Organophosphate di- and tri-esters in indoor and outdoor dust from China and its implications for human exposure

Journal Pre-proofs Organophosphate Di- and Tri-esters in Indoor and Outdoor Dust from China and its Implications for Human Exposure Yu Wang, Yiming Yao, Xiaoxin Han, Wenhui Li, Hongkai Zhu, Lei Wang, Hongwen Sun, Kurunthachalam Kannan PII: DOI: Reference:

S0048-9697(19)34493-6 https://doi.org/10.1016/j.scitotenv.2019.134502 STOTEN 134502

To appear in:

Science of the Total Environment

Received Date: Revised Date: Accepted Date:

25 July 2019 12 September 2019 15 September 2019

Please cite this article as: Y. Wang, Y. Yao, X. Han, W. Li, H. Zhu, L. Wang, H. Sun, K. Kannan, Organophosphate Di- and Tri-esters in Indoor and Outdoor Dust from China and its Implications for Human Exposure, Science of the Total Environment (2019), doi: https://doi.org/10.1016/j.scitotenv.2019.134502

This is a PDF file of an article that has undergone enhancements after acceptance, such as the addition of a cover page and metadata, and formatting for readability, but it is not yet the definitive version of record. This version will undergo additional copyediting, typesetting and review before it is published in its final form, but we are providing this version to give early visibility of the article. Please note that, during the production process, errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain.

© 2019 Elsevier B.V. All rights reserved.

Organophosphate Di- and Tri-esters in Indoor and Outdoor Dust from China and its Implications for Human Exposure

Yu Wanga,b, Yiming Yaoa, Xiaoxin Hana, Wenhui Lib,c, Hongkai Zhub, Lei Wanga, Hongwen Suna,* and Kurunthachalam Kannanb,d,* a

MOE Key Laboratory of Pollution Processes and Environmental Criteria, College of Environmental Science and Engineering, Nankai University, Tianjin 300350, China

b

Wadsworth Center, New York State Department of Health, Albany, New York 12201, United States c

Civil and Environment Engineering School, University of Science and Technology Beijing, Beijing 100083, China

d

Department of Environmental Health Sciences, School of Public Health, State University of New York at Albany, Albany, New York 12201, United States

Corresponding Authors: *Wadsworth Center, Empire State Plaza, P.O. Box 509, Albany, NY 12201, USA. Tel.: +1 518 474 0015; fax: +1 518 473 2895. E-mail: [email protected] (K. Kannan). *Nankai University, 38 Tongyan Road, Jinnan District, Tianjin 300350, China. Tel.: +86 22 23509241. E-mail: [email protected] (H. Sun).

For

submission

to:

Science

of

1

the

Total

Environment

1

Graphical abstract

2

2

3

Highlights

4

Organophosphate tri- and di-esters are ubiquitous in dust samples from China

5

TCIPP and DPHP are the predominant tri- and di-esters, respectively, in dust

6

TEHP and TPHP are the sources of their corresponding diesters in dust

7

Human exposure levels of DEP and DPHP through dust are notable

8 9

3

10

ABSTRACT

11

Organophosphate (OP) esters are emerging environmental contaminants, but little is known

12

about their occurrence in dust. In this study, 19 OP triesters and their 11 diester degradation

13

products were measured in indoor dust (ID) and outdoor dust (OD) collected from China. ∑OP

14

triester concentrations in ID (median: 2380 ng/g dry weight [dw]) were an order of magnitude

15

higher than those in OD (446 ng/g dw). The median concentrations of ∑OP diesters in ID and

16

OD were 260 and 96.8 ng/g dw, respectively. Dust samples collected from eastern and southern

17

China contained higher concentrations of ∑OP di- and tri-esters than those from the other

18

regions. Dust from the most urbanized areas in China including Beijing, Shanghai, and

19

Guangzhou exhibited the highest concentrations of ∑OP di- (>1000 ng/g dw) and triesters

20

(>4000 ng/g dw). We also found notable concentrations of emerging aryl-OP triesters in dust

21

(3.85-10.6 ng/g dw). Significant correlations existed between the concentrations of bis(2-

22

ethylhexyl) phosphate (BEHP) and tris(2-ethylhexyl) phosphate (TEHP) (rho = 0.672-0.691, p <

23

0.01), as well as DPHP and triphenyl phosphate (TPHP) (rho = 0.537-0.766, p < 0.01) in dust

24

samples, indicating that OP diesters originated from the degradation of triesters. High molar

25

concentration ratios of DEP to triethyl phosphate (TEP) and DPHP to TPHP/ethylhexyl diphenyl

26

phosphate (EHDPP) suggested that these OP triesters degrade in dust readily. Significant

27

correlations were found between the concentrations of ∑OP di- (R2 = 0.390, p < 0.05) and tri-

28

esters (R2 = 0.475, p < 0.01) in paired indoor-outdoor dust samples, which suggested that indoor

29

dust was the source of OP esters to the outdoor environment. The estimated daily intake (EDI) of

30

∑OP diesters through dust ingestion was 0.21 ng/kg bw/d for adults and 2.59 ng/kg bw/d for

31

children. The exposure levels of OP diesters, DEP and DPHP, were comparable to those of their

32

parent triester compounds.

4

33

Keywords: organophosphate triesters, OPFRs, flame retardants, dust, human exposure

5

34

1. Introduction

35

Organophosphate (OP) triesters are used in many commercial products, including PVC

36

plastics, electronics, and textiles (van der Veen et al., 2012; Wei et al., 2015). As alternatives to

37

brominated flame retardants (BFRs), the production of OP triesters has increased significantly in

38

recent years, with the global annual production estimated at 680,000 tons in 2015 (Wang et al.,

39

2015). Because OP triesters are used as additives in products, they can be emitted into the

40

environment through volatilization and leaching (van der Veen et al., 2012). OP triesters have

41

been detected in water (Kim et al., 2018; Kim et al., 2017; Shi et al., 2016), outdoor air (Castro-

42

Jimenez et al., 2014; Castro-Jimenez et al., 2016; Moller et al., 2011), indoor air (Tokumura et

43

al., 2017; Yadav et al., 2017), indoor dust (Hoffman et al., 2015; Kademoglou et al., 2017),

44

sediment (Tan et al., 2016; Zhong et al., 2018), and soil (Mihajlovic et al., 2011; Yadav et al.,

45

2018). Some OP triesters, especially chlorinated ones, have been reported to exhibit reproductive,

46

carcinogenic and neurotoxic effects (van der Veen et al., 2012; Wei et al., 2015).

