Occurrence, distribution and health risk assessment of organophosphate esters in outdoor dust in Nanjing, China: Urban vs. rural areas

Chemosphere 231 (2019) 41e50

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Occurrence, distribution and health risk assessment of organophosphate esters in outdoor dust in Nanjing, China: Urban vs. rural areas Yiqun Chen a, Qin Zhang b, Tingwen Luo c, Liqun Xing a, d, *, Huaizhou Xu b, e, ** a

State Key Laboratory of Pollution Control and Resource Reuse, School of the Environment, Nanjing University, Nanjing, 210023, China Nanjing Institute of Environmental Sciences, Ministry of Environmental Protection, No.8 Jiangwangmiao Street, Nanjing, 210042, China Key Laboratory of Urban Land Monitoring and Simulation, Ministry of Land Resource of China, Shenzhen Research Centre of Digital City Engineering, Shenzhen, 518037, China d Nanjing University & Yancheng Academy of Environmental Protection Technology and Engineering, Yancheng, 224000, China e Shen Shan Smart City Research Institute Co., Ltd, Technology Incubator Base, Chuangfu Road, Shenzhen-Shanwei Special Cooperation Zone, 516473, China b c

h i g h l i g h t s

g r a p h i c a l a b s t r a c t


Composition profiles of 13 OPEs between urban and rural dust were compared.
Significantly higher concentrations of OPEs in urban areas than in rural areas.
Concentration of SOPEs and six individual OPEs varied with seasons.
OPEs sources were traced by PCAMLR and spearman correlations.

a r t i c l e i n f o

a b s t r a c t

Article history: Received 24 February 2019 Received in revised form 13 May 2019 Accepted 16 May 2019 Available online 18 May 2019

With increasing use of organophosphate esters (OPEs) largely due to the phasing out of various brominated flame retardants, much more attention has been paid to their occurrence, distribution and potential health risks. In this study, we investigated the occurrence and distribution characteristics associated to their potential health risks of selected 13 OPEs in outdoor dust with a comparison between urban and rural areas in Nanjing, China as well as seasonal variations. Ten out of 13 OPEs showed higher concentrations in urban dust than those in rural dust (p < 0.05). Six OPEs congeners exhibited significantly different concentrations with seasonal variations (p < 0.01) in rural dust. Halogenated OPEs were the dominant group in both urban (median: 56.8%) and rural (median: 45.9%) dust, and tris(2chloroisopropyl) phosphate (TCPP) was found to be the most abundant OPE in both urban (median: 48.7%) and rural (median:26.4%) dust. Principal component analysis with multiple linear regression (PCA-MLR) and spearman correlations showed the different sources of OPEs in urban and rural dust. The non-carcinogenic (Hazard Index, HI < 1.62  105) and carcinogenic risks (CR < 2.28  109) of SOPEs

Handling Editor: R Ebinghaus Keywords: Organophosphate esters Outdoor dust Seasonal variations Human risk assessment Rural areas Urban areas

* Corresponding author. State Key Laboratory of Pollution Control and Resource Reuse, School of the Environment, Nanjing University, Nanjing, 210023, China. . ** Corresponding author. Nanjing University & Yancheng Academy of Environmental Protection Technology and Engineering, Yancheng, 224000, China. E-mail addresses: [email protected] (L. Xing), [email protected] (H. Xu). https://doi.org/10.1016/j.chemosphere.2019.05.135 0045-6535/© 2019 Elsevier Ltd. All rights reserved.

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Y. Chen et al. / Chemosphere 231 (2019) 41e50

were much lower than the theoretical threshold of risk, revealing a negligible risk to local residents from the exposure of OPEs in outdoor dust. © 2019 Elsevier Ltd. All rights reserved.

1. Introduction The use of organophosphate esters (OPEs) has rapidly increased after the phasing out of previously widely used on polybrominated diphenyl ether (PBDEs) (Li et al., 2016; Marklund et al., 2003; Rauert et al., 2018). OPEs are extensively used as additives in diverse material such as plasticizers, anti-foaming, furniture, textile, electronics, and vehicles and so on (Salamova et al., 2014; Shi et al., 2016; Yadav et al., 2017) due to their good flame retardancy and environment-friendly properties. In recent years, the global consumption of OPEs has increased from 102,000 tons in 1992 to 370,000 tons in 2013 and to 680,000 tons in 2015 (Li et al., 2018a; Wang et al., 2018a). As one of the largest consuming countries of OPEs, the consumption of OPEs increased from 70,000 tons in 2007 to 179,000 tons in 2012 with an annual increase rate of more than 20% (Li et al., 2018a). In many cases, OPEs were added in to the matrix through physical rather than chemical bonded to the material, thus they can be readily released to environment via volatilization, leaching and abrasion during the progress of production, transportation, application and disposal (Cui et al., 2017; Ivana et al., 2011; Shi et al., 2016; Yadav et al., 2017). They have been ubiquitously detected in many environmental samples, such as indoor and outdoor air (Wang et al., 2018b; Zhou et al., 2017; Vykoukalov
a ET AL., 2017), atmosphere (Bi et al., 2010; Rauert et al., 2018), wastewater and sludge (Kim et al., 2017; Pang et al., 2016), surface water (Mcdonough et al., 2018; Shi et al., 2016; Xing et al., 2018), deposit sediment (Cao et al., 2017a) and soils (KurtKarakus et al., 2018; Wan et al., 2016), and even in global oceanic
nez et al., 2016; Mo €ller et al., 2012) atmosphere (Castro-Jime proving long-range transportation to remote regions. The pollution of OPEs had been widely spread in the whole world and caused global concern. Dust is a good carrier of pollutants due to the large specific surface area, thus it plays a significant role in the transportation and distribution of pollutants from air to other environment matrices through atmospheric deposition (Cao et al., 2017a; Li et al., 2018c). Humans are exposed to OPEs via ingestion, inhalation and dermal contact of indoor and outdoor dust, which is considered to be one of the important exposure sources, especially for young children and groups working or living in areas with high OPEs levels (Xu et al., 2018). Many researchers have investigated OPEs concentrations in indoor dust and their associated health risks to human in indoor dust due to high concentrations and long-time exposure duration (Khairy and Lohmann, 2018; Li et al., 2018b; Zhu et al., 2015). To date, researchers have investigated seasonal variation on surface water (Shi et al., 2016), air (Li et al., 2018a), global atmosphere (Rauert et al., 2018) and indoor dust (Cao et al., 2014b). Besides, the concentration and profiles of rural and urban dust and human hair of Chongqing were investigated (He et al., 2018). To our knowledge, a limited number of studies have reported the occurrence, distributions and related health risk of OPEs in outdoor dust in Eastern China. Especially, there is few study that has compared the distribution, seasonal variations and their related health risk of OPEs in dust between urban and rural areas in Nanjing. It is reported that the physicochemical characteristics (e.g., macro-element, ions and deposition rate) of dust were different between indoor and outdoor as well as urban and rural (Xu et al., 2018; Yang et al., 2016). Khan

