persistent toxic substances such as PBDEs inevitably emitted into the environment, and have created a particular type of PBDEs emission source (Wang et al., 2005). Except for Guiyu, there are other big e-waste cycling sites in China, such as Taizhou (Han et al., 2009) and Qingyuan (Luo et al., 2009). 2.3. Production In China, there were enormous domestic demand for brominated flame retardants (including PBDEs), it was reported that the demand increased at an annual rate of 8% (Mai et al., 2005). A portion of BFRs currently used in China is believed to be imported from other countries (although the exact amount is unknown), because three of the largest BFR manufacturers in the world (i.e., Great Lakes Chemical, Indianapolis, IN; Albemarle Chemical, Richmond, VA; and Dead Sea Chemical, Beer-Sheva, Israel) all have distributors in China to sell BFRs (Mai et al., 2005). Except for import, there were lots of PBDEs production factories in China, mainly in the east of China, especially in Shandong and Jiangsu Province (Jin et al., 2008). The domestic production of BFRs was 10 000 tons in 2000 (Mai et al., 2005). Jiangsu Province is an important domestic BFR production base, and all of the three technical PBDE products are manufactured there. At the same time, as it is a major production center for electronics, textiles and the chemical industry, the annual consumption of PBDEs is very high (JCA, 2006). In 2006, the productions of the technical Deca-BDE

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mixture and decabromodiphenylethane (DBDPE) in China were 15 000 and 11 000 tons, respectively (Hu et al., 2010). The release of PBDEs to the environment can occur during the initial synthesis, incorporation into products. Hence, increasing production inevitably resulted in continuously increasing PBDE levels in the environment (Jin et al., 2011), which have become another particular source of PBDEs in China. Additional information is needed for a more accurate estimation of emissions from these sources.

3. PBDEs in the environment materials of China 3.1. Air The fate, transport, and removal of persistent organic pollutants (POPs) from the atmosphere are strongly influenced by their gasparticle partitioning (Cetin and Odabasi, 2007). Like other semivolatile organic compounds, PBDEs are partitioned between the gaseous and particulate phases in air, and are likely to undergo air surface exchange and long-range atmospheric transport (Wania and Dugani, 2003). Because of their potential for long-range atmospheric transport, world-wide attempts have been made toward air monitoring of PBDEs on different scales from local to global. In China, a few of regional monitoring on BFRs was performed, focusing on the economically developed eastern and southern coastal areas and big cities (Chen et al., 2006a,b; Qiu et al., 2010; Yu et al., 2011). Qiu et al. analyzed for 33 PBDE congeners in air samples from Taihu Lake, South China, found that BDE-209 was most abundant (average 41% of total PBDEs), followed by BDE-47 (17%) and BDE-28 (15%). The annual concentration of total air PBDEs was 220 pg m 3, which was nearly double that in the US urban regions, indicating that the pollution from PBDEs was heavy in this area. In addition, according to the fact that Tetra- and Tri-BDE congeners (including BDE-47, -28, -49, -66, and -17) were relative abundance rather than BDE-99, the author suggested that a specific Penta-BDE formulation might be produced and/or consumed in this region (Qiu et al., 2010). Yu et al. (2011) analyzed PBDEs in air of Shanghai, where conP centrations of PBDEs exhibited the changing trend of industrial area > urban areas, the highest level of particulate PBDEs was 744 ± 152 pg m 3. Compared with similar data in other areas of the world, PBDEs in Shanghai were at medium pollution level. The total air concentration of PBDEs in Beijing was in the range of 57–470 pg m 3 (Hu et al., 2011), in Hongkong was 33.8– 358 pg m 3 (Deng et al., 2007), in the pearl river delta (PRD) was 61–1383 pg m 3 (Zhang et al., 2009), in Fengjiang, Taizhou was 135.1–678.5 pg m 3 (Han et al., 2009). Chen et al. (2006a,b) reported that PBDEs concentrations in urban air of Guangzhou are even close to those extremely high values found in a range of special working environments in Sweden, in an electronics recycling facility (Sjödin et al., 2001). The higher PBDE levels observed at these Chinese cites indicate that PBDEs has been a severe environmental problem in China, especially in certain developed areas like Shanghai, Guangzhou, PRD, Taihu Lake basin, etc. Chen et al. (2008) measured 10 PBDEs (BDE-28, -47, -66, -100, 99, -85, -154, -153, -138, and -183) and BDE-209 in the indoor (home and workplace) and outdoor air in Guangzhou from October P 2004 to April 2005. The PBDEs and BDE-209 concentration ranges were 125.1–2877 and 39–11 468 pg m 3, respectively for home air, 181.3–8315 and 80.1–13 732 pg m 3 for office air, 322.1–2437 and 73.1–8194 pg m 3 for air in other workplaces, and 203.2–2426 and 1082–49 937 pg m 3 for outdoor air. The P mean values of PBDEs for home and workplace microenvironments are slightly higher than those for outdoor air, without a clear P cut indoor–outdoor gradient. Concentrations of PBDEs found in the workplace of Guangzhou are lower than those found in the

