Measurement of legacy and emerging flame retardants in indoor dust from a rural village (Kopawa) in Nepal: Implication for source apportionment and health risk assessment

Ecotoxicology and Environmental Safety 168 (2019) 304–314

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Measurement of legacy and emerging flame retardants in indoor dust from a rural village (Kopawa) in Nepal: Implication for source apportionment and health risk assessment

T

⁎

Ishwar Chandra Yadava,b, , Ningombam Linthoingambi Devic, Vipin Kumar Singhd, Jun Lia, Gan Zhanga a

State Key Laboratory of Organic Geochemistry, Guangzhou Institute of Geochemistry, Chinese Academy of Sciences, Guangzhou 510640, PR China Department of International Environmental and Agricultural Science (IEAS), Tokyo University of Agriculture and Technology (TUAT) 3-5-8, Saiwai-Cho, Fuchu-Shi, Tokyo 1838509, Japan c Centre for Environmental Sciences, Central University of South Bihar, SH-7, Gaya-Panchanpur, Post-Fatehpur, P.S-Tekari, District-Gaya 824236, Bihar, India d Department of Botany, Banaras Hindu University, Varanasi 221005, Uttar Pradesh, India b

A R T I C LE I N FO

A B S T R A C T

Keywords: Brominated flame retardants Organophosphate esters Consumer materials House dust Nepal Health risk exposure

Under the Stockholm Convention, signatory countries are obliged to direct source inventories, find current sources, and provide ecological monitoring evidence to guarantee that the encompassing levels of persistent organic pollutants (POPs) are declining. However, such monitoring of different types of POPs are to a great degree constrained in most developing countries including Nepal and are primarily confined to suspected source area/ densely populated regions. In this study, 9 polybrominated diphenyl ethers (PBDEs), 2 dechlorane plus (DPs), 6 novel brominated flame retardants (NBFRs) and 8 organophosphate ester flame retardants (OPFRs) were investigated in indoor dust from a rural area (Kopawa) in Nepal in order to evaluate their occurrence/level, profile, spatial distribution and their sources. Additionally, health risk exposure was estimated to anticipate the possible health risk to the local population. The results showed that OPFRs was the most abundant FR measured in the dust. The concentration of ∑8OPFRs was about 2, 3 and 4 orders of magnitude higher than the ∑6NBFRs, ∑9PBDEs, and ∑2DPs, respectively. Tris (methylphenyl) phosphate (TMPP) and Tris (2-ethylhexyl) phosphate (TEHP) were the most abundant OPFRs analyzed in the dust; while decabromodiphenyl ethane (DBDPE) exceeded among NBFRs. Likewise, 2,2′,3,3′,4,4′,5,5′,6,6′-decabromodiphenylether (BDE-209) was the most identified chemical among PBDEs. The total organic carbon (TOC) content in dust was significantly and positively connected with octa-BDEs (Rho = 0.615, p < 0.01), BTBPE (Rho = 0.733, p < 0.01), TPHP (Rho = 0.621, p < 0.01), TEHP (Rho = 0.560, p < 0.01) and TMPPs (Rho = 0.550, p < 0.01), while black carbon (BC) was either weakly related or not related, suggesting little or no impact of BC in the distribution of FRs. Principal component analysis indicated the contribution from commercial penta-, octa- and deca-BDEs formulation, the adhesive substance, food packaging and paints, and degradation of BDE-209 as the essential sources of FRs. Health risk exposure estimates showed that dermal absorption via dust as the primary route of FRs intake. The estimated daily exposure of PBDEs, NBFRs and OPFRs were 2–10 orders of magnitude lower than their corresponding reference dose (RfD), suggesting insignificant risk. However, other routes such as inhalation and dietary intake might still be significant in the case of Kopawa which should be tested in future.

1. Introduction Flame retardants (FRs) are a class of synthetic compounds which are added to consumer materials, for instance, carpets, electronic appliances, clothing and textiles, thermal insulation and cable coatings to slow ignition and the spread of fire (Alaee et al., 2003). Since the 1990s,

polybrominated diphenyl ethers (PBDEs) are one of the most broadly utilized chemicals among the different class of FRs. Worldwide studies on PBDEs have indicated persistent, bio-accumulative and toxic characteristics of PBDEs (Darnerud, 2003; Law et al., 2014). Consequently, in 2004, the commercial Penta-BDE and Octa-BDE mixtures were eliminated from North American and European markets for further

⁎ Corresponding author at: State Key Laboratory of Organic Geochemistry, Guangzhou Institute of Geochemistry, Chinese Academy of Sciences, Guangzhou 510640, PR China.

https://doi.org/10.1016/j.ecoenv.2018.10.089 Received 1 September 2018; Received in revised form 21 October 2018; Accepted 24 October 2018 0147-6513/ © 2018 Elsevier Inc. All rights reserved.

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situated in Rautahat district of central region of Nepal, approximately 92 km south of capital city Kathmandu (Fig. S1, Supplementary information). It is located between 26° 52' 36.84"N to 85° 13' 17.25"E at an altitude of 83 m above the mean sea level (MSL). As per the national census of 2011, Kopawa had the population of about 4074 (CBS/NPCS, 2011). A total of 21 indoor dusts were collected from 21 different houses of the village in July 2015. The characteristics of dust sampled houses are given in Table S1, Supplementary information. Dust samples were collected by sweeping of the kitchen room, study room, bedroom, living room and passage of concerned households. The dust samples were manually collected using house broom into a dustpan. Approximately 50 g of dust were acquired and placed in a zipper bag prior to bringing to the laboratory. At the laboratory, dust samples were sieved with mess size of 500 µm and stored at − 20 °C until analysis.

trading. They were enlisted in the persistent organic pollutants (POPs) list of the Stockholm Convention, limiting their application in signatory countries (Stockholm Convention, 2009a, 2009b). Later, the production of the deca-BDE mixture was also suspended at the end of 2012 (USEPA, 2013). However, the discontinuation of such PBDE mixtures has delighted the increased use of existing chemicals or development of alternative chemicals to displace PBDEs for meeting fire resistant standard. Because of governmental restriction on the commercial application of PBDEs, they were displaced with alternatives halogenated flame retardants, commonly referred to as "emerging" flame retardants (EFRs). These EFRs are subcategorized into groups, i.e. novel brominated flame retardant (NBFR), dechlorane plus (DP) and organophosphate ester flame retardant (OPFR) (Stapleton et al., 2008; Wang et al., 2011; van der Veen and de Boer, 2012). NBFR comprises 2, 3, 4, 5, 6-pentabromoethylbenzene (PBEB), hexabromobenzene (HBB), 2-Ethylhexyl2,3,4,5-tetrabromobenzoate (EH-TBB), 1,2-bis (2,4,6-tribromophenoxy) ethane (BTBPE), bis (2-ethylhexyl)-3,4,5,6-tetrabromophthalate (BEH-TEBP), decabromodiphenyl ethane (DBDPE). DP includes syn- and anti-dechlorane plus. OPFR represents a group of halogenated or non-halogenated compounds with tri-ester structures, comprising tri-n-butyl phosphate (TNBP), tris(2-chloroethyl) phosphate (TCEP), tris(1-chloro-2-propyl) phosphate (TCIPP), triphenyl phosphate (TPHP), and tris(1,3-dichloro-2-propyl) phosphate (TDCIPP), and few others (van der Veen and de Boer, 2012). After the worldwide restriction on the usages of PBDEs, penta-BDE was replaced by EH-TBB and BEH-TEBP, octa-BDE by BTBPE, and deca-BDE by DBDPE and DP. However, the displacements of PBDEs by EFRs have shown similar environmental fate and human health impact. Evidence suggests that the usages of EFRs in consumer materials have been expanded rapidly in recent year (Brommer and Harrad, 2015; Yadav et al., 2017a; Cao et al., 2014; Tajima et al., 2014; Newton et al., 2015; Cequier et al., 2014). It is estimated that around 70 different EFR chemicals have been recognized as an industrial formulation (Covaci et al., 2011). These include a variety of brominated or chlorinated substances as well as phosphate substance. Indoor dust acts as a repository sink for many organic pollutants, including PBDEs and EFRs. Dust from residential houses, offices, and stores are considered as the significant sources of different chemicals to human via ingestion (Alves et al., 2014; Jones-Otazo et al., 2005). Dust in the indoor environment originated from consumer products or/and building materials may act as an important vector for different types of chemical to human exposure. Chemicals related to dust can enter the body by means of ingestion after hand-to-mouth contact, inhalation of re-suspended dust, or direct contact through the skin (Whitehead et al., 2011). Past examinations have shown the ubiquitous distribution of PBDEs and EFRs in the indoor dust (Harrad et al., 2010; Tajima et al., 2014; Newton et al., 2015; Cequier et al., 2014; Yadav et al., 2017a). These investigations recommended indoor dust as a critical exposure pathway for PBDEs and perhaps for other FRs. While a couple of previous studies reported an elevated level of legacy and emerging FRs from suspected source areas/densely populated regions of Nepal (Yadav et al., 2017a, 2017b, 2018a, 2018b, 2018c), the fate and transport, and sources of legacy and EFRs are extremely limited in remote areas. This study aims to investigate the environmental concentration and sources of PBDEs, NBFR, DP, and OPFR in indoor dust from a rural village (Kopawa) in Nepal, characterize spatial variation, and to investigate human exposure risk via dust intake. Additionally, total organic carbon (TOC) and black carbon (BC) in dust were evaluated to assess the role of TOC/BC in the dissemination of FRs.

