Passive sampling of polybrominated diphenyl

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Engineering, Huaqiao University, Xiamen 361021, China. 2. Institute of .... cause the Bflying saucer^ housing of the PUF-PAS can ade- quately dampen the wind ...
Passive sampling of polybrominated diphenyl ethers in indoor and outdoor air in Shanghai, China: seasonal variations, sources, and inhalation exposure Wenliang Han, Tao Fan, Binhua Xu, Jialiang Feng, Gan Zhang, Minghong Wu, Yingxin Yu & Jiamo Fu Environmental Science and Pollution Research ISSN 0944-1344 Volume 23 Number 6 Environ Sci Pollut Res (2016) 23:5771-5781 DOI 10.1007/s11356-015-5792-9

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Author's personal copy Environ Sci Pollut Res (2016) 23:5771–5781 DOI 10.1007/s11356-015-5792-9

RESEARCH ARTICLE

Passive sampling of polybrominated diphenyl ethers in indoor and outdoor air in Shanghai, China: seasonal variations, sources, and inhalation exposure Wenliang Han 1 & Tao Fan 1 & Binhua Xu 2 & Jialiang Feng 2 & Gan Zhang 3 & Minghong Wu 2 & Yingxin Yu 2 & Jiamo Fu 2

Received: 21 April 2015 / Accepted: 10 November 2015 / Published online: 20 November 2015 # Springer-Verlag Berlin Heidelberg 2015

Abstract Ninety-seven seasonal, passive indoor and outdoor air samples were collected in Shanghai to study polybrominated diphenyl ethers (ΣPBDEs, 16 congeners including BDE-209), their concentrations, composition profiles, seasonal variations, influencing factors, emission sources, and human inhalation exposure. In summer, median indoor concentrations of Σ 15 PBDEs (excluding BDE-209) were 82 pg m−3 in offices and 30 pg m−3 in homes, ∼3 times the winter concentrations. The average summer concentration of 130 pg m−3 BDE-209 in homes was higher than that in offices (which was 90 pg m−3); in winter, home and office concentrations were similar (46 and 47 pg m−3, respectively). For outdoor air, the median concentration of Σ15PBDEs in summer (12 pg m−3) was twice the winter concentration (6 pg m−3), while the summer median concentration of BDE-209 (398 pg m − 3 ) was half the winter concentration (794 pg m−3). Higher concentrations of Σ15PBDEs indoors compared with outdoors showed that the lower brominated BDEs found were mainly from indoor sources. Meanwhile, Responsible editor: Roland Kallenborn Electronic supplementary material The online version of this article (doi:10.1007/s11356-015-5792-9) contains supplementary material, which is available to authorized users. * Jialiang Feng [email protected] 1

Institute of Environmental and Resources Technology, Department of Environmental Science and Engineering, College of Chemical Engineering, Huaqiao University, Xiamen 361021, China

2

Institute of Environmental Pollution and Health, Shanghai University, Shanghai 200444, China

3

Guangzhou Institute of Geochemistry, Chinese Academy of Sciences, Guangzhou 510640, China

the much lower indoor concentration of BDE-209 compared with the outdoors showed that BDE-209 came mainly from outdoor sources. The data set also indicated that electric/ electronic appliances were the main sources of indoor ΣPBDEs, and old appliances emitted more lower brominated BDEs, while industrial emissions should be the main source of the outdoor BDE-209. Median daily human exposures to Σ15PBDEs and BDE-209 through inhalation were estimated to be 0.23 and 1.73 ng day −1 in winter and 0.65 and 2.28 ng day−1 in summer for adults. The human inhalation exposure to ΣPBDEs (3.44 ng day −1 for adults and 1.33 ng day−1 for toddlers) was comparable to that from eating contaminated fish for both toddlers and adults in Shanghai. Keywords Polybrominated diphenyl ethers (PBDEs) . Passive air sampling . Indoor air . Outdoor air . Shanghai

Introduction Polybrominated diphenyl ethers (PBDEs) are ubiquitous in the environment (Law et al. 2014; Lohmann et al. 2013) due to both the extensive usage of these materials as flame retardant additives and their persistency (Abbasi et al. 2015). Concentrations of PBDEs in abiotic environmental media and biological and human samples have increased rapidly in the past two to three decades (Law et al. 2014; Linares et al. 2015), and their potential adverse health effects have been well documented (Lai et al. 2011; Linares et al. 2015). The usage of Penta-BDE and Octa-BDE products has been banned by the European Union since the summer of 2004, and their production in the USA has been voluntarily halted since December 2004 (La Guardia et al. 2006). In 2009, these products were listed as persistent organic pollutants (POPs) under the Stockholm convention. Unfortunately, many products