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UK: 2300 pg m 3 level (mean of six samples) in workplace samples (Wijesekera et al., 2002). The arithmetic and geometric mean concentration of BDE-209 in domestic environments were 1974 and 338.6 pg m 3 (39–11 468 pg m 3), respectively; in offices, the values were 2448 and 341 pg m 3 (80.1–13 732 pg m 3), respectively, in other workplaces, the values were, respectively, 1569 and 668.7 pg m 3 (73.1–8194 pg m 3); in outdoor environments, the results were 13 192 and 7481 pg m 3 (1082–49 937 pg m 3). The concentrations of BDE-209 in outdoor air were higher than in domestic and workplace microenvironments. The primary indoor emission sources for PBDEs in Guangzhou are suggested to originate from the relatively old electronic/electrical appliances, especially computers. In addition, China has several BFR (including PBDEs) manufacturing plants mostly located in the east of China, especially in Shandong and Jiangsu Provinces (Jin et al., 2008). Actually, PBDEs concentrations in air of PBDEs production areas were much higher than those economically developed areas (Jin et al., 2011). Up to now, there were few studies on PBDEs levels in air of production areas. Recently, a survey has been carried out on the levels of PBDEs in air of Laizhou Bay, a major PBDE production area in China (Jin et al., 2011). A total of 46 air samples were analyzed, the conP centrations range for PBDEs was 0.017–1.17 ng m 3 in the gaseous phase and 0.5–161.1 ng m 3 in the particulate phase. The concentrations in the particulate phase were higher than those in the gaseous phase among all samples, and higher brominated congener class was mainly found in the particulate phase. For all samples, the observed concentrations at the production source emission site were significantly higher than other sampling sites. It is worth noting that the concentrations of PBDEs in outdoor air in this study were similar with that of indoor air of an electronics recycling facility (as high as 650 ng m 3 for BDE-209) in US (Cahill et al., 2007), indicating that the air of production area was significantly polluted. Except for PBDEs production areas, PBDEs concentrations in air of e-waste recycling sites were also extremely high (Chen et al., 2011). It was reported that in Guiyu town the mean air concentration of BDE-47 was 2748 pg m 3 in summer and 6146 pg m 3 in winter, BDE-99 was 1656 pg m 3 in summer and 4911 pg m 3 in winter (Chen et al., 2011). The average concentration of BDE-47 in Guiyu was even about 2–5 times of that in the recycling hall of the recycling electrical plant of Sweden (Sjödin et al., 2001). 3.2. Soil Soil is a major reservoir and sinks for organic pollutants because of its sorption quality and holding capacity (Dalla Vallea et al., 2005). In contrast to other POPs, such as PCBs, PAHs or OCPs, only very limited data are available for PBDE contamination in soil (Zou et al., 2007). Extremely high levels of PBDEs contamination have been reported in soil at e-waste recycling sites of Guiyu (Leung et al., 2007) and Qingyuan (Luo et al., 2009) in China. In Qingyuan, the total PBDE concentrations ranged from 191 to 9156 ng g 1 dw in road soils and from 2.9 to 207 ng g 1 dw in farmland soils 2 km from an e-waste recycling workshop, respectively. In Guiyu, the total PBDE concentrations in soils were in the range of 2720– 4250 ng g 1 dw. In addition, PBDEs were also measured in soil samples from a PBDE production area of Laizhou Bay in China, it was reported that the concentrations ranged from 73– 2629 ng g 1 dw (Jin et al., 2011). It was interesting to find that BDE-209 was the most dominant congener among e-waste recycling sites and production area, indicating the prevalence of commercial Deca-BDE. However, signature congeners from commercial Penta- and Octa-BDE were also found in e-waste sites. Except for typical e-waste recycling sites, PBDEs were also found in urban soils in China, and some were in high levels. For

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example, a study conducted in Shanghai shows that the concentraP tions of PBDEs in urban soils ranged from 23.6 to 3797 ng kg 1 with a mean value of 735 ng kg 1, BDE-209 was the predominant congener in a range of 0.291–2910 ng kg 1 with the mean of 477 ng kg 1 (Jiang et al., 2010a,b). Sun et al. (2009) studied the status of PBDEs in soils irrigated by sewage or wastewater in the east of Beijing, where PBDEs concentrations ranged from 501.9 to 3310.7 ng kg 1 dw with the mean of 1538 ng kg 1. According to Zou et al. (2007), the concentrations of RPBDEs in the surface soils from PRD ranged from 130 to 3810 ng kg 1 with an average of 1020 ng kg 1. Compared to those big cites or developed regions, PBDEs levels in soils of Taiyuan (a city in central of China) and Harbin (a city in northeastern China) were rather low, with the total PBDEs concentrations ranged from 0.016 to 211.416 ng g 1 and from 2.45 to 55.9 ng kg 1 dw with a mean of 26.3 ng kg 1 dw, respectively (Li et al., 2008; Wang et al., 2009a,b,c). PBDEs levels in those two cities were not only lower than those in other areas in China, but also lower than those in background soils from the United Kingdom and Norway (Hassanin et al., 2004). It was obvious that PBDEs concentrations in soils of different cities and areas of China varied greatly, which may be attributed to the degree of urbanization and the distribution of industrial plants in these cities. In addition, in all of the Chinese cities and areas mentioned above, BDE-209 was the predominant congener, which was in agreement with the fact that commercial decabromodiphenyl ether (DeBDE) mixtures were dominant technical PBDEs mixtures used in China (Jiang et al., 2010a,b). PBDEs have many physicochemical properties of POPs, for example, they can undergo long-range atmospheric transport from polluted environments to pristine areas far away from any anthropogenic activities (Wang et al., 2009a,b,c). PBDEs were also detected in soils of the Tibetan Plateau, the highest plateau in the world. The concentrations ranged from 4.3 to 34.9 ng kg 1 dw with an average at 11.1 ng kg 1 dw. BDE-47 was the main congener, accounting for about 40% of PBDEs, while BDE-209 was not detected in any of the samples. This could be explained by the fact that higher brominated congeners (i.e., BDE-206, BDE-207, BDE208, and BDE-209) were bound to the atmospheric particulate phase, while lower brominated congeners (i.e., BDE-28, BDE-47) were bound to atmospheric gaseous phase due to their different physicochemical properties. Thus, lower brominated congeners are more propitious for long-range migration than higher brominated congeners. 3.3. Water Recently, the distribution and partition of PBDEs in water from the Pearl River Estuary were reported (Chen et al., 2011). The P PBDEs concentrations, referring to the sum of BDE-28, -47, -66, -99, -100, -153, -154, and -183 in the dissolved phase, ranged from 2.15 to 127 pg L 1 with a mean of 29.0 pg L 1. Dissolved PBDE concentrations in the flood season (with a mean of 13.6 pg L 1 in May 2005 and mean of 5.80 pg L 1 in July 2006, respectively) were lower than concentrations in the dry season (mean of 69.4 pg L 1 in October 2005 and 27.0 pg L 1 in July 2006, respectively). This difference can be attributed to the dilution of a large amount of fresh water flowing upstream from the Pearl River because the rainfall was higher in the flood season than in the dry season. The concenP trations of PBDEs and BDE-209 in the particle phase were between 6.20 and 77.6 pg L 1 and between 0.27 and 5690 pg L 1, respectively. BDE-209 was the dominant congener in the particle phase, comprising approximately 95% of the total PBDEs. It was believed that there was an exponential increase in the concentrations of PBDEs in the marine environment over the past two decade (Ikonomou et al., 2002). Wurl et al. (2006) determined