et al. (2016) has compared the differences of the flame retardants occurrence, distribution and associated health risk in indoor and outdoor dust of industrial, rural and background zones of Pakistan. Ren et al. (2016) reported that the composition profile for OPEs concentrations in total suspended particulates was different between the urban and rural areas and originated from different sources. Hoffman et al. (2017) found that season could be an important factor for OPEs exposure. The aim of this study was to generate new knowledge about the differences of occurrence, spatial distribution and health risk of OPEs in outdoor dust between urban and rural areas in Nanjing, China. Meanwhile, life-style in rural areas, dependence on land for growing field crops, was highly related with seasonal change and has higher frequency to contact with outdoor dust/soil, especially for agricultural use of plastics, for example agricultural plastic film, pesticide bottle, and fertilizer plastic bag, etc. Therefore, seasonal variation characteristics of OPEs in outdoor dust from rural areas were investigated. During the process of study, different kinds of exposure areas including school, village, chemical industry park, high density of human and traffic flow were considered to make a comprehensive overview of OPEs distribution in outdoor dust. The sources identification of OPEs was achieved by correlation coefficient and principal component analysis with multiple linear regression (PCA-MLR) model, which is often used to trace the source of measured pollutant (Jiang et al., 2016; Liu et al., 2018). In addition, risk assessment for different exposed population groups (children, adults and outdoor workers such as professional street sweepers) related to non-carcinogenic and carcinogenic health risk based on three pathways (ingestion, inhalation and dermal contact) were conducted according to hazard index (HI) and carcinogenic risk (CR) methods (Li et al., 2018b, 2018c; Xu et al., 2018). The results could fill the knowledge gaps of the occurrence, temporal and spatial distributions of OPEs in dust, and provide an insight into the risk of OPEs in China. 2. Materials and methods 2.1. Chemicals and reagents A total of 13 OPEs were investigated in this study including six alkyl OPEs, four halogenated OPEs and three aryl OPEs, which were obtained from J&K Chemical, Ltd. (St. Louis, MO, USA). The six alkyl OPEs were trimethyl phosphate (TMP), tripropyl phosphate (TPP), tri-n-butyl Phosphate (TnBP), triethyl phosphate (TEP), tris(2butoxyethyl)phosphate (TBEP) and tris(2-ethylhexyl) phosphate (TEHP), four halogenated OPEs contained tris(2-chloroethyl) phosphate (TCEP), tris(2-chloroisopropyl) phosphate (TCPP), tris(1,3-dichloroisopropyl) phosphate (TDCPP) and tris(2,3dibromopropyl) phosphate (TDBPP), and three aryl OPEs contained triphenyl phosphate (TPhP), 2-ethylhexyl diphenyl phosphate (EHDPP) and tricresyl phosphate (TCP). The mass-labelled standards for TnBP-d27 and TPhP-d15 were purchased from Cambridge Isotope Laboratories (Tewksbury, MA, USA). Deionized water (18.2 MU cm resistivity at 25  C) produced by a Milli-Q water purification system (Millipore, Bedford, MA) was used in all the experiments. Methanol, ethyl acetate and n-hexane were purchased from Sigma-Aldrich (Pittsburgh, PA, USA), Acetone were purchased

Y. Chen et al. / Chemosphere 231 (2019) 41e50

from Merck (Germany). All solvents used throughout this study were HPLC grade. The detailed information of all the target chemicals is listed in Table S1. 2.2. Sample collection Dust samples from ten sampling sites in rural areas, including seven countrysides named Shi Pai, Shui Qutou, Tai Jiazhuang, Qiang Wuli, Xin hua, Hou Lanxiang and Xiang Kou and three different stage schools named Kindergarten, Yaxi Primary school and Yaxi Middle school, were collected in four different months, March 2016, December 2016, July 2017 and October 2017 represented four different seasons in Nanjing, respectively. Here, it is necessary to declare that the wet-dry seasons alternation during these sampling campaigns is not clear as Nanjing being located in eastern China. A total of eight samples in urban areas, including five subway entrances of heavy transportation named Jiang Wangmiao, Ma Qun, Da Chang, Zhong Huamen and Gu Lou, two residence communities named Tian Runcheng and Dongjiao town, and one Chemical Industry Park, were collected in October 2017. The sampling sites in urban and rural areas were shown in Fig. S1. In each sample site, we collected four samples and mixed them together as one sample. Each dust sample was collected from an area of a circle with a semidiameter of 300 m, in which the dust was mostly made of atmospheric deposition. Samples were collected with a polyethylene brush and tray, and then transferred into a pre-cleaned stainless steel contains. After collection, each sample was sieved (100-mesh) and the fraction below 250 mm was collected and stored at 20  C before the extraction of organic solvents. All dust samples were pretreated within five days.

2.4. Quality assurance and quality control Any plastic and rubber materials were avoided to minimize different OPEs contamination of the samples during sampling, storage, transport, and extraction. Similar to our previous studies (Xing et al., 2018), strict protocols were established to ensure reliable results. Solvent blanks, standard and procedure blanks were simultaneously run in sequence to check for background contamination, peak identification and quantification. The quantification of OPEs was conducted using an internal standard method and the correlation coefficients (R2) of the calibration curves were greater than 0.994. Three levels of 13 OPEs were spiked into 1.0 g of the matrix blank (fine black soil of China with almost no target compounds) with triplicates. Extraction recoveries for 13 OPEs were within the satisfactory range of 73%e115% and the relative standard deviations (RSDs) of all the OPEs were under 10%, results were show in Table S3. TCEP and TPhP were detected in the field blanks. Limit of detection (LOD) was calculated from the mean of field blanks (n ¼ 6)þ3  SD. For other compounds which were not detected in the blanks, the instrument detection limit (IDL) was used for calculation of LODs. The IDL corresponded to the amount of analyte that generated a signal to noise ratio of 3:1. The method detection limit (MDL) experiment was not conducted, and IDLs were used to as the MDL in this study. The calculated LODs of 13 OPEs were between 0.03 ng/g and 1.59 ng/g.

2.5. Human health risk assessment The chronic daily intake (CDI) values for OPEs with the dust via ingestion, inhalation and dermal contact are calculated according to the following equations (Li et al., 2018b, 2018c; Xu et al., 2018).

