Environmental Pollution 225 (2017) 252e260

Contents lists available at ScienceDirect

Environmental Pollution journal homepage: www.elsevier.com/locate/envpol

Occurrence, profiles, and toxic equivalents of chlorinated and brominated polycyclic aromatic hydrocarbons in E-waste open burning soils* Chiya Nishimura a, Yuichi Horii b, Shuhei Tanaka a, c, Kwadwo Ansong Asante d, Florencio Ballesteros Jr. e, Pham Hung Viet f, Takaaki Itai g, Hidetaka Takigami h, Shinsuke Tanabe g, Takashi Fujimori a, i, * a

Department of Environmental Engineering, Graduate School of Engineering, Kyoto University, 615-8540, Kyoto, Japan Center for Environmental Science in Saitama, 347-0115, Saitama, Japan c Department of Technology and Ecology, Graduate School of Global Environmental Studies, Kyoto University, 606-8501, Kyoto, Japan d CSIR Water Research Institute, P.O. Box AH 38, Achimota, Accra, Ghana e Department of Chemical Engineering, University of the Philippines Diliman, Quezon City, 1101, Metro Manila, Philippines f Center for Environmental Technology and Sustainable Development, Hanoi University of Science, 334 Nguyen Trai, Hanoi, Viet Nam g Center for Marine Environmental Studies, Ehime University, 790-8577, Matsuyama, Japan h Center for Material Cycles and Waste Management Research, National Institute for Environmental Studies, 305-8506, Tsukuba, Japan i Department of Global Ecology, Graduate School of Global Environmental Studies, Kyoto University, 615-8540, Kyoto, Japan b

a r t i c l e i n f o

a b s t r a c t

Article history: Received 12 August 2016 Received in revised form 27 October 2016 Accepted 29 October 2016 Available online 24 March 2017

We conducted this study to assess the occurrence, profiles, and toxicity of chlorinated polycyclic aromatic hydrocarbons (Cl-PAHs) and brominated polycyclic aromatic hydrocarbons (Br-PAHs) in e-waste open burning soils (EOBS). In this study, concentrations of 15 PAHs, 26 Cl-PAHs and 14 Br-PAHs were analyzed in EOBS samples. We found that e-waste open burning is an important emission source of Cl-PAHs and Br-PAHs as well as PAHs. Concentrations of total Cl-PAHs and Br-PAHs in e-waste open burning soil samples ranged from 21 to 2800 ng/g and from 5.8 to 520 ng/g, respectively. Compared with previous studies, the mean of total Cl-PAH concentrations of the EOBS samples in this study was higher than that of electronic shredder waste, that of bottom ash, and comparable to fly ash from waste incinerators in Korea and Japan. The mean of total Br-PAH concentrations of the EOBS samples was generally three to four orders of magnitude higher than those in incinerator bottom ash and comparable to incinerator fly ash, although the number of Br-PAH congeners measured differed among studies. We also found that the Cl-PAH and Br-PAH profiles were similar among all e-waste open burning soil samples but differed from those in waste incinerator fly ash. The profiles and principal component analysis results suggested a unique mechanism of Cl-PAH and Br-PAH formation in EOBS. In addition, the Cl-PAHs and Br-PAHs showed high toxicities equivalent to PCDD/Fs measured in same EOBS samples when calculated based on their relative potencies to benzo[a]pyrene. Along with chlorinated and brominated dioxins and PAHs, Cl-PAHs and Br-PAHs are important environmental pollutants to investigate in EOBS. © 2016 Elsevier Ltd. All rights reserved.

Keywords: PAHs Chlorinated PAHs Brominated PAHs Electronic waste Open burning soil

1. Introduction With the rapid growth of technology and electronics industries, as well as greater electronic product consumption and shortening

*

This paper has been recommended for acceptance by Charles Wong. * Corresponding author. Department of Environmental Engineering, Graduate School of Engineering, Kyoto University, 615-8540, Kyoto, Japan. E-mail address: [email protected] (T. Fujimori). http://dx.doi.org/10.1016/j.envpol.2016.10.088 0269-7491/© 2016 Elsevier Ltd. All rights reserved.

product lifespans, electronic waste (e-waste) is increasing worldwide. The total annual global volume of e-waste was ~41.8 million metric tons in 2014, and is expected to reach 50 million metric tons in 2018 (Balde et al., 2015). E-waste, which contains various valuable metals, is an important source of metal resources when properly recycled. However, some e-waste collected for recycling in industrialized countries is transported to developing countries due to their lower labor costs and less stringent environmental regulations (Terazono et al., 2006; Amoyaw-Osei et al., 2011), where it is

C. Nishimura et al. / Environmental Pollution 225 (2017) 252e260

treated with informal e-waste recycling techniques such as dismantling, acid leaching, and open burning (Balde et al., 2015; Brigden et al., 2005). These methods result in severe environmental pollution of nearby environments (Tue et al., 2010; Eguchi et al., 2015; Wittsiepe et al., 2015). Open burning of e-waste, which is an informal technique to reduce the waste mass and to recover valuable metals from ewaste, has become a serious source of environmental pollution in ecosystem near recycling areas. Previous studies have reported high concentrations of toxic organic aromatic chlorinated compounds, such as polychlorinated dibenzo-p-dioxins (PCDDs), polychlorinated dibenzofurans (PCDFs), and dioxin-like polychlorinated biphenyls (DL-PCBs), in e-waste open burning soils (EOBS) in China (Leung et al., 2007; Shen et al., 2009; Ren et al., 2015). In our previous research, we found that chlorinated dioxin concentrations in EOBS samples collected in Vietnam exceeded environmental standards (Nishimura et al., 2014). However, the percentage contribution of chlorine from chlorinated dioxins and chlorobenzenes of total organic chlorine (TOCl) was low (
0.1%). Even after accounting for naturally produced unidentified chlorinated organic compounds, > 99.3% of TOCl was attributed to unidentified anthropogenic chlorinated organic compounds generated by ewaste open burning (Nishimura et al., 2014). In addition, a higher concentration of PAHs than chlorinated dioxins has been detected in EOBS (Shen et al., 2009; Yu et al., 2006). This suggests that high concentrations of Cl-PAHs (i.e., substituted PAHs), could be detected among the unidentified chlorinated compounds. Because brominated flame-retardants are present in electronic products, brominated compounds can also be emitted when the waste is burned. Recent studies have detected brominated organic compounds, in particular brominated dioxins such as polybrominated dibenzo-p-dioxins (PBDDs) and polybrominated dibenzofurans (PBDFs), in EOBS samples (Zennegg et al., 2009; Fujimori et al., 2016; Tue et al., 2016). This suggests that BrPAHs could also be detected in EOBS. Selected studies have reported that Cl-PAHs and Br-PAHs (Cl-/ Br-PAHs) are produced during solid waste combustion (Horii et al., 2008; Miyake et al., 2012; Wang et al., 2013) and they are mutagenic, with toxicities similar to PCDD/Fs (Kido et al., 2013; Ohura et al., 2007, 2009; Horii et al., 2009a). Although Cl-PAHs have been identified in dust from workshop floors from e-waste recycling facilities and soils near these facilities in China (Ma et al., 2009), Cl-/Br-PAHs in EOBS have not been investigated. From the above points, it is important to investigate Cl-/Br-PAHs in EOBS. In this study, in addition to 15 PAHs, we determined 26 ClPAHs and 14 Br-PAHs to investigate the occurrence, profiles, and toxicity of Cl-/Br-PAHs in EOBS. 2. Materials and methods 2.1. Sample collection and sample preparation Soil samples collected at sites in Vietnam (January 2011), the Philippines (August 2010), and Ghana (August 2013) are described in Table 1 and illustrated in Fig. S1. A reference soil sample (control) was collected in Duong Quang, Hanoi, Vietnam, where e-waste recycling is not conducted. At each site, blackened EOBS samples were collected directly beneath combusted residue using a shovel, packed in plastic bags, and stored in a cooler. Two blackened surface soil samples (GH-1 and GH-2) were collected from Agbogbloshie Market, Accra, Ghana, which is the largest informal e-waste recycling site in Ghana; numerous small e-waste recycling workshops also surround the market. Wires and cables are burned along the edge of the market, although connected materials, including printed circuit boards, and other materials, including plastics, are