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containing Penta- and Octa-BDE are still in use in many places. The production of Deca-BDE has been banned in Canada and restricted for usage in the European Union (Zhang et al. 2011), but it is still being used in many countries, including China. Recent studies showed that BDE-209 can degrade to lower brominated BDEs in the environment and can be metabolized in living organisms; the toxicity of the hydroxylated metabolites of PBDEs is of great concern (Lai et al. 2011; Yu et al. 2010). Thus, PBDE pollution will be an environmental problem for quite a long time, even after they are completely banned. People spend up to 90 % of their time indoors in spaces often characterized by limited ventilation and emissions of hazardous compounds from office equipment, furniture, building materials, and home appliances (Barro et al. 2009; Krol et al. 2011). The limited ventilation acts to concentrate these emissions (Garcia-Jares et al. 2009). Many semi-volatile organic compounds (SVOCs) have primarily indoor sources and much higher indoor concentrations than outdoors (Besis and Samara 2012; Bohlin et al. 2014b). Unlike other POPs, such as PCBs, of which the primary exposure comes through diet, residential exposure can be the dominant means of human exposure to PBDEs (Dodson et al. 2015). Chen et al. (2008) studied PBDEs in indoor and outdoor air in Guangzhou in South China using high volume air samplers. High volume samplers, originally designed for outdoor use, are bulky and noisy and are, therefore, not suitable for indoor sampling (Bohlin et al. 2007). Passive air samplers (PUFPAS)—in which a polyurethane foam (PUF) disk is used to collect air samples through diffusive adsorption—are silent and low cost and thus suitable for simultaneous long-term, multi-site indoor monitoring (Hazrati and Harrad 2006). Furthermore, recent studies have shown that PUF-PAS samplers collect both gas- and particle-phase pollutants at similar sampling rates (Bohlin et al. 2014a; Harner et al. 2014; Harner et al. 2013; Markovic et al. 2015), so the pollutants sampled by PUF-PAS can be compared with those collected by active air samplers. Shanghai is the largest city in China, with a population of about 23 million; human inhalation exposure to PBDEs and consequent health risks are of great concern due to the intensive manufacturing and usage of PBDEcontaining products. However, data for PBDEs in indoor and outdoor air in Shanghai were very limited. Li et al. (2015) reported PBDE concentrations in outdoor total suspended particles (TSP) and gas phase samples in Shanghai based on short-term sampling at one urban and one rural site. Due to the potential huge temporal and spatial variations in the concentrations of airborne pollutants, sampling over longer periods at more sites was needed to give a time-weighted average concentration of PBDEs in the atmosphere. Moreover, as people spend most of their time indoors, it is hard to assess human

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inhalation exposure without information on PBDE concentrations in indoor air. In this study, PUF-PAS samplers were used to investigate the concentrations, congener profiles, seasonal variations, influencing factors, emission sources, and human inhalation exposure to PBDEs in both indoor and outdoor air in Shanghai. To our knowledge, this is the first research on PBDEs conducted concurrently in indoor and outdoor air in China, using passive air samplers.

Materials and methods Sampling Twenty homes and 20 offices as well as 10 outdoor sites were selected in the urban and suburban areas of Shanghai for passive air sampling; descriptions of the sampling sites are listed in Table S1-3. The sampling sites were scattered across 12 districts in Shanghai: Luwan (LW), Jing’an (JA), Changning (CN), Xuhui (XH), Zhabei (ZB), Putuo (PT), Yangpu (YP), Hongkou (HK), Minhang (MH), Pudong (PD), Baoshan (BS), and Jiading (JD). Samples were collected for about 56 days in January–February of 2008 (winter) and July–August of 2008 (summer); a total of 97 samples were collected. A questionnaire was used at each sampling site to collect information on the indoor environment. Prior to sampling, polyurethane foam disks (PUF disks, 14 cm diameter×1.35 cm thick; density 0.0213 g cm−3) were Soxhlet extracted with a methanol and acetone/hexane mixture (1:1) for 48 h each, sequentially. The pre-cleaned polyurethane foam disks were individually wrapped in pre-cleaned aluminum foil and sealed with double layers of polyethylene bags. The samplers were cleaned and rinsed with solvent prior to use and were assembled at the sampling sites to avoid contamination during transportation. After sampling, the PUF disks were individually wrapped in pre-cleaned aluminum foil, sealed with double layers of polyethylene bags, and stored at −20 °C until analysis. The extraction and cleanup were carried out within 1 month after sampling.