prevailing concentrations of PBDEs in the sea-surface microlayer and their enrichment relative to bulk seawater in the coastal environment of Hong Kong, China. Samples were collected in March 2005 at five sample locations and analyzed for eight congeners of primary interest, i.e., BDE-28, -47, -99, -100, -153, -156, -183 and P -209. Concentration ranges of PBDEs in the dissolved phase (DP, defined as sum of truly dissolved and colloidal phase) and suspended particulate matter of seawater were 31.1–118.7 pg L 1 (mean 70.7 pg L 1), and 25.7–32.5 pg L 1 (mean 28.1 pg L 1), respectively. Concentrations of PBDEs were general low and below detection limits in samples of an oceanic character and highest in the sheltered waters of Victoria Harbor. The congeners BDE-28, 47, -100 and -183 were most abundant, where BDE-209 was detected only in trace amounts. It was suggested that Hongkong’s marine waters show relatively low levels of PBDE contamination, and these compounds may originate from the disposal of electronic waste in South China, as well as untreated discharge of wastewater locally. 3.4. Sediment Chen et al. (2006a,b) collected 32 surface sediment samples from the intertidal zone of the Yangtze River Delta (YRD), of which 12 were taken within and outside of the YRD in April 2002, 15 from Hangzhou Bay, and five from the Qiantang River adjacent to the city of Hangzhou City in July 2004. 13 PBDEs congers were measured in those samples, i.e., BDE -7, -11, -15, -17, -28, -47, P -66, -100, -99, -154, -153, -183, and -209. The PBDEs (sum of PBDE congeners except for BDE-209) concentrations varied from non-detectable (at three sites) to 0.55 ng g 1 dw with a mean value of 0.15 ng g 1. The BDE-209 concentrations ranged from 0.16 to 94.6 ng g 1 with an average of 13.4 ng g 1, and were 1–3 orders P of magnitude higher than those of PBDEs. Relatively high conP centrations of PBDEs (0.1–0.55 ng g 1) were observed in the Qiantang River and along the southern shore of the YRD, followed by those found around the upper Hangzhou Bay. The lowest P PBDEs (non-detectable-0.01) concentrations occurred in the north shore of outer Hangzhou Bay, and they were even lower than those outside the YRD. P A study conducted in Bo sea, China revealed that PBDEs (including BDE-17, -28, -47, -66, -71, -85, -99, -100, -138, -153, -154, -183 and -190) and BDE-209 ranged from 0.07 to 5.24 ng g 1 dw (median 0.16 ng g 1), 0.30–2776 ng g 1 with a median value of 2.29 ng g 1 in sediment, respectively. The four most abundant congeners were BDE-47 (40.3%), -99 (22.5%), -71 (8.9%) and -28 (5.8%) in sediment (Wang et al., 2009a,b,c). Zhao et al. (2011) investigated the concentrations, compositional profiles, possible sources of PBDEs in sediments of the Daliao River Estuary. The total concentrations of PBDEs (BDE-15, -28, -47, -49, -66, -99, -100, -153, -183 and -209) ranged from 0.13 to 1.98 ng g 1 dw, with a mean value of 0.63 ng g 1 dw. P The PBDEs (BDE-28, -47, -99, -100, -153, -154, -183, -206, -207, -208, and -209) concentrations in river sediments from Laizhou Bay in China were ranged from 1.3 to 1800 ng g 1 dw (Jin et al., 2008). There were significant positive correlations for BDE-28 and BDE-100 (r = 0.945, p < 0.01), BDE-47 and BDE-99 (r = 0.879, p < 0.01), BDE-153 and BDE-154 (r = 0.934, p < 0.01), Nona-BDEs and BDE-209 (r = 0.934, p < 0.01). PBDEs concentrations in surface sediments from Baiyangdian Lake and its inflowing river (Fuhe River) in North China were P investigated (Hu et al., 2010). The concentrations of PBDEs1 (including BDE-28, -47, -99,-100, -153, -154, and -183), OctaBDE, Nona-BDE, and Deca-BDE in sediments of Fuhe River were in the range of 0.13–6.39, 0.27–2.92, 5.07–34.9, and 11.8– P 292.7 ng g 1 dw, respectively. The concentrations of PBDEs1, Octa-BDE, Nona-BDE, and Deca-BDE in sediments of Baiyangdian