2.3. Sample pretreatment and analyses All samples were extracted according to the method described by our previous studies with some modifications (Xing et al., 2018). Briefly, accurately weighed 1.0 g dust sample was spiked with 20 ng TnBP-d27 as surrogate and the mixture was then incubated overnight. After that, the samples were extracted with 10 mL Acetone: n-hexcane (v:v, 1:3) followed by ultrasonic extraction for 30 min, then the extracts were centrifuged (3000 rpm, 15 min) and transferred to 40 mL amber vials. This process was performed twice. The total extract was concentrated to about 1 mL using a rotary evaporator. Then the analytes were diluted to 100 mL with deionized water for solid-phase extraction (SPE). After activated with 5 mL ethyl acetate, 5 mL methanol and 5 mL ultrapure water, Oasis HLB cartridges (500 mg, 6 mL, Waters) were passed through by filtrates at a flow rate of about 4 mL/min. After that, the cartridges were washed with 2  5 mL ethyl acetate at a flow rate of approximately 1 mL/min and then concentrated to about 200 mL with a gentle stream of nitrogen. The final extracts were spiked with 10 ng of TPhP-d15 before analyzed by HPLC-MS/MS. The extracts were finally analyzed using the HPLC-MS/MS (LCAgilent Technologies 1290 Infinity, MS-AB SCIEX QTRAP 4500; CA). Chromatographic analysis was performed with the ZORBAX Eclipse Plus C18 column (150 mm  2.1 mm, 3.5 mm; Agilent) at a constant temperature of 30  C, with the injection volume of 5 mL. 0.02% (v/v) formic acid (A) and methanol (B) was used as the mobile phases for the separation of analytes at a flow rate of 0.3 mL/min. The starting composition of mobile phase was 30% A, held for 8 min. Then the linear gradient was decrease from 30% A to 5% A at 0.1 min, and mobile phase at 5% A was held for 7.9 min finally return to the starting composition of 30% A at 4 min. The quantification and confirmation ions for each analyte in multiple reaction monitoring mode are given in Table S2 in Supplementary Information.

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CDIingestion ¼

Ci  IRing  ED  EF  CF AT  BW

CDIinhalation ¼

CDIdermal

Ci  IRing  ED  EF AT  BW  RPE

contact

¼

Ci  ED  EF  SA  AF  ABS  CF AT  BW

(1)

(2)

(3)

where Ci is the concentrations of individual OPEs in dust (ng/g); IRing is ingestion rate of dust (mg/day); IRinh is inhalation rate of dust (m3/day); ED is the exposure duration (day); EF is the exposure frequency (day/year); CF is the conversion factor (106 kg/mg); AT is the average time (day), be equal to ED for the non-carcinogens and 70 years for carcinogens; BW is the body weight (kg); RPE is the particle emission rate; SA is the surface area of the skin that contacts dust (cm2); AF is the skin adherence factor (mg/cm2); ABS is the dermal adsorption fraction of individual OPEs. The parameters used in health risk and CDI calculation are listed in Table S4 and Table S5. The non-carcinogenic risk (non-CR) assessment was based on the Hazard index (HI) and Hazard Quotient (HQ). The HI of multiple OPEs for non-CR and CR via ingestion, inhalation and dermal contact pathways was calculated according the following equations (Li et al., 2018b).

HI ¼ HQing þ HQinh þ HQderm ¼

Xn CDIing þ CDIinh 1 RfDi  CDIderm þ RfDi  GIABSi

(4)

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Y. Chen et al. / Chemosphere 231 (2019) 41e50

CR ¼ CRing þ CRinh þ CRderm ¼

X n   CDIing þ CDIinh  SFOi 1  SFOi þ CDIderm  GIABSi (5)

where GIABSi and SFOi is respective gastrointestinal absorption factor and corresponding oral cancer slope factor of individual OPEs. Under most regulatory programs, the HI value of less than 1 for OPEs exposures is considered acceptable for non-CR; while CR value below 1.0  106 indicates negligible cancer risk, whereas a value between 1.0  106 and 1.0  104 suggests potential cancer risk, and a value above 1.0  104 is an indication of high-potential risk (Li et al., 2018b, 2018c).

2.6. Data analysis All the statistical analyses were performed by using statistical computing software R version 3.5.1 (R Development Core Team, http://www.r-project.org/). Nonparametric Kruskal-Wallis oneway analysis of variance on ranks was performed to detect the significant differences. Half of the MDL was used for the concentration values below the detection limit according to previous

studies (Mishra et al., 2015; Sun et al., 2016). Differences were considered significant at p < 0.05. Spearman correlation analysis and PCA-MLR were conducted to establish the relationship and sources of OPEs. It is well known that low concentration lead to the low detection frequencies (DFs). The DFs are usually lower in rural than in the urban area mainly due to lower concentration level than method detection limit. 3. Results and discussion 3.1. Concentrations and distribution of OPEs in dust The concentrations of 13 detected individual OPEs and SOPEs from urban and rural areas in Nanjing in October 2017 were summarized in Table 1 and Fig. 1. Most of the investigated OPEs were found both in urban and rural areas, but higher detection frequencies (DFs) were found in urban dust. Eleven congeners out of the 13 OPEs (except TMP and TPP) were detected with a DF of 100% in urban dust. Eight congeners out of 13 OPEs (except TMP, TPP, TnBP, TDBPP and EHDPP) were detected with a DF of 100% in rural dust. The DFs of TDBPP and EHDPP were 100% in urban dust, but declined to 10% and 20% in rural dust, respectively. The DF of TBEP (100%) in our study is consist with that in road dust in Beijing (Li et al., 2018c), but much higher than that in sub-urban (4%) and urban (2%) sites in Shanghai (Ren et al., 2016). The DFs of TMP

Table 1 Summary for OPEs concentrations in outdoor dust (ng/g dw) in urban and rural areas in October 2017. Analytes

TMP TEP TPP TnBP TBEP TEHP TCEP TCPP TDCPP TDBPP TPhP EHDPP TCP Alkyl OPEs Halogenated OPEs Aryl OPEs P OPEs

Urban areas (n ¼ 8)

Rural areas (n ¼ 10)

DF (%)

Mean

Median

Range

DF (%)

Mean

Median

Range

ND 100 25 100 100 100 100 100 100 100 100 100 100 100 100 100 100

ND 4.56 0.02 2.98 11.03 27.21 7.81 90.65 2.61 2.17 17.92 3.04 4.39 174.39 45.81 103.24 174.39

ND 4.43 ND 2.82 5.35 26.75 6.85 81.99 2.77 1.29 16.59 2.14 2.94 157.25 45.46 93.18 157.25

ND 1.17e7.30 ND-0.11 0.85e6.03 1.25e33.18 15.15e39.18 3.21e15.35 29.61e216.77 1.17e3.46 0.65e4.27 4.37e36.57 0.47e6.12 1.59e10.17 66.79e367.01 22.64e82.00 36.21e232.46 66.79e367.01

12.5 100 30 90 100 100 100 100 100 10 100 20 100 100 100 100 100

0.22 2.52 0.02 2.23 0.11 2.27 3.25 7.14 0.88 0.06 2.08 0.06 0.16 33.18 16.53 2.43 20.99

ND 2.49 ND 1.33 0.11 1.64 1.85 4.12 0.80 ND 1.16 ND 0.12 18.64 15.28 1.88 20.08

ND-2.20 1.04e4.74 ND-0.16 ND-8.66 0.03e0.20 1.02e5.06 0.57e12.77 3.21e16.16 0.45e1.50 ND-0.56 0.58e7.23 ND-0.34 0.02e0.65 5.45e99.95 8.96e30.13 0.30e7.06 10.89e31.27

DF: Detection Frequency. ND: Not Detected.