47

OP triesters are readily metabolized in the human body to their corresponding diester

48

metabolites (Wang et al., 2019). The occurrence of OP diesters in foodstuffs (He et al., 2018a;

49

Wang and Kannan, 2018), human urine (Saillenfait et al., 2018), and chorionic villi (Zhao et al.,

50

2017) has been documented. It has been shown that some OP diesters were more toxic than their

51

parent triesters (Su et al., 2014). OP triesters can be hydrolyzed in water (Su et al., 2016), and the

52

degradation of non-chlorinated OP triesters in lake water has been documented (Regnery et al.,

53

2010). Furthermore, notable levels of OP diesters in sewage sludge suggested biodegradation of

54

triesters (Fu et al., 2017; Wang et al., 2018a). Therefore, it is reasonable to state that OP triesters

55

undergo degradation in the environment. OP diesters have been found in sewage sludge from

56

China and the USA, and degradation of triesters was suggested as a source of diesters in the

6

57

environment (Fu et al., 2017; Wang et al., 2018a). OP diesters, bis(2,4-di-tert-butylphenyl)

58

phosphate (BDtBPP) and diphenyl phosphate (DPHP) were found in indoor dust samples from

59

Canada, Spain, and the Netherlands (Bjornsdotter et al., 2018; Liu et al., 2018). Co-occurrence of

60

OP di- and triesters were reported in dust samples from southern China and the midwestern USA

61

(Tan et al., 2019). However, to the best of our knowledge, no earlier studies have measured OP

62

diesters in dust samples collected nationwide in China.

63

OP triesters have been reported to occur in indoor dust at remarkable concentrations in the

64

range of 1000-1000000 ng/g dw (He et al., 2015; Peng et al., 2017; Wu et al., 2016). Indoor dust

65

ingestion is considered an important pathway of human exposure to OP triesters (Cequier et al.,

66

2014; He et al., 2015; Xu et al., 2016; Zheng et al., 2015). High concentrations of OP triesters

67

have been reported in outdoor dust from electronic or plastic waste recycling areas in China

68

(Wan et al., 2016; Wang et al., 2018b). Furthermore, dust samples collected from China have

69

shown some of the highest concentrations of OP triesters (Ali et al., 2017). However, studies in

70

China were limited to localized regions like Beijing (Wu et al., 2016) and the Pearl River delta

71

(Tan et al., 2017; Zheng et al., 2015). There has been no previous study reporting the occurrence

72

and distribution of both OP triesters and diesters across China. Nationwide monitoring surveys of

73

OP tri- and di-esters in indoor and outdoor dust will be useful to assess sources, inventory,

74

transport and fate of these chemicals in the environment.

75

Recently, aryl-OP triesters such as cresyl diphenyl phosphate (CDPP), t-butylphenyl diphenyl

76

phosphate (BPDP) and tris(p-tert-butylphenyl) phosphate (TBPHP), which are used in hydraulic

77

fluids and in PVC plastics (WHO 1997), have received considerable attention (Bjornsdotter et al.,

78

2018; Tan et al., 2018). Similar to monomeric OP triesters, oligomeric OP triesters such as

79

resorcinol bis(diphenyl phosphate) (RDP) and bisphenol A bis(diphenyl phosphate) (BDP) are

7

80

used as flame retardants in plastic and electronic products (Ballesteros-Gomez et al., 2014), and

81

high concentrations of such oligomeric OP triesters have been measured in sewage sludge in

82

China (Liang et al., 2018). For a comprehensive investigation of the nationwide survey of OP

83

triesters and their diester degradation products, we analyzed 19 OP triesters and 11 OP diesters in

84

141 indoor and outdoor dust samples collected across China, with the objectives of elucidating

85

spatial distribution, correlations between parent OP esters and their degradation products, and

86

potential for degradation of OP triesters in dust. We also estimated human exposure to OP di-

87

and tri-esters through indoor and outdoor dust ingestion.

88

2. Materials and methods

89

2.1. Target chemicals

90

The four groups of OP triesters analyzed in this study were as follows: 7 alkyl-OP triesters:

91

trimethyl phosphate (TMP), triethyl phosphate (TEP), tripropyl phosphate (TPP), tri-n-butyl

92

phosphate (TNBP), tri-iso-butyl phosphate (TIBP), tris(2-butoxyethyl) phosphate (TBOEP),

93

tris(2-ethylhexyl) phosphate (TEHP); 3 Cl-OP triesters: tris(2-chloroethyl) phosphate (TCEP),

94

tris(2-chloroisopropyl) phosphate (TCIPP), tris(1,3-dichloro-2-propyl) phosphate (TDCIPP); 7

95

aryl-OP triesters: triphenyl phosphate (TPHP), trimethylphenyl phosphate (TMPP), 2-

96

ethylhexyl diphenyl phosphate (EHDPP), CDPP, isodecyl diphenyl phosphate (IDDP), BPDP,

97

TBPHP; and 2 oligomeric OP triesters: RDP and BDP. Because studies on the occurrence of

98

CDPP, IDDP, BPDP and TBPHP are scarce, they are regarded as emerging aryl-OP triesters.

99

Nine deuterated triester compounds, TMP-d9, TEP-d15, TPP-d21, TNBP-d27, TEHP-d51

100

TCEP-d12, TCIPP-d18, TDCIPP-d15, and TPHP-d15 were used as surrogate standards.

101

In addition to 19 triesters, three groups of OP diesters were included in the analysis: 6 alkyl-

102

OP diesters: diethyl phosphate (DEP), dipropyl phosphate (DPP), dibutyl phosphate (DNBP),

8

103

diisobutyl phosphate (DIBP), bis(butoxyethyl) phosphate (BBOEP), bis(2-ethylhexyl) phosphate

104

(BEHP); 3 Cl-OP diesters: bis(2-chloroethyl) phosphate (BCEP), bis(1-chloro-2-propyl)

105

phosphate (BCIPP), bis(1,3-dichloro-2-propyl) phosphate (BDCIPP); and 2 aryl-OP diesters:

106

DPHP, bis(methylphenyl) phosphate (BMPP). DNBP-d18, DIBP-d14, BBOEP-d8, BEHP-d34,

107

BCEP-d8, BCIPP-d12, BDCIPP-d10, DPHP-d10, and BMPP-d14 were used as surrogate

108

standards.

109

Information with regard to chemicals and reagents used in the analysis is presented in the

110

Supplementary Information (SI). Detailed information on each of the OP di- and tri-esters,

111

including physicochemical properties, is listed in Table S1.

112

2.2. Sample collection

113

Collection of indoor and outdoor dust samples was performed from March to August, 2017. In

114

total, 44 indoor dust and 97 outdoor dust were collected across the mainland China. There were

115

31 paired indoor/outdoor dust, which were collected in close proximity to each other. Samples

116

were collected from both urban and rural areas from each region except for those from

117

Northwestern and Southwestern China, where only urban samples were collected. Detailed

118

information on all sampling sites is presented in Table S2.

119

Indoor dust samples were collected in homes. A vacuum cleaner (Yangzi Co., Zhejiang, China)

120

was used to collect the dust from floors in homes. Dust was then transferred to a glass jar.

121

Outdoor dust samples were collected from outside of the building walls, window sills and

122

surfaces of other objects 1 m above the ground by sweeping them into a glass bottle with a

123

woolen brush.

124

A pre-cleaned vacuum cleaner was used for the collection of each dust sample. All the

125

sampling tools and glassware were rinsed with methanol prior to use. We also analyzed the tools

9

126

and glassware prior to use and found no detectable levels of both OP di- and triesters. All

127

samples were transported to the laboratory within 3 days of sampling, freeze-dried, homogenized,

128

sieved using a stainless-steel sieve (<2.0 mm), and stored at −20 °C until further analysis.