253

mixed and burned (Brigden et al., 2008; Otsuka et al., 2012). Two blackened surface soil samples (PHI-1 and PHI-2) were collected from e-waste open burning sites on the bank of the Marilao River in Caloocan, Metro Manila, Philippines, where informal e-waste recycling activities are conducted (Fujimori et al., 2012). The types of e-waste burned there were unknown. Two blackened surface soil samples (VN-1 and VN-2) were collected at an informal open burning site for wires and cables in Bui Dau, Hanoi, Vietnam, a town known for informal e-waste recycling. Activities are often conducted in the backyards of homes, but have been observed around rice fields as well (Tue et al., 2010; Fujimori et al., 2013). Samples were air-dried for 1 week, sieved (<2 mm), and homogenized in the laboratory before being stored in a refrigerator until further analysis. 2.2. Target compounds PAHs were assessed among 16 USEPA priority PAHs; however, acenaphthylene was omitted due to its low fluorescence. Fig. S2 shows the abbreviations and structures of the target and parent PAHs. PAH calibration mix was purchased from Sigma-Aldrich (St, Louis, MO). In total, 26 mono-to tetra- Cl-PAHs were examined in this study, including: monochlorofluorenes (ClFle), mono-to trichlorophenanthrenes (ClnPhe, n ¼ 1e3), mono-to tetrachloroanthracenes (ClnAnt, n ¼ 1e4), monoand dichlorofluoranthenes (ClnFlu, n ¼ 1e2), mono-to tetrachloropyrenes (ClnPyr, n ¼ 1e4), mono- and dichlorochrysenes (ClnChr, n ¼ 1e2), mono- and dichlorobenz[a]anthracenes (ClnBaA, n ¼ 1e2), and monochlorobenzo[a]pyrene (ClBaP). For some ClPAHs, the substitution position has not been identified due to a lack of analytical standards. In addition, 14 mono- and di-Br-PAHs were assessed, including: monobromophenanthrene (BrPhe), mono-to dibromoanthracenes (BrnAnt, n ¼ 1e2), monobromopyrene (BrPyr), mono-to dibenzo[a]anthracenes (BrnBaA, n ¼ 1e2), and monobromobenzo[a]pyrene (BrBaP). 9-ClPhe were purchased from Acros Organics (Geel, Belgium). 2-ClAnt, 9-ClAnt, 1,5,9-Cl3Ant, 1,5,9,10-Cl4Ant and 9,10-Br2Ant were obtained from Sigma-Aldrich (St, Louis, MO). 9-BrPhe, 1-BrAnt, 2-BrAnt, 9-BrAnt, 1,5-Br2Ant, and 2,6-Br2Ant were purchased from Tokyo Chemical Industry Co., Ltd. (Tokyo, Japan). The remaining Cl-/Br-PAHs were synthesized in Japan. They were identified by GC-MS and 1H NMR spectroscopy (Ohura et al., 2005, 2009). The purities of the synthesized standards were >95% (determined by GC-MS by using areas of GC-MS chromatogram). 2.3. Sample preparation and analysis To analyze the 15 priority PAHs, 5 mL of acetone was added to 1 g of soil sample (1 g) for ultrasonic extraction. The resulting extracts were centrifuged at 2500 rpm for 5 min in a centrifugal separator, and then dissolved into 500 mL of ultra-pure water. This process was conducted twice. The solution was cleaned and concentrated by passing through a cartridge (InertSep RP-1; GL Sciences, Tokyo, Japan) at a flow rate of 7 mL/min. After dehydration, the cartridge was eluted with 10 mL of dichloromethane. The elute was concentrated to <1 mL under a nitrogen stream and reconstituted to 1.0 mL with acetonitrile. The PAHs were analyzed using high performance liquid chromatography with fluorescence (GL-7400 Series; GL Sciences). A blank test was performed to check contamination and cross-contamination. Additional recovery test was conducted for calibration using Akadama soil. Mean recoveries (n ¼ 5 determination) were 80.8% for naphthalene, 46.9% for acenaphthene, 70.2% for fluorene, 74.4% for phenanthrene, 70.8% for anthracene, 73.1% for fluoranthene, 82.9% for pyrene, 64.6% for benz [a]anthracene, 69.9% for chrysene, 54.8% for benzo[b]fluoranthene,

254

C. Nishimura et al. / Environmental Pollution 225 (2017) 252e260

Table 1 Description of the sample locations. sample name

sample type

City Country

year

Description

Control

reference soil

Duong Quang Vietnam Accra Ghana

2011

Caloocan, the Philipines

2010

Bui Dau Vietnam

2011

 Collected from a reference site in Duong Quang, in Hung Yung Province, northern Vietnam.  e-waste recycling activities weren't conducted.  The GH-1 and GH-2 samples were collected from the largest e-waste recycling site, Agbogbloshie, Accra, Ghana.  There are nmerous small e-waste recycling workshop in the field. Wires and cables were burnt by the edge of the field. Other material such as plastic were mixed and burnt there.  The PHI-1 and PHI-2 samples were collected from e-waste open burning sites by the Marilao river bank in Calooncan, the Philippines.  E-waste burnt there was unknown.  The VN-1 and VN-2 samples collected at open burning sites of wires and cables in Bui Dau, a small village known for its e-waste recycling activities, Hung Yung Province, northern Vietnam.  E-waste recycling activities were family-based and conducted in the backyard of their house. Open burning activities of them were often observed along rice fields.

GH-1 GH-2

PHI-1 PHI-2

e-waste open burning soil

VN-1 VN-2

2013

Open burning soil samples were collected directly from beneath combusted residue at each site. All samples were collected by a shovel, packed in plastic bags and stored in a cooler. They were then air-dried for 1 week, sieved (<2 mm), and homogenized in a laboratory prior to analyses.