Fig. 1. Concentration levels and distribution characteristics of OPEs ((a) barplot of OPEs concentrations, (b) boxplot of OPEs concentrations) in different dust samples from urban and rural areas outdoors (*p < 0.05, **p < 0.01, ***p < 0.001).

Y. Chen et al. / Chemosphere 231 (2019) 41e50

(98.5%) and EHDPP (60%) in road dust in Beijing is much higher than those in our study, while the DF of TPP (0.0%) is lower than that in our study (Li et al., 2018c). The SOPEs concentrations ranged from 66.8 to 367 ng/g dw (dry weight) with a median of 157 ng/g dw in urban dust and ranged from 10.9 to 31.3 ng/g dw with a median of 20.1 ng/g dw in rural dust, respectively. As shown in Fig. 1, the concentrations of individual and sum of OPEs detected in urban dust were approximately eight times higher than those in rural dust, presenting significant differences (p < 0.05) except TMP, TPP and TnBP, which is consistent with previous studies (He et al., 2018; Ren et al., 2016). The concentrations of SOPEs in road/street dust in Beijing (Li et al., 2018c) or Chongqing (He et al., 2018) are much higher than that in our study as well as alkyl-, halogenated- and aryl-OPEs, which may be due to different sampling sites and OPEs sources. Generally, urban areas with higher population density consumed more plastics and commercial materials than rural areas. It was demonstrated that OPEs contents in urban area were likely related to the intensity of industrial, commercial activities as well as population density (Cequier et al., 2014; Cui et al., 2017; Li et al., 2018c). The concentrations of SOPEs at subway entrances ranged from 93.0 ng/g dw for Zhong Huamen to 367 ng/g dw for Gu Lou. The concentrations of SOPEs at residence communities were 130 ng/g dw for Tian Runcheng and 247 ng/g dw for Dongjiao town, respectively. The concentration of SOPEs at chemical industry park was 98.1 ng/g dw. It was noted that Gu Lou, as the center of the city with high population density and heavy traffic, had highest concentration of SOPEs than any other sites, which is agreed with the previous studies (Cequier et al., 2014; Cui et al., 2017; Li et al., 2018c). The concentrations of SOPEs in chemical industry park, residence community and subway entrance further revealed the density of population and traffic condition. In case of rural areas, the concentrations of SOPEs at school ranged from 16.7 ng/g dw to 31.3 ng/ g dw, and the concentrations of SOPEs at countryside ranged from 10.9 ng/g dw for Xiang Kou to 26.8 ng/g dw for Shi Pai. These results indicated that the OPEs were almost equally used for flame retardants and/or plasticizers. However, it is reported that the concentration levels of OPEs in primary school dust were in the range of 10-10,000 ng/g dw (Cequier et al., 2014), which showed relatively higher OPE levels than that in street dust due to be a repository for organic compounds released from both indoors and outdoors (Cao et al., 2014a; He et al., 2018). Analogue profiles of OPEs in dust from urban and rural areas were illustrated in Fig. S2. Similar distribution patterns were exhibited with halogenated OPEs occupying a large proportion, followed by alkyl OPEs and aryl OPEs both in urban and rural areas. Halogenated OPEs were the predominant compounds both in urban dust (56.8%) and rural dust (46.0%), followed by alkyl OPEs (urban dust 27.8%, rural dust 35.7%) and aryl OPEs (urban dust 13.8%, rural dust 7.78%). It is consistent with the previous study in road dust (Li et al., 2018c). The composition distribution percentages in urban dust varied less than that in rural dust. TCPP was the dominant pollutant both in urban dust (48.7%) and rural dust (26.4%), followed by TEHP (17.8%) and TPhP (9.3%) in urban dust, and followed by TEP (12.6%), TCEP (10.6%) and TEHP (9.88%) in rural dust. It is well known that TCPP and TCEP are applied as the most popular organophosphate flame retardants in flexible and rigid polyurethane material, such as upholstery foam, mattress, automotive steering, wheel head and other car accessories (Marklund et al., 2003; Shi et al., 2016; Wei et al., 2015). It is reported that TCPP represents approximately 80% of the chlorinated phosphorus flame retardants (Van and De, 2012). Over 98% of TCPP was used as flame retardant additive in building insulation materials and household products including polyurethane foam furniture, indicating that emissions from these materials may significantly

45

contribute to the occurrence of OPEs in indoor and outdoor air (Zhou et al., 2017). It seems to be reasonable that TCPP was usually found as the dominant compound in some media such as road dust, suspended particulates and indoor and outdoor air (Li et al., 2018c; Ren et al., 2016; Zhou et al., 2017). In comparison, alkyl OPEs occupied the most proportion, followed by halogenated OPEs and aryl OPEs in Yaxi middle school, Xiangkou, Qiang Wuli and Xinhua from rural areas. The proportion of TBEP was lower in rural areas than that in urban areas, while the proportions of TEP, TnBP and TCEP were higher in rural areas than those in urban areas. It is also reported that TCEP occupied high proportion in street dust of Chongqing (He et al., 2017). The dominant OPEs and proportions from different areas varied due to different pollution sources of OPEs, different life-style and consumptions. TnBP in rural areas might attributed to the input from the urban areas atmosphere deposition (Wei et al., 2015) and wide use of them as hydraulic fluids in vehicles (Marklund et al., 2005; van der Veen and de Boer, 2012). 3.2. Seasonal variation of OPEs in dust sample To get a better comprehension of the seasonal variability of OPEs, samples from four different months in rural areas were further analyzed (Fig. 2). Generally, volatile organic compounds (VOCs) have higher concentrations in summer due to the higher temperature (Faiz et al., 2017), while it may also elevate the transfer of VOCs from particle phase into gas phase. The highest concentrations of SOPEs were found in July for Shui Qutou and Yaxi middle school, in which TnBP with medium vapor pressure 1.10  103 pa at 25  C among investigated OPEs was the most abundance compound accounting for more than 50%. There was no obvious seasonal variability for the concentration of SOPEs trend, which is consistent with the findings by Li et al. (2018a). While the levels of SOPEs in July and March were usually higher than October and December in most sampling sites. It is reported that organophosphate flame retardants exposure could be temperature-dependent (Hoffman et al., 2017). It was demonstrated that PBDEs in outdoor air concentration increased during spring due to snow melted (Gouin et al., 2005). The concentration of SOPEs increased with wet deposition, because snow and rainfall could transfer air OPEs to dust. Besides, OPEs can be absorbed to dust from aqueous release (Zeng et al., 2015) due to physicochemical properties of hydrophobicity, hardly desorption from particles and poor biodegradability. There were significant differences (p < 0.01) for TPP, TnBP, TCEP, TPhP, EHDPP, TCP and SOPEs as well as Salkyl-OPEs and Saryl-OPEs among four months, but no significant differences (p > 0.05) for the rest individual OPEs and halogenated OPEs. Therefore, it is a comprehensive result dependence on the impact of temperature on volatilization from surface of targets and particle phase into gas phase. It is inferred that different temperature may have different impacts on different OPEs with different properties (e.g., vapor pressure and octanol-water partition coefficient). 3.3. Probable source identification Spearman rank correlation coefficient and PCA-MLR were widely applied for source identification and relationship determination in dust according to previous studies (Jiang et al., 2016; Khan et al., 2016; Liu et al., 2018; Wang et al., 2018b) based on the measured concentrations of OPEs. The mean percentage contribution of each factors can be calculated according to previous studies based on PCA-MLR model (Jiang et al., 2016; Liu et al., 2018). The variation diagram with three principal components (PCs) for both urban and rural areas was shown in Fig. 3. It shows that PC1, PC2 and PC3 account for 61.5%, 11.7% and 7.61% of the variance,