129

2.3. Sample preparation

130

Samples were extracted using an accelerated solvent extractor (ASE-200, DIONEX,

131

Sunnyvale, CA, USA) for the extraction of OP triesters as reported previously (Kim et al., 2017).

132

OP diesters in dust samples were extracted by mechanical oscillation, followed by

133

ultrasonication. Detailed information on sample preparation is presented in the SI.

134

2.4. Instrumental analysis

135

High-performance liquid chromatography (HPLC, Agilent 1100 series; Agilent Technologies,

136

Santa Clara, CA) coupled with electrospray triple quadrupole mass spectrometry (ESI-MS/MS,

137

API 2000; Applied Biosystems, Foster City, CA) was used in the identification and

138

quantification of target compounds. Detailed information on instrumental analysis is presented in

139

the SI and MS/MS parameters of the target compounds are presented in Tables S3 and S4.

140

2.5. Quality assurance and quality control

141

An 11-point calibration standard ranging in concentrations from 0.1 to 300 ng/mL was used in

142

the calculation of target chemicals’ concentrations in samples. The regression coefficients of

143

calibration curves were >0.99. The limits of detection (LODs) were set at a signal-to-noise ratio

144

(of the lowest point of the calibration standard) of 3 and the limits of quantitation (LOQs) at 10

145

(Table S5). Absolute recoveries of surrogate standards (44.0-86.1% for OP triesters; 52.2-99.7%

146

for OP diesters) are shown in Table S5. Quantification was by isotope dilution, which

147

compensated for the low recoveries observed for some OP compounds. All glassware and

148

polypropylene tubes were pre-cleaned using hexane, acetone, methanol and acetonitrile in

10

149

sequence, prior to use. Trace levels of TNBP, TCIPP, TEHP, TMPP, and IDDP (0.06-0.72 ng) as

150

well as DNBP, DIBP BEHP, and DPHP (0.01-0.04 ng) were found in procedural blanks. No OP

151

di- and tri-esters were found in travel and field blank samples. Recoveries of target chemicals in

152

spiked sample matrix were calculated as the difference in analyte’s concentration between pre-

153

spiked (before extraction) and post-spiked (before injection) samples (both after subtraction of

154

matrix blank). Average recoveries were 67−123% for OP triesters (n = 8) and 72−113% for OP

155

diesters (n = 8) (Table S5). A midpoint calibration standard was injected after every 20 samples

156

to monitor for drift in instrumental sensitivity.

157

2.6. Statistical analysis

158

All data outliers were checked by box-plot analysis using SPSS statistics ver 20. The outliers

159

were defined as data points with a value below Q1 - 1.5 IQR or above Q3 + 1.5 IQR with Q1 as

160

the 25th percentile, Q3 as the 75th percentile, and IQR as the interquartile range (IQR = Q3 - Q1).

161

The Spearman’s rank correlation analysis was performed using SPSS statistics ver 20. Prior to

162

correlation analysis, outliers and OP esters with low detection frequencies (< 50%) were

163

removed from the data set. Heatmaps were created by R software (The R Foundation, Vienna,

164

Austria) for the presentation of Spearman’s rank correlation matrix. Concentrations below the

165

method detection limits (MDLs) were assigned a value of 1/2 MDL. Surfer 11 (Golden Software,

166

Golden, CO, USA) was used for Kriging analysis. All the plots were created by Origin Pro 8.5

167

(Origin Lab Corporation, Northampton, MA, USA).

168

3. Results and discussion

169

3.1. Concentrations of OP triesters in indoor and outdoor dust

170

Eighteen of nineteen OP triesters (with the exception of TPP) were found in indoor dust

171

(detection frequency, DF=11-100%) and outdoor dust samples (DF=2-100%) (Table 1). TPP was

11

172

reported to occur rarely in water and dust, in previous studies (Shi et al., 2016; Tan et al., 2017),

173

and it was not detected in dust in our study. OP triester concentrations in indoor dust (median:

174

2380, range: 292-9540 ng/g dw) were 5-10 fold higher than those in outdoor dust (446, 99.7-

175

5960 ng/g dw) (Table 1).

176

OP triester concentrations in indoor dust from China were comparable to those reported from

177

Belgium (median; 4870 ng/g dw) (Van den Eede et al., 2011), New Zealand (5510 ng/g dw) (Ali

178

et al., 2012), and Canada (5870 ng/g dw) (Vykoukalova et al., 2017). However, indoor dust from

179

the USA (9820 ng/g dw) (Stapleton et al., 2009; Li et al., 2019a) and Japan (32490 ng/g dw)

180

(Tajima et al., 2014) contained much higher concentrations of OP triesters than those found in

181

China (Table S6). Recent studies (2017-2018) showed elevated OP triester concentrations in

182

indoor dust from Australia (40000 ng/g dw) (He et al., 2018b), Germany (14000 ng/g dw) (Zhou

183

et al., 2017), and the USA (18000 ng/g dw) (Vykoukalova et al., 2017; Li et al., 2019b), which

184

suggest extensive use of OP triesters in more developed countries. The concentrations of OP

185

triester in indoor dust were similar to those reported earlier for Chinese house dust (3120 ng/g

186

dw) (He et al., 2016) but lower than those reported for electronic waste (E-waste) workshops

187

(25000 ng/g dw) (He et al., 2015) and cars (92300 ng/g dw) (He et al., 2018c). The Cl-OP

188

triester, TCIPP (DF=100%, median: 690, range: 7.41-6730 ng/g dw), showed the highest

189

concentration, followed by TCEP (DF=100%, 239, 3.52-3600 ng/g dw) and TEHP (DF=98%,

190

147, 1.01-1510 ng/g dw), which is consistent with the patterns reported earlier (Xu et al., 2016;

191

Yang et al., 2014). Among the emerging aryl-OP triesters, CDPP (median: 10.6,
192

dw) and IDDP (3.85,
193

materials (van der Veen et al., 2012), presented notable concentrations.

12

194

OP triesters have been reported to undergo long-range atmospheric transport (Castro-Jimenez

195

et al., 2016; Moller et al., 2012; Rauert et al., 2018) and outdoor dust can sequester contaminants

196

and settle in soil through dry deposition (Sabin et al., 2006). However, data on the occurrence of

197

OP triesters in outdoor dust are limited. We found that outdoor dust contained TCIPP (85.8,

198

12.8-5050 ng/g dw) at notable concentrations. Similarly, TCEP, TDCIPP, and TNBP were

199

present at several tens of nanograms per gram (27.5-53.6 ng/g dw) (Table 1). The concentrations

200

of OP triesters in outdoor dust were comparable to those reported for street dust (702 ng/g dw)

201

(He et al., 2017), but lower than those reported from a waste recycling area in China (2690 ng/g

202

dw) (Wang et al., 2018b). The concentrations of most OP triesters decreased significantly

203

between indoor and outdoor dust.