52.7% for benzo[k]fluoranthene, 51.6% for benzo[a]pyrene, 31.7% for dibenz[a,h]anthracene, 34.0% for benzo[g,h,i]perylene, and 36.0% for indeno[1,2,3-cd]pyrene. We confirmed that each coefficient of variation was less than 10% under this pretreatment. The 26 Cl-PAHs and 14 Br-PAHs were analyzed following a previously described method with several modifications (Horii et al., 2009b). Briefly, 1 g of soil sample was homogenized with 10 g of anhydrous sodium sulfate and extracted for 18 h in a Soxhlet extractor using 360 mL of a mixture of dichloromethane and nhexane (3:1, v/v). After the extraction, sample aliquots were spiked with 2.5 ng of each 13C-labeled PAH standard (13C-labeled PAH, US EPA 16 PAH Cocktail; Cambridge Isotope Laboratory, Tewksbury, MA, USA). In this study, internal standards were spiked after extraction, because concentration orders of Cl-/Br-PAHs in EOBS couldn't be estimated and the amount of sample was not enough. Therefore, extracts had to be prepared first and internal standard was spiked into sample aliquots. The extracts were concentrated and reconstituted with hexane. Each solution was purified and fractionated in an activated carbon cartridge column (Carboxen1016, 200 mg; Supelco Inc., Bellefonte, PA, USA) connected to a silica gel cartridge column (Supelclean LC-Si, 2 g; Supelco Inc.). The cartridge columns were eluted with 20 mL of 10% dichloromethane/hexane. The silica gel cartridge was removed and the activated carbon cartridge was reversed and eluted with 120 mL of toluene. The toluene fraction including the Cl-/Br-PAHs was concentrated and spiked with 2.5 ng of d12-chrysene as a recovery standard to a total volume of 100 mL. The Cl-/Br-PAH concentrations were determined with high-resolution gas chromatography/highresolution mass spectrometry (JMS-800D; JEOL, Tokyo, Japan). Recoveries of 13C-labeled PAHs used as internal standards were 78% for 13C-fluorene (13C-Flu), 83% for 13C-chrysene (13C-Chr), and 79% for 13C-benzo[a]pyrene (13C-BaP).

PAHs ¼ 200e600 ng/g), contaminated soil (S16 PAHs ¼ 600e1000 ng/g), and heavily contaminated soil (S16 PAHs > 1000 ng/g). Based on this classification system, the control in this study was uncontaminated, GH-1 and GH-2 were weakly contaminated, and PHI-1, PHI-2, VN-1, and VN-2 were heavily contaminated, although we measured only 15 of the 16 priority PAHs. Moreover, according to the Dutch List (VROM, 2000), the target value of S10 PAH (i.e., naphthalene, anthracene, phenanthrene, fluoranthene, benz[a]anthracene, chrysene, benzo[a]pyrene, benzo[ghi]perylene, benzo[k]fluoranthene, indeno[1,2,3-cd] pyrene) concentration is 20 ng/g and the intervention value of the S10 PAH is 1000 ng/g. The S10 PAH concentrations of the soil samples in this study were one to two orders of magnitude higher than the target value. In particular, the PHI and VN samples were 2.2- and 5.2-fold higher than intervention value, respectively

3. Results and discussion 3.1. PAHs 3.1.1. PAH concentrations Table S1 lists the concentrations of the individual PAHs and sum of the 15 PAH (S15 PAH) concentrations in the soil samples. The S15 PAH concentrations were 30 ng/g in the control sample, 390 and 710 ng/g at GH-1 and GH-2, 2900 and 4000 ng/g at PHI-1 and PHI-2, and 7200 and 6500 ng/g at VN-1 and VN-2. Maliszewska-Kordy-bach (1996) suggested four contamination classifications based on total PAH concentration: uncontaminated soil (S16 PAHs < 200 ng/g), weakly contaminated soil (S16

Fig. 1. (a) Total concentrations of 15 PAHs (ng/g) and (b) compositions of PAH homologs to total 15 PAHs (%) in this study and previous studies, including a burned plastic dump site in Guiyu (Leung et al., 2006), e-waste open burning soils (EOBS) in Guiyu (Yu et al., 2006) and Hong Kong (Lopez et al., 2011), electronic shredder waste (ESW) and soil around chemical plants (Ma et al., 2009), and bottom and fly ash from waste incinerators (Horii et al., 2008).

C. Nishimura et al. / Environmental Pollution 225 (2017) 252e260

(Table S1). Fig. 1 shows the S15 PAH concentrations and compositions of PAH homologs in the samples from this study and previous studies. The S15 PAH concentration in the control in this study (30 ng/g) was within a similar range as previously reported in reference soils, including 0.19 ng/g in Shanghai (Ma et al., 2009), 126 ng/g in Hong Kong (Lopez et al., 2011), and 125 ng/g in Guiyu (Leung et al., 2006). This indicates that our control was a valid reference soil. The mean S15 PAH concentration in GH (550 ng/g), where open burning of plastic e-waste parts and cables has been reported, was equivalent to that of a burned plastic dump site (428 ng/g) in Guiyu (Leung et al., 2006). Moreover, the compositions of PAH homologs of the S15 PAHs in GH were similar to those in the burned plastic dump site (Leung et al., 2006), where four-ringed PAHs predominated (Fig. 1(b)). The mean S15 PAH concentrations in PHI (3400 ng/g) and VN (6800 ng/g) were greater than those of EOBS samples in Guiyu (2049 ng/g) (Yu et al., 2006) and Hong Kong (977 ng/g) (Lopez et al., 2011). Moreover, the composition of PAH homologs in PHI and VN were similar to those of the Chinese EOBS samples, where three-ringed PAHs predominated. Compared with other sources of PAHs, the mean S15 PAH concentration in this study (3600 ng/g) was not as high as those in soils around chemical plants (19,000 ng/g) in China (Ma et al., 2009) and fly ash in waste incinerators (28,182 ng/g) in Korea (Horii et al., 2008). However, it was greater than those of e-waste shredding samples (2575 ng/g) from e-waste recycling facilities (Ma et al., 2009) and bottom ash (416 ng/g) from waste incinerators in Korea (Horii et al., 2008). This suggests that PAHs are important pollutant at e-waste open burning sites. 3.1.2. PAH source Previous studies have used low-to-high molecular weight (LMW/HMW) PAH ratio and isomer ratios of phenanthrene/ anthracene (Phe/Ant) and fluoranthene/pyrene (Flu/Pyr) to identify possible PAH sources (Lopez et al., 2011; Leung et al., 2006). HMW PAHs are the sum of four to six-ring PAHs, while LMW PAHs are the sum of two and three-ring PAHs. Pyrogenic sources typically have LMW/HMW ratios <1, Phe/Ant ratios <10, and Flu/Pyr ratios >1, while petrogenic sources typically have LMW/HMW ratios >1, Phe/ Ant ratios >10, and Flu/Pyr ratios <1. Table S1 lists the LMW and HMW PAH concentrations and LMW/HMW, Phe/Ant, Flu/Pyr ratios calculated in this study. We hypothesized that the PAH ratios of the soil samples would be consistent with pyrogenic sources, because incomplete combustion of PVC for wire coatings, casings and other plastic parts from e-waste, should be the main source of PAHs at EOBS. However, only GH-2 and VN-2 were identified as pyrogenic sources based on their LMW/HMW ratios. Moreover, only GH-2 had a Phe/Ant ratio consistent with pyrogenic sources (Phe/Ant ¼ 6.1) and only PHI-1 had a Flu/Pyr ratio consistent with pyrogenic sources (Flu/ Pyr ¼ 5.9). The other EOBSs, especially PHI-2 and VN-1, were identified as petrogenic sources. These results indicate that the PAHs in EOBS are strongly influenced by petrogenic sources, although pyrogenic sources contribute to a lesser extent. Similarly, Lopez et al. (2011) found PAH profiles indicating that EOBS were influenced by both pyrogenic and petrogenic sources. During sample collection in Vietnam, we observed that gasoline was used to assist with e-waste ignition, which could explain the sources revealed by the PAH profiles in this study. Especially for sites VN-1 and VN-2, we found that fluoranthene was not detected while pyrene concentrations were high (~700 ng/g). Both were collected at an open burning site mainly for wires and cables. Conesa et al. (2013) conducted simulated open burning for cables under some conditions. We found that the concentration of fluoranthene was lower when PVC cables including copper were combusted than