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Y. Chen et al. / Chemosphere 231 (2019) 41e50

Fig. 2. Temporal variation of concentration and distribution characteristics of OPEs ((a) barplot of OPEs concentrations, (b) boxplot of OPEs concentrations) in different dust samples from rural areas.

Fig. 3. Plot of principal component scores of OPEs as variables shown in green text with three and two components for outdoors dust samples from both urban and rural areas. (For interpretation of the references to colour in this figure legend, the reader is referred to the Web version of this article.)

with the three PCs accounted for 80.7% of the total variances. The samples clustered into two main groups, where dust samples from urban areas (darkblue) were mostly located on the left-side of PC1, and dust samples from rural areas (orange) were located on the right-side of PC1. Consistent with the findings of Wong et al. (2017), there are no obvious patterns observed in the relative abundances with a predominant OPE for PC1 due to diverse application of OPEs in consumer products on the market, in which the relative abundances of TBEP, TEHP, TCPP, TDBPP, TPhP, EHDPP and TCP were similar. While relative abundances were similar with predominance of TnBP and TPP for PC2. PCA-MLR based on the measured concentrations of OPEs in urban and rural areas was respectively performed (except TMP in urban areas that was not detected in all dust samples). The number of factors to be retained according to Kaiser's criteria (Kaiser, 1958), in which PCs with eigenvalues>1 were considered as the most significant factors. The three principal components (PC1, PC2 and PC3) explained 58.9%, 18.2% and 11.3% of the total variance for urban areas, respectively; while the five principal components (PC1, PC2, PC3, PC4 and PC5) explained 34.7%, 20.6%, 12.6%, 9.63% and 8.27% of the total variance in rural areas, respectively (Table S6). The

emission source of road dust was considered from several pathways, such as industrial discharge, residential sources, the car dust and engine oil and tires (Li et al., 2018c). TBEP, TCPP, TPhP and TDCPP were usually detected in car dust with a highest concentration of 1100 mg/g dominated by TDCPP (Brandsma et al., 2014). TBEP originates from floor wax, plastics or rubber products, and the commonly used chlorinated OPEs such as TCPP, TCEP and TDCPP originates from rigid and flexible polyurethane foams (Marklund et al., 2003; Wang et al., 2018b; van der Veen and de Boer, 2012). It is known that TCEP is contained in many products such as building materials, paints and soft foams (Wei et al., 2015). The sources TnBP, TDCPP, TPhP, TCPP and TCP were mainly assigned to the diffusion-dominated mass transfer of atmospheric pollutants from industrial area (Ren et al., 2016). Thus, PC1 appears to be indicative of multiple sources of emissions with percentage contribution of 72.6% for urban areas and 6.88% for rural areas based on PCA-MLR, respectively. In urban areas, PC2 was characterized by high loading of TnBP, suggesting traffic emission (contribution 23.0%) (Marklund et al., 2005; Zhou et al., 2017), and PC3 was characterized by high loading of TCEP, suggesting polyurethane foams (contribution 4.34%) (Marklund et al., 2003; Wei

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et al., 2015; van der Veen and de Boer, 2012). In rural areas, PC2 was characterized by high loadings of TPP and TCP suggesting industrial emission transfer (contribution 16.1%); PC3 was characterized by high loading of TDBPP, suggesting residential sources (contribution
et al., 2017); PC4 was not characterized by 4.34%) (Vykoukalova some OPEs, suggesting no specific source (contribution 30.9%); and PC5 was characterized by high loadings of TnBP and TCEP, suggesting traffic emission and polyurethane foams (contribution 23.5%) (Marklund et al., 2003; Marklund et al., 2005; van der Veen and de Boer, 2012; Zhou et al., 2017). In rural areas, the use of plastics in agricultural production may be transferring by atmospheric and deposit somewhere in dust, however, more information is needed to trace the sources. Correlation coefficients, as a non-parametric measure of correlation, uses ranks to calculate the correlation between concentrations of OPEs with >90% DF (Table S7), indicating the existence of the same predominant sources, similar properties, and/or similar emission pathways of both compounds. The results showed significant correlations (p < 0.05) between most OPEs pairs except TnBP and TDBPP in urban areas, while significant correlations (p < 0.05) were only found in three pairs: TBEP and TnBP, TEHP and TCP, TPhP and TCP in rural areas. To our best knowledge, the potential sources of OPEs in outdoor environment can be put down to the release of OPEs from the direct discharge from the industry, residential sources, mobile sources as well as the diffuse discharge from road soil (Li et al., 2018c; Marklund et al., 2003; Wei et al., 2015; Zhou et al., 2017). Road dust was contaminated by traffic and the OPEs released from vehicles which reflect the application of OPEs in vehicles (Cao et al., 2017b; Li et al., 2018c; Marklund et al., 2003). Taking TnBP as an example, it is one of the main ingredients (65e79%) in several commercial hydraulic fluids used in vehicles and high amounts were detected in indoor dust and air of cars, thus hydraulic fluids and aircraft were regarded as a source of OPEs in outdoor environments (Marklund et al., 2005; Zhou et al., 2017). Building materials such as polyolefin covering, insulation hard foam and acoustic ceilings were considered as the main sources of TEP, TCEP and TDCPP in indoor dust, and that can also contaminate outdoor dust through the dust exchange between outdoor and indoor environment.