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Lake were in the range of 0.05–5.03, 0.44–0.75, 2.10–4.19, and 4.35–19.3 ng g 1 dw, respectively. The levels of PBDEs in Fuhe River were significantly higher than those in Baiyangdian Lake (p < 0.05). For the lowly brominated congeners (tri- to heptaBDE), BDE-47 and BDE-99 were the most abundant, which contributed 52.1% and 44.1% to the sum of tri- to hepta-BDEs in the sediments from Baiyangdian Lake and Fuhe River, respectively. The compositional patterns of PBDEs in Baiyangdian Lake sediments indicated that technical Deca-BDE mixture was the major pollutant sources with a minor contribution of Penta-BDE mixture. In August through October 2002, Mai et al. gathered 66 surface sediment samples from PRD and adjacent South China Sea and analyzed 10 PBDE congeners (BDE-28, -47, -66, -100, -99, -154, -153, P 138, -183, and -209). The concentrations of BDE-209 and PBDEs (defined as the sum of all targeted PBDE congeners except for BDE-209) ranged from 0.4 to 7340 and from 0.04 to 94.7 ng g 1, respectively. Up to now, this was the highest value of BDE-209 in sediment been reported around the world. Analyses of two short sediment cores collected from the Pearl River Estuary showed that concentrations of BDE-209 rapidly increased in the upper layers of both cores, coincident with the growth of the electronics manufacturing capacities in the PRD region (Mai et al., 2005). It was worth noting that although PBDEs concentrations varied in sediments from different river/lake in the mainland of China, BDE-209 was always the most predominant congener, with contributions to the total PBDEs exceed 70% in all studies mentioned above (Mai et al., 2005; Chen et al., 2006a,b; Jin et al., 2008; Hu et al., 2010; Zhao et al., 2011). This was consistent with the fact that technical Deca-BDE mixtures are presently the dominant technical PBDE mixtures used in China. However, the phenomenon was not observed in the sediment from Hongkong. Liu et al. (2005) measured the concentration of PBDEs in 13 sediment samples P taken from Hong Kong marine waters, the PBDEs in sediments 1 ranged between 1.7 and 53.6 ng g dw, with the highest concentrations located around the most heavily populated areas of Victoria Harbour and Sai Kung. It should be noted that the concentrations of PBDEs in sediment in Hong Kong are higher than those in other areas, but the concentration of BDE-209 was only 2.92 ng g 1 dw, far lower than the level at other locations, which was contrary to the PBDEs profile in sediment in PRD (Mai et al., 2005). Hence, PBDEs in the sediment in Hongkong could not completely attribute to PBDEs discarded from PRD, and there may be other sources of PBDEs.

4. PBDEs in biological samples of China PBDEs have been of environmental concern for their bioaccumulation and toxicities (Hites, 2004), and lots of studies on PBDEs in animals have been undertaken since the mid-1980s (Lawa et al., 2003). In China, studies were mainly focus on aquatic species, which are good bio-indicators of environmental pollution because they concentrate bioaccumulative pollutants in their bodies from water, and sediment, in additional to uptake from diet (Wang et al., 2007a,b). Qin et al. (2011) analyzed PBDEs in chicken tissues and eggs from wenling, Taizhou, southeast coast of China. As well as Guiyu of Guangdong Province, the region near Wenling is one of the largest e-wastes recycle process areas. The mean PBDEs concentrations in tissues ranged from 15.2 to 3138.1 ng g 1 lipid weight (lw) and in egg the concentration was 563.5 ng g 1 lw. The results showed P that the level of PBDEs in the chicken tissue was 2–3 orders of magnitude higher than those reported in the literature. BDE-209 P was the predominant congener (82.5–94.7% of PBDEs) in all P chicken tissues except in brain (34.7% of PBDEs), which indicated that Deca-BDE was major pollution source in this area. The dietary

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PBDEs intake of the local residents from chicken muscle and egg, assuming only local bred chickens and eggs were consumed, ranged from 2.2 to 22.5 ng (d kg) 1 with a mean value of 13.5 ng (d kg) 1. E-waste contaminants can enter aquatic systems via leaching from dumpsites where processed or unprocessed e-waste may have been deposited. Luo et al. (2007) reported that carp from the Nanyang river, near Guiyu, were bioaccumulating PBDEs to concentrations of up to 766 ng g 1. Jiang et al. (2010a,b) measured 39 PBDEs congeners in the muscle tissue of three species of fish (Sciaenops ocellatus, Sparus macrocephalus, and Lateolabrax japonicus) and four species of shellfish (Tegillarca granosa, Cyclina sinensis, Sinonovacula constricta, and Ostrea cucullata) that were collected downstream of e-waste recycling plants in Taizhou, China. The P PBDEs ranged from 545.4 to 1688.7 ng kg 1 ww, and the mean concentration was 1382.6 ng kg 1 ww in fish and 858.1 ng kg 1 ww P in shellfish. The mean PBDEs level in all species of fish was at least two times higher than that in the majority of the shellfish species, except S. constricta. These differences may be explained by the feeding ecology of the organisms (Guo et al., 2007). The fish typically feed on small fish, shrimp, and shellfish whereas the shellfish feed primarily on diatoms and organic debris. In 2005, Jin et al. began monitoring for PBDEs in shellfish samples surrounding the PBDEs manufacturing plants in Laizhou Bay of Shandong Province in P China. It was reported that the PBDEs concentrations in shellfish samples from Laizhou Bay were in the range of 230–720 ng g 1 lipid. BDE-209 was the predominant congener in all analyzed samples, consistent with the fact that Deca-BDE technical mixtures are the dominant PBDEs product in Laizhou Bay (Jin et al. 2008). PBDEs were also detected in frogs in Taizhou. Liu et al. (2011a,b,c) collected 58 adult frogs (Rana limnocharis) from a rice field in Taizhou, China in the early May in 2009 and the early May in 2010, and detected PBDEs in the whole frogs and various tissues (brain, liver, testis and egg), in which PBDEs concentrations ranged from 17.10 to 141.11 ng g 1 ww. BDE-209 was the predominant congener in various tissues, followed by BDE-47 and BDE-28. From 2004 to 2007, Gao et al. collected sixteen species of aquatic biota from the lower reach of the Yangtze River including fishes, crabs and shrimps and analyzed for 13 PBDE congeners. All the PBDE congeners except BDE-17 were detectable in the P samples, the PBDEs ranged from 3.52 to 603.69 ng g 1 lipid (0.032–62.69 ng g 1 ww), with a mean of 44.04 ng g 1 lipid (2.69 ng g 1 ww). The predominant congeners were BDE-47, BDE-28, BDE-154, BDE-100 and BDE-153 (Gao et al. 2009). Xian et al. (2008) reported that the concentrations of PBDEs in muscle of freshwater fishes from the Yangtze River ranged from 18 to 1100 ng g 1 lipid ww. BDE-15, BDE-28 and BDE-47 were the predominant congeners in the fishes. It should be noted that the two studies in Yangtze River both present particular congener profile in aquatic biota, indicating a specific commercial PBDE formulation (probably made in China) might have been used in the Yangtze River Delta region. In addition, it was found that the PBDE levels in liver and eggs were significantly higher than that in muscle in fish, which was in agreement with a previous study (Covaci et al., 2005). Liu et al. (2011a,b,c) measured PBDEs levels in marine fish from four areas of China (South China Sea, Bohai Sea, East China Sea, and P Yellow Sea), PBDEs (BDE-28, -47, -99, -100, -153, -154, -183 and -209) in all samples ranged from 0.3 ng g 1 ww to 700 ng g 1 ww, with median and mean values of 85 ng g 1 ww and 200 ng g 1 ww, respectively. BDE-209 and BDE-47 were the major congeners in all samples, contributing 54% and 19% to the total concentration, P respectively. The average level of PBDEs in samples from South China Sea, Bohai Sea, East China Sea, and Yellow Sea were 0.8, 36, 375, 388 ng g 1 ww, respectively. PBDE levels in marine fish from East China Sea and Yellow Sea were notably higher than those