204

Occurrence of aryl- and oligomeric OP triesters, including CDPP, IDDP, RDP and BDP, has

205

been rarely reported in the literature (Bjornsdotter et al., 2018; Christia et al., 2018; Kademoglou

206

et al., 2017). In China, aryl- and oligomeric OP triesters were reported to occur in sewage sludge

207

(CDPP: 3.80 ng/g dw, BDP: 2.06-5.82 ng/g dw, RDP: 0.44-3.45 ng/g dw) (Gao et al., 2016;

208

Liang et al., 2018). The concentrations reported in sewage sludge were comparable to those

209

found in indoor dust in our study. The emerging aryl- and oligomeric OP triesters were used in

210

hydraulic fluids and PVC plastics (CDPP, BPDP, TBPHP) (WHO 1997) as well as in electronic

211

products (RDP and BDP) (Ballesteros-Gomez et al., 2014). Moreover, the production of these

212

novel OP triesters are increasing as they are replacement for the traditional OP triesters (Tan et

213

al., 2018). In this study, the emerging aryl- and oligomeric OP triesters were found in dust

214

samples at notable concentrations. To the best of our knowledge, ours is the first report of

215

emerging aryl- and oligomeric OP triesters in indoor and outdoor dust samples collected across

216

China.

13

217

The OP triester concentration profile in dust was in the order of: Cl- > alkyl- > aryl- >

218

oligomeric OP triesters. TCIPP was the most abundant compound among the 19 OP triesters

219

analyzed in dust. The global production of Cl-OP triesters (TCIPP: 40000, TDCIPP: 8000

220

tons/year) was significantly higher than that of the other OP triesters (TBOEP: <6000, TEHP:

221

<5000 tons/year) (van der Veen et al., 2012). Cl-OP triesters also degrade more slowly in water

222

(Cristale et al., 2017; Regnery et al., 2010; Su et al., 2016) and during biological and physico-

223

chemical treatments (Liang et al., 2016) than aryl- and alkyl-OP triesters, which could explain

224

high concentrations of these compounds found in dust.

225

3.2. Concentrations of OP diesters in indoor and outdoor dust

226

The DF of OP diesters ranged from 27% to 100% in indoor dust and from 20% to 100% in

227

outdoor dust (Table 2). Cl-OP diesters exhibited low DF in both indoor and outdoor dust samples

228

(<43%). Median concentrations of OP diesters in indoor dust (260, range: 3.25-3000 ng/g dw)

229

were three times higher than those measured in outdoor dust (96.8, 8.24-2000 ng/g dw) (Table 2).

230

Cl-OP triesters were the predominant OP esters found in this study, whereas their degradation

231

products (i.e., diesters) were detected at low concentrations in both indoor and outdoor dust.

232

DPHP and DEP were the most abundant OP diesters found at median concentrations of 47.5

233

(0.33-2810) and 38.9 (
234

(
235

Low concentrations of Cl-OP diesters found in this study support the notion that Cl-OP

236

triesters are not easily degraded in the atmospheric environment, as has been reported for these

237

chemicals in the aquatic environment (Cristale et al., 2017; Regnery et al., 2010; Su et al., 2016).

238

The predominance of DEP and DPHP in dust indicates degradability of the short-chain triester

239

TEP and aryl-OP triester TPHP in dust. In addition to the formation from the degradation of

14

240

triesters, BEHP and DNBP can be emitted directly from consumer products (Lemire et al., 1986;

241

Quintana et al., 2006), which could explain the higher concentrations of BEHP and DNBP found

242

in dust samples.

243

Thus far, the data on the occurrence of OP diester in environmental matrices is limited. OP

244

diesters were reported to occur in sewage sludge samples from China (Fu et al., 2017) and the

245

USA (Wang et al., 2018a) at median concentrations of 82.9 (17.0-1300) and 78.4 (22.9-1990)

246

ng/g dw, respectively, which were comparable to those found in our dust samples. DPHP was

247

measured in indoor dust from Spain, the Netherlands, southern China, and midwestern USA

248

(31.2–79,661 ng/g dw; range) (Bjornsdotter et al., 2018; Tan et al., 2019), at concentrations

249

higher than those found in our study, which suggested regional difference in concentrations. In

250

general, the occurrence of OP diesters in dust suggests that triesters are degraded in the

251

environment, and that contributes to human exposure to OP diesters.

252

3.3. Spatial distribution of OP di- and tri-esters in dust in China

253

Geographically, China is divided into six administrative divisions as North China (NC),

254

Northeast China (NE), East China (EC), South Central China (SC), Southwest China (SW), and

255

Northwest China (NW) (Figure S1). Kriging interpolation analysis was used to present spatial

256

distribution of OP di- and tri-esters in outdoor dust samples collected across China (Figure 1).

257

OP di- and tri-ester concentrations in indoor dust were similar among EC, NC, SC, and NE,

258

whereas the concentrations found in NW and SW regions were lower than those in the other

259

regions. However, the overall differences in OP di- and triester concentrations in indoor dust

260

between more developed (EC, SC) and remote regions (NW, SW) of China were modest (Figure

261

1, Tables S7 and S8). For outdoor dust, ∑OP di- and tri-ester concentrations in NE, EC and SC

262

were higher than those in NW and SW China (Figure 1). Some dust samples collected from areas

15

263

located along the east coast of EC and the most urbanized areas including Beijing in NC,

264

Shanghai in EC, and Guangzhou in SC showed the highest concentrations (∑OP triesters: >4000

265

ng/g dw; ∑OP diesters: >1000 ng/g dw). On the contrary, ∑OP di- and tri-ester concentrations in

266

remote areas of SW China were the lowest (∑OP triester: <300 ng/g dw; ∑OP diester: <50 ng/g

267

dw).

268

There was no significant difference in the distribution profile of OP di- and tri-esters among

269

the six divisions (Figure S2). The geographic and socio-economic differences among the six

270

divisions in China were not associated with OP di- and tri-ester profiles. Cl-OP triesters

271

accounted for 66.3±11.1%, and 51.4±5.4% of ∑OP triester concentrations whereas BEHP, DPHP,

272

and DEP collectively accounted for 78.5±8.2%, and 81.2±3.8% of ∑OP diester concentrations in

273

indoor and outdoor dust samples, respectively, for all six geographic divisions (Figure S3).

274

Between indoor and outdoor dust samples, OP diester profiles were similar. However, TCIPP

275

accounted for a higher proportion of ∑OP triester concentrations in indoor dust (ID: 42.8±18.4%)

276

than that in outdoor dust (OD: 20.1±9.0%).

277

The concentrations of OP esters measured in indoor and outdoor dust samples from each

278

region were further divided into urban and rural areas. The median concentrations of ∑OP

279

triester in urban indoor (1600 ng/g dw) and outdoor dust samples (288 ng/g dw) were

280

comparable to those in rural samples (indoor dust, 921 ng/g dw; outdoor dust, 301 ng/g dw)

281

(Figure S4). Similarly, no significant difference (p > 0.05) was found in the concentrations of OP

282

diesters between dust samples from urban and rural areas. The OP ester composition profiles

283

were similar between rural and urban areas (Figure S4). The major OP diester compounds,

284

BEHP, DPHP, and DEP accounted for 70~80% of ∑OP diester concentrations found in dust

285

samples from rural and urban areas. TCIPP and TCEP were the dominant OP triesters,

16

286

accounting for 65% of the total concentrations in indoor dust and 45% in outdoor dust from rural

287

and urban areas.