255

when only PVC coating were combusted. This suggests that lower concentration of fluoranthene seems to be specific for combustion of wires and cables. However, it didn't apply to pyrene. So, further studies are needed to find possible factors other than burnt materials (e.g. fuel used for e-waste open burning, combustion temperature). 3.2. Cl-PAHs 3.2.1. Cl-PAH concentrations Table S2 lists the concentrations of the individual Cl-PAHs and the sum of 26 Cl-PAH (S26 Cl-PAH) concentrations in soil samples measured in this study. The S26 Cl-PAH concentration was 0.096 ng/g in the control, and 200-fold greater at GH-1 and GH-2 (29 and 21 ng/g), three orders of magnitude higher at PHI-1 and PHI-2 (110 and 250 ng/g), and four orders of magnitude higher at VN-1 and VN-2 (2800 and 1800 ng/g). These results demonstrate that e-waste open burning results in Cl-PAHs pollution in soils. The large differences in the Cl-PAH levels at the different sites could not be explained and need further assessment of the sites and the processes. Fig. 2 shows a comparison of the Cl-PAHs measured in this study and those in previous studies. Of the 26 Cl-PAHs measured in this study, six (1,5,9-Cl3Ant, 1,5,9,10-Cl4Ant, Cl2Pyr-1, Cl2Pyr-2, Cl3-Pyr, and Cl4-Pyr) had not been measured in previous studies; therefore, we compared the S20 Cl-PAH concentrations in this study with previous studies. The S20 Cl-PAH concentration in the control in this study (0.096 ng/g) was similar to that of urban soil (not detected), rural soil (0.19 ng/g), and agricultural soil (0.15 ng/g) in China used as references (Ma et al., 2009), indicating that our control was a valid reference. Compared with other polluted sites reported in previous studies, the mean S20 Cl-PAH concentration (760 ng/g) of the EOBS samples in this study was higher than sediment samples near industrial areas (Horii et al., 2009b) and samples near a large e-waste recycling facility (Ma et al., 2009). Horii et al. (2009b) reported that the mean S20 Cl-PAH concentrations were 0.58 ng/g in sediment samples collected in Tokyo Bay, 1.1 ng/g in the Saginaw River watershed in USA, 8.8 ng/g in sediment samples collected near a former chlor-alkali plant in Brunswick in USA, and 1.9 ng/g in sediment samples collected in New Bedford Harbor in USA. Ma et al. (2009) reported that the S20 Cl-PAH concentrations were 87.5 ng/g (46.0e111 ng/g) in leaf samples, 103 ng/g (37.2e139 ng/g) in floor dust samples, and 26.8 ng/g (not detectede96.4 ng/g) in soils around an e-waste recycling facility in China. Moreover, they found that the S20 Cl-PAH concentration was 88.0 ng/g in soils around chemical industrial facilities (see C-soil, Fig. 2). Compared with these results, the soil pollution from e-waste burning was severe at EOBS in our study. The mean S20 Cl-PAH concentration (760 ng/g) of the EOBS samples in this study was 13-fold higher than that of electronic shredder waste (59.1 ng/g) (Ma et al., 2009), two orders of magnitude higher than that in bottom ash (7.4 ng/g), and comparable to fly ash (1930 ng/g) from waste incinerators in Korea (Horii et al., 2008). They were similar to concentrations measured in bottom ash and fly ash from waste incinerators in Japan (Miyake et al., 2012; Wang et al., 2013), although these results are not included in Fig. 2 since the concentrations of individual congeners were unavailable and the S20 Cl-PAH concentrations could not be calculated. These results suggest that e-waste open burning is a relevant source of Cl-PAHs. 3.2.2. Cl-PAH source Fig. 2(c) shows the compositions of each Cl-PAH based on ring number in this study and previous studies. The profiles were

256

C. Nishimura et al. / Environmental Pollution 225 (2017) 252e260

Fig. 2. (a)(b) Mean S20 chlorinated polycyclic aromatic hydrocarbons (Cl-PAH) concentrations (ng/g) (±standard deviation) and (c) composition of Cl-PAHs by number of rings to S20 Cl-PAHs (%) in this study and previous studies (Ma et al., 2009; Horii et al., 2009b, 2008).

similar among all EOBS samples in this study. Three-ring Cl-PAHs (ClPhe, ClAnt, and ClFlu) were the predominant type of Cl-PAH, accounting for 53e89% (mean, 68%) of S20 Cl-PAH concentrations. Meanwhile, four-ring Cl-PAHs (ClPry, ClChr, and ClBaP) accounted for 8.3e37% (mean, 26%), and two-ring (ClFle) and fivering (ClBaP) Cl-PAHs accounted for only 0.86 and 5.6% of S20 ClPAHs, respectively. This profile differed from those of e-waste shredding recycling facilities (Ma et al., 2009) and fly ash (Horii et al., 2008), in which four- and five-ring Cl-PAHs were more prominent than three-ring Cl-PAHs. Since these profiles are specific to EOBSs and differ from other emission sources, e-waste open burning is a unique emission source of Cl-PAHs that differs from unburned e-waste recycling and waste incinerators. Previous studies calculated concentration ratios of several ClPAHs normalized to 1-ClPyr and 3-ClFlu (6-ClBaP/1-ClPyr, 3-ClFlu/ 1-ClPyr, 7-ClBaA/1-ClPyr, 6-ClBaP/3-ClFlu, 1-ClPyr/3-ClFlu, and 7ClBaA/3-ClFlu) to identify Cl-PAH sources for fly ash (Horii et al., 2008), e-waste recycling facility samples (Ma et al., 2009), urban air (Kitazawa et al., 2006), and samples collected in road tunnels (Nilsson and Oestman, 1993). The 6-ClBaP/3-ClFlu ratio was used as a suitable indicator of Cl-PAH source, because 3-ClFlu has a higher photostability than other Cl-PAHs (Kitazawa et al., 2006; Ohura et al., 2005). On the basis of these ratios, Ma et al. (2009) defined four potential emission sources of Cl-PAHs: (i) e-waste recycling facilities, including soils and dust from workshop floors and ewaste shredder dust; (ii) municipal solid waste incineration, including fly ash; (iii) chemical industrial activities involving chlorine; and (iv) automobile exhaust including the air in road tunnels. In this categorization, Ma et al. (2009) reported the urban air samples reported in Kitazawa et al. (2006) were categorized into