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In addition, OPEs in dust can transform to other environment media, raindrop eluted the water-solubility OPEs from dust entry to surface water and ground water finally gathering in rivers, lakes and reservoir. However, the low vapor pressures of OPEs make it hard to release to the air from aquatic phase (Regnery and Püttmann, 2009). It is also reported that the volatilization of OPEs from soils to the air is presumed to be low as same to their low vapor pressures and high dust adsorption coefficient (WHO, 2000). Therefore, the release of OPEs from dust to air was difficult to occur because of the low Henry's law constants and high dust adsorption coefficient. Due to the low Henry's law constants, the release of OPEs from the aquatic phase to the air is not likely occur (Regnery and Püttmann, 2009). The snow samples and air samples collected in Northern Sweden indicated the transformation of OPEs via atmosphere after emission from urban areas by traffic (Marklund et al., 2005). Emphatically, regardless of short half-lives, OPEs in dust could undergo a long range atmosphere transform and
nez et al., 2016; occurred in global oceanic atmosphere (Castro-Jime €ller et al., 2012). Mo 3.4. Health risk assessment The available parameters related to health including ten OPEs for non-carcinogenic risk and six OPEs for carcinogenic risk are listed in Table S4 and Table S5. The results of individual and total sum HIs and carcinogenic risks of health-related OPEs for children, adults and outdoor workers (as a special group) are shown in Fig. 4 and Fig. 5. The non-carcinogenic (HI < 1.62  105) and carcinogenic (CR < 2.28  109) risk values of OPEs were all relatively lower than the theoretical risk threshold values and indicating 3e5 orders of magnitude lower than the acceptable risk level and suggesting a negligible risk to local residents from the exposure of OPEs in outdoor dust. Those results were agreed with road dust in Beijing (Li et al., 2018c), but much lower than that of indoor dust (Li et al., 2018b). However, the values of both HI and CR in urban areas were one order higher than that in rural areas. Differently, TCPP was always the primary contributors to the non-carcinogenic risk and TDBPP was primary contributors to carcinogenic risk in urban areas, respectively. Whereas primary contributors were not always

Fig. 4. Non-carcinogenic HI related to OPEs ((a) barplot of HI, (b) boxplot of HI) in different dust samples from urban and rural areas outdoors. (Note: Ch, children; Au, adults; Sp, special group).

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Fig. 5. CR related to OPEs ((a) barplot of CR, (b) boxplot of CR) in different dust samples from urban and rural areas outdoors.

the same OPE congener in rural areas, TCPP as the dominant risk contributor in seven out of ten sites, TCEP as the dominant risk contributor in two out of ten sites, and TnBP as the dominant risk contributor in one out of ten sites for non-carcinogenic risk; TCEP as the dominant risk contributor in seven out of ten sites, TCPP, TDCPP and TnBP as the dominant risk contributor in one out of ten sites for carcinogenic risk, respectively. As shown in Fig. 4, the median HI values of SOPEs in urban areas were 6.47  106 for children, 1.44  106 for adults, and 5.77  106 for special group, respectively; the median HI values of SOPEs in rural areas were 9.97  107 for children, 2.22  107 for adults, and 8.62  107 for special group, respectively. The same risk levels were also found in road dust in Beijing (Li et al., 2018c), but lower than that caused by indoor dust (Li et al., 2018b). It is well known that the risk threshold values is 1 for HI, adverse health risk may occur if HI > 1 (Li et al., 2018b, 2018c). The HI values of SOPEs were in the order of children > special group (outdoor worker) > adult. Higher risks in urban areas were found due to higher concentrations compared to rural areas. In addition, some studies indicated that children had higher frequency to intact with the outdoor dust (Wei et al., 2015). Hence, as the sensitive subpopulation we should pay more attention on children for the chemical health effect. Outdoor workers usually spend much longer time in outdoor environment than normal adult, causing similar risk as children. Gu Lou presented the highest risk in the urban areas probably due to high density of traffic and population (Cequier et al., 2014; Cui et al., 2017; Li et al., 2018c). As shown in Fig. 5, the median CR values of the SOPEs in urban areas were 1.67  1010 for children, 1.84  1010 for adults, and 7.37  1010 for special group, respectively; the median CR values of SOPEs in rural areas were 1.05  1011 for children, 1.25  1011 for adults, and 5.00  1011 for special group, respectively. The risk threshold value is 1.0  106 for CR (Li et al., 2018b, 2018c). Different from the HI values, the CR values of SOPEs were in the order of special group (outdoor worker) > adult > children. Outdoor workers were posed for higher carcinogenic risk due to higher exposure frequency in outdoor environment, while children and adult presented similar risk levels as reported by Li et al. (2018a). In spite of the fact that the health risk of OPEs in outdoor dust is negligible, some other exposure scenarios by dust should be

considered due to higher OPEs concentration levels in indoor dust or air and more time spending in indoor environment (residence, office and school, etc.) (Cao et al., 2014a; He et al., 2018; Khairy and Lohmann, 2018; Zhou et al., 2017). It is reported that indoor activities comprised between 64.0% and 92.7% of the total health risk incurred during daily indoor and outdoor activities (Liu et al., 2016). The results in this study further fill the knowledge gaps of the risk level of OPEs through outdoor dust in China.

4. Conclusions In this study, we explored the concentration, seasonal variations and health risk assessments of OPEs in outdoor dust from urban and rural areas in Nanjing, China. Halogenated OPEs were found as dominant group in both urban and rural dust, and TCPP was found to be the most abundant OPE in both urban and rural dust. The concentrations of ten out of 13 OPEs (expect TMP, TPP, and TnBP) and SOPEs in urban dust were significantly higher than those in rural dust. The highest concentration of SOPEs was shown at Gu Lou form urban areas, which is the center of the city with high density population and traffic. In rural areas, six out of ten OPEs concentrations including TPP, TnBP, TCEP, TPhP, EHDPP, TCP and SOPEs presented significantly differences with seasonal variations. PCA-MLR and spearman correlations showed different sources of OPEs in rural and urban areas, the sources of OPEs in rural is more complicated. Negligible non-carcinogenic risks and accepted carcinogenic risks of SOPEs were suggested through outdoor dust in urban and rural areas.

Author information The authors declare no competing financial interest.

Acknowledgments This work was funded by the National Natural Science Foundation of China (Grant 21507036).