2.2. Determination of TOC and BC TOC and BC content in dust were analyzed based on techniques recommended by Ramu et al. (2010), and Gustafsson et al. (2001), respectively. The detailed procedure about sample pre-treatment and analysis have been described in our previous study (Yadav et al., 2018d). Briefly, about 2–3 g of sieved and homogeneously mixed dust was treated with 3 mL of 10% HCl to remove inorganic carbon. The dust was then washed with Milli-Q water (three times) and dried overnight in an oven at 45 °C. A portion of this was utilized to determine TOC by Elemental Carbon-Hydrogen-Nitrogen Analyzer (Elementar VARIO EL III). For BC analysis, approximately 2–3 g of sieved and well-mixed dust was burnt in a muffle furnace to thermal oxidation (375 °C, 18 h) under steady air flow. Later the dust sample was mixed with 1 N HCl to remove inorganic carbon. The residual organic carbon in dust was analyzed as BC by Elemental Carbon-Hydrogen-Nitrogen Analyzer following chemo-thermal oxidation (CTO-375) technique.

2.3. Extraction and cleanup The extraction and cleanup method for the analysis of FR chemicals in the dust has been explained earlier (Yadav et al., 2017a, 2018a). Concisely, around 10 g of homogenized and sieved dust was treated with a known amount of recovery standards (20 ng of PCB 209 for PBDEs, DPs and NBFRs) and 1000 ng of deuterated tris(2-chloroethyl) phosphate (TCEP-d12). The dust samples were then soxhlet extracted with dichloromethane (DCM) for 24 h. Small chips of copper were added to the extracting flask to remove the elemental sulphur present in dust. Prior to adding in the extraction flask, small copper chips were cleaned and activated with HCl. On rotary evaporation (Heildolph 4000, Germany), the extracted samples were concentrated and reduced to a final approximate volume of 2–3 mL. Later, they were solvent exchanged with 0.5 mL of hexane. The sample extract was then subjected to solid phase extraction (Supelclean Envi Florisil SPE 6 mL (1 g) (SUPELCO, USA). Before SPE, the florisil cartridges were cleaned and conditioned with 6 mL ethyl acetate, 6 mL hexane/DCM (8:2, v/v), and 10 mL of hexane. Finally, the extracted samples were passed through the SPE column and first part was eluted with 6 mL of 8:2 hexane: DCM, and was taken for PBDEs, DPs and NBFRs analysis. The extract was then purified by a multilayer silica gel/alumina column followed by anhydrous sodium sulfate. The extract was then injected to GC-MS with the addition of 10 ng of 13C-PCB-141 as an internal standard. The second fraction that contained OPFRs was eluted with 20 mL of ethyl acetate, evaporated until dryness under nitrogen flow, and the residue was re-dissolved in 200 µL of iso-octane. The resulting fraction was transferred to GC vials for GC-MS analysis. Prior to GC-MS injection, a known amount (1000 ng) of hexamethylbenzene (HMB) was added as an internal standard for quantification.

2. Materials and methods 2.1. Study area and sampling Kopawa village is a fifth-order administrative division and is 305

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with Kaiser Normalization was applied to distinguish the sources of FRs.

2.4. GC-MS analysis The GC-MS operating condition including the injection temperature for the analysis of FRs chemicals is discussed separately in our previous studies (Yadav et al., 2017a, 2017b, 2018a). Concisely, PBDEs, NBFRs and DPs concentration were determined by an Agilent 7890 GC attached with an Agilent 5975 C MSD (Mississauga, ON) operated in electrochemical negative ionization (ECNI) mode, with HP-5MS section (15 m × 0.25 mm i.d.×0.1 µm film thickness) column. One µL of sample was injected in GC in splitless mode with 295 °C of injector temperature. Helium (He) was utilized as a carrier gas with a flow rate of 1.2 mL/min. GC oven temperature was set to an initial temperature of 110 °C for 5 min and raised to 200 °C at a rate of 20 °C/min (held for 4.5 min); increased at a rate of 10 °C/min to the final temperature of 310 °C (hold for 12 min). The transfer line temperature was 280 °C while 230 °C of ion source temperatures was maintained. Six different compounds of NBFRs (BEH-TEBP, PBEB, BTBPE, EH-TBB, HBB, and DBDPE), 9 congeners of PBDEs (BDE-28, BDE-35, BDE-47, BDE-99, BDE-100, BDE-153, BDE-154, BDE-183 and BDE-209), and 2 congeners of DPs (anti- and syn-isomer) were targeted and quantified. The full name and physicochemical characteristics of individual FRs have been detailed elsewhere (Yadav et al., 2017b). For the determination of OPFRs, the second part of the extract was subjected to an Agilent 7890 A GC attached with an Agilent 7000 A GC/ MS operated in EI mode. The transfer line and ion source temperature were maintained at 280 °C and 230 °C, respectively. The GC oven temperature started at 60 °C for 1 min, increased to 220 °C at a rate of 30 °C min−1 (held for 0 min), then to 300 °C at a rate of 5 °C min−1 (held for 15 min). Eight OPFRs such as TCEP, TDCIPP, TCIPPs (mix of three isomers), TNBP, TPHP, EHDPHP, TEHP and TMPPs (mix of three isomers) were targeted and quantified. The full name and physicochemical characteristics of single OPFRs have been detailed elsewhere (Yadav et al., 2017a).

2.7. Health risk assessment Dust ingestion and dermal absorption via dust are the two important routes of human exposure of FRs in the indoor environment released from consumer materials. The daily intake of individual FRs (PBDE, DP, NBFR and OPFR) to local population was estimated through dust ingestion and dermal contact via dust routes as per USEPA risk assessment model (USEPA, 2011, 2014; Abdallah et al., 2016; Cequier et al., 2014; Pawar et al., 2017; Yadav et al., 2017a). Mathematically, the daily human exposure of FRs through dust ingestion and dermal absorption via dust can be estimated by following Eqs. (1) and (2), respectively.

Dust ingestion exposure =

CD × DI BW

Dermal absorption via dust =

(1)

CD × DAS × ESA × AF BW

(2)

Where, CD is the concentration of PBDEs/DPs/NBFRs/OPFRs in dust (ng/g). DI represents the daily intake of dust (60 mg/day for children and 30 mg/day for the adult), DAS refers to the dust adhered to skin rate (0.01 mg/cm2), and ESA denotes exposed skin area (4970 cm2/day and 8620 cm2/day for children and adult, respectively). AF talks about absorption factor (0.17%), while BW denotes the body weight of children (15 kg) or adult (70 kg). 3. Results and discussion 3.1. Overall comments on FRs Results of the legacy FRs and EFRs measured in dust in this study are summarized in Table 1. All the four classes of FRs were detected in

2.5. QA/QC

Table 1 Statistical summary of FRs in the dust (ng/g).

As described previously, strict QA/QC criteria were followed for the extraction and analysis of FRs (Yadav et al., 2017a, 2018a). Ten laboratory blank and three field blank samples were extracted and analyzed together with dust samples to assess the possible cross-contamination and GC-MS detection limit. All the compounds of interest were either not detected or less detected in blank samples. The method detection limit (MDL) reads the average plus 3 times standard deviation of the all blank samples. In absence of FRs in the blank, the MDL was calculated as the 3 times signal to noise ratio of the lowest spiked standard. The MDL ranged from 0.19 to 0.27 pg/g, 0.18–0.29 pg/g, 6.1–13.3 pg/g, and 0.38–27.93 ng/g for PBDEs, DPs, NBFRs, and OPFRs, respectively. The average recovery rate for the PCB 209 and TCEP-d12 were 92–121% and 80–124%, respectively. The relative standard deviations were less than 15% in both cases. For statistical analysis, the blank concentration was subtracted from PBDEs, DPs, NBFRs, and OPFRs measured values but not modified for recovery standard. 2.6. Statistical analysis General statistics (minimum, maximum, mean, median, and standard deviation) of GC-MS quantified data were made in MS Excel 2010. The graphical presentation (box and whisker plots, profile and relative abundance) and spatial distribution map were performed with an IBM SPSS statistics (version 21) and ESRI-Arc GIS geospatial (version 10.3) software, respectively. GC-MS measurement data with below detection limit (BDL) concentration were set as zero for computation and analysis purpose. Spearman's rank correlation coefficient was used to investigate the relationship of TOC and BC with FR compounds, and interrelationship among different FRs. Principal component analysis (PCA)

Compounds

Min

Max

Median

BDE-28 BDE-35 BDE-47 BDE-100 BDE-99 BDE-154 BDE-153 BDE-183 BDE-209 ∑PBDEs syn-DP anti-DP ∑DPs PBEB HBB EH-TBB BTBPE BEH-TEBP DBDPE ∑NBFR TNBP TCEP TCIPPs TDCIPP TPHP EHDPHP TEHP TMPPs ∑OPFRs ∑FRs

0.02 0.02 0.01 0.01 0.01 0.01 0.01 0.03 0.02 0.14 nd nd nd nd 0.05 nd 0.01 0.24 1.30 1.6 14.7 0.6 16.4 0.1 0.8 19.1 24.2 127.5 203.4 205