municipal solid waste incineration because they have a similar profile of the ratios to municipal solid waste incineration. Leaf samples around the e-waste recycling facilities were categorized into automobile exhaust sources because the profile of isomer ratios in the leaf samples was similar to that in the road-tunnel air reported in Nilsson and Oestman (1993). They discussed that similarities in profile with leaves from the e-waste recycling operations presumably indicate the influence of environmental partitioning properties of Cl-PAHs (i.e., preferential partitioning of 3ClFlu and 1-ClPyr to air (Nilsson and Oestman, 1993)), since automobile exhaust is the major source of Cl-PAHs in road-tunnel air. We calculated these ratios for EOBS samples and performed a principal component analysis (PCA) on the previous dataset and our dataset using the six ratios of the selected Cl-PAHs normalized to 1-ClPyr and 3-ClFlu. The PCA scores plot are shown in Fig. 3. Two factors explained 88.0% of the original data variance governing the concentration ratio distributions; PC1 (53.8% of total variance) was influenced mainly by e-waste recycling activity, as almost all of the EOBS samples and electronic shredder waste (see ESW, Fig. 3) were located in the positive region, while PC2 (34.2%) was influenced mainly by industrial activity, as industrial samples, including fly ash and soils, were located in the positive region. The results of the control are not shown in Fig. 3 because only two of the ratios (3ClFlu/1-ClPyr and 1-ClPyr/3-ClFlu) could be calculated. The results of the two ratios in the control were most similar to those of urban air. In general, our PCA results are consistent with the previous categorizations by Ma et al. (2009). There were four distinct clusters in the scores plot and most of the EOBS samples were located in close proximity to each other, with the exception of GH-2 and PHI1. The large difference between the majority of the EOBS samples

C. Nishimura et al. / Environmental Pollution 225 (2017) 252e260

257

concentrations were 0.08 ng/g (<0.14e0.65 ng/g) in bottom ash and 200 ng/g (<0.14e1235 ng/g) in fly ash from waste incinerators in Korea. Wang et al. (2013) reported that mean S9 Br-PAH concentrations were 0.048 ng/g (0.011e0.13 ng/g) in bottom ash and 666 ng/g (0.061e3270 ng/g) in fly ash from waste incinerators in Japan. Ni and Zeng (2012) reported that a mean S6 Br-PAH concentration was 12 ng/g (0.099e10 ng/g) in fly ash from garbage incinerators in China. In general, the EOBS samples had Br-PAH concentrations three to four orders of magnitude higher than those in incinerator bottom ash and comparable to incinerator fly ash, although the number of Br-PAH congeners measured differed among studies. This indicates that e-waste open burning is an important source of Br-PAHs. Fig. 4 shows a comparison of the EOBS samples in this study with soil samples collected from different areas reported by Ni and Zeng (2012). Comparing the concentrations of five Br-PAHs (9-BrPhe, 9-BrAnt, 9,10-Br2Ant, 1-BrPyr, and 7BrBaA) in a previous study and this study, the mean S5 Br-PAH concentration decreased in the following order: EOBS (61 ng/ g) > traffic (1.0 ng/g) > commercial (0.97 ng/g) > residential (0.63 ng/g) > industrial ¼ orchard (0.35 ng/g) > forestry (0.28 ng/ g) > agricultural (0.20 ng/g) > greenbelt (0.11 ng/g) > the control in this study (not detected, < 0.15 ng/g). This indicates that e-waste open burning is an important emission source of Br-PAHs as well as Cl-PAHs.

Fig. 3. Principal component analysis (PCA) scores plot of two principal components from the dataset based on information on selected Cl-PAHs normalized to 1-ClPyr and 3-ClFlu in this study and previous studies (Horii et al., 2008; Ma et al., 2009; Kitazawa et al., 2006; Nilsson and Oestman, 1993).

and GH-2 and PHI-1 was due to differences in the waste material compositions burned at these sites. Among e-waste recycling activities, open burning and non-burning processes, such as e-waste shredding, formed separate clusters (Fig. 3). This suggests that combustion is an important source of Cl-PAHs. Among combusted samples, the EOBS samples formed a different cluster from fly ash on the PCA plot. This might result from the specific smouldering of PVC and other halogen containing plastic at lower temperature compared to waste incineration and related generated ashes. Basic studies in the formation of Cl-PAHs in e-waste open burning are needed to better understand their formation path and differences to other processes. The EOBS samples were located near leaf samples collected around an e-waste recycling facility (see leaf, Fig. 3) on the PCA plot. This result was consistent with the result of the ClPAH profiles discussed in Fig. 2 (c). Finally, the EOBS samples were located near air sample in road tunnel (see tunnel air, Fig. 3), which was reported to be categorized into automobile exhaust sources in Ma et al. (2009), on the PCA plot. That is probably because gasoline is often used to ignite and maintain open burning fires. 3.3. Br-PAHs 3.3.1. Br-PAH concentrations Table S3 shows individual Br-PAH concentrations and the sum of 14 Br-PAHs (S14 Br-PAHs) in soil samples measured in this study. Br-PAHs were undetected (<0.15 ng/g) in the control. The S14 BrPAH concentrations were 12 and 5.8 ng/g at GH-1 and GH-2, 60 and 32 ng/g at PHI-1 and PHI-2, and 520 and 84 ng/g at VN-1 and VN-2. The mean S14 Br-PAH concentration was 120 ng/g, suggesting that Br-PAHs were generated anthropogenically from ewaste open burning. Horii et al. (2008) reported that mean S11 Br-PAH