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in the other two areas. This may be connected with e-waste cycling and production in Zhejiang and Shandong Province. From 2003 to 2004, 15 fish samples were gathered in the PRD and Daya Bay, and analyzed for PBDEs levels. The total PBDE concentrations in biota were much higher in the PRD than in Daya Bay. The median total PBDE concentrations of 19 fish samples collected from the estuary were 1.9 ng g 1 ww compared with 0.1 ng g 1 ww of 22 fish samples from Daya Bay. BDE-47 was the predominant PBDE congener. It is noteworthy that BDE-209 concentrations were relatively low, comprised only <3% of the total PBDEs in the fishes. The result was in consistent with those in the Yangtze River Delta, while contrary to those in Taizhou and four areas of China (South China Sea, Bohai Sea, East China Sea, and Yellow Sea), indicating that there were different sources of PBDEs input in those areas above mentioned. PBDEs were also found in eggs of waterbirds in South China (Lam et al., 2007), concentrations of total PBDEs in Ardeid eggs from Hong Kong, Xiamen and Quanzhou ranged from 140–1000, 30–550 and 140–380 ng g 1, lipid, respectively. Marked concentrations of higher-brominated diphenyl ethers (BDE-183, -196, -197, -206, -207 and -209) were detected in the egg samples from all sites, probably indicating uptake of BDE-209 by top predators, and that there is widespread occurrence of BDE-209 in South China. A study conducted in typical freshwater cultured fish ponds of South China found that tri- to Deca-BDEs were detected in all samples analyzed, with mean concentrations of 21–20 ng g 1 lipid in fish. BDE-47 was predominant in fish samples, whereas BDE-209 not. It was suggested that fish feed, as well as pond water at a lesser degree, may have been the major source of PBDEs in freshwater farmed fish (Zhang et al., 2010). As mentioned above, PBDEs can undergo long-range transport, potentially impacting remote alpine ecosystems and species. Remote high mountains with low temperature can act as condensers for selected atmospheric contaminants (Wania and Westgate, 2008). The Tibetan Plateau is the largest and highest plateau in the world, the lakes on the Plateau may be especially sensitive to atmospherically deposited contaminants because of the thin soil and sparse vegetation have little retention of organic chemicals. Yang et al. (2011) gathered 60 fish samples were collected from eight alpine lakes across the Tibetan Plateau and analyzed PBDEs levels in those samples. Concentrations of PBDEs in fish were in the range of 0.01–2.1 ng g 1 ww with a mean of 0.13 ng g 1 ww, which was lower than those reported for European mountains (range, 0.07– 1.1 ng g 1 ww; mean, 0.38 ng g 1 ww), Canada Rocky mountain (range, 0.17–52 ng g 1 ww; mean, 5.6 ng g 1 ww) and Western US Parks (range, 0.18–5.7 ng g 1 ww; mean, 1.1 ng g 1 ww) (Vives et al., 2004; Demers et al., 2007; Ackerman et al., 2008).