Standards Sixteen PBDE standards (BDE-17, BDE-28, BDE-71, BDE47, BDE-66, BDE-77, BDE-100, BDE-99, BDE-118, BDE85, BDE-154, BDE-153, BDE-138, BDE-183, BDE-190, and BDE-209) were purchased from AccuStandards (New Haven, CT, USA); 13C12-labeled internal standard (13C12-CB-208) and surrogates (13C12-CB-141, 13C12-CB-209) were purchased from Cambridge Isotope Laboratories (Andover, MA, USA).

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Extraction and cleanup The extraction protocol used was a modified version of that reported by Chen et al. (2006). In brief, the samples were spiked with 13C12-CB-141 and 13C12-CB-209 and Soxhlet extracted with a mixture of acetone/hexane (1:1) for 48 h. The concentrated extracts were cleaned and fractionated through acidic/basic multilayer silica/alumina columns, and the fraction containing the PBDEs was eluted with 70 mL hexane/methylene chloride (1:1). The final extracts were concentrated under a gentle N2 stream to 50 μL, and a known amount of internal standard (13C12-CB-208) was added prior to instrumental analysis. Instrumental analysis PBDE concentrations were measured by GC-MS (Agilent 6890N Gas Chromatograph coupled with Agilent 5975 Mass Selective Detector) with negative chemical ionization (NCI) in selected ion monitoring (SIM) mode. PBDE congeners, except for BDE-209, were determined with a DB-5MS capillary column (30 m length×0.25 mm i.d×0.25 μm film thickness). The column temperature was initiated at 110 °C (held for 1 min), increased to 180 °C at 20 °C min−1 (held for 1 min), then to 280 °C at 2 °C min−1 (held for 1 min), and finally to 305 °C at 5 °C min−1 (held for 15 min). The pulsed splitless mode was used for manual injection of 2 μL samples. Methane was used as a chemical ionization moderating gas and helium as the carrier gas at a flow rate of 1 mL min−1. For BDE-209, a short DB-5MS capillary column (9 m×0.25 mm i.d×0.1 μm film thickness) was used to reduce its degradation in the column. The oven temperature was programmed from 110 °C (held for 1 min) to 300 °C at a rate of 10 °C min−1 (held for 6 min). The ion source, inlet, and interface temperatures were set to 250, 290, and 285 °C, respectively. The 6-level linear calibration curve method was used to calculate the relative response factors of each PBDE congener to the internal standard.

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Σ15PBDEs (15 congeners without BDE-209), 0.6 m3 day−1 for BDE-209 in the indoor environment, and 6.7 ± 1.5 m3 day−1 for ΣPBDEs (16 congeners including BDE209) in the outdoor environment. Recent studies further showed that the PUF-PAS sampling rates for gas- and particle-phase pollutants were similar (Bohlin et al. 2014a; Harner et al. 2014; Harner et al. 2013; Markovic et al. 2015). The average R e q of 0.9 for Σ 1 5 PBDEs and 0.6 m3 day−1 for BDE-209 in the indoor environment and 6.7 m3 day−1 for ΣPBDEs in the outdoor environment were used in this study. The indoor Req in this study was lower than the values of 1.1–1.9 m3 day−1 for a similar passive air sampler reported by Hazrati and Harrad (2007), which may have been because of the difference between the ventilation conditions of the sites in this study compared with those of Hazrati and Harrad. The outdoor Req was within the range of other studies (2–24 m3 day−1, while in most cases 3–10 m3 day−1, Pozo et al. 2009; Tuduri et al. 2006). The higher sampling rate in the outdoor air compared with the indoor air was mainly due to the higher outdoor wind speed (Tuduri et al. 2006). It should be noted that the PAS sampling rates would be affected by both air circulation and wind speed; thus, the sampling rates in different homes or outdoor sites might be different. However, the variations in sampling rates should be low because the Bflying saucer^ housing of the PUF-PAS can adequately dampen the wind effect (Markovic et al. 2015; Tuduri et al. 2006). It was for this dampening effect that fully sheltered PUF-PAS (double-dome) samplers were chosen for this study instead of the partly sheltered ones with a more open design and just a top cover, though the sampling rate of the partly sheltered PAS was higher (Bohlin et al. 2014b; Harrad and Abdallah 2008). The average outdoor wind speeds during our summer and winter sampling periods in Shanghai were almost the same at 3.25 and 3.15 m s−1 (weather conditions such as wind speed and ambient temperature were gathered from http://www.wunderground.com). The results from passive sampling in this study should reflect the distribution of PBDEs in Shanghai, despite the aforementioned uncertainties in the sampling rates.