288

3.4. Correlation between OP di- and tri-esters in dust

289

Significant correlations were found between BEHP and TEHP (rho = 0.672, p < 0.01), DPHP

290

and TPHP (rho = 0.537, p < 0.01), and BMPP and TMPP (rho = 0.614, p < 0.01) in indoor dust

291

samples (Figure 2, Table S9). Similarly, BEHP and DPHP were significantly correlated with

292

their triester parent compounds, TEHP (rho = 0.691, p < 0.01) and TPHP (rho = 0.766, p <

293

0.01)/EHDPP (rho = 0.613, p < 0.01) in outdoor dust samples. Thus, BEHP and DPHP in dust

294

samples are thought to originate from the degradation of corresponding parent OP triesters.

295

Furthermore, DPHP showed significant correlations with CDPP and IDDP, suggesting that these

296

aryl-OPs could also degrade to DPHP in dust. Among other OP di- and tri-esters, Spearman’s

297

rank correlations were generally weak but significant, which suggested that these diesters

298

originate predominantly from the degradation of triesters and also from the direct environmental

299

releases. For example, degradation of TNBP may not be the only source of DNBP in the outdoor

300

environment (rho = 0.122, p > 0.05), because DNBP is also used as plasticizer and metal

301

extractant (Quintana et al., 2006). It is known that BEHP and DNBP can be emitted directly from

302

consumer products (Lemire et al., 1986; Quintana et al., 2006), whereas information on the

303

industrial and commercial use of other OP diesters is limited. The results of our study suggest

304

that the sources of BEHP and DPHP in dust were degradation of corresponding triesters as well

305

as direct releases.

306

Aryl-OP triesters, including TPHP, TMPP, EHDPP, CDPP, and IDDP showed significant

307

correlations (rho = 0.362-0.695, p < 0.05) with each other in indoor dust samples (Table S10),

308

including a strong correlation (rho = 0.695, p < 0.01) between CDPP and TMPP. TPHP and

17

309

CDPP are considered as demethylated products of TMPP, with similar chemical properties and

310

usage patterns (hydraulic fluids, PVC plastics), which would explain significant correlations

311

found among these three aryl-OP triesters. TPHP has been found as an impurity in EHDPP

312

mixtures (Ballesteros-Gomez et al., 2015), which might explain significant correlation between

313

them in indoor and outdoor dust samples (rho = 0.602-0.693, p < 0.01) (Tables S10 and S11).

314

Samples collected in the vicinity of each other or at adjacent sites were considered as paired

315

samples (Table S2) for the correlation analysis of OP di- and triester concentrations between

316

indoor and outdoor dust samples. There were significant correlations between the concentrations

317

of ∑OP triesters (R2 = 0.475, p < 0.01) and diesters (R2 = 0.390, p < 0.05) in paired samples of

318

indoor-outdoor dust (Figure 3) in linear fitting analysis. Similar significant correlations were

319

observed for major OP triesters (TCIPP: R2 = 0.574, p < 0.01; TEHP: R2 = 0.395, p < 0.05) and

320

OP diesters (DPHP: R2 = 0.374, p < 0.05; DEP: R2 = 0.320, p < 0.05) in paired samples of

321

indoor-outdoor dust (Figure S5). These correlations between indoor and outdoor dust suggest the

322

dispersion of OP esters via dust from indoor to the outdoor environment.

323

Molar concentration ratios of Cl-OP diesters to their parent Cl-OP triesters were close to zero

324

(Figure 4), which suggest limited degradation of Cl-OP triesters in dust. On the contrary, molar

325

concentration ratios of DEP to TEP, DNBP to TNBP, and DPHP to TPHP+EHDPP were close to

326

or higher than one, which suggested higher degradability of precursors of DEP, DIBP and DPHP

327

in both indoor and outdoor dust samples. Further, the high molar concentration ratios of DNBP

328

to TNBP and BEHP to TEHP can be ascribed to environmental degradation of parent triesters

329

(Lemire et al., 1986; Quintana et al., 2006). In general, short-chain OP triester (TEP) and aryl-

330

OP triester (TPHP) are thought to be easily degraded in dust. Cl-OP triesters are more persistent

331

than the other OP esters.

18

332

The reported median concentration ratios of DPHP and BEHP to their parent triesters (TPHP

333

and TEHP) were 1.1 and 1.0 in dust from southern China, respectively (Tan et al., 2019), which

334

were similar to those found in our study. Moreover, direct uses and the degradation of OP

335

triesters were suggested as sources of OP diesters (Tan et al., 2019). Future studies on the

336

application of OP diesters and the degradation pathways of OP triesters are needed to elucidate

337

the sources of diesters in dust.

338

3.5. Human exposure to OP di- and tri-esters via dust

339

Indoor dust ingestion is an important pathway for human exposure to OP esters. Dermal

340

absorption and inhalation of dust can contribute to OP esters exposure (Hou et al., 2016; Wei et

341

al., 2015). We estimated the daily intake (EDI) of OP esters through indoor and outdoor dust

342

dermal contact, inhalation and ingestion as shown in eqn 1-4:

343 344

𝐸𝐷𝐼 = 𝐷𝐼𝑑𝑒𝑟 +𝐷𝐼𝑖𝑛ℎ +𝐷𝐼𝑖𝑛𝑔 Dermal contact (DIder):

345 346

𝐷𝐼𝑑𝑒𝑟 =

349

𝐶𝑑 × 𝑆𝐴 × 𝐴𝐵𝐹 × 𝐴𝐹 × 𝐸𝐹 × 𝐶𝐹 𝐵𝑊

(2)

Inhalation (DIinh):

347 348

(1)

𝐷𝐼𝑖𝑛ℎ =

𝐶𝑑 × 𝑅𝑖𝑛ℎ × 𝐸𝐹 𝐵𝑊 × 𝑅𝑃𝐸

(3)

Ingestion (DIing): 𝐷𝐼𝑖𝑛𝑔 =

𝐶𝑑 × 𝑅𝑖𝑛𝑔 × 𝐸𝐹 × 𝐶𝐹 𝐵𝑊

(4)

350

where, Cd is the concentration of OP esters measured in dust (μg/kg); SA is the skin contact

351

surface area (cm2); AF is the skin adherence factor for dust (mg/cm2); ABF is the dermal

352

absorption factor; EF is the exposure frequency (minutes/day); CF is the conversion factor

19

353

(1×10−6 kg/mg); BW is the average body weight (kg); Ring is the ingestion rate (mg/day); Rinh is

354

the inhalation rate (m3/day); RPE is the particle emission factor (m3/kg).

355

The exposure parameters were selected by following the recommendations of the U.S.

356

Environmental Protection Agency (EPA) (United States Environmental Protection Agency 2011).

357

The median concentrations for an average exposure scenario and the 95th percentile for a high

358

exposure scenario were considered in this calculation. Detailed information used for each

359

parameter in the exposure analysis is listed in Table S12.