5. PBDEs in human samples of China 5.1. Breast milk PBDEs make their way into the human body primarily through food intake, and ingestion of dust through inhalation (Hites, 2004). For breastfeeding infants, human milk has been estimated as the main exposure source of PBDEs (Roosens et al., 2010). Previous studies indicated that PBDEs concentrations in human milk have increased exponentially over the last 2–3 decades (Hites, 2004). The fetus and developing children are far more sensitive than the adults to the effects of many chemicals. Thus, the accumulation of organic contaminants including PBDEs in human milk has been a matter of growing concern in the world. In 2007, a national survey of PBDEs levels in human milk was carried out for the first time to evaluate the background body bur-

den of general population and breastfeeding infant exposure in China (Zhang et al., 2011). Seven PBDE congeners (BDE-28, BDE47, BDE-99, BDE-100, BDE-153, BDE-154 and BDE-183) were measured in 24 pooled human milk samples comprised of 1237 individual samples from 12 provinces in China. The concentrations of P PBDEs ranged from 0.85 to 2.97 ng g 1 lipid with the mean of 1.49 ng g 1 lipid. BDE-28, BDE-47 and BDE-153 were predominant P PBDE congeners accounting for nearly 70% of PBDEs. No significantly statistical relationship were observed between concentrations of PBDEs in milk samples and maternal age, dietary habits as well as geographical locations (costal or inland, rural or urban), indicating that Chinese general population is probably exposed to PBDEs with multiple sources and pathways. The estimated daily intakes of BDE-47, BDE-99 and BDE-153 for the Chinese nursing infant were much lower than corresponding threshold reference values suggested by USEPA. Compared to previous studies around the world, PBDEs levels observed in this study are slightly lower than those from Sweden (Lind et al., 2003), UK (Kalantzi et al., 2004), Denmark and Finland (Main et al., 2007), Italy (Ingelido et al., 2007), Spain (Schuhmacher et al., 2007), Korea (Haraguchi et al., 2009) and France (Antignac et al., 2009), and much lower than that from Australia (Toms et al., 2007), Canada (She et al., 2007) and USA (Dunn et al., 2010), comparable to those from Norway (Thomsen et al., 2010), Japan (Haraguchi et al., 2009), Poland (Jaraczewska et al., 2006), Indonesia (Sudaryanto et al., 2008), Germany (Vieth et al., 2005) and Czech Republic (Kazda et al., 2004), and only higher than that from Vietnam (Haraguchi et al., 2009). 5.2. Blood Bi et al. (2006) analyzed PBDEs levels (BDE-28, -47, -99, -100, 153, -154, -183) in 21 paired human fetal and maternal serum in P South China, found that the concentrations of PBDEs varied from 1 1 1.5 to 12 ng g lipid (median 3.9 ng g lipid), 1.6–17 ng g 1 lipid (median 4.4 ng g 1 lipid) in fetal and maternal serum, respectively. BDE-47 and -153 were the most abundant congeners. BDE-47 accounted for 35.4 ± 11.9%, 30.1 ± 10.1% of the total mass of PBDEs in fetal and maternal serum, respectively; while BDE-153 accounted for another 22.8 ± 11.9%, 29.4 ± 12.2%, respectively. PBDEs levels in this study were similar to those found in Korean municipal waste incinerators workers (sum of BDE-28, -47, -99, -100, 153, -154, -183, mean 19.3 ng g 1 lipid) (Kim et al., 2005), but much lower than those in human blood from Guiyu, the mean PBDEs concentration as high as 91 ng g 1 lipid in adult serum (Bi et al., 2007), which was similar to those in human milk from the US and Canada (96 ng g 1 lipid, (She et al., 2007)). Qu et al. (2007) studied PBDEs levels in serum of e-waste workers and other residents from Guiyu, the result showed that the median serum PBDEs concentrations were 126 ng L 1 and 35 ng L 1, respectively, compared to referents from a nearby town who had just 10 ng L 1. In addition, there were two studies on PBDEs levels in children’s blood gathered from Dalian (Chen et al., 2010) and e-waste recycling sites of Zhejiang Province (Shen et al. 2010). Children appear to be particularly suitable for such monitoring programs, as they are not directly exposed to occupational pollution; thus, children normally reflect present trends in environmental exposure more accurately than adults (Pérez-Maldonado et al., 2009). Chen et al. (2010) studied the levels and patterns of PBDEs in children’s plasma from Dalian, China. 17 PBDE congeners (BDE-30, -28, -35, -37, -75, -47, -66, -100, -99, -116, -155, -154, -153, -183, -181, -190 and -209) in 29 plasma samples were measured. PBDEs concentrations were in the range of not detected to 188.37 ng g 1 lipid with the mean of 40.08 ng g 1 lipid. BDE-153 was the dominant congener, followed by BDE-99, -47, and -183. No significant difference in PBDE concentrations was observed between male (n = 15)