Calibration of PAS sampling rates Active air samplers were used to calibrate the PAS sampling rates; one indoor and one outdoor sampling site were selected for the calibration study. Gaseous and particle phase samples were collected weekly during the winter PAS sampling period (8 weeks). The equivalent air sampling rate of PAS (Req) was calculated as Req =M/(CAΔt), where M was the mass (pg) of PBDEs captured by the PUF-PAS (Table S4), while CA was the concentration (pg m−3) of PBDEs from the active air sampler and Δt was the sampling time (days) of PUF-PAS. In accordance with the results of Hazrati and Harrad (2007), the Req values for individual BDE congeners were different, but the variability was relatively low: 0.9±0.2 m3 day−1 for

QA/QC Field and procedure blanks were analyzed, and only low levels of BDE-47 (18 pg/PUF) and BDE-209 (52 pg/PUF) were detected (n=8) in the field blanks. The blank levels were less than 10 % of the concentrations in most of the samples. The mean recovery of the 16 PBDEs measured was 97±7 % for 3–7 Br BDEs and 88±17 % for BDE-209 in the spiked blank samples (n=7). Duplicate samples showed deviations of less than 12 % for the 16 PBDEs congeners. Surrogate recoveries were 97±12 % for 13C12-CB-141 and 13C12-CB-209 in all the samples (n=114). Concentrations reported were blank-

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subtracted, but not recovery-corrected. The instrumental detection limits (IDLs) of PBDE congeners were determined and are given in Table S5.

Results and discussion Concentrations and seasonal variations of PBDEs Σ15PBDE and BDE-209 concentrations in both indoor and outdoor air followed a log-normal distribution for both seasons (Kolmogorov-Smirnov test, SPSS 16.0, α=0.05). The median concentrations of Σ 15 PBDEs in the offices (82 pg m−3 in summer, 29 pg m−3 in winter, Table 1) in the urban area of Shanghai were about 3 times their concentration in homes in both seasons (30 pg m−3 in summer and 10 pg m−3 in winter), quite similar to other reported studies (Chen et al. 2008; Gevao et al. 2006; Harrad et al. 2006). This is attributable to the intensive usage of electronic office equipment such as copiers, computers, scanners, and printers. Higher concentrations of Σ15PBDEs in summer than in winter could be expected because PBDEs evaporate more readily in higher summer temperatures. Indoor concentrations of BDE-209 followed the same seasonal trend as Σ15PBDEs. Office concentrations of BDE-209 Table 1 Concentrations of PBDEs in indoor and outdoor air in Shanghai (pg m−3) Summer

Winter

Home Office Outdoor Home Office Outdoor Σ15PBDE Median 30 GMa

38 Mean 58 SD 55 Min 7 Max 212 BDE-209 Median 130 GM 138 Mean 213 SD 225 Min 29 Max 744 ΣPBDE Median 188 GM 199 Mean 271 SD 236 Min 45 Max 813 a

82

12

10

29

6

91 175 344 19 1535 90 107 138 111 39 412 218 229 313 384 92 1793

13 14 4 6 22 398 322 454 332 51 945 411 339 468 335 58 965

11 20 32 2 145 46 56 67 42 15 145 64 73 87 58 18 271

28 44 64 4 304 47 49 75 68 5 271 84 92 118 85 25 331

6 7 2 4 12 794 610 823 556 113 1629 799 618 830 557 117 1635

GM represents geometric mean

in winter (47 pg m−3) were comparable to those in homes (46 pg m −3), while summer concentrations were lower (90 pg m−3) in offices than in homes (130 pg m−3). Modern office airspaces are usually closed and are air-conditioned for a longer time than homes, so less exchange of air with the outdoors was to be expected (Barro et al. 2009). Spearman correlations (α=0.05, SPSS 16.0) were carried out between Σ15PBDEs and BDE-209 in homes or offices for each season, and no significant correlations were found (home: r=0.10, p=0.68 in summer, r=0.20, p=0.42 in winter; office: r=−0.17, p=0.50 in summer, r=0.001, p=1.00 in winter), indicating that the emission sources of Σ15PBDEs and BDE-209 were different (Dodson et al. 2015; Schecter et al. 2005). Significant correlations were found between the winter and summer Σ15PBDE concentrations at homes (r=0.68, p= 0.001) and in offices (r=0.86, p1 were identified, which explained 82 % of the total variance (Table 2). The congeners related to component 1 (C1) with factor loadings greater than 0.8 were BDE-47, BDE-100, BDE-99, BDE-85, BDE-154, and BDE153, which are the main components of commercial PentaBDE products (Besis and Samara 2012; La Guardia et al. 2006). The high contribution of C1 to the total variance indicated that the emission of the lower brominated congeners could easily be affected by factors such as the ambient temperature and the working temperature of the electrical/