1.96 0.04 1.30 0.07 0.14 0.04 0.61 0.10 0.89 0.05 0.25 0.04 1.03 0.05 0.70 0.12 56.1 1.13 63.0 1.64 9.29 0.06 17.6 0.42 26.9 0.48 0.68 0.02 0.79 0.12 3.02 0.12 3.15 0.18 538 1.30 4580 49.7 5130 51.4 58.2 17.9 5290 17.5 137 26.4 19.5 0.60 217 16.8 32.3 25.6 869 81.3 233,000 12,900 240,000 13,000 245,000 13,100

nd: not detected; CV: coefficient of variation. 306

Mean

SD

0.15 0.42 0.13 0.27 0.05 0.04 0.15 0.15 0.11 0.19 0.06 0.07 0.11 0.22 0.18 0.19 6.70 13.2 7.65 14.7 0.81 2.01 1.47 3.82 2.28 5.83 0.06 0.15 0.24 0.21 0.28 0.64 0.42 0.79 28.6 117 488 1090 518 1210 21.6 10.0 328 1170 37.5 27.9 4.9 8.1 48.0 62.0 25.9 3.4 166 222 35,900 61,500 36,500 63,000 37,000 64,200

CV (%) 272 206 80 98 179 106 194 104 197 – 247 260 – 235 89 229 188 408 223 – 46 356 74 165 129 13 134 171 – –

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concentration in dust in this study was 22 and 500 times lower than those reported in house dust of Western Australia (22.5 ng/g) (Stasinska et al., 2013) and Istanbul in Turkey (574 ng/g)(Kurt-Karakus et al., 2017). The concentration of BDE-209 in dust in this study (1.13 ng/g) was consistent with the BDE-209 level reported in dust from Pretoria in South Africa (3.47 ng/g) (Kefeni et al., 2014). Much higher concentrations of BDE-209 than this study have been reported in Guangzhou in China (2640 ng/g) (Chen et al., 2011), California in USA (2500 ng/g) (Whitehead et al., 2013), and Birmingham in UK (8100 ng/ g) (Harrad et al., 2008). This could be because of the high degree of urbanization and industrialization in those big cities, as dust from the urban area had been reported to contain significantly high BDE-209 than the rural area (Stasinska et al., 2013; Zhu et al., 2015). Because, rural areas are known to contains relatively less household product/ consumer materials than urban area thereby releasing lesser PBDEs (Zhu et al., 2015). Although the production of FRs is not known in Nepal, the presence of PBDE in house dust might be due to an import of PBDEs-containing goods and materials, leading to their subsequent release (Yadav et al., 2018a). In this study, the influence of penta- and octa-BDEs was low in all dust samples, while deca-BDE (BDE-209 in this study) was responsible for more than 69% of the PBDE content (Fig. S2, Supplementary information). Relatively low level of penta- and octa-BDEs compared to octa-BDE in dust in this study may be because of global phase-out of penta- and octa-BDEs (Dodson et al., 2012; Bjorklund et al., 2012). Meanwhile, DBDPE is a major substituent of BDE-209, the ratio of BDE209/DBDPE is utilized to mark the phase-out of deca-BDE (EBFRIP, 2008). In this study, the BDE/DBDPE ratio was below 1 for house dust, indicating low usage of the deca-BDE formulation. The profile of the individual PBDE compounds in dust in this study is illustrated in Fig. 2. ∑9PBDEs in dust was mostly dominated by BDE-209 and accounted for 69% of the ∑9PBDEs. The abundance of BDE-209 (82–93%) has been also reported previously in house dust from Australia (Stasinska et al.,

dust samples. DP was measured in 16 out of 21 samples with a detection frequency of (DF) 76%. PBEB and EH-TBB were the two NBFR which relatively less detected in the dust with a DF of 95%. The concentration of ∑4FRs in dust ranged from 205 to 245,000 ng/g (median 13,100 ng/ g). Among different classes of FRs, OPFRs was identified as the most abundant FR, followed by NBFR, PBDE, and DP, and accounted for 99.6%, 0.40%, 0.01% and 0.004% of ∑4FRs, respectively. Overall, the concentrations of ∑8OPFRs in dust were about 2, 3 and 4 orders of magnitude higher than the ∑6NBFRs, ∑9PBDEs, and ∑2DPs, respectively. The concentration of ∑8OPFRs, ∑6NBFRS, ∑9PBDEs, and ∑2DPs in dust ranged from 203 to 24,0000 ng/g (median 13,000 ng/g), 1.6–5130 ng/ g (median 51.4 ng/g), 0.14–63.0 ng/g (median 1.64 ng/g), and nd26.9 ng/g (median 0.48 ng/g), respectively. The highest concentration of OPFRs in house dust indicates the replacement of legacy FRs by EFRs, especially OPFR (Dodson et al., 2012; Bjorklund et al., 2012). Several folds higher concentrations of ∑OPFRs than ∑PBDEs and ∑NBFRs have been also reported in indoor dust from Norway (Cequier et al., 2014), USA (Schreder and La Guardia, 2014) and Spain (Cristale et al., 2016). 3.2. Polybrominated diphenyl ethers (PBDEs) All the 9 compounds of PBDEs were detected in 21 dust samples with 100% DF. The concentration of individual PBDEs measured in dust in this study has been presented in Fig. 1. It is evident from the Fig. 1 that BDE-209 was the most abundant compound followed by BDE-183 measured in the dust, and ranged from 0.02 to 56.1 ng/g (median 1.13 ng/g) and 0.03–0.70 ng/g (median 0.12 ng/g), respectively. Significantly high level of BDE-209 in dust in this study indicates their occurrence in the dust with the commercial deca-BDE formulation which might have originated from local used consumer materials (BSEF, 2013). The level of individual PBDEs measured in dust in this study was compared with similar studies around the world and are presented in Table S2, Supplementary information. The BDE-209

Fig. 1. Box and whiskers plots showing the concentration of PBDEs, DPs, NBFRs, and OPFRs in the dust (ng/g) from Kopawa. 307

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Fig. 2. Profile of the PBDEs, DPs, NBFRs, and OPFRs in house dust of Kopawa village.

g), KPD-16 (22.9 ng/g) and KPD-20 (22.8 ng/g). Higher concentration of ∑9PBDEs was also prominent at KPD-4 (16.5 ng/g). KPD-11 and KPD16 sites are associated with relatively high socioeconomic group attributing to use of wide variety of consumer materials (furniture, electrical and electronic items), KPD-20 were from lower socioeconomic group indicating usage of old items (furniture or construction materials). The concentration of BDE-209 was also very high at KPD-11 indicating local usage of consumer items (BSEF, 2013). The coefficients of variation (CV) of the PBDE compounds are used to compare the variability of PBDEs with respect to space. In this study, the CV of the PBDEs ranged from 80% to 272%. Among different isomers of PBDEs, BDE-28, BDE-35, BDE-209, and BDE-153 showed much higher variability in the dust. This indicates the area-specific source is responsible for the elevated level of PBDEs. BDE-47 was the least variable compound in the dust with the CV of 80%.

2013), Germany (Fromme et al., 2014), Egypt (Hassan and Shoeib, 2015), and China (Zhu et al., 2015). BDE-183 was the second most abundant chemical measured in the dust, followed by BDE-100, and accounted for 7% and 6% of ∑9PBDEs, respectively. BDE-183 was also second highest brominated FRs reported in indoor dust from China (Zhu et al., 2015). The median concentrations of PBDEs in dust decrease in the order of BDE-209 > BDE-183 > BDE-100 > BDE-35 > BDE99 > BDE-153 > BDE-154 > BDE-47 > BDE-28. Despite, BDE-35 is present only in little amount in deca-BDE formulation, slightly higher concentration of BDE-35 in dust in this study than BDE-47 and BDE-99 could be possibly due to debromination of highly brominated technical PBDEs mixture (Watanabe and Tatsukawa, 1987; Soderstrom et al., 2004). A similar trend of PBDE in indoor air has been reported previously (Yadav et al., 2017b). The spatial distribution of the ∑9PBDEs, ∑2DPs, ∑6NBFRs, and ∑8OPFRs in the dust of the Kopawa has been shown in Fig. 3 and S3. Highest concentrations of ∑9PBDEs were measured at KPD-11 (58.3 ng/ 308

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Fig. 3. Spatial map showing the distribution of PBDEs, DPs, NBFRs, and OPFRs in house dust from Kopawa village, Nepal.

dust, followed by BEH-TEBP, and ranged from 1.3 to 4580 ng/g (median 49.7 ng/g) and 0.24–538 ng/g (median 1.30 ng/g), respectively. The median concentration of DBDPE in dust in this study was 3–6 times lower than those reported in Oslo in Norway (147 ng/g) (Cequier et al., 2014) and Barcelona in Spain (307 ng/g) (Cristale et al., 2016) (Table S3). The profile of the NBFRs as presented in Fig. 2, indicated the abundance of DBDPE and accounted for 97% of the ∑6NBFRs in the dust. The concentration of DBDPE was also significantly higher than the BDE-209 indicating electrical and electronic items as the important source. This is because DBDPE is an alternative substituent to BDE-209, which is known to have been used in electrical and electronic materials (Betts, 2009; Schlummer et al., 2007). BEH-TEBP was the second most abundant NBFR measured in the dust and accounted for 3% of the ∑6NBFRs. The PBEB, HBB, and EH-TBB were the least detected chemicals in the dust. Similar proportions of DBDPE and BEH-TEBP have been reported in the house dust of Barcelona in Spain (Cristale et al., 2016), California in the USA (Dodson et al., 2012), 23 provinces in China (Qi et al., 2014) and Birmingham in United Kingdom (Tao et al., 2016). The median concentration of NBFRs in dust decreased in the order of DBDPE > BEH-TEBP > BTBPE > EH-TBB > HBB > PBEB. The spatial distribution of 6 NBFRs measured in dust in this study is illustrated in Fig. 3 and S3. It is clear from Fig. 3 that the highest concentration of ∑6NBFRs was detected at KPD-11 (5120 ng/g), KPD-4 (2240 ng/g) and KPD-5 (1350 ng/g). Among NBFRs, BEH-TEBP, EHTBB, PBEB, and DBDPE were the most varied chemical in the dust with the CV of 408%, 229%, 235%, and 223%, respectively. HBB was least varied chemical in dust suggesting the identical distribution of NBFRs in the dust.