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and female (n = 14) children’s plasma. In addition, Shen et al. (2010) analyzed concentrations and congener specific profiles of PBDEs in 22 pooled blood samples representing 66 children (age: 5–11 years; boys: n = 36; girls: n = 30) in Luqiao (n = 21) (e-waste recycling region), and Longyou (n = 18) and Tiantai (n = 27) (control areas) in Zhejiang Province, China. The highest concentration P (32.1 ± 17.5 ng g 1 lipid) of PBDEs was detected in Luqiao (range 1 from 8.14 to 65.2 ng g lipid), which was significantly (p < 0.001) higher than that in Longyou (12.1 ± 7.59 ng g 1 lipid) and Tiantai (8.43 ± 3.99 ng g 1 lipid). In terms of PBDE congener patterns, among all 15 detectable congeners in this study, PBDE-99, -47, -153, -154, -15 and -183 were the main congeners, accounting for 69.4% of total PBDEs. It was worth noting that although the children’s blood PBDE levels in Dalian and Zhejiang were much lower than those in human blood in Guiyu (Bi et al., 2007), but were higher than those found in maternal serum in South China (Bi et al., 2006). 5.3. Hair Human hair has been considered as a suitable non-destructive indicator for the study of environmental or occupational exposure to many chemical pollutants (Schramm, 1997). In China, only a few studies reported PBDEs concentrations in human hair, and the results were varied greatly. Zhao et al. gathered 48 hair samples from residents (including 44 villager workers engaged in primitive recycling operations from the disassembly sites, and four local residents from the control site) around an e-waste disassembly site in Zhejiang and analyzed 12 PBDE congeners. The highest level of PBDEs (29.64 ng g 1 dw) in hair sample was found in the disassembly site Xinqiu, which are two times more than those observed in hair from the control site (Zhao et al., 2008). The result was much lower than that in another study (Ma et al., 2011), in which hair samples collected from e-waste recycling workers (n = 23 males, n = 4 females) were analyzed to assess occupational exposures to PBDEs, and hair samples from a reference population composed of residents of Shanghai (n = 11) were analyzed for comparison. The mean concentration of PBDEs (range, 22.8–1020 ng g 1 dw; mean, 157 ng g 1 dw) found in hair samples from e-waste recycling workers was approximately three times higher than the mean P determined for the reference samples. High PBDEs concentra1 tions (range, 18.1–9400 ng g ; mean, 871 ng g 1 dw for eight PBDE congeners) was also reported previously in hair samples from male e-waste recycling workers from Taizhou (Wen et al., 2008). Kang et al. collected 18 hair samples from Hongkong and analyzed some PBDE congeners, BDE-47, -99, -100, and -183 deP tected in most samples. PBDEs was ranged from 1.49 to 1 7.45 ng g with the median and mean of 3.40, 3.71, respectively. BDE-47 and BDE-99 were the dominant congeners detected in human hair samples ranging from 0.86 to 5.24 ng g 1 with a median of 2.25 ng g 1 and 0.22 to 1.47 ng g 1 with a median of 0.54 ng g 1, respectively (Kang et al., 2011). Obviously, PBDEs concentrations in human hair in Hongkong were much lower than those found in ewaste recycling workers who are occupationally exposed to elevated levels of PBDEs in the mainland of China (Zhao et al., 2008; Ma et al., 2011). It was reported that there were evident differences in the congener profiles of PBDEs in hair from e-waste recycling workers and general people (Ma et al., 2011). The congener profiles of PBDEs in hair from e-waste recycling workers were dominated by BDE-209, for example, in the e-waste recycling site in eastern China, BDE209 accounted for 82.1 ± 16.5% of the total PBDE concentrations in hair for males, and 92.3 ± 5.5% for females. The next commonest congeners, BDE-47 and BDE-99 accounted for 82.1 ± 16.5% of the total PBDE concentrations for males, and 92.3 ± 5.5% for females. Whereas the profiles in the reference-population samples showed

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comparable levels of BDE-47 (44.6 ± 25% for males and 34.9 ± 11.6% for females) and BDE-209 (43.5 ± 30.5% for males and 46.2 ± 19.1% for females). The difference in the apportionment of BDE-47 and BDE-209 between e-waste workers and non-ewaste workers suggests elevated exposures to BDE-209 (i.e., Deca-BDE) in e-waste recycling workers (Ma et al., 2011). 5.4. Tissues Data of PBDEs in human tissues are very limited in China. Recently, Ren et al. (2011) measured 6 PBDE congeners (BDE-47, -66, -99, -100, -153, and -154) in 130 placenta samples (cases: n = 80, control: n = 50) collected from Shanxi Province, China, the concentrations ranged from 0.37–0.94 ng g 1 lipid with the median 0.54 ng g 1 lipid and from 0.34–0.79 ng g 1 lipid with the median 0.54 ng g 1 lipid in case samples and control samples, respectively. The concentrations in this study were at least ten times lower than those in studies conducted in Europe and the United States (Gómara et al., 2007; Frederiksen et al., 2009; Dassanayake et al., 2009). Another report was on PBDEs burden in human tissues from cancer patients living in the e-waste disassembly sites (Zhao et al., 2009). In this study, 12 PBDE congeners were detected in P kidney, liver, and lung samples, the results showed PBDEs except for BDE-209 in kidney, liver, and lung samples were in the range of 56.49–410.61 ng g 1 with the median concentration of 186.09 ng g 1, 47.69–541.10 ng g 1 with the median concentration of 192.61 ng g 1, 64.42–333.70 ng g 1 with the median concentration of 171.96 ng g 1, respectively. Concentrations of BDE-209 in kidney, liver, and lung samples were in the range of 1.00– 72964.52 ng g 1 with the median concentration of 191.27 ng g 1, 1.00–61624.65 ng g 1 with the median concentration of 118.10 ng g 1, 1.00–38887.60 ng g 1 with the median concentration of 270.00 ng g 1, respectively. There was a comparable PBDEs level among these collected kidney (182.3 ng g 1 lipid), liver (174.1 ng g 1 lipid) and lung tissue samples (174.2 ng g 1 lipid) (p > 0.05). BDE-47 and BDE-28 were the most predominant PBDE congeners, accounting for >17% and >10% of the total PBDEs observed in the collected samples. BDE-209 were detected in >70% of the samples. PBDEs concentrations in liver samples of this study (174.1 ng g 1 lipid) were much higher than those observed in Belgium (3.6 ng g 1 lipid) (Covaci et al., 2002), Sweden (4.5–18 ng g 1 lipid) (Meironyté et al., 2001) and USA (23.1 ng g 1 lipid) (Schecter et al., 2007). 5.5. Semen Recently, PBDEs were found for the first time in human semen samples (n = 101) from Taizhou, China (Liu et al. 2011a,b,c), with concentrations varied from 15.8 to 86.8 pg g 1 ww (median 31.3 pg g 1 ww), which was about two times lower than those in human blood samples. A correlation of RPBDEs concentration was found between paired semen and in blood. The results suggest that semen could be used to detect PBDE burden in human body as a non-invasive matrix. In addition, the levels of BDE-209 and BDE-153, especially the latter, were much higher in blood than in semen, while the levels of BDE-28, BDE-47 and BDE-99 were comparable in the two matrices, suggesting that low brominated congeners could be more easily transferred to semen than high brominated congeners. 6. Conclusions From analysis above, it was obvious that PBDEs has become a new pollutants are of great concern in China, typically in southeast