Office Summer

400

200

0 Office Winter

300

200

100

W 2 -J O A -X O H1 -X O H2 -X O H3 -X O H4 -X H O 5 -P D O -B O S -Z B O 1 -Z B O 2 -Z B3 O -Z B O 4 -Z B5 O -Z B O 6 -Z B O 7 -Z B8 O

-L O

-L

W

1

0 O

-3

Concentrations (pg m )

Fig. 3 Concentrations and profiles of ΣPBDEs in indoor air of offices in Shanghai, China (Abbreviations in the horizontal axis represent the districts where the sampled Offices (O-) located)

Concentrations (pg m )

BDEs in summer. Our data also showed that more lower brominated BDEs were emitted in offices than at homes. BDE-47, BDE-99, BDE-28, and BDE-17 were also abundant in the indoor air (14±19 % in summer and 7±15 % in winter of the ΣPBDEs in homes, 19±23 % in summer, and 18 ±23 % in winter in offices), revealing that lower brominated BDEs were still being released into the air even after the phase out of the usage of Penta-BDE products as flame retardants (Besis and Samara 2012; Law et al. 2014). In contrast, BDE183 was only detected in a few offices (0.9 % of the ΣPBDEs in winter and 1.5 % in summer), indicating the rare usage of Octa-BDE (Besis and Samara 2012; La Guardia et al. 2006). Principal component analysis (PCA) was conducted to investigate the relationship between the PBDE congeners. The concentration data of PBDE congeners in the indoor air of

BDE-209 BDE-190 BDE-183 BDE-138 BDE-153 BDE-154 BDE-85 BDE-118 BDE-99 BDE-100 BDE-77 BDE-66 BDE-47 BDE-71 BDE-28 BDE-17

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Table 2 Principal component analysis of PBDE congeners in indoor air in Shanghai, China (n=79) Component matrix C1 (39.1 %)

C2 (19.3 %)

C3 (16.4 %)

BDE47

0.903

−0.240

0.097

0.180

BDE100

0.906

−0.414

−0.017

−0.041

BDE99 BDE85

0.902 0.940

−0.423 −0.256

−0.020 −0.148

−0.042 −0.072

BDE154 BDE153

0.952 0.911

−0.248 −0.117

−0.134 −0.133

−0.071 −0.103

conducted to investigate possible influencing factors/sources of PBDEs in indoor and outdoor air, such as the quantity of electrical/electronic appliances and the time for which they were used, as well as interior furnishing and decorations.

C4 (7.4 %)

BDE138

0.457

0.627

−0.485

0.011

BDE183 BDE190

0.333 0.324

0.745 0.768

−0.494 −0.464

−0.179 −0.145

BDE17 BDE28

0.214 0.207

0.228 0.311

0.716 0.806

−0.269 −0.271

BDE71 BDE66 BDE77

0.139 0.569 0.302

0.275 0.291 0.602

0.462 0.624 0.283

−0.069 0.220 0.394

BDE118 BDE209

0.424 0.233

0.544 0.183

0.102 0.009

−0.166 0.828

electronic products. The congeners related to C2 were BDE183, BDE-190, and BDE-138, which are the main components of commercial Octa-BDE products (Besis and Samara 2012; La Guardia et al. 2006). The congeners related to C3 were BDE-17 and BDE-28, which may be due to their different physical chemical properties as compared to other congeners. A factor score plot (Fig. S1-c) showed that the score values of C3 (BDE-17, 28) for the summer indoor samples had a wider range of variation than the others. The congener in C4 was BDE-209, which could be interpreted as the contribution of Deca-BDE. The factor score plot (Fig. S1-c) showed that the variation of score values for the outdoor sites were mainly on C4 (BDE-209). The results of the factor analysis indicated the different environmental behavior of commercial Penta-, Octa-, and Deca-BDE flame retardants. As compared to the indoor environment, lower brominated BDEs appeared in very low concentrations in the outdoor air (Fig. 4), and BDE-209 accounted for 97±3 % of the ΣPBDEs in summer and 99±1 % in winter. The obvious difference in PBDE composition in indoor and outdoor air indicated different sources. The dominance of BDE-209 in outdoor air showed that Deca-BDE is practically the only commercial PBDE being used in Shanghai today. Factors affecting PBDE levels in indoor air in Shanghai The study found distinct differences between concentrations of ΣPBDE in homes and offices, but no regular spatial variation could be identified (Figs. 2 and 3). Statistical analysis was