3.3.2. Br-PAH source Horii et al. (2008) calculated the concentration ratios of several Br-PAHs normalized to 1-BrPyr (6-BrBaP/1-BrPyr, 7-BrBaA/1-BrPyr, and 9-BrPhe/1-BrPyr) to identify the source of Br-PAHs. We calculated these ratios of our EOBS samples. However, we couldn't perform inter-study comparison, because 6-BrBaP was not measured or 9-BrPhe was not detected in other previous studies. In general, in this study, 6-BrBaP/1-BrPyr and 7-BrBaA/1-BrPyr were <1 and 9-BrPhe/1-BrPyr > 1. These results differed from those of fly ash reported by Horii et al. (2008), in which all three ratios were <1 and were within a similar concentration range. Fig. 4(b) shows a comparison of the profiles of five Br-PAH congeners measured in the EOBS samples in this study with those of soils and fly ash in a previous study (Ni and Zeng, 2012). Based on the results, 9-BrPhe and 9,10-Br2Ant were the dominant congeners in the EOBS samples. However, the Br-PAH profiles in soils reported by Ni and Zeng (2012) differed from those in our study; 7-BrBaA, the least congener in our study, was the dominant congener (>60%) in non-residential soils, while those of 9-BrPhe and 9,10-Br2Ant were <10%. Similar results were obtained from the comparison with fly ash reported by Ni and Zeng (2012), who found that 7-BrBaA was the dominant congener and 9-BrPhe was not detected. This difference in profiles suggests that they represent different sources of Br-PAHs from EOBSs. 3.4. Toxic equivalents of PAHs, Cl-PAHs, and Br-PAHs 3.4.1. Based on relative potency to 2,3,7,8-tetrachlorodibenzo-pdioxin Some PAHs and Cl-/Br-PAHs can reportedly bind to and activate the aryl hydrocarbon receptor (AhR). Villeneuve et al. (2002) reported the relative potencies to 2,3,7,8-tetrachlorodibenzo-pdioxin (REPTCDD) of 7 PAHs obtained from in vitro bioassay using a recombinant rat hepatoma cells (H4IIE-luc cell). Horii et al. (2009a) also reported REPTCDD values of two Cl-PAHs (6-ClChr and 7-ClBaA) and two Br-PAHs (7-BrBaA and 4,7-Br2BaP) determined by in vitro assay using H4IIE-luc cell. Using these reported REPTCDD values, we estimated the toxic equivalent (TEQ) concentrations of PAHs and Cl-/Br-PAHs in soil samples using (E1):

258

C. Nishimura et al. / Environmental Pollution 225 (2017) 252e260

Fig. 4. (a) S5 brominated polycyclic aromatic hydrocarbons (Br-PAH) concentrations (ng/g) and (b) compositions of individual Br-PAHs to S5 Br-PAH concentration (%) in this study and a previous study (Ni and Zeng, 2012). The five Br-PAHs included 9-bromophenenthrene (9-BrPhe), 9-bromoanthracene (9-BrAnt), 9,10-dibromoanthracene (9,10-Br2Ant), 1bromopyrene (1-BrPyr), and 7-bromobenz[a]anthracene (7-BrBaA).

TEQ ðpg$TEQ =gÞ ¼

X

Ci  REPTCDD; i  1000

(E1)

where Ci is the concentration of individual compounds (ng/g). Table S4 lists their REPTCDD values and TEQ concentrations. Almost all the change amounts of the TEQ concentrations were negligible when the results were displayed to 2 significant digits. Therefore, our result was displayed without using LOD values. Total TEQ concentration of the 7 PAHs based on REPTCDD, (TEQS7PAHs-TCDD) was 0.13 pg TEQ/g in Control. The mean TEQS7PAHsTCDD were 1.6 pg TEQ/g in GH, 2.1 pg TEQ/g in PHI, and 9.0 pg TEQ/g in VN. Compared with previous studies for e-waste related samples, the mean TEQS7PAHs-TCDD values for EOBS sample in our study were lower than that for soils collected from e-waste recycling sites in China (15.66 pg TEQ/g) reported in Shen et al. (2009) and that for soils collected from an e-waste recycling facility in China (19.1 pg TEQ/g) reported in Ma et al. (2009). Since 6-ClChr and 7-ClBaA were undetected in the control, their total TEQ concentration based on REPTCDD (TEQS2 Cl-PAHs) could not be calculated. The mean TEQS2 Cl-PAHs were 0.017 pg TEQ/g in GH, 0.17 pg TEQ/g in PHI, and 3.1 pg TEQ/g in VN. Compared with previous studies, they were two to four orders of magnitude lower than TEQ concentrations of PCDDs (490 pg TEQ/g) and DL-PCBs (200 pg TEQ/g) in EOBS in Ghana (Fujimori et al., 2016) and two to three orders of magnitude lower than TEQ concentrations of PCDDs (1100 pg TEQ/g) and DL-PCBs (170 pg TEQ/g) in the same EOBS as VN-1 (Nishimura et al., 2014). Horii et al. (2009b) reported that TEQS2 Cl-PAHs in sediment samples were 0.0011 pg TEQ/g in Tokyo Bay and 0.0024 pg TEQ/g in Saginaw River, as well as 0.055 pg TEQ/g near a former chlor-alkali plant in Brunswick in USA. They also reported that these TEQS2 Cl-PAHs were approximately five orders of magnitude lower than those of PCDD/Fs and two to three orders of magnitude lower than those of DL-PCBs. These results were in agreement with our results. We could not calculate total TEQ concentration of Br-PAHs using REPTCDD (TEQS2 Br-PAHs) for the control, since 7-BrBaA and 4,7Br2BaP were undetected. However, the mean TEQ S2 Br-PAHs was

0.00089 pg TEQ/g in GH, 0.0045 pg TEQ/g in PHI, and 0.11 pg TEQ/g in VN. Fujimori et al. (2016) reported TEQ concentrations of brominated dioxins estimated from toxic equivalency factors (TEFs) of similarly substituted PCDD/Fs listed by the World Health Organization (WHO), as the WHO and United Nations Environmental Programme recommends the use of similar interim TEFs for brominated and brominated/chlorinated congeners in human risk assessments. They also reported that mean TEQ concentrations were 4.2 pg TEQ/g for PBDDs and 3737 pg TEQ/g for PBDFs in EOBS samples in Ghana. The TEQS2 Br-PAHs in our study were approximately one to four orders of magnitude lower than that of the PBDDs and four to seven orders of magnitude lower than that of the PBDFs.

3.4.2. Based on relative potency to Benzo[a]pyrene Ohura et al. (2007) reported the relative potencies to benzo[a] pyrene (REPBaP) of 7 individual PAHs, 17 individual Cl-PAHs and 9 individual Br-PAHs determined in a YCM3 cell bioassay. In the YCM3 cell, the potency of TCDD was 60 times higher than that of BaP (Kawanishi et al., 2003). Therefore, another equation (E2) was suggested to calculate their TEQ concentrations equivalent to TCDD as follows:

TEQ ðpg$TEQ =gÞ ¼

X

 Ci  REPBaP; i 60  1000

(E2)

where Ci is the concentration of individual compounds (ng/g). Table S5 lists their REPBaP and TEQ concentrations based on REPBaP/ 60. When using REPBaP/60, total TEQ concentration of the 7 PAHs (TEQS7PAHs-BaP) was 56 pg TEQ/g in Control. The mean TEQS7PAHs-BaP were 1900 pg TEQ/g in GH, 14,000 pg TEQ/g in PHI, and 40,000 pg TEQ/g in VN. Total TEQ concentration of 17 Cl-PAHs using REPBaP (TEQS17 Cl-PAHs) was 0.14 pg TEQ/g in the control. The mean TEQS17 Cl-PAHs were 95 pg TEQ/g in GH, 680 pg TEQ/g in PHI, and 9900 pg TEQ/g in VN. Total TEQ concentration of 9 Br-PAHs using REPBaP (TEQS9 Br-PAHs) could not be calculated because no Br-PAH congener was detected. The mean TEQS9 Br-PAHs were 1.6 pg TEQ/g in GH, 7.5 pg TEQ/g in PHI, and 120 pg TEQ/g in VN.