360

Cl-OP triesters showed the highest exposure doses, among all OP triesters analyzed. The EDI

361

values of Cl-OP triesters were 1.20 ng/kg bw/day for adults and 15 ng/kg bw/day for children via

362

indoor dust, and 7.34E-2 ng/kg bw/day for adults and 0.40 ng/kg bw/day for children via outdoor

363

dust (Table 3). The total OP triester intake estimates for adults via indoor dust were 1.93 ng/kg

364

bw/d, whereas those for children were 24.0 ng/kg bw/d. Dust ingestion was the dominant

365

pathway of human exposure, accounting for over 98% of OP triesters exposure via dust. Notably,

366

the calculated EDI values from outdoor environment were less than 5% of the indoor values. The

367

reference doses (RfD) of nine OP triesters (Table S13) have been reported in a previous study (Li

368

et al., 2018). The EDIs of OP triesters calculated in this study were much lower than the RfD

369

values by 4-6 orders of magnitude, which indicated a low risk of these chemicals to human

370

health at the current levels of exposure through dust.

371

The average EDI values estimated in this study for OP triesters (adults: 0.55, children: 6.84

372

ng/kg bw/d) were comparable to those reported in Pakistan (adults: 0.18, children: 2.70 ng/kg

373

bw/d) (Ali et al., 2013) and the UK (adults: 0.32, children: 1.91 ng/kg bw/d) (Kademoglou et al.,

374

2017), but much lower than those reported from Belgium (adults: 2.8, children: 32.0 ng/kg bw/d)

375

(Van den Eede et al., 2011) and southern China (adults: 9.45, children: 23.0 ng/kg bw/d) (Zheng

20

376

et al., 2015). These regional differences in dust ingestion exposure might be related to

377

differences in the usage of OP esters and regulations on the use of OP flame retardants among

378

these countries (Hou et al., 2016). Furthermore, the EDI of OP triesters for Chinese adults via

379

dust intake was similar to that reported from the inhalation of atmospheric particulate matter

380

(adults: 0.85 ng/kg bw/d) (Yang et al., 2014) and ingestion of water (adults: 0.22-1.25 ng/kg

381

bw/d) (Kim et al., 2018), and accounted for 10% of the dietary intake (adults: 5.45 ng/kg bw/d)

382

(Kim et al., 2011), 37% of the intake from hand-to-mouth contact (adults: 1.50 ng/kg/bw/d) (Tan

383

et al., 2018), suggesting that dust ingestion was an important exposure pathway for OP triesters

384

in humans.

385

OP diesters, the degradation products of OP triesters, were also detectable in dust samples in

386

this study, and their respective EDI values via dust were 0.21 and 2.59 ng/kg bw/d for adults and

387

children through indoor dust, and 3.19E-2 and 0.17 ng/kg bw/d for adults and children through

388

outdoor dust (Table 3). The EDI of alkyl-OP diesters and aryl-OP diesters were 10-fold higher

389

than those of Cl-OP diesters through dust. Human exposure to Cl-OP triesters was significantly

390

higher than that of Cl-OP diesters via indoor dust ingestion whereas exposure to DEP, DIBP, and

391

DPHP was comparable to that of their parent OP triesters, TEP, TIBP, and TPHP (Table 3).

392

Therefore, indoor dust is a significant pathway of human exposure to OP diesters, especially to

393

DEP, DIBP, and DPHP.

394

A previous study reported that foodstuff was a source of OP diester intake in Australia and the

395

EDI of DPHP from diet was higher (mean of 71 ng/kg bw/d) than that of its parent OP triester,

396

TPHP (He et al., 2018a). Ingestion of DPHP through indoor dust in Spain at doses as high as

397

0.96 ng/kg bw/d was reported for toddlers (Bjornsdotter et al., 2018). The former study reported

398

EDI of seven OP diesters for toddlers with a median value of 8.4 ng/kg bw/d in southern China,

21

399

and 54.4 ng/kg bw/d in midwestern USA, respectively (Tan et al., 2019). Overall, we show that

400

dust intake is another source of exposure to OP diesters.

401

4. Conclusions

402

In summary, OP triesters and diesters were ubiquitously detected in indoor dust and outdoor

403

dust samples from China. Cl-OP triesters were found at notable concentrations in comparison to

404

other OP triesters, suggesting that these halogenated flame retardants require further studies to

405

understand their future trends and potential health risks. OP diesters were found in dust samples,

406

which suggested that they originate from the degradation of triesters. Aryl and alkyl- triesters,

407

TEP, TPHP/EHDPP appear to be more easily degraded in dust than Cl-OP triesters. A positive

408

correlation between indoor-outdoor pairs of dust samples suggested the migration of OP esters

409

from indoor environment to outdoor environment. Indoor dust ingestion was found to be an

410

important pathway of human exposure to OP diesters.

411 412

Corresponding Author

413

*Wadsworth Center, Empire State Plaza, P.O. Box 509, Albany, NY 12201, USA. Tel.: +1

414

518 474 0015; fax: +1 518 473 2895. E-mail address: [email protected] (K.

415

Kannan).

416

*Nankai University, 38 Tongyan Road, Jinnan District, Tianjin 300350, China. Tel.: +86 22

417

23509241. E-mail: [email protected] (H. Sun).

418

Notes

419 420

The authors declare no competing financial interest. Acknowledgments

22

421

This study was supported in part by National Natural Science Foundation of China (Nos.

422

41573097; 41603101; 21777075), Ministry of Education of China (No. T2017002, IRT13024),

423

973 program (2014CB441104). All samples were analyzed at Wadsworth Center.

424

Appendix A. Supplementary data

425

Sections on chemicals, reagents, and sample preparation; Tables including summary of sample

426

information, instrumental analysis information, concentrations of OP triester and diester in

427

indoor dust, and outdoor dust samples from six geographical divisions of China, Spearman’s

428

rank correlation coefficient analysis; Figures: composition profiles and linear correlations of OP

429

triester and diester in indoor dust and outdoor dust samples.