Usage of electrical/electronic appliances The lowest level of ΣPBDEs was observed in a dormitory room on the campus of Shanghai University (H-ZB1, 45 pg m−3 in summer and 18 pg m−3 in winter). The room contained only beds, desks, and chairs. Another room on the same campus (H-ZB2) had a TV set and one desktop computer. There, ΣPBDE concentrations were much higher (4 times higher in winter and 18 times higher in summer than in HZB1), indicating that indoor emissions from the usage of electrical/electronic appliances are an important source of PBDEs. Further analysis showed that summer concentrations of Σ15PBDE in homes were significantly correlated with the age of TV sets (r=0.81, p=0.0003, Spearman correlation, α= 0.05), revealing that the old TV sets contained more lower brominated BDEs (Guo et al. 2015). Σ15PBDE concentrations in offices were significantly correlated with the number of personal computers (PCs) (summer: r=0.60, p=0.015; winter: r=0.51, p=0.043) and moderately correlated with the use time of PCs (summer: r=0.47, p= 0.075; winter: r=0.42, p=0.116), showing that PCs in use were an important source of Σ15PBDE (Betts 2006; Hazrati and Harrad 2006). The seasonal change in ΣPBDE concentrations in a school office with four partitions (O-BS) supported the contribution of ΣPBDEs in indoor air from the usage of computers. In winter, the PAS sampler was put in the partition with 4 desktop computers operating for more than 10 h day−1, while in summer, air was sampled in a partition with only one intermittently used computer: levels of ΣPBDEs at O-BS were higher in winter than in summer, which was the opposite of the results from other offices. Air conditioners have been found to have an obvious influence on the levels of ΣPBDEs in indoor air. BDE-209 concentrations in offices were significantly correlated with the running time of air conditioners in summer (r=0.54, p= 0.046), indicating that air conditioning in cooling mode could significantly elevate indoor BDE-209 concentrations. The influence of air conditioners on levels of ΣPBDEs in indoor air was especially strong when they were in the heating mode. Concentrations of ΣPBDEs at H-CN were the highest among all homes (271 pg m−3 in winter, 733 pg m−3 in summer), and the fraction of Σ15PBDEs was higher than in the other homes in winter. The resident of H-CN worked at home during the sampling periods; thus, the TV, computer, and air conditioner stayed on for longer everyday compared to other homes (although the exact time was not recorded). A similar phenomenon was also seen at office O-ZB4. The air turbulence caused by the air conditioning probably increased the sampling rate

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Concentrations (pg m )

Fig. 4 Concentrations and profiles of ΣPBDEs in outdoor air in Shanghai, China (Abbreviations in the horizontal axis were the same as that in Fig. 1)

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Outdoor Summer

800 400

20 10

-3

Concentrations (pg m )

0 1600

Outdoor Winter

1200 800 400

10

BDE-209 BDE-190 BDE-183 BDE-138 BDE-153 BDE-154 BDE-85 BDE-118 BDE-99 BDE-100 BDE-77 BDE-66 BDE-47 BDE-71 BDE-28 BDE-17

5 0 ZB

despite the dampening effect of the Bflying saucer^ structure of the PUF-PAS sampler, while the higher Σ15PBDE fraction in winter indicated the higher emission rate of PBDEs due to running air conditioners in heating mode. The proportion of BDE-209 in ΣPBDEs varied greatly between different homes (38–96 % in winter and 22–97 % in summer). Higher proportions of lower brominated BDEs were usually detected in homes/offices with old electrical/ electronic appliances, such as in H-ZB4, H-PT2, and H-CN (Fig. 2, Table S1), indicating that old electrical appliances may contain relatively more Penta-BDE (Betts 2006; Guo et al. 2015; Hazrati and Harrad 2006). Σ15PBDEs in offices O-ZB1-3 supported the finding that old electrical/electronic equipment emits more lower brominated BDEs. The main difference between the sites O-ZB1-3 was that the computers and air conditioner at O-ZB1 were new while those at O-ZB3 were the oldest. Concentrations of ΣPBDEs and the proportion of lower brominated BDEs followed the order of O-ZB3>O-ZB2>O-ZB1. Site O-PD belonged to a small company; it was the only office without air conditioning but had many old computers and printers. The concentration of ΣPBDEs in summer at O-PD (1793 pg m−3) was the highest among all the offices sampled and was about 12 times the concentration in winter. The PBDE profile showed that Penta-BDE was the major component (73 %) at O-PD in summer, indicating that many lower brominated BDEs were emitted from the Bhot^ computers and printers under high summer temperatures. It should be noted that the higher sampling rate caused by the usage of fans in O-PD would also make the summer concentration appear to be high. Our results raised the concern that many electrical/ electronic appliances containing Penta- and Octa-BDE are still in use despite the fact that usage of lower brominated BDEs as