3.3. Dechlorane plus (DPs) DP is a chlorinated additive FR which is commonly used in wire coatings, electrical and electronic materials (Weil and Levchik, 2004). It is an important substituent of deca-BDE. DP is ubiquitous in the environment and has the high potential for long-range atmospheric transport (Sverko et al., 2011; Xian et al., 2011). Individually, syn-DP measured only 52% of the dust, while anti-DP was identified in 71% of the dust samples. A similar proportion of syn- and anti-DP has been reported in house dust from Egypt (Hassan and Shoeib, 2015). The ratio of syn- to anti-isomers in a technical DP varies from 0.33 to 0.67 depending on the manufacturer (Zhu et al., 2007). In this study, the ratio of syn- and anti-DP ranged from 0.15 to 2.54. This larger variation in the isomeric ratio of DP attributed to the differential degradation rate of syn- and anti-DP due to indoor light, UV light and similar effect (Hassan and Shoeib, 2015). A similar variation in the ratio of syn- and anti-DP have been previously reported in house dust from Canada (Shoeib et al., 2012) and other environmental matrices (Sverko et al., 2008, 2011). The ∑2DPs concentrations measured in dust in this study (0.48 ng/g) were in good agreement with that reported in Cairo in Egypt (0.31 ng/ g) (Hassan and Shoeib, 2015), but 14 and 36 times lower than those reported in Vancouver in Canada (6.8 ng/g) (Shoeib et al., 2012) and California in USA (17.5 ng/g) (Dodson et al., 2012) (Table S3, Supplementary information). The higher concentration of ∑2DPs in these countries may be due to a relatively high degree of urbanization. Urban houses are more exposed to high level of DPs due to atmospheric transport and deposition from outdoor source (Li et al., 2015a, 2015b). A similar concentration of DPs has been reported in earlier (Wang et al., 2011; Zheng et al., 2010). The profiling of DPs showed that anti-DP was most prevalent in the dust, and accounted for 88% of the ∑2DPs (Fig. 2). ∑2DPs was abundantly detected at KPD-11 which was associated with the house having the high socioeconomic group (Fig. 3 and S3).

3.5. Organophosphate ester flame retardants (OPFRs) The concentrations of ∑8OPFRs together with individual OPFR measured in dust in this study are summarized in Table 1 and Fig. 1. The ∑8OPFRs in dust ranged from 203 to 240,000 ng/g (median 13,000 ng/g). The ∑8OPFRs in dust in this study was comparable to those reported in house dust from Flemish in Belgium (13,100 ng/g) (Van den Eede et al., 2011) and Oslo in Norway (Cequier et al., 2014)

3.4. Novel brominated flame retardants (NBFRs) The inter-quartile range and median concentration of the individual NBFR measured in dust in this study are shown in Fig. 1. Among different NBFRs, DBDPE was identified as the most abundant NBFR in the 309

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(Table S4). However, this ∑8OPFRs level in dust was 17–18 times greater than that reported in house dust from urban Nepal (Yadav et al., 2017a)(Table S4). ∑OPFRs in dust in this study was also 2–70 times greater than those reported in Egypt (189 ng/g) (Abdallah and Covaci, 2014), New Zealand (1738 ng/g) (Ali et al., 2012), and Southern China (7480 ng/g) (He et al., 2015)., The ∑8OPFRs level observed in dust in this study was 7 times lower than that reported Japan (97,000 ng/g) (Mizouchi et al., 2015). Generally, the arylated-OPFR was the most abundant compound measured in the dust at all study sites (Fig. S4, Supplementary information). Alkylated-OPFRs (sum of TPHP, EHDPHP, and TMPPs) was the second most abundant after arylated-OPFR. Chlorinated-OPFR (sum of TCEP, TCIPPs, and TDCIPP) was the least abundant chemicals measured in the dust. Both alkylated- and arylated- OPFRs are applied as plasticizers in different consumer materials, for instance, PVC, synthetic rubber, synthetic resin and other similar products (WHO, 2000; van der Veen and de Boer, 2012). TMPP and TEHP were the most abundant chemicals measured in the dust, and ranged from 128 to 233,000 ng/g (median 12,900 ng/g) and 24.2–869 ng/g (median 81.3 ng/g), respectively. The profile of the individual OPFRs presented in Fig. 2. stated that TMPPs was the most prominent OPFR, followed by TEHP, TCIPPs, and EHDPHP, and accounted for 98.5%, 0.62%, 0.20% and 0.19% of ∑8OPFRs, respectively. The median concentration of OPFRs in dust decreased in the order of TMPPs > TEHP > TCIPPs > EHDPHP > TNBP > TCEP > TPHP > TDCIPP. In this study, the highest level of TMPP and TEHP in dust could be because of the fact that these FRs (TMPP and TEHP) are the general contaminants present in all environments (WHO, 2000). They can release to the environment mainly from end point use (Lassen and Lokke, 1999). It can also evaporate from upholstery fabric (MRI, 1979). TMPP and TEHP are the main components utilized in PVC and hydraulic fluid as FR (OEHHA, 2011). TMPP is used as plasticizer in floor and wall coverings (WHO, 1990). Small amount of TMPP is also utilized in synthetic leather, shoes, lacquer and varnishes as additive (WHO, 1990). Hence, the elevated level of TMPP in this study could be from consumer materials utilized in rural houses as no other source exists. Besides farming, majority of the villagers go to Nepalese city/foreign country in search of better livelihood. This has helped to improve their financial condition in past few years. Multiple sources of income encourage them to buy and bring many household materials such as cellphones, TV, radio and several consumer items to decorate their houses. The median concentration of TMPP in this study (12,900 ng/g) was 40–50 times greater than those reported in Barcelona in Spain (278 ng/g) (Cristale et al., 2016) and Oslo in Norway (307 ng/g) (Cequier et al., 2014). Highest level of TMPP ranging from 12,500 to 59,800 ng/g in dust has been previously reported from Belgium (van den Eede et al., 2011) and Japan (Araki et al., 2014). The spatial distribution map of OPFRs measured in dust in this study has been shown in Fig. 3 and S3. The highest concentration of ∑8OPFRs was detected at KPD-1 (23,4000 ng/g), KPD-4 (158,000 ng/g) and KPD9 (123,000 ng/g). Higher concentration of OPFRs was also measured at KPD-2 (56,000 ng/g) and KPD-6 (52,600 ng/g). The lowest level of ∑8OPFRs was detected at KPD-14 (217 ng/g). TCEP, TMPPs, and TDCIPP were the most varied chemicals in the dust with a high CV of 356%, 171%, and 165%, respectively. The concentration of EHDPHP in dust was least variable (CV=13%) and was uniformly distributed.

was positively and significantly correlated with octa-BDEs (Rho = 0.615, p < 0.01), BTBPE (Rho = 0.733, p < 0.01), TPHP (Rho = 0.621, p < 0.01), TEHP (Rho = 0.560, p < 0.01) and TMPPs (Rho = 0.550, p < 0.01). However, BC content in dust was either poorly related or not related with FR compounds, suggesting little or no role in the distribution of FRs. The poor relationship of BC with PBDE and other FRs could be because of low content of BC in dust/soil. The low BC/TOC might inhibit sorption to BC because of natural attenuation. Further, BC contain only 4% of TOC in dust/soil (Cornelissen et al., 2005). Another possible reason of weak relationship of BC with FR could be because of the fact that the CTO-375 method incapable of detecting all types of BC (Hammes et al., 2007). This correlation results indicated significance of amorphous organic matter rather than carbonaceous materials in retention of PBDE in dust. A similar relationship of TOC with other POPs in soil has been reported previously (Yadav et al., 2017c; Nam et al., 2008). Octa-BDE was positively and significantly correlated with decaBDEs (Rho = 0.519, p < 0.01), BTBPE (Rho = 0.531, p < 0.01) and DBDPE (Rho = 0.524, p < 0.01). A similar relationship of octa-BDE with BTBPE has been reported earlier (Ali et al., 2013). Deca-BDE was also positively and significantly related with syn-DP (Rho = 0.932, p < 0.01), anti-DP (Rho = 0.915, p < 0.01), EH-TBB (Rho = 0.880, p < 0.01), BEH-TEBP (Rho = 0.868, p < 0.01) and DBDPE (Rho = 0.841, p < 0.01). Likewise, syn-DP and anti-DP were significantly correlated with EH-TBB, BEH-TEBP, and DBDPE. EH-TBB and BEHTEBP are two major components of Firemaster (FM®550) which is an alternate replacement to PBDEs (Stapleton et al., 2008). These indicated DPs, EH-TBB, BEH-TEBP and DBDPE can have similar source of origin. This relationship is consistent with previous studies (Zhang et al., 2011; Yadav et al., 2017a). The two isomers of DP (syn- and anti) were positively correlated (Rho = 0.981, p < 0.01) to each other which is likely in the fact that both have same primary source of origin and have similar properties and environmental fate. 3.7. Source apportionment study A PCA was applied to the whole data set of BFRs (PBDEs, DPs and NBFRs) and OPFR separately to investigate the possible source. A varimax rotation with Kaiser Normalization was used to simplify the factor coefficient. The results of PCA together with factor loading obtained after varimax orthogonal rotation of BFRs data in dust are summarized in Table 2. Altogether, five major factors were extracted with Table 2 Principal component analysis (PCA) of BFR, DP and NBFR in dust. Compounds