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Appendix A. Supplementary data Supplementary data to this article can be found online at https://doi.org/10.1016/j.chemosphere.2019.05.135. References Bi, X., Simoneit, B.R.T., Wang, Z.Z., Wang, X., Sheng, G., Fu, J., 2010. The major components of particles emitted during recycling of waste printed circuit boards in a typical e-waste workshop of South China. Atmos. Environ. 44, 4440e4445. Brandsma, S.H., Boer, J.D., Velzen, M.J.M.V., Leonards, P.E.G., 2014. Organophosphorus flame retardants (PFRs) and plasticizers in house and car dust and the influence of electronic equipment. Chemosphere 116, 3e9. Cao, D., Guo, J., Wang, Y., Li, Z., Liang, K., Corcoran, M.B., Hosseini, S., Bonina, S.M., Rockne, K.J., Sturchio, N.C., 2017a. Organophosphate esters in sediment of the great lakes. Environ. Sci. Technol. 51, 1441e1449. Cao, Z., Xu, F., Covaci, A., Wu, M., Wang, H., Yu, G., Wang, B., Deng, S., Huang, J., Wang, X., 2014a. Distribution patterns of brominated, chlorinated, and phosphorus flame retardants with particle size in indoor and outdoor dust and implications for human exposure. Environ. Sci. Technol. 48, 8839e8846. Cao, Z., Xu, F., Covaci, A., Wu, M., Yu, G., Wang, B., Deng, S., Huang, J., 2014b. Differences in the seasonal variation of brominated and phosphorus flame retardants in office dust. Environ. Int. 65, 100e106. Cao, Z., Zhao, L., Kuang, J., Chen, Q., Zhu, G., Zhang, K., Wang, S., Wu, P., Zhang, X., Wang, X., 2017b. Vehicles as outdoor BFR sources: Evidence from an investigation of BFR occurrence in road dust. Chemosphere 179, 29e36.
nez, J., Gonzalez-Gaya, B., Pizarro, M., Casal, P., Pizarro-Alvarez, C., Castro-Jime Dachs, J., 2016. Organophosphate ester flame retardants and plasticizers in the global oceanic atmosphere. Environ. Sci. Technol. 50, 12831e12839.
, R.M., Becher, G., Thomsen, C., 2014. Cequier, E., Ionas, A.C., Covaci, A., Marce Occurrence of a broad range of legacy and emerging flame retardants in indoor environments in Norway. Environ. Sci. Technol. 48, 6827e6835. Cui, K., Wen, J., Zeng, F., Li, S., Zhou, X., Zeng, Z., 2017. Occurrence and distribution of organophosphate esters in urban soils of the subtropical city, Guangzhou, China. Chemosphere 175, 514e520. Faiz, Y., Siddique, N., He, H., Sun, C., Waheed, S., 2017. Occurrence and profile of organophosphorus compounds in fine and coarse particulate matter from two urban areas of China and Pakistan. Environ. Pollut. 233, 26e34. Gouin, T., Daly, T.H.L., Wania, F., Mackay, D., Jones, K.C., 2005. Variability of concentrations of polybrominated diphenyl ethers and polychlorinated biphenyls in air: implications for monitoring, modeling and control. Atmos. Environ. 39, 151e166. He, M.J., Lu, J.F., Ma, J.Y., Wang, H., Du, X.F., 2018. Organophosphate esters and phthalate esters in human hair from rural and urban areas, Chongqing, China: concentrations, composition profiles and sources in comparison to street dust. Environ. Pollut. 237, 143e153. He, M.J., Yang, T., Yang, Z.H., Li, Q., Wei, S.Q., 2017. Occurrence and distribution of organophosphate esters in surface soil and street dust from Chongqing, China: implications for human exposure. Arch. Environ. Contam. Toxicol. 73, 1e13. Hoffman, K., Butt, C.M., Webster, T.F., Preston, E.V., Hammel, S.C., Makey, C., Lorenzo, A.M., Cooper, E.M., Carignan, C., Meeker, J.D., 2017. Temporal trends in exposure to organophosphate flame retardants in the United States. Environ. Sci. Technol. Lett. 4, 112e118. Ivana, M., Mirjana Vojinovic, M., Elke, F., 2011. Application of Twisselmann extraction, SPME, and GC-MS to assess input sources for organophosphate esters into soil. Environ. Sci. Technol. 45, 2264e2269. Jiang, J.J., Lee, C.L., Brimblecombe, P., Vydrova, L., Fang, M.D., 2016. Source contributions and mass loadings for chemicals of emerging concern: Chemometric application of pharmaco-signature in different aquatic systems. Environ. Pollut. 208, 79e86. Kaiser, H.F., 1958. The varimax criterion for analytic rotation in factor analysis. Psychometrika 23, 187e200. Khairy, M.A., Lohmann, R., 2018. Selected organohalogenated flame retardants in Egyptian indoor and outdoor environments: levels, sources and implications for human exposure. Sci. Total Environ. 633, 1536e1548. Khan, M.U., Li, J., Zhang, G., Malik, R.N., 2016. New insight into the levels, distribution and health risk diagnosis of indoor and outdoor dust-bound FRs in colder, rural and industrial zones of Pakistan. Environ. Pollut. 216, 662e674. Kim, U.-J., Oh, J.K., Kannan, K., 2017. Occurrence, removal, and environmental emission of organophosphate flame retardants/plasticizers in a wastewater treatment plant in New York State. Environ. Sci. Technol. 51, 7872e7880. Kurt-Karakus, P., Alegria, H., Birgul, A., Gungormus, E., Jantunen, L., 2018. Organophosphate ester (OPEs) flame retardants and plasticizers in air and soil from a highly industrialized city in Turkey. Sci. Total Environ. 625, 555e565. Li, J., Dong, H., Li, X., Han, B., Zhu, C., Zhang, D., 2016. Quantitatively assessing the health risk of exposure to PAHs from intake of smoked meats. Ecotoxicol. Environ. Saf. 124, 91e95. Li, J., Tang, J., Mi, W., Tian, C., Emeis, K.-C., Ebinghaus, R., Xie, Z., 2018a. Spatial distribution and seasonal variation of organophosphate esters in air above the bohai and yellow seas, China. Environ. Sci. Technol. 52, 89e97. Li, J.F., Zhang, Z.Z., Ma, L.Y., Zhang, Y., Niu, Z.G., 2018b. Implementation of USEPA RfD and SFO for improved risk assessment of organophosphate esters