23

References Ali, N., Ali, L., Mehdi, T., Dirtu, A.C., Al-Shammari, F., Neels, H., Covaci, A., 2013. Levels and profiles of organochlorines and flame retardants in car and house dust from Kuwait and Pakistan: implication for human exposure via dust ingestion. Environ. Int. 55, 62-70 Ali, N., Dirtu, A.C., Van den Eede, N., 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., Shahzad, K., Rashid, M.I., Shen, H., Ismail, I.M.I., Eqani, S., 2017. Currently used organophosphate and brominated flame retardants in the environment of China and other developing countries (2000-2016). Environ. Sci. Pollut. Res. 24, 18721-18741 Ballesteros-Gomez, A., Brandsma, S.H., de Boer, J., Leonards, P.E., 2014. Analysis of two alternative organophosphorus flame retardants in electronic and plastic consumer products: resorcinol bis-(diphenylphosphate) (PBDPP) and bisphenol A bis (diphenylphosphate) (BPABDPP). Chemosphere 116, 10-14 Ballesteros-Gomez, A., Erratico, C.A., Eede, N.V., Ionas, A.C., Leonards, P.E., Covaci, A., 2015. In vitro metabolism of 2-ethylhexyldiphenyl phosphate (EHDPHP) by human liver microsomes. Toxicol. Lett. 232, 203-212 Bjornsdotter, M.K., Romera-Garcia, E., Borrull, J., de Boer, J., Rubio, S., Ballesteros-Gomez, A., 2018. Presence of diphenyl phosphate and aryl-phosphate flame retardants in indoor dust from different microenvironments in Spain and the Netherlands and estimation of human exposure. Environ. Int. 112, 59-67 Castro-Jimenez, J., Berrojalbiz, N., Pizarro, M., Dachs, J., 2014. Organophosphate ester (OPE) flame retardants and plasticizers in the open Mediterranean and Black Seas atmosphere. Environ. Sci. Technol. 48, 3203-3209 Castro-Jimenez, J., Gonzalez-Gaya, B., Pizarro, M., Casal, P., Pizarro-Alvarez, C., Dachs, J., 2016. Organophosphate ester flame retardants and plasticizers in the global oceanic atmosphere. Environ. Sci. Technol. 50, 12831-12839 Cequier, E., Ionas, A.C., Covaci, A., Marce, 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 Christia, C., Poma, G., Besis, A., Samara, C., Covaci, A., 2018. Legacy and emerging organophosphomicronrus flame retardants in car dust from Greece: Implications for human exposure. Chemosphere 196, 231-239 Cristale, J., Dantas, R.F., De Luca, A., Sans, C., Esplugas, S., Lacorte, S., 2017. Role of oxygen and DOM in sunlight induced photodegradation of organophosphorous flame retardants in river water. J. Hazard. Mater. 323, 242-249 Fu, L., Du, B., Wang, F., Lam, J.C.W., Zeng, L., Zeng, E.Y., 2017. Organophosphate triesters and diester degradation products in municipal sludge from wastewater treatment plants in China: Spatial patterns and ecological implications. Environ. Sci. Technol. 13614-13623 Gao, L., Shi, Y., Li, W., Liu, J., Cai, Y., 2016. Occurrence and distribution of organophosphate triesters and diesters in sludge from sewage treatment plants of Beijing, China. Sci. Total Environ. 544, 143-149

24

He, C., Wang, X., Tang, S., Thai, P., Li, Z., Baduel, C., Mueller, J.F., 2018a. Concentrations of organophosphate esters and their specific metabolites in food in southeast Queensland, Australia: Is eietary exposure an important pathway of organophosphate esters and their metabolites? Environ. Sci. Technol. 52, 12765-12773 He, C., Wang, X., Thai, P., Baduel, C., Gallen, C., Banks, A., Bainton, P., English, K., Mueller, J.F., 2018b. Organophosphate and brominated flame retardants in Australian indoor environments: Levels, sources, and preliminary assessment of human exposure. Environ. Pollut. 235, 670679 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 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, 349-361 He, R., Li, Y., Xiang, P., Li, C., Zhou, C., Zhang, S., Cui, X., Ma, L.Q., 2016. Organophosphorus flame retardants and phthalate esters in indoor dust from different microenvironments: Bioaccessibility and risk assessment. Chemosphere 150, 528-535 He, R.W., Li, Y.Z., Xiang, P., Li, C., Cui, X.Y., Ma, L.Q., 2018c. Impact of particle size on distribution and human exposure of flame retardants in indoor dust. Environ. Res. 162, 166172 Hoffman, K., Garantziotis, S., Birnbaum, L.S., Stapleton, H.M., 2015. Monitoring indoor exposure to organophosphate flame retardants: hand wipes and house dust. Environ. Health Perspect. 123, 160-165 Hou, R., Xu, Y., Wang, Z., 2016. Review of OPFRs in animals and humans: Absorption, bioaccumulation, metabolism, and internal exposure research. Chemosphere 153, 78-90 Kademoglou, K., Xu, F., Padilla-Sanchez, J.A., Haug, L.S., Covaci, A., Collins, C.D., 2017. Legacy and alternative flame retardants in Norwegian and UK indoor environment: Implications of human exposure via dust ingestion. Environ. Int. 102, 48-56 Kim, J.W., Isobe, T., Chang, K.H., Amano, A., Maneja, R.H., Zamora, P.B., Siringan, F.P., Tanabe, S., 2011. Levels and distribution of organophosphorus flame retardants and plasticizers in fishes from Manila Bay, the Philippines. Environ. Pollut. 159, 3653-3659 Kim, U.J., Kannan, K., 2018. Occurrence and distribution of organophosphate flame retardants/plasticizers in surface waters, tap water, and rainwater: Implications for human exposure. Environ. Sci. Technol. 52, 5625-5633 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, 7872-7880 Lemire, A.E., Janzen, A.F., Marat, K., 1986. A 31P and 15N NMR study of the synergistic extraction of uranyl nitrate by tributylphosphate and Di-2-ethylhexyl phosphoric acid. Inorg. Chim. Acta 114, 211-214 Li, J., Zhang, Z., Ma, L., Zhang, Y., Niu, Z., 2018. Implementation of USEPA RfD and SFO for improved risk assessment of organophosphate esters (organophosphate flame retardants and plasticizers). Environ. Int. 114, 21-26 Li, W., Wang, Y., Asimakopoulos, A.G., Covaci, A., Gevao, B., Johnson-Restrepo, B., Kumosani, T.A., Malarvannan, G., Moon, H-B., Nakata, H., Sinha, R.K., Tran, T.M., Kannan, K., 2019a. 25

Organophosphate esters in indoor dust from 12 countries: Concentrations, composition profiles, and human exposure. Environ. Int. (in press). Li, W., Wang, Y. and Kannan, K., 2019b. Occurrence, distribution and human exposure to 20 organophosphate esters in air, soil, pine needles, river water, and dust samples collected around an airport in New York state, United States. Environ. Int. 131, 105054. Liang, K., Liu, J., 2016. Understanding the distribution, degradation and fate of organophosphate esters in an advanced municipal sewage treatment plant based on mass flow and mass balance analysis. Sci. Total Environ. 544, 262-270 Liang, K., Shi, F., Liu, J., 2018. Occurrence and distribution of oligomeric organophosphorus flame retardants in different treatment stages of a sewage treatment plant. Environ. Pollut. 232, 229235 Liu, R., Mabury, S.A., 2018. Unexpectedly High Concentrations of a Newly Identified Organophosphate Ester, Tris(2,4-di- tert-butylphenyl) Phosphate, in Indoor Dust from Canada. Environ. Sci. Technol. 52, 9677-9683 Mihajlovic, I., Miloradov, M.V., Fries, E., 2011. Application of Twisselmann extraction, SPME, and GC-MS to assess input sources for organophosphate esters into soil. Environ. Sci. Technol. 45, 2264-2269 Moller, A., Sturm, R., Xie, Z., Cai, M., He, J., Ebinghaus, R., 2012. Organophosphorus 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, 31273134 Moller, A., Xie, Z., Caba, A., Sturm, R., Ebinghaus, R., 2011. Organophosphorus flame retardants and plasticizers in the atmosphere of the North Sea. Environ. Pollut. 159, 3660-3665 Peng, C., Tan, H., Guo, Y., Wu, Y., Chen, D., 2017. Emerging and legacy flame retardants in indoor dust from East China. Chemosphere 186, 635-643 Quintana, J.B., Rodil, R., Reemtsma, T., 2006. Determination of phosphoric acid mono- and diesters in municipal wastewater by solid-phase extraction and ion-pair liquid chromatographytandem mass spectrometry. Anal. Chem. 78, 1644-1650 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, 2777-2789 Regnery, J., Puttmann, W., 2010. Occurrence and fate of organophosphorus flame retardants and plasticizers in urban and remote surface waters in Germany. Water Res. 44, 4097-4104 Sabin, L.D., Lim, J.H., Venezia, M.T., Winer, A.M., Schiff, K.C., Stolzenbach, K.D., 2006. Dry deposition and resuspension of particle-associated metals near a freeway in Los Angeles. Atmos. Environ. 40, 7528-7538 Saillenfait, A.M., Ndaw, S., Robert, A., Sabate, J.P., 2018. Recent biomonitoring reports on phosphate ester flame retardants: a short review. Arch. Toxicol. 92, 2749-2778 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, 1-10 Stapleton, H.M., Klosterhaus, S., Eagle, S., Fuh, J., Meeker, J.D., Blum, A., Webster, T.F., 2009. Detection of organophosphate flame retardants in furniture foam and US house dust. Environ. Sci. Technol. 43, 7490-7495