PT

CN

XH1 XH2

BS

JD

PD1 PD2

YP

additives has been stopped for several years. Many obsolete electrical/electronic appliances such as computers, TV sets, and air conditioners containing Penta-BDE find their way to homes through the second-hand market (Babbitt et al. 2011; Williams et al. 2008). Penta- and Octa-BDE from those old appliances will remain as sources of human exposure for a longer time than expected. Interior furnishing and furniture Samples were collected at a newly decorated home (H-PT1), which had new electrical appliances and furniture but was uninhabited, to investigate the impact of furnishings and furniture on the indoor ΣPBDEs. The concentrations of ΣPBDEs (mainly BDE-209) at H-PT1 were just slightly higher than the lowest level found, at H-ZB1, suggesting that electrical appliances not in use and decoration materials did not have a great impact on the indoor PBDE level, or that the newly purchased electrical appliances did not contain PBDEs. A similar situation was also found in office O-ZB5, a newly decorated and furnished but unoccupied office at Shanghai University. The concentration of ΣPBDEs at O-ZB5 (92 pg m−3 in summer and 28 pg m−3 in winter, Fig. 3) was the lowest in all the sampled offices. Penta-BDE is usually added to foam pieces in furniture (Abbasi et al. 2015). However, the concentrations of PentaBDE (sum of BDE-47, 100, 99, 85, 154, 153) in this study were not significantly correlated with the number of foam pieces in either the homes or the offices (judged at significance level of α=0.05). Further analysis showed that individual PBDE congeners of Penta-BDE were also not significantly correlated with the number of foam pieces, suggesting that foam pieces in chairs and sofas in homes or offices were not an important

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source of indoor Penta-BDE. This was consistent with reports that flame retardants are not added to foam in house furniture in China, to decrease production costs (Chen et al. 2008).

absorption of the inhaled PBDE was assumed, representing an upper limit of uptake. Human daily exposure to ΣPBDEs was calculated using the following formula:

Factors affecting PBDE levels in outdoor air

Σexposurei ¼ ½ðC w F w Þ þ ðC h F h Þ þ ðC o F o ÞRr

Eighteen outdoor air samples were collected at urban and suburban sites in Shanghai. Concentrations and profiles of ΣPBDEs are shown in Fig. 4. The lowest levels of ΣPBDEs were found at PD1 in winter, which was located in a suburban area, indicating relatively low concentrations of ΣPBDEs in suburbs. In order to confirm this spatial variation, we took an additional sample in summer at a site closer to the urban area, though in the same district (PD2). The levels of ΣPBDEs at PD2 were about 3.3 times of that at PD1 and comparable to other urban areas, implying much higher emission of ΣPBDEs in the urban area than in the suburban areas (Hearn et al. 2012; Melymuk et al. 2014; Melymuk et al. 2012; Wang et al. 2012). Sampling height obviously influenced the ΣPBDE concentrations. XH1 and XH2 were two sampling sites close to each other in the urban area, but the sampling height of XH1 was about 35 m while that of XH2 was about 4 m. ΣPBDE levels at XH2 were about 3 times those at XH1, indicating groundlevel sources of PBDEs such as vehicle emissions and resuspended road dust. ΣPBDE levels were relatively higher in traditional industrial areas such as BS, YP, and JD (Fig. 4), indicating that industrial activities were important sources of airborne ΣPBDEs (Hearn et al. 2012). The seasonal variation of ΣPBDEs at sampling site ZB would be an indication of industrial emissions: The ΣPBDE level in winter was about 10 times that in summer. ZB was near the border of the residential and commercial areas, many industrial manufacturing plants were scattered to the north, and the traditional industrial zones (such as BS and JD) were to its north/northwest. In the summer, when the prevailing wind was from southeast, the ΣPBDE levels at ZB reflected the ΣPBDE levels of the central urban area; in the winter, the prevailing north/northwest wind would bring pollutants from the industrial areas to the sampling site. At site PT, the levels of ΣPBDEs were higher in summer than in winter, which was different from the seasonal trend at the other sites. A chemical plant, about 1.5 km to the south of the sampling site, should be responsible for the higher summer concentrations at PT. Human inhalation exposure to PBDEs in Shanghai Samples collected by PUF-PAS were considered to be representative of inhaled PBDEs because they trapped pollutants in both the gas and particle phase (Bohlin et al. 2014a; Harner et al. 2014; Harner et al. 2013; Markovic et al. 2015). Due to the lack of absorption efficiency data for PBDEs, 100 %