BDE-28 BDE-35 BDE-47 BDE-100 BDE-99 BDE-154 BDE-153 BDE-183 BDE-209 syn-DP anti-DP PBEB HBB EH-TBB BTBPE BEH-TEBP DBDPE Eigen values % of variance Cumulative %

3.6. Effect of TOC and BC A Spearman's rank correlation analysis was performed on whole data set of FRs to investigate the inter- and intra-relationship among different FRs together with TOC and BC in the dust. The TOC and BC content in dust sample ranged from 0.09% to 3.65% (median 1.03%) and 0.014–0.153% (median 0.08%), respectively. The Spearman's rank correlation coefficient matrices of individual FR have been summarized in Table S5, Supplementary information. The result indicated that TOC 310

Factor component Factor 1

Factor 2

Factor 3

Factor 4

Factor 5

− 0.057 − 0.081 − 0.037 0.222 0.969 0.090 0.068 0.713 0.917 0.983 0.988 − 0.066 − 0.181 0.991 0.054 0.986 0.918 7.447 43.808 43.808

0.979 0.961 0.272 0.770 0.061 0.229 0.058 0.010 − 0.041 0.056 − 0.015 0.219 − 0.141 − 0.013 − 0.020 − 0.001 − 0.046 3.725 21.913 65.721

− 0.021 − 0.024 0.211 0.352 − 0.010 0.686 0.232 0.662 0.146 0.022 0.117 − 0.094 − 0.253 0.040 0.952 − 0.066 0.283 1.879 11.054 76.775

0.103 0.214 0.810 0.326 0.121 0.432 0.225 0.036 − 0.014 − 0.021 − 0.021 0.844 − 0.334 − 0.021 − 0.083 − 0.056 − 0.032 1.148 6.752 83.527

− 0.037 − 0.108 0.270 0.187 − 0.061 0.179 0.719 0.174 0.188 − 0.005 − 0.033 − 0.210 0.686 − 0.062 − 0.078 − 0.062 − 0.067 1.044 6.143 89.670

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contaminants and consumer products (Ballesteros-Gomez, Brandsma, de Boer, 2014). Hence, Factor 1 is referred to as contribution from PVC made consumer products. Factor 2 accounted for 35.3% of total variance in data with highest loading on low mass OPFR (TNBP, TCEP, and TCIPP). High loading on TNBP, TCEP, and TCIPP suggests their co-occurrence in the dust. TNBP is significant constituents of plastics, vinyl resins, cellulose esters and natural gums (WHO, 1991). TCIPP is the major replacement of TCEP (Schindler and Forster, 2009), and can release from their usages in furniture upholstery and PVC wallpaper (Lassen and Lokke, 1999; European Union, 2008). Therefore, Factor 2 is referred to originate from consumer materials. Factor 3 explained 17.7% of the OPFR data set with high loading on TDCIPP (0.863) and negative loading on EHDPHP (-0.760). TDCIPP is primarily used as plasticizers, lacquer and anti-foaming agents (Marklund et al., 2005). EHDPHP is the major constituents of food packaging and paints (USFDA, 2006; Brommer, 2014). Hence, Factor 3 is identified as originated from food packaging, paints and antifoaming agents.

Eigenvalue greater than 1 which accounted for 89.7% of the total variation in BFR dataset. Factor 1 accounted for 43.8% of total variance in BFR data, and was highly loaded with BDE-99 (0.969), BDE-183 (0.713), BDE-209 (0.917), syn-DP (0.983), anti-DP (0.988), EH-TBB (0.991), BEH-TEBP (0.986), and DBDPE (0.918). BDE-99 is the significant constituent of commercial penta-BDE (La Guardia et al., 2006). BDE-183 is the component of the octa-BDE which releases from the handling of e-waste materials (Cahill et al., 2007). High loading of DBDPE together with BDE-209 indicates their similar sources in the dust. BDE-209 is an important constituent of the deca-BDE formulation (Zhang et al., 2014). BDE-209 also results from usages of consumer materials such as electrical and electronic items, plastics, textiles, polystyrene and polyamides (Wu et al., 2015). Hence, Factor 1 was identified as contribution from commercial PBDE formulations. Factor 2 explained 21.9% of variation in data with highest loading on low mass PBDE i.e. BDE-28 (0.979), BDE-35 (0.961), and BDE-100 (0.770). BDE-28 and BDE-100 are the constituents of penta-BDEs, while BDE-35 is rarely detected in the environment. Part of BDE-28, BDE-35 and BDE-100 also results from degradation/debromination of BDE-209 (La Guardia et al., 2006; Tokarz et al., 2008). This is true in case of Nepal because Nepal being a tropical and warm region, consequently receives more sunshine. These light irradiation and pressure differences can significantly stimulate the degradation process of BDE-209 (Harrad and Abdallah, 2011). Hence, Factor 2 is referred to as originate d from degradation/debromination of BDE-209. Factor 3 explored 11% of the variation in BFR data and was highly loaded with BTBPE (0.952) and moderately with BDE-154 (0.686) and BDE-183 (0.662). BTBPE is greatly utilized in polystyrene, thermoset resin, thermoplastics, coatings, and polycarbonate (WHO, 1997). BDE154 and BDE-183 are the components of technical octa-BDE. Octa-BDE is mainly used thermosetting plastic and polycarbonate. Therefore, factor 3 can be considered as emission from polystyrene, thermoplastics and polycarbonate materials Factor 4 and 5 accounted for 6.75% and 6.1% variation in BFR dataset with significant loadings on PBEB (0.844), BDE-47 (0.810), HBB (0.686) and BDE153 (0.719), respectively. PBEB and HBB are widely applied in varieties of polymers (Covaci et al., 2011). PBEB is also utilized in polyurethane foam, adhesive, and the circuit board (Hoh et al., 2005). Therefore, Factor 4 and 5 is commonly referred to as contribution from polymers and adhesive substance. For OPFR, a total of three major factors were extracted with Eigen value greater than 1 after varimax rotation with Kaiser Normalization. The factor loading of OPFR dataset following varimax rotation is presented in Table 3. Factor 1 contributed 40.4% of the variation in OPFR data and was highly loaded with TPHP (0.975), TEHP (0.998) and TMPP (0.997). TEHP and TMPP are the general contaminants present in all environment (Cristale et al., 2016). Because of its plasticizing property, TPHP is mainly used in unsaturated polyester resins and in PVC (Wei et al., 2015). Additionally, TPHP also results from e-waste

3.8. Health risk assessment To perform a preliminary evaluation of human exposure to FRs, we assumed 100% absorption of the intake. The daily exposures of selected FRs (PBDEs, BDE-209, DPs, NBFRs, DBDPE, and OPFRs) to adult and children population were estimated through dust ingestion and dermal absorption via dust and are summarized in Fig. 4 and Table S6. It is imperative to note that the main route of FRs to both adult and children population was dermal absorption via dust, followed by dust ingestion. The health exposure estimates showed that children (median 0.092 ng/ kg BW/day, 0.03 ng/kg BW /day, 2.9 ng/kg BW /day and 735 ng/kg BW /day for PBDEs, DPs, NBFRs, and OPFRs, respectively) are more vulnerable than adult population via dermal absorption of dust. This finding is consistent with the human exposure results of FRs reported in urban dust from Nepal (Yadav et al., 2017a). Cequier et al. (2014) also reported dermal absorption via dust as the primary exposure of OPFRs in Norwegian women. In this study, when the estimated exposure of PBDEs, NBFRs and OPFRs was compared with corresponding reference dose (RfD) suggested by Ali et al. (2012), they were 8–9, 9–10, and 2–7 orders of magnitude lower than the RfD, respectively (Table S6), indicating insignificant risk to local population. A similar exposure risk of FRs has been reported previously in indoor dust from Norway, Pakistan and Kuwait (Ali et al., 2012, 2013; Cequier et al., 2014). 4. Conclusions In this study, four different class of FRs was investigated in indoor dust from a remote village of Nepal to mark the contamination level, profile, spatial distribution and their sources and possible health risk. Results showed that the concentration of ∑PBDEs, ∑DPs and ∑NBFRs determined in dust in this study was several folds lower than those reported worldwide in urban house dust while ∑OPFRs were comparable to/or greater than other countries. This difference indicated indoor dust in Kopawa is less polluted by FRs than the urban area. OPFR was the most prominent FRs identified in the dust, followed by NBFRs, PBDEs, and DP. The concentration of ∑9PBDEs in dust was mostly dominated by BDE-209, while DBDPE exceeded in ∑6NBFRs. TMPP and TEHP were more prominent phosphate FRs analyzed in the dust. DP was the least detected chemical in the dust. TOC content in dust was positively and significantly correlated with the compounds of PBDEs, NBFR and OPFRs, while BC was either poorly related or not related with FRs, suggesting little or no role in the distribution of FR in house dust. PCA indicated the contribution from commercial penta-, octa- and deca-BDEs formulation, the adhesive substance, food packaging and paints, and degradation of BDE-209 as the possible sources of FRs. The dermal contact via dust was identified as the principal route of FRs