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region. E-waste recycling and PBDEs production have heavily contaminated the circumjacent environment, resulting high exposure of these pollutants and serious health risks to the workers and residents, especially for the children. Restricts on e-waste recycling and PBDEs production are needed urgently to manage environmental risks. Up to now, researches on PBDEs in China are very limited and mainly restricted to southeast region, where e-waste recycling sites concentrated. In the future, studies on PBDEs should be enhanced in China, roundly research their occurrence in marine and terrestrial living organisms, transfer from food web to human, potential health risks to human, and elimination of these chemicals. Acknowledgment This work was supported by Central Level, Scientific research institutes for Basic R&D Special Fund Business (2011YSKY06) and National Natural Science Foundation Grant (21077004 and 20807003). References Ackerman, L.K., Schwindt, A.R., Massey Simonich, S.L., Koch, D.C., Blett, T.F., Schreck, C.B., Kent, M.L., Landers, D.H., 2008. Atmospherically deposited PBDEs, pesticides, PCBs, and PAHs in western US National Park fish: concentrations and consumption guidelines. Environ. Sci. Technol. 42, 2334–2341. Antignac, J.P., Cariou, R., Zalko, D., Berrebi, A., Cravedi, J.P., Maume, D., Marchand, P., Monteau, F., Riu, A., Andre, F., 2009. Exposure assessment of French women and their newborn to brominated flame retardants: determination of tri-to decapolybromodiphenylethers (PBDE) in maternal adipose tissue, serum, breast milk and cord serum. Environ. Pollut. 157, 164–173. Bi, X.H., Qu, W.Y., Sheng, G.Y., Zhang, W.B., Mai, B.X., Chen, D.J., Yu, L., Fu, J.M., 2006. Polybrominated diphenyl ethers in south China maternal and fetal blood and breast milk. Environ. Pollut. 144, 1024–1030. Bi, X.H., Thomas, G.O., Jones, K.C., Qu, W.Y., Sheng, G.Y., Martin, F.L., Fu, J.M., 2007. Exposure of electronics dismantling workers to polybrominated diphenyl ethers, polychlorinated biphenyls, and organochlorine pesticides in south China. Environ. Sci. Technol. 41, 5647–5653. BSEF (Bromine Science and Environmental Forum), Fact Sheet: Brominated Flame Retardant Deca-BDE, 2004. . BSEF (Bromine Science and Environmental Forum), Total Market Demand, 2006. . BSEF (Bromine Science and Environmental Forum), 2010. . Cahill, T.M., Groskova, D., Chales, M.J., Sanborn, J.R., Denison, M.E., Baker, L., 2007. Atmospheric concentrations of polybrominated diphenyl ethers at near-source sites. Environ. Sci. Technol. 41, 6370–6377. Cetin, B., Odabasi, M., 2007. Particle-phase dry deposition and air–soil gas-exchange of polybrominated diphenyl ethers (PBDEs) in Izmir, Turkey. Environ. Sci. Technol. 41, 4986–4992. Chen, S.J., Gao, X.J., Mai, B.X., Chen, Z.M., Luo, X.J., Sheng, G.Y., Fu, J.M., Zeng, E.Y., 2006a. Polybrominated diphenyl ethers in surface sediments of the Yangtze river delta: levels, distribution and potential hydrodynamic influence. Environ. Pollut. 144, 951–957. Chen, L.G., Mai, B.X., Bi, X.H., Chen, S.J., Wang, X.M., Ran, Y., Luo, X.J., Sheng, G.Y., Fu, J.M., Zeng, E.Y., 2006b. Concentration levels, compositional profiles, and gasparticle partitioning of polybrominated diphenyl ethers in the atmosphere of an urban city in south China. Environ. Sci. Technol. 40, 1190–1196. Chen, L.G., Mai, B.X., Xu, Z.C., Peng, X.C., Han, J.L., Ran, Y., Sheng, G.Y., Fu, J.M., 2008. In- and outdoor sources of polybrominated diphenyl ethers and their human inhalation exposure in Guangzhou, China. Atmos. Environ. 42, 78–86. Chen, C., Chen, J.W., Zhao, H.X., Xie, Q., Yin, Z.Q., Ge, L.K., 2010. Levels and patterns of polybrominateddiphenyl ethers in children’s plasma from Dalian, China. Environ. Int. 36, 163–167. Chen, M.Y., Yu, M., Luo, X.J., Chen, S.J., Mai, B.X., 2011. The factors controlling the partitioning of polybrominated diphenyl ethers and polychlorinated biphenyls in the water-column of the Pearl river estuary in south China. Mar. Pollut. Bull. 62, 29–35. Cobbing, M., Toxic Tech: Not in Our Backyard. Uncovering the Hidden Flows of e-waste. Report from Greenpeace International. http://www.greenpeace.org/ raw/content/belgium/fr/press/reports/toxic-tech.pdf, Amsterdam, 2008. Covaci, A., de Boer, J., Ryan, J., Voorspoels, S., Schepens, P., 2002. Distribution of organobrominated and organochlorinated contaminants in Belgian human adipose tissue. Environ. Res. 88, 210–218. Covaci, A., Bervoets, L., Hoff, P., Voorspoels, S., Voets, J., Van Campenhout, K., Blust, R., Schepens, P., 2005. Polybrominated diphenyl ethers (PBDEs) in freshwater mussels and fish from Flanders, Belgium. J. Environ. Monit. 7, 132–136. Dalla Vallea, M., Juradob, E., Dachsb, J., Sweetmana, A.J., Jonesa, K.C., 2005. The maximum reservoir capacity of soils for persistent organic pollutants: implications for global cycling. Environ. Pollut. 134, 153–164.

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