C. Nishimura et al. / Environmental Pollution 225 (2017) 252e260

The TEQ concentrations based on REPBaP/60 as well as REPTCDD generally followed the order: STEQPAHs > STEQClPAHs > STEQBrPAHs. However, the TEQ concentrations using REPBaP/60 were generally three or four orders of magnitude higher than those using REPTCDD in EOBSs in our study. This may have been due to differences between the bioassays used and different numbers of congeners used to calculate the REPs, because no standard bioassay to evaluate their toxicity has been established. There is more REP data available for TEQS17 Cl-PAHs and TEQS9 Br-PAHs, and the mean TEQS17Cl-PAHs of the EOBS was one fifth that of fly ash and five-fold higher than the maximum value in bottom ash. The TEQS17 Cl-PAHs at VN-1, VN-2, and PHI-2 exceeded the Environmental Quality Standard (1000 pg TEQ/g), accounting for 3.9e54% of the TEQ concentration due to chlorinated dioxins (i.e., the sum of the TEQ concentrations of PCDD/Fs and DL-PCBs) measured in the same soil samples (Nishimura et al., 2014). We did not analyze more Cl-/Br-PAHs due to a lack of analytical standards, although we observed several unidentified peaks while monitoring molecular ions. When more Cl-/Br-PAHs can be identified and more REPs of Cl-/Br-PAHs can be determined in the future, Cl-/Br-PAHs will be greater risk factors than our estimated calculations. 4. Conclusion In summary, we found the occurrence of Cl-/Br-PAHs in e-waste open burning soils. Our results indicate that e-waste open burning is likely to be one of the important emission sources of Cl-/Br-PAHs. Furthermore, the Cl-/Br-PAH profiles were similar among EOBS samples but differed from those of waste incinerator ash. The profiles and PCA results suggest that there is a unique mechanism of Cl-/Br-PAHs formation in EOBSs. The Cl-/Br-PAHs in the EOBS samples had high toxicities equivalent to PCDD/Fs when calculated using REPBaP/60. In contrast, toxic equivalent for PAHs and Cl-/BrPAHs using REPTCDD was orders of magnitude lower than that for PCDD/Fs measured in the same EOBS samples, but that a range of halogenated aromatics have not been assessed for REPTCDD in this study. Along with dioxins and PAHs, Cl-/Br-PAHs are important environmental pollutants to investigate in EOBSs. Acknowledgments We acknowledge financial support from a Grand-in-Aid for Young Scientist (A) (26701012) from JSPS, the Ministry of Education, Culture, Sports, Science and Technology (MEXT), Japan. This study was partly supported by the MEXT, Japan, to a project on Joint Usage/Research Center e Leading Academia in Marine and Environmental Research (LaMer), Ehime University. Appendix A. Supplementary data Supplementary data related to this article can be found at http:// dx.doi.org/10.1016/j.envpol.2016.10.088. References Amoyaw-Osei, Y., Agyekum, O.O., Pwamang, J.A., Mueller, E., Fasko, R., Schluep, M., 2011. Ghana e-waste country assessment. SBC e-waste Africa Proj. 1e123. Balde, C.P., Wang, F., Kuehr, R., Huisman, J., 2015. The global e-waste monitor e 2014. United Nations University, IAW-SCYCLE, Bonn, Germany. Brigden, K., Labunska, I., Santillo, D., Allsopp, D., 2005. Recycling of Electronic Wastes in China and India: Workplace and Environmental Contamination (Report, Greenpeace International). Brigden, K., Labunska, I., Santillo, D., Johnston, P., 2008. Chemical contamination at e-waste recycling and disposal sites in Accra and Korforidua, Ghana. Greenpeace Research Laboratories Technical Note 1e24. Conesa, J.A., Egea, S., Molto, J., Ortno, N., Font, R., 2013. Decomposition of two type