Table 3 Principal component analysis of OPFR in dust. Compounds

TNBP TCEP TCIPPs TDCIPP TPHP EHDPHP TEHP TMPPs Eigen values % of variance Cumulative %

Factor component Factor 1

Factor 2

Factor 3

0.268 − 0.134 0.037 0.027 0.975 − 0.114 0.998 0.997 3.231 40.386 40.386

0.889 0.960 0.895 0.416 0.159 0.522 − 0.031 − 0.027 2.826 35.323 75.709

− 0.184 0.016 0.274 0.863 0.083 − 0.760 0.007 0.042 1.415 17.688 93.397

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Fig. 4. Box and whisker's plot showing the exposure of selected FRs to adult and children through ingestion and dermal absorption via dust.

exposure to both adult and children population. Relatively children were more prone to FRs exposure than the adult. The estimated daily exposure of PBDE, NBFR, and OPFR were 7–10 orders of magnitude lower than their corresponding RfD, suggesting insignificant risk. However, other routes of exposure might still be prominent in the case of Kopawa which could be examined in the future.

Araki, A., Saito, I., Kanazawa, A., Morimoto, K., Nakayama, K., Shibata, E., Tanaka, M., Takigawa, T., Yoshimura, T., Chikara, H., Saijo, Y., Kishi, R., 2014. Phosphorus flame retardants in indoor dust and their relation to asthma and allergies of inhabitants. Indoor Air 24, 3–15. Ballesteros-Gomez, Brandsma, S.H., de Boer, J., 2014. Direct probe atmospheric pressure photoionization/atmospheric pressure chemical ionization high-resolution mass spectrometry for fast screening of flame retardants and plasticizers in products and waste. Anal. Bioanal. Chem. 406 (11), 2503–2512. Betts, K., 2009. Glut of data on “new” flame retardant documents its presence all over the world. Environ. Sci. Technol. 43, 236–237. Bjorklund, J.A., Sellstrom, U., de Wit, C.A., Aune, M., Lignell, S., Darnerud, P.O., 2012. € Comparisons of polybrominated diphenyl ether and hexabromocyclododecane concentrations in dust collected with two sampling methods and matched breast milk samples. Indoor Air 22, 279–288. Brommer, S., Harrad, S., 2015. Sources and human exposure implications of concentrations of organophosphate flame retardants in dust from UK cars, classrooms, living rooms, and offices. Environ. Int. 83, 202–207. https://doi.org/10.1016/j.envint. 2015.07.002. Brommer, S., 2014. Characterizing Human Exposure to Organophosphate Ester Flame Retardants. A Ph.D. Thesis Submitted to Birmingham Division of Environmental Health and Risk Management College of Life and Environmental Sciences. School of Geography, Earth and Environmental Sciences, The University of Birmingham, Edgbaston, B15 2TT, United Kingdom. BSEF, 2013. About Decabromo Diphenyl Ether (Deca-BDE). Bromine Science and Environmental Forum. 〈http.//www.bsef.com/our-substances/deca-bde/aboutdecabde〉 (Accessed on April 2013). Cahill, T.M., Groskova, D., Charles, M.J., Sanborn, J.R., Denison, M.S., Baker, L., 2007. Atmospheric concentrations of polybrominated diphenyl ethers at near-source sites. Environ. Sci. Technol. 41, 6370–6377. Cao, Z., Xu, F., Covaci, A., Wu, M., Wang, H., Yu, G., Wang, B., Deng, S., Huang, J., Wang, X., 2014. Distribution patterns of brominated, chlorinated, and phosphorus flame retardants with the particle size in indoor and outdoor dust and implications for human exposure. Environ. Sci. Technol. 48, 8839–8846. CBS/NPCS, 2011. Nepal Population and Housing Census 2011 (National Report). Central Bureau of Statistics, National Planning Commission Secretariat Government of Nepal. Available at: 〈http://cbs.gov.np/sectoral_statistics/population/national_report〉. Cequier, E., Ionas, A.C., Covaci, A., Marcé, R.M., Becher, G., Thomsen, C., 2014. Occurrence of a broad range of legacy and emerging flame retardants in indoor environments in Norway. Environ. Sci. Technol. 48, 6827–6835. Chen, L.G., Huang, Y.M., Xu, Z.C., Wen, L.J., Peng, X.C., Ye, Z.X., et al., 2011. Human exposure to PBDEs via house dust ingestion in Guangzhou, South China. Arch. Environ. Con. Toxicol. 60, 556–564. Cornelissen, G., Gustafsson, O., Bucheli, T.D., Jonker, M.T.O., Koelmans, A.A., van Noort, P.C.M., 2005. Extensive sorption of organic compounds to black carbon, coal, and kerogen in sediments and soils: mechanisms and consequences for distribution, bioaccumulation, and biodegradation. Environ. Sci. Technol. 39, 6881–6895. Covaci, A., Harrad, S., Abdallah, M.A.E., Ali, N., Law, R.J., Herzke, D., de Wit, C.A., 2011. Novel brominated flame retardants: a review of their analysis, environmental fate, and behavior. Environ. Int. 37, 532–556. Cristale, Joyce, Hurtado, A., Gomez-Canela, C., Lacorte, S., 2016. Occurrence and sources of brominated and organophosphorus flame retardants in dust from different indoor environments in Barcelona, Spain. Environ. Res. 149, 66–76.

Acknowledgment ICY is thankful to Chinese Academy of Science for providing financial assistance in the form of CAS fellowship (2014FFZB0017) for International Young Scientist. Authors are also thankful to the residents of the Kopawa village for their assistance in the collection of house dust samples. This study was partly supported by the CAS Belt & Road Initiative No. 132744KYSB20170002 (Southern Contaminants Program). Appendix A. Supplementary material Supplementary data associated with this article can be found in the online version at doi:10.1016/j.ecoenv.2018.10.089. References Abdallah, M.A.E., Covaci, A., 2014. Organophosphate flame retardants in indoor dust from Egypt: implications for human exposure. Environ. Sci. Technol. 48, 4782–4789. Abdallah, M.A.E., Pawar, G., Harrad, S., 2016. Human dermal absorption of chlorinated organophosphate flame retardants: implications for human exposure. Toxicol. Appl. Pharmacol. 291, 28–37. Alaee, M., Arias, P., Sjödin, A., Bergman, Å., 2003. An overview of commercially used brominated flame retardants, their applications, their use patterns in different countries/regions and possible modes of release. Environ. Int. 29, 683–689. Ali, N., Dirtu, A.C., Eede, N.V.D., Goosey, E., Harrad, S., Neels, H., t Mannetje, A., Coakley, J., Douwes, J., Covaci, A., 2012. Occurrence of alternative flame retardants in indoor dust from New Zealand: indoor sources and human exposure assessment. Chemosphere 88, 1276–1282. Ali, N., Ali, L., Mehdi, T., Dirtu, A.C., Shammari, A., Neels, F.H., Covaci, A., 2013. Levels and profiles of organochlorines and flame retardants in a car and house dust from Kuwait and Pakistan: implication for human exposure via dust ingestion. Environ. Int. 55, 62–70. Alves, A., Kucharska, A., Erratico, C., Xu, F., Hond, E.D., Koppen, G., Vanermen, G., Covaci, A., Voorspoels, S., 2014. Human biomonitoring of emerging pollutants through noninvasive matrices: state of the art and future potential. Anal. Bioanal. Chem. 406, 4063–4088. https://doi.org/10.1007/s00216-014-7748-1.