49

(organophosphate flame retardants and plasticizers). Environ. Int. 114, 21e26. Li, W., Shi, Y., Gao, L., Wu, C., Liu, J., Cai, Y., 2018c. Occurrence, distribution and risk of organophosphate esters in urban road dust in Beijing, China. Environ. Pollut. 241, 566e575. Liu, X.H., Liu, Y., Lu, S.Y., Guo, X.C., Lu, H.B., Qin, P., Bi, B., Wan, Z.F., Xi, B.D., Zhang, T.T., Liu, S.S., 2018. Occurrence of typical antibiotics and source analysis based on PCA-MLR model in the East Dongting Lake, China. Ecotoxicol. Environ. Saf. 163, 145e152. Liu, Y.Z., Ma, J.W., Yan, H.X., Ren, Y.Q., Wang, B.B., Lin, C.Y., Liu, X.T., 2016. Bioaccessibility and health risk assessment of arsenic in soil and indoor dust in rural and urban areas of Hubei province, China. Ecotoxicol. Environ. Saf. 126, 14e22. €ller, A., Sturm, R., Xie, Z., Cai, M., He, J., Ebinghaus, R., 2012. Organophosphorus Mo flame retardants and plasticizers in airborne particles over the northern pacific and Indian ocean toward the polar regions: Evidence for global occurrence. Environ. Sci. Technol. 46, 3127e3134. Marklund, A., Andersson, B., Haglund, P., 2003. Screening of organophosphorus compounds and their distribution in various indoor environments. Chemosphere 53, 1137e1146. Marklund, A., Barbro Andersson, A., Haglund, P., 2005. Traffic as a source of organophosphorus flame retardants and plasticizers in snow. Environ. Sci. Technol. 39, 3555e3562. Mcdonough, C., Silva, A.O.D., Sun, C., Cabrerizo, A., Adelman, D., Soltwedel, T., Bauerfeind, E., Muir, D.C.G., Lohmann, R., 2018. Dissolved organophosphate esters and polybrominated diphenyl ethers in remote marine environments: arctic surface water distributions and net transport through fram strait. Environ. Sci. Technol. 6208e6216. Mishra, N., Bartsch, J., Ayoko, G.A., Salthammer, T., Morawsk, L., 2015. Volatile Organic Compounds: characteristics, distribution and sources in urban schools. Atmos. Environ. 106, 485e491. Pang, L., Yuan, Y., He, H., Liang, K., Zhang, H., Zhao, J., 2016. Occurrence, distribution, and potential affecting factors of organophosphate flame retardants in sewage sludge of wastewater treatment plants in Henan Province, Central China. Chemosphere 152, 245e251. Rauert, C., Schuster, J.K., Eng, A., Harner, T., 2018. Global atmospheric concentrations of brominated and chlorinated flame retardants and organophosphate esters. Environ. Sci. Technol. 52, 2777e2789. Regnery, J., Püttmann, W., 2009. Organophosphorus flame retardants and plasticizers in rain and snow from Middle Germany. Clean. - Soil, Air, Water 37, 334e342. Ren, G., Chen, Z., Feng, J., Wen, J., Zhang, J., Zheng, K., Yu, Z., Zeng, X., 2016. Organophosphate esters in total suspended particulates of an urban city in East China. Chemosphere 164, 75e83. Salamova, A., Hermanson, M.H., Hites, R.A., 2014. Organophosphate and halogenated flame retardants in atmospheric particles from a European Arctic site. Environ. Sci. Technol. 48, 6133e6140. Shi, Y., Gao, L., Li, W., Wang, Y., Liu, J., Cai, Y., 2016. Occurrence, distribution and seasonal variation of organophosphate flame retardants and plasticizers in urban surface water in Beijing, China. Environ. Pollut. 209, 1e10. Sun, J., Wu, F.K., Hu, B., Tang, G.Q., Zhang, J.K., Wang, Y.S., 2016. VOC characteristics, emissions and contributions to SOA formation during hazy episodes. Atmos. Environ. 141, 560e570. van der Veen, I., de Boer, j., 2012. Phosphorus flame retardants: properties, production, environmental occurrence, toxicity and analysis. Chemosphere 88, 1119e1153.  Melymuka, L., Be Vykoukalov
a, M., Venier, M., Vojta, S., canov
aa, J., Romanakb, K.,
nov
Prokes, R., Okeme, J.O., Saini, A., Diamond, M.L., Kla a, J., 2017. Organophosphate esters flame retardants in the indoor environment. Environ. Int. 106, 97e104. Wan, W., Zhang, S., Huang, H., Wu, T., 2016. Occurrence and distribution of organophosphorus esters in soils and wheat plants in a plastic waste treatment area in China. Environ. Pollut. 214, 349e353. Wang, X.L., Zhu, L.Y., Zhong, W.J., Yang, L.P., 2018a. Partition and source identification of organophosphate esters in the water and sediment of Taihu Lake, China. J. Hazard Mater. 360, 43e50. Wang, Y., Sun, H., Zhu, H., Yao, Y., Chen, H., Ren, C., Wu, F., Kannan, K., 2018b. Occurrence and distribution of organophosphate flame retardants (OPFRs) in soil and outdoor settled dust from a multi-waste recycling area in China. Sci. Total Environ. 625, 1056e1064. 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, 29e46. WHO, 2000. Flame Retardants: Tris(2-Butoxyethyl) Phosphate, Tris(2-Ethylhexyl) Phosphate and Tetrakis(hydroxymethyl) Phosphonium Salts. Environmental Health Criteria 218. World Health Organization, Geneva. Wong, F., Suzuki, G., Michinaka, C., Yuan, B., Takigami, H., de Wit, C., 2017. Dioxinlike activities, halogenated flame retardants, organophosphate esters and chlorinated paraffins in dust from Australia, the United Kingdom, Canada, Sweden and China. Chemosphere 168, 1248e1256. Xing, L., Zhang, Q., Sun, X., Zhu, H., Zhang, S., Xu, H., 2018. Occurrence, distribution and risk assessment of organophosphate esters in surface water and sediment from a shallow freshwater Lake, China. Sci. Total Environ. 636, 632e640. Xu, H., Guinot, B., Cao, J., Li, Y., Niu, X., Ho, K.F., Shen, Z., Liu, S., Zhang, T., Lei, Y., 2018. Source, health risk and composition impact of outdoor very fine particles (VFPs) to school indoor environment in Xi'an, Northwestern China. Sci. Total Environ.

50

Y. Chen et al. / Chemosphere 231 (2019) 41e50

612, 238e246. Yadav, I.C., Devi, N.L., Zhong, G., Li, J., Zhang, G., Covaci, A., 2017. Occurrence and fate of organophosphate ester flame retardants and plasticizers in indoor air and dust of Nepal: Implication for human exposure. Environ. Pollut. 229, 668e678. Yang, Y., Liu, L., Xiong, Y., Zhang, G., Wen, H., Lei, J., Guo, L., Lyu, Y., 2016. A comparative study on physicochemical characteristics of household dust from a metropolitan city and a remote village in China. Atmos. Pollut. Res. 7, 1090e1100. Zeng, X., Liu, Z., He, L., Cao, S., Han, S., Yu, Z., Sheng, G., Fu, J., 2015. The occurrence and removal of organophosphate ester flame retardants/plasticizers in a

municipal wastewater treatment plant in the Pearl River Delta, China. Environ. Lett. 50, 1291e1297. Zhou, L., Hiltscher, M., Gruber, D., Püttmann, W., 2017. Organophosphate flame retardants (OPFRs) in indoor and outdoor air in the Rhine/Main area, Germany: comparison of concentrations and distribution profiles in different microenvironments. Environ. Sci. Pollut. Res. 24, 10992e11005. Zhu, N.Z., Liu, L.Y., Ma, W.L., Li, W.L., Song, W.W., Qi, H., Li, Y.F., 2015. Polybrominated diphenyl ethers (PBDEs) in the indoor dust in China: levels, spatial distribution and human exposure. Ecotoxicol. Environ. Saf. 111, 1e8.