259

of electronic wires considering the effect of the metal in the production of pollutant. Chemosphere 91, 118e123. Eguchi, A., Nomiyama, K., Tue, N.M., Trang, P.T.K., Viet, P.H., Takahashi, S., Tanabe, S., 2015. Residue profiles of organohalogen compounds in human serum from ewaste recycling sites in North Vietnam: association with thyroid hormone levels. Environ. Res. 137, 440e449. Fujimori, T., Takaigami, H., Agusa, T., Eguchi, A., Bekki, K., Yoshida, A., Terazono, A., Ballesteros Jr., F.C., 2012. Impact of metals in surface matrices from formal and informal electronic-waste recycling around Metro Manila, the Philippines, and intra-Asian comparison. J. Hazard. Mater 221e222, 139e146. Fujimori, T., Takigami, H., Takaoka, M., 2013. Organochlorines in surface soil at electronic-waste wire burning sites and metal contribution evaluated using quantitative X-ray speciation. J. Phys. Conf. Ser. 430, 012094. Fujimori, T., Itai, T., Goto, A., Asante, K.A., Otsuka, M., Takahashi, S., Tanabe, S., 2016. Interplay of metals and bromine with dioxin-related compounds concentrated in e-waste open burning soil from Agbogbloshie in Accra. Ghana. Environ. Pollut. 209, 155e163. Horii, Y., Ok, G., Ohura, T., Kannan, K., 2008. Occurrence and profiles of chlorinated and brominated polycyclic aromatic hydrocarbons in waste incinerators. Environ. Sci. Technol. 42, 1904e1909. Horii, Y., Khim, J.S., Higley, E.B., Giesy, J.P., Ohura, T., Kannan, K., 2009a. Relative potencies of individual chlorinated and brominated polycyclic aromatic hydrocarbons for induction of aryl hydrocarbon receptor-mediated responses. Environ. Sci. Technol. 43, 2159e2165. Horii, Y., Ohura, T., Yamashita, N., Kannan, K., 2009b. Chlorinated polycyclic aromatic hydrocarbons in sediments from industrial areas in Japan and the United States. Arch. Environ. Contam. Toxcol. 57, 651e660. Kawanishi, M., Sakamoto, M., Ito, A., Kishi, K., Yagi, T., 2003. Construction of reporter yeast for mouse aryl hydrocarbon receptor ligand activity. Mutat. Res. 540, 99e105. Kido, T., Sakakibara, H., Ohura, T., Guruge, K.S., Kojima, M., Hasegawa, J., Iwamura, T., Yamanaka, N., Masuda, S., Sakaguchi, M., Amagai, T., Shimoi, K., 2013. Evaluation of chlorinated benz[a]anthracene on hepatic toxicity in rats and mutagenic activity in Salmonella typhimurium. Environ. Toxicol. 28, 21e31. Kitazawa, A., Amagai, T., Ohura, T., 2006. Temporal trends and relationships of particulate chlorinated polycyclic aromatic hydrocarbons and their parent compounds in urban air. Environ. Sci. Technol. 40, 4592e4598. Leung, A., Cai, Z.W., Wong, M.H., 2006. Environmental contamination from electronic waste recycling at Guiyu, southeast China. J. Mater. Cycles Waste Manag. 8, 21e33. Leung, A.O.W., Luksemburg, W.J., Wong, A.S., Wong, M.H., 2007. Spatial distribution of polybrominated diphenyl ethers and polychlorinated dibenzo-p-dioxins and dibenzofurans in soil and combusted residue at Guiyu, an electronic waste recycling site in southeast China. Environ. Sci. Technol. 41, 2730e2737. Lopez, B.N., Man, Y.B., Zhao, Y.G., Zheng, J.S., Leung, A.O.W., Yao, J., Wong, M.H., 2011. Major pollutants in soils of abandoned agricultural land contaminated by ewaste activities in Hong Kong. Arch. Environ. Contam. Toxicol. 61, 101e114. Ma, J., Horii, Y., Cheng, J., Wang, W., Wu, Q., Ohura, T., Kannan, K., 2009. Chlorinated and parent polycyclic aromatic hydrocarbon in environmental samples from an electronic waste recycling facility and a chemical industrial in China. Environ. Sci. Technol. 43, 643e649. Maliszewska-Kordy-bach, B., 1996. Polycyclic aromatic hydrocarbons in agricultural soils in Poland: preliminary proposals for criteria to evaluate the level of soil contamination. Appl. Goechem. 11, 121e127. Miyake, Y., Tang, L., Horii, Y., Nojiri, K., Ohtsuka, N., Fujimine, Y., Amagai, T., 2012. Concentration profiles of halogenated polycyclic aromatic hydrocarbons in flue gas, bottom ash, and fly ash from waste incinerator. Organohalogen Compd. 74, 636e639. Ni, H.G., Zeng, E.Y., 2012. Environmental and human exposure to soil chlorinated brominated polycyclic aromatic hydrocarbons in an urbanized region. Environ. Toxicol. Chem. 31 (7), 1494e1500. Nilsson, U.L., Oestman, C.E., 1993. Chlorinated polycyclic aromatic hydrocarbons: method of analysis and their occurrence in urban air. Environ. Sci. Technol. 27, 1826e1831. Nishimura, C., Fujimori, T., Suzuki, G., Agusa, T., Takaoka, M., Takahashi, S., Tue, N.M., Viet, P.H., Tanabe, S., Takigami, H., 2014. Percentage contributions of chlorinated aromatic compounds in electronic waste open burning soils. Organohalogen Compd. 76, 7038e7042. Ohura, T., Kitazawa, A., Amagai, T., Makino, M., 2005. Occurrence, profiles, and photostabilities of chlorinated polycyclic aromatic hydrocarbons associated with particulates in urban air. Environ. Sci. Technol. 39, 85e91. Ohura, T., Morita, M., Makino, M., Amagai, T., Shimoi, K., 2007. Aryl hydrocarbon receptor-mediated effects of chlorinated polycyclic aromatic hydrocarbons. Chem. Res. Toxicol. 20, 1237e1241. Ohura, T., Sawada, K., Amagai, T., Shinomiya, M., 2009. Discovery of novel halogenated polycyclic aromatic hydrocarbons in urban particulate matters: occurrence, photostability, and AhR activity. Environ. Sci. Technol. 43, 2269e2275. Otsuka, M., Itai, T., Asante, K.A., Muto, M., Tanabe, S., 2012. Trace element contamination around the e-waste recycling site at Agbogbloshie, Accra city, Ghana. Interdiscip. Stud. Environ. ChemistrydEnviron. Pollut. Ecotoxicol 6, 161e167. Ren, M., Tang, Y.H., Peng, P.A.H., Cai, Y., 2015. PCDD/Fs in air and soil around an ewaste dismantling area with open burning of insulated wires in south China. Bull. Environ. Contam. Toxicol. 94, 647e652. Shen, C.F., Chen, Y.X., Huang, S.B., Wang, Z.J., Yu, C.N., Qiao, M., Xu, Y.P., Setty, K.,

260

C. Nishimura et al. / Environmental Pollution 225 (2017) 252e260

Zhang, J.Y., Zhu, Y.F., Lin, Q., 2009. Dioxin-like compounds in agricultural soils near e-waste recycling sites from Taizhou area, China: chemical and bioanalytical characterization. Environ. Int. 35, 50e55. Terazono, A., Murakami, S., Abe, N., Inanc, B., Moriguchi, Y., Sakai, S., Kojima, M., Yoshida, A., Li, J., Yang, J., Wong, M.H., Jain, A., Kim, I., Peralta, G., Lin, C., Mungcharoen, T., Williams, E., 2006. Current status and research on E-waste issues in Asia. J. Mater. Cycles Waste Manag. 8, 1e12. Tue, N.M., Sudaryanto, A., Minh, T.B., Isobe, T., Takahashi, S., Viet, P.H., Shinsuke, T., 2010. Accumulation of polychlorinated biphenyls and brominated flame retardants in breast milk from women living in Vietnamese e-waste recycling sites. Sci. Total Environ. 408, 2155e2162. Tue, N.M., Goto, A., Takahashi, S., Itai, T., Asante, K.A., Kunisue, T., Tanabe, S., 2016. Release of chlorinated, brominated and mixed halogenated dioxin-related compounds to soils from open burning of e-waste in Agbogbloshie (Accra, Ghana). J. Hazard. Mater 302, 151e157. Villeneuve, D.L., Khim, J.S., Kannan, K., Giesy, J.P., 2002. Relative potencies of individual polycyclic aromatic hydrocarbons to induce dioxinlike and estrogenic

responses in three cell lines. Environ. Toxicol. 17, 128e137. VROM, 2000. Circular on Target Values and Intervention Values for Soil Remediation. Spatial Planning and Environment. Ministry of Housing, Netherlands. Wang, Q., Miyake, Y., Amagai, T., Horii, Y., Nojiri, K., Ohtsuka, N., 2013. Profiles and distribution of halogenated polycyclic aromatic hydrocarbons from waste incinerator. Organohalogen Compd. 75, 27e30. Wittsiepe, J., Fobil, J.N., Till, H., Burchard, G.D., Wilhelm, M., Feldt, T., 2015. Levels of polychlorinated dibenzo-p-dioxins, dibenzofurans (PCDD/Fs) and biphenyls (PCBs) in blood of informal e-waste recycling workers from Agbogbloshie, Ghana, and controls. Environ. Int. 79, 65e73. Yu, X.Z., Gao, Y., Wu, S.C., Zhang, H.B., Cheung, K.C., Wong, M.H., 2006. Distribution of polycyclic aromatic hydrocarbons in soils at Guiyu area of China, affected by recycling of electronic waste using primitive technologies. Chemosphere 65, 1500e1509. Zennegg, M., Yu, X., Wong, M.H., Weber, R., 2009. Fingerprints of chlorinated, brominated and mixed halogenated dioxins at two e-waste recycling sites in Guiyu/China. Organohalogen Compd. 71, 2248e2252.