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Ecotoxicology and Environmental Safety 168 (2019) 304–314

I.C. Yadav et al. Darnerud, P., 2003. Toxic effects of brominated flame retardants in man and in wildlife. Environ. Int. 29, 841–853. Dodson, R.E., Perovich, L.J., Covaci, A., Ionas, A.C., Dirtu, A.C., Brody, J.G., Rudel, R.A., 2012. After the PBDE phase-out: a broad suite of flame retardants in repeat house dust samples from California. Environ. Sci. Technol. 46, 13056–13066. European Union, 2008. Tris(2-chloro-1-methyl ethyl)Phosphate (TCPP) Risk Assessment report [online]. Available from: 〈https://echa.europa.eu/documents/10162/13630/ trd_rar_ireland_tccp_en.pdf〉. (Accessed 5 May 2012). Fromme, H., Hilger, B., Kopp, E., Miserok, M., Volkel, W., 2014. Polybrominated diphenyl ethers (PBDEs), hexabromocyclododecane (HBCD) and “novel” brominated flame retardants in house dust in Germany. Environ. Int. 64, 61–68. Gustafsson, O., Bucheli, T.D., Kukulska, Z., Andersson, M., Largeau, C., Rouzaud, J.N., Reddy, C.M., Eglinton, T.I., 2001. Evaluation of a protocol for the quantification of black carbon in sediments. Glob. Biogeochem. Cycles 15, 881–890. Hammes, K., Schmidt, M.W.I., Smernik, R.J., Currie, L.A.R., Ball, W.P., Nguyen, T.H., Louchouarn, P., Houel, S., Gustafsson, O., Elmquist, M., Cornelissen, G., Skjemstad, J.O., Masiello, C.A., Song, J., Peng, P., Mitra, S., Dunn, J.C., Hatcher, P.G., Hockaday, W.C., Smith, D.M., Hartkopf-Fro¨der, C., Bo¨hmer, A., Lu¨ er, B., Huebert, B.J., Amelung, W., Brodowski, S., Huang, L., Zhang, W., Gschwend, P.M., FloresCervantes, D.X., Largeau, C., Rouzaud, J.-N., Rumpel, C., Guggenberger, G., Kaiser, K., Rodionov, A., Gonzalez-Vila, F.J., Gonzalez-Perez, J.A., de la Rosa, J.M., Manning, D.A.C., Lo´ pez-Cape, l., E., Ding, L., 2007. Comparison of quantification methods to measure fire-derived (black/elemental) carbon in soils and sediments using reference materials from soil, water, sediment and the atmosphere. Glob. Biogeochem. Cycles 21, GB3016. Harrad, S., Abdallah, M.A.E., 2011. Brominated flame retardants in dust from UK cars within vehicle spatial variability, evidence for degradation and exposure implications. Chemosphere 82, 1240–1245. Harrad, S., Ibarra, C., Abdallah, M.A.-E., Boon, R., Neels, H., Covaci, A., 2008. Concentrations of brominated flame retardants in dust from United Kingdom cars, homes, and offices: causes of variability and implications for human exposure. Environ. Int. 34, 1170–1175. Harrad, S., Goosey, E., Desborough, J., Abdallah, M.A., Roosens, L., Covaci, A., 2010. Dust from U.K. primary school classrooms and daycare centers: the significance of dust as a pathway of exposure of young U.K. children to brominated flame retardants and polychlorinated biphenyls. Environ. Sci. Technol. 44 (11), 4198–4202. Hassan, Y., Shoeib, T., 2015. Levels of polybrominated diphenyl ethers and novel flame retardants in microenvironment dust from Egypt: an assessment of human exposure. Sci. Total Environ. 505, 47–55. He, C.T., Zheng, J., Qiao, L., Chen, S.J., Yang, J.Z., Yuan, J.G., Yang, Z.Y., Mai, B.X., 2015. Occurrence of organophosphorus flame retardants in indoor dust in multiple microenvironments of southern China and implications for human exposure. Chemosphere 133, 47–52. Hoh, E., Zhu, L., Hites, R.A., 2005. Novel flame retardants, 1,2- bis(2,4,6-tribromophenoxy)-ethane and 2,3,4,5,6-pentabromoethylbenzene, in the United States' environmental samples. Environ. Sci. Technol. 39, 2472–2477. Jones-Otazo, H.A., Clarke, J.P., Diamond, M.L., Archbold, J.A., Ferguson, G., Harner, T., 2005. Is house dust the missing exposure pathway for PBDEs? An analysis of the urban fate and human exposure to PBDEs. Environ. Sci. Technol. 39, 5121–5130. Kefeni, K.K., Okonkwo, J.O., Botha, B.M., 2014. Concentrations of polybromobiphenyls and polybromodiphenyl ethers in home dust: relevance to socio-economic status and human exposure rate. Sci. Total Environ. 470–471, 1250–1256. Kurt-Karakus, P.B., Alegria, H., Jantunen, L., Birgul, A., Topcu, A., Jones, K.C., Turgut, C., 2017. Polybrominated diphenyl ethers (PBDEs) and alternative flame retardants (NFRs) in the indoor and outdoor air and indoor dust from Istanbul-Turkey: levels and an assessment of human exposure. Atmos. Pollut. Res. 8 (5), 801–815. La Guardia, M.J., Hale, R.C., Harvey, E., 2006. Detailed polybrominated diphenyl ether (PBDE) congener composition of the widely used penta-, octa-, and deca-PBDE technical flame-retardant mixtures. Environ. Sci. Technol. 40, 6247–6254. Lassen, C., Lokke, S., 1999. Danish Environmental Protection Agency (EPA), Brominated Flame Retardants: Substance Flow Analysis and Assessment of Alternatives 1999. Law, R., Covaci, A., Harrad, S., Herzke, D., Abdallah, M.A., Fernie, K., Toms, L.M.L., Takigami, H., 2014. Levels and trends of PBDEs and HBCDs in the global environment: status at the end of 2012. Environ. Int. 65, 147–158. Lei, J.Q., 2015a. Polybrominated diphenyl ethers (PBDEs) in soil and outdoor dust from a multi-functional area of Shanghai: levels, compositional profiles, and interrelationships. Chemosphere 118, 87–95. Li, W.L., Qi, H., Ma, W.-L., Liu, L.-Y., Zhang, Z., Zhu, N.Z., Mohammed, M.O.A., Li, Y.F., 2015b. Occurrence, behavior and human health risk assessment of dechlorane plus and related compounds in indoor dust of China. Chemosphere 134, 166–171. Marklund, A., Andersson, B., Haglund, P., 2005. Organophosphorus flame retardants and plasticizers in air from various indoor environments. J. Environ. Monit. 7, 814–819. Mizouchi, S., Ichiba, M., Takigami, H., Kajiwara, N., Takamuku, T., Miyajima, T., Kodama, H., Someya, T., Ueno, D., 2015. Exposure assessment of organophosphorus and organobromine flame retardants via indoor dust from elementary schools and domestic houses. Chemosphere 123, 17–25. MRI, 1979. Assessment of the Need for Limitation on Triaryl and Trialkyl/aryl Phosphates. Draft Final Report. Midwest Research Institute, Kansas City (EPA contract 68-01-4313). Nam, J.J., Gustafsson, O., Kurt-Karakus, P., Breivik, K., Steinnes, E., Jones, K.C., 2008. Relationships between organic matter, black carbon and persistent organic pollutants in European background soils: implications for sources and environmental fate. Environ. Pollut. 156, 809–817. Newton, S., Sellström, U., De Wit, C.A., 2015. Emerging flame retardants, PBDEs, and HBCDDs in indoor and outdoor media in Stockholm, Sweden. Environ. Sci. Technol. 49, 2912–2920.