where Σexposurei is the daily human exposure through inhalation, Cw/h/o is the ΣPBDE concentration in the workplace/home/ outdoor air, respectively, Rr is the inhalation rates, which are about 16 m3 day−1 for adults and 8 m3 day−1 for toddlers (USEPA 2011), and Fw/h/o is the respective fraction of time spent at the workplace/home/outdoors. Taking weekends into consideration, Fw, Fh, and Fo for adults were 23.8 % (5.7 h day−1), 67.9 % (16.3 h day−1), and 8.3 % (2 h day−1), and 0, 91.7 % (22 h day−1), and 8.3 % (2 h day−1) respectively for toddlers (Currado and Harrad 1998; US-EPA 2011). The median daily human exposures to Σ15PBDEs and BDE-209 through inhalation for adults thus calculated were: 0.23 and 1.73 ng day−1 in winter and 0.65 and 2.28 ng day−1 in summer. The exposures for toddlers were: 0.15 and 1.04 ng day−1 in winter and 0.23 and 1.22 ng day−1 in summer. The inhalation exposure level to Σ15PBDEs in Shanghai is comparable to that in Birmingham, England (0.82 ng day−1) and Ottawa, Canada (2 ng day−1), but is higher than in Kuwait (0.14 ng day−1) (Gevao et al. 2006; Harrad et al. 2006; Wilford et al. 2004). The inhalation exposure to Σ15PBDEs in Shanghai varied greatly from site to site, the high-to-low ratio was about 57 in winter and 51 in summer, with a maximum exposure of 2.75 ng day −1 in winter and 8.18 ng day−1 in summer. The inhalation exposure to BDE209 showed a variation of about 14-fold in winter and 21-fold in summer, with a maximum exposure of 4.77 and 10.90 ng day−1 seasonally. In the total inhalation exposure, 83 % of the total came from BDE-209 in winter and 67 % in summer. Although BDE-209 was assumed to have low bioavailability because of its high molecular weight and high hydrophobicity, it has been detected in the blood of occupationally exposed and unexposed individuals (Bennett et al. 2015; Bi et al. 2007; Watkins et al. 2011). Moreover, BDE-209 can be metabolized into OH-PBDEs in the human body, which might inflict more adverse health effects (Yu et al. 2010). Due to the relatively high concentration of PBDEs in fish compared with other food, fish was believed to be a key route of dietary exposure to PBDEs, accounting for ∼50 % of the bioaccessible PBDEs through dietary intake (Domingo 2012; Yu et al. 2011). Comparison of the inhalation exposure with fish eating was conducted to assess the relative importance of inhaled and eaten PBDEs. Human exposure to PBDEs via fish eating was estimated using the reported PBDE concentrations in Chinese consumer fish (156 pg g−1, wet weight, Meng et al. 2007) and freshwater fish from the Taihu Lake of China (531 pg g−1, Yu et al. 2012). Daily fish intake (13.8 g day−1 for adults and 3.9 g day−1 for toddlers) was based on US-EPA recommendations (US-EPA 2011). The calculated daily intake

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of ΣPBDEs via inhalation for adults and toddlers (3.44 and 1.33 ng day−1) was comparable with that from eating fish (4.74 and 1.34 ng day−1, calculated by averaging the studies of Meng et al. and Yu et al.), demonstrating the importance of human inhalation exposure to PBDEs for both toddlers and adults in Shanghai.

Conclusions The different seasonal variations and congener compositions of ΣPBDEs in indoor and outdoor air in Shanghai revealed that their respective emission sources were different. Indoor emissions were one major source of lower brominated BDEs in and near cities, while outdoor emissions were the main sources of BDE-209, though BDE-209 in indoor air could be affected by outdoor air. Indoor air concentrations of ΣPBDEs were significantly influenced by the electrical/ electronic appliances in use but not by interior furnishing and decorations. Industrial emissions should be the key sources of ΣPBDEs in outdoor air. Ventilation by opening window would decrease the indoor concentration of Σ15PBDEs but increase BDE-209. Penta- and Octa-BDE from old appliances still in use will remain as sources of human exposure for a longer time than expected. The inhalation exposure to ΣPBDEs for adults in Shanghai was 2.6±2.1 and 4.3±5.6 ng day−1 in winter and summer, respectively, comparable to exposure via fish eating. Acknowledgments This study was financially supported by the Natural Science Foundation of China (41203077, 20877052, 41173097), the Natural Science Foundation of Fujian Province, China (2011J05112), the Science and Technology Project of Quanzhou, China (2012Z85), and the Program for Changjiang Scholars and Innovative Research Team in University (IRT13078), for which the authors are grateful. We deeply appreciate the advice and comments given by Prof. Ming Fang and the anonymous reviewers.

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