OEHHA, 2011. Chemical for CIC consultation: tris (2-ethylhexyl) phosphate [online]. Available from: 〈http://www.oehha.org/prop65/public_meetings/CIC101211/ 101211Tris2ethylhexylphosphate.pdf〉. (Accessed 11 November 2013). Pawar, G., Abdallah, M.A.E., de Saa, E.V., Harrad, S., 2017. Dermal bioaccessibility of flame retardants from indoor dust and the influence of topically applied cosmetics. J. Expo. Sci. Environ. Epidemiol. 27 (1), 100–105. Qi, H., Li, W.L., Liu, L.Y., Zhang, Z.F., Zu, N.Z., Song, W.W., Ma, W.L., Li, Y.F., 2014. Levels, distribution and human exposure of new non-BDE brominated flame retardants in the indoor dust of China. Environ. Pollut. 195, 1–8. Ramu, K., Isobe, T., Takahashi, S., Kim, E.Y., Min, B.Y., We, S.U., Tanabe, S., 2010. Spatial distribution of polybrominated diphenyl ethers and hexabromocyclododecanes in sediments from coastal waters of Korea. Chemosphere 79, 713–719. Schindler, B.K., Forster, K., 2009. Quantification of two urinary metabolites of organophosphorus flame retardants by solid-phase extraction and gas chromatographytandem mass spectrometry. Anal. Bioanal. Chem. 395 (4), 1167–1171. Schlummer, M., Gruber, L., Mäurer, A., Wolz, G., van Eldik, R., 2007. Characterisation of polymer fractions from waste electrical and electronic equipment (WEEE) and implications for waste management. Chemosphere 67, 1866–1876. Schreder, E.D., La Guardia, M.J., 2014. Flame retardant transfers from U.S. households (dust and laundry wastewater) to the aquatic environment. Environ. Sci. Technol. 48, 11575–11583. Shoeib, M., Harner, T., Webster, G.M., Sverko, E., Cheng, Y., 2012. Legacy and current use flame retardants in house dust from Vancouver, Canada. Environ. Pollut. 169, 175–182. Soderstrom, G., Sellstrom, U., de Wit, C.A., Tysklind, M., 2004. Photolytic debromination of decabromodiphenyl ether (BDE 209). Environ. Sci. Technol. 38, 127–132. Stapleton, H.M., Allen, J.G., Kelly, S.M., Konstantinov, A., Klosterhaus, S., Watkins, D., McClean, M.D., Webster, T.F., 2008. Alternate and new brominated flame retardants detected in U.S. House dust. Environ. Sci. Technol. 42, 6910–6916. Stasinska, A., Reid, A., Hinwood, A., Stevenson, G., Callan, A., Odland, J., Heyworth, J., 2013. Concentrations of polybrominated diphenyl ethers (PBDEs) in residential dust samples from Western Australia. Chemosphere 91, 187–193. Stockholm Convention, 2009a. UNEP/POPS/POPRC.4/14 listing of hexabromodiphenyl ether and heptabromodiphenyl ether. Stockholm Convention, 2009b. UNEP/POPS/POPRC.4/18 listing of tetrabromodiphenyl ether and pentabromodiphenyl ether. Sverko, E., Tomy, G.T., Reiner, E.J., Li, Y.F., Mc Carry, B.E., Arnot, J.A., Law, R.J., Hites, R.A., 2011. Dechlorane plus and related compounds in the environment: a review. Environ. Sci. Technol. 45 (12), 5088–5098. Sverko, E.D., Tomy, G.T., Marvin, C.H., Zaruk, D., Reiner, E., Helm, P.A., 2008. Dechlorane plus levels in the sediment of the Lower Great Lakes. Environ. Sci. Technol. 42, 361–366. Tajima, S., Araki, A., Kawai, T., Tsuboi, T., Ait Bamai, Y., Yoshioka, E., Kanazawa, A., Cong, S., Kishi, R., 2014. Detection and intake assessment of organophosphate flame retardants in house dust in Japanese dwellings. Sci. Total Environ. 478, 190–199. https://doi.org/10.1016/j.scitotenv.2013.12.121. Tao, F., Abdallah, M.A.E., Harrad, S., 2016. Emerging and legacy flame retardants in UK indoor air and dust: evidence for replacement of PBDEs by emerging flame retardants? Environ. Sci. Technol. 50, 13052–13061. Tokarz III, J.A., Ahn, M.Y., Leng, J., Filley, T.R., Nies, L., 2008. Reductive debromination of polybrominated diphenyl ethers in anaerobic sediment and a biomimetic system. Environ. Sci. Technol. 42, 1157–1164. USEPA, 2011. Exposure Factors Handbook. 2011 Edition. (Washington DC, 2011). USEPA, 2013. EPA Announces Chemicals for Risk Assessment in 2013, Focus on Widely Used Flame Retardants. United States Environmental Protection Agency, Washington, D.C., United States (〈http://yosemite.epa.gov/opa/admpress.nsf/ bd4379a92ceceeac8525735900400c27/c6be79994c3fd08785257b3b0054e2fa!〉 (Accessed on March 2015). USEPA, 2014. Risk assessment guidance for superfund volume I human health evaluation manual (Part A). 〈https://www.epa.gov/risk/risk-assessment-guidancesuperfundrags-part/〉 (Accessed January 2014). USFDA, 2006. Total Dietary Study. Market Baskets 1991- through 2003-, U.S. Food and Drug Administration. Available at: 〈http://www.fda.gov/downloads/Food/ FoodSafety/FoodContaminantsAdulteration/TotalDietStudy/UCM184304.pdf〉. Van den Eede, N., Dirtu, A.C., Neels, H., Covaci, A., 2011. Analytical developments and preliminary assessment of human exposure to organophosphate flame retardants from indoor dust. Environ. Int. 37, 454–461. https://doi.org/10.1016/j.envint.2010. 11.010. van der Veen, I., de Boer, J., 2012. Phosphorus flame retardants: properties, production, environmental occurrence, toxicity, and analysis. Chemosphere 88, 1119–1153. https://doi.org/10.1016/j.chemosphere.2012.03.067. Wang, J., Tian, M., Chen, S.J., Zheng, J., Luo, X.J., An, T.C., Mai, B.X., 2011. Dechlorane plus in house dust from e-waste recycling and urban areas in South China: source, degradation, and human exposure. Environ. Toxicol. Chem. 30 (9), 1965–1972. https://doi.org/10.1002/etc.587. Watanabe, I., Tatsukawa, R., 1987. Formation of brominated dibenzofurans from the photolysis of flame retardant decabromobiphenyl ether in hexane solution by UV and sun light. Bull. Environ. Contam. Toxicol. 39 (6), 953–959. Wei, G.L., Li, D.Q., Zhuo, M.N., Liao, Y.S., Xie, Z.Y., Guo, T.L., Li, J.J., Zhang, S.Y., Liang, Z.Q., 2015. Organophosphorus flame retardants and plasticizers: sources, occurrence, toxicity and human exposure. Environ. Pollut. 196, 29–46. Weil, E.D., Levchik, S.V., 2004. Commercial flame retardancy of polyurethanes. J. Fire Sci. 22, 183–210. Whitehead, T., Metayer, C., Buffler, P., Rappaport, S.M., 2011. Estimating exposures to indoor contaminants using residential dust. J. Expo. Sci. Environ. Epidemiol. 21, 549–564.

313

Ecotoxicology and Environmental Safety 168 (2019) 304–314

I.C. Yadav et al.

Ecotoxicol. Environ. Saf. 144, 498–506. Yadav, I.C., Devi, N.L., Li, J., Zhang, G., 2018a. Environmental concentration and atmospheric deposition of halogenated flame retardants in soil from Nepal: source apportionment and soil-air partitioning. Environ. Pollut. 23, 642–654. Yadav, I.C., Devi, N.L., Li, J., Zhang, G., 2018b. Polycyclic aromatic hydrocarbons in house dust and surface soil in major urban regions of Nepal: implication on source apportionment and toxicological effect. Sci. Total Environ. 616–617, 223–235. Yadav, I.C., Devi, N.L., Zhong, G., Li, J., Zhang, G., Covaci, A., 2018c. Concentration and spatial distribution of Organophosphate Esters in the soil-sediment profile of Kathmandu Valley, Nepal: implication for risk assessment. Sci. Total Environ. 613–614, 502–512. Yadav, I.C., Devi, N.L., Li, J., Zhang, G., 2018d. Organophosphate ester flame retardants in Nepalese soil: spatial distribution, source apportionment and air-soil exchange assessment. Chemosphere 190, 114–123. Zhang, S., Xu, X., Wu, Y., Ge, J., Li, W., Huo, X., 2014. Polybrominated diphenyl ethers in residential and agricultural soils from an electronic waste polluted region in South China: distribution, compositional profile, and sources. Chemosphere 102, 55–60. Zhang, X.L., Luo, X.J., Liu, H.Y., Yu, L.H., Chen, S.J., Mai, B.X., 2011. Bioaccumulation of several brominated flame retardants and dechlorane plus in waterbirds from an ewaste recycling region in South China: associated with the trophic level and diet sources. Environ. Sci. Technol. 45, 400–405. Zheng, J., Wang, J., Luo, X.J., Tian, M., He, L.Y., Yuan, J.G., Mai, B.X., Yang, Z.Y., 2010. Dechlorane Plus in human hair from an E-waste recycling area in South China: Comparison with dust. Environ. Sci. Technol. 44 (24), 9298–9303. Zhu, J., Feng, Y.L., Shoeib, M., 2007. Detection of Dechlorane Plus in residential indoor dust in the city of Ottawa, Canada. Environ. Sci. Technol. 41 (22), 7694–7698.

Whitehead, T.P., Brown, F.R., Metayer, C., Park, J.S., Does, M., Petreas, M.X., 2013. Polybrominated diphenyl ethers in residential dust: sources of variability. Environ. Int. 57–58, 11–24. WHO, 1990. Tricresyl Phosphate, Environmental Health Criteria. World Health Organization, Geneva, Switzerland, pp. 110. WHO, 1991. Tributyl Phosphate. Environmental Health Criteria112. World Health Organization, Geneva, Switzerland. WHO, 1997. Environmental Health Criteria 192. Flame Retardants: a General Introduction. World Health Organization, Geneva, Switzerland. 〈http://www. inchem.org/documents/ehc/ehc/ehc192.htm〉. WHO, 2000. Environmental Health Criteria 209, Flame Retardants: tris(2-butoxyethyl) Phosphate, Tris(2-ethylhexyl) Phosphate and Tetrakis(hydroxymethyl) Phosphonium Salts. World Health Organization, Geneva, Switzerland. Wu, M.H., Pei, J.C., Zheng, M., Tang, L., Bao, Y.Y., Xu, B.T., Sun, R., Sun, Y.F., Xu, G., Lei, J.Q., 2015. Polybrominated diphenyl ethers (PBDEs) in soil and outdoor dust from a multi-functional area of Shanghai: levels, compositional profiles and interrelationships. Chemosphere 118, 87–95. Xian, Q., Siddique, S., Li, T., Feng, Y.L., Takser, L., Zhu, J., 2011. Sources and environmental behavior of Dechlorane Plus-a review. Environ. Int. 37 (7), 1273–1284. Yadav, I.C., Devi, N.L., Zhong, G., Li, J., Zhang, G., Covaci, A., 2017a. Occurrence and fate of Organophosphate Ester Flame Retardants and Plasticizers in indoor air and dust of Nepal: implication for human exposure. Environ. Pollut. 229, 668–678. Yadav, I.C., Devi, N.L., Li, J., Zhang, G., 2017b. Occurrence and source apportionment of halogenated flame retardants in the indoor air of Nepalese cities: implication on human health. Atmos. Environ. 161, 122–131. Yadav, I.C., Devi, N.L., Li, J., Zhang, G., 2017c. Polychlorinated biphenyls in Nepalese surface soils: spatial distribution, air-soil exchange, and soil-air partitioning.

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