26

Su, G., Crump, D., Letcher, R.J., Kennedy, S.W., 2014. Rapid in vitro metabolism of the flame retardant triphenyl phosphate and effects on cytotoxicity and mRNA expression in chicken embryonic hepatocytes. Environ. Sci. Technol. 48, 13511-13519 Su, G.Y., Letcher, R.J., Yu, H.X., 2016. Organophosphate flame retardants and plasticizers in aqueous solution: pH-dependent hydrolysis, kinetics, and pathways. Environ. Sci. Technol. 50, 8103-8111 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 Tan, H., Chen, D., Peng, C., Liu, X., Wu, Y., Li, X., Du, R., Wang, B., Guo, Y., Zeng, E.Y., 2018. Novel and traditional organophosphate esters in house dust from South China: Aassociation with hand wipes and exposure estimation. Environ. Sci. Technol. 52, 11017-11026 Tan, H., Peng, C., Guo, Y., Wang, X., Wu, Y., Chen, D., 2017. Organophosphate flame retardants in house dust from South China and related human exposure risks. Bull. Environ. Contam. Toxicol. 99, 344-349 Tan, H., Yang, L., Yu, Y., Guan, Q., Liu, X., Li, L., Chen, D., 2019. Co-Existence of Organophosphate Di- and Tri-Esters in House Dust from South China and Midwestern United States: Implications for Human Exposure. Environ. Sci. Technol. 53, 4784-4793 Tan, X.X., Luo, X.J., Zheng, X.B., Li, Z.R., Sun, R.X., Mai, B.X., 2016. Distribution of organophosphorus flame retardants in sediments from the Pearl River Delta in South China. Sci. Total Environ. 544, 77-84 Tokumura, M., Hatayama, R., Tatsu, K., Naito, T., Takeda, T., Raknuzzaman, M., Al-Mamun, M.H., Masunaga, S., 2017. Organophosphate flame retardants in the indoor air and dust in cars in Japan. Environ. Monit. Assess. 189, 48 United States Environmental Protection Agency. 2011. Exposure factors handbook: 2011 Edition. Washington, DC. 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 van der Veen, I., de Boer, J., 2012. Phosphorus flame retardants: properties, production, environmental occurrence, toxicity and analysis. Chemosphere 88, 1119-1153 Vykoukalova, M., Venier, M., Vojta, S., Melymuk, L., Becanova, J., Romanak, K., Prokes, R., Okeme, J.O., Saini, A., Diamond, M.L., Klanova, J., 2017. Organophosphate esters flame retardants in the indoor environment. Environ. Int. 106, 97-104 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, 349-353 Wang, R., Tang, J., Xie, Z., Mi, W., Chen, Y., Wolschke, H., Tian, C., Pan, X., Luo, Y., Ebinghaus, R., 2015. Occurrence and spatial distribution of organophosphate ester flame retardants and plasticizers in 40 rivers draining into the Bohai Sea, north China. Environ. Pollut. 198, 172178 Wang, Y., Kannan, K., 2018. Concentrations and dietary exposure to organophosphate esters in foodstuffs from Albany, New York, United States. J. Agric. Food Chem. 66, 13525-13532. Wang, Y., Kannan, P., Halden, R.U., Kannan, K., 2018a. A nationwide survey of 31 organophosphate esters in sewage sludge from the United States. Sci. Total Environ. 655, 446-453 27

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, 1056-1064 Wang, Y., Li, W., Martínez-Moral, M-P., Sun, H., Kannan, K., 2019. Metabolites of organophosphate esters in urine from the United States: Concentrations, temporal variability, and exposure assessment. Environ. Intern. 122, 213-221 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 WHO. 1997. Flame Retardants: A General Introduction. World Health Organization. Geneva, Switzerland. http://www.inchem.org/documents/ehc/ehc/ehc192.htm Wu, M., Yu, G., Cao, Z., Wu, D., Liu, K., Deng, S., Huang, J., Wang, B., Wang, Y., 2016. Characterization and human exposure assessment of organophosphate flame retardants in indoor dust from several microenvironments of Beijing, China. Chemosphere 150, 465-471 Xu, F., Giovanoulis, G., van Waes, S., Padilla-Sanchez, J.A., Papadopoulou, E., Magner, J., Haug, L.S., Neels, H., Covaci, A., 2016. Comprehensive study of human external exposure to organophosphate flame retardants via air, dust, and hand wipes: The importance of sampling and assessment strategy. Environ. Sci. Technol. 50, 7752-7760 Yadav, I.C., Devi, N.L., Li, J., Zhang, G., 2018. Organophosphate ester flame retardants in Nepalese soil: Spatial distribution, source apportionment and air-soil exchange assessment. Chemosphere 190, 114-123 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, 668-678 Yang, F., Ding, J., Huang, W., Xie, W., Liu, W., 2014. Particle size-specific distributions and preliminary exposure assessments of organophosphate flame retardants in office air particulate matter. Environ. Sci. Technol. 48, 63-70 Zhao, F., Chen, M., Gao, F., Shen, H., Hu, J., 2017. Organophosphorus flame retardants in pregnant women and their transfer to chorionic villi. Environ. Sci. Technol. 51, 6489-6497 Zheng, X., Xu, F., Chen, K., Zeng, Y., Luo, X., Chen, S., Mai, B., Covaci, A., 2015. Flame retardants and organochlorines in indoor dust from several e-waste recycling sites in South China: composition variations and implications for human exposure. Environ. Int. 78, 1-7 Zhong, M., Wu, H., Mi, W., Li, F., Ji, C., Ebinghaus, R., Tang, J., Xie, Z., 2018. Occurrences and distribution characteristics of organophosphate ester flame retardants and plasticizers in the sediments of the Bohai and Yellow Seas, China. Sci. Total Environ. 615, 1305-1311 Zhou, L., Hiltscher, M., Puttmann, W., 2017. Occurrence and human exposure assessment of organophosphate flame retardants in indoor dust from various microenvironments of the Rhine/Main region, Germany. Indoor Air. 27, 1113-1127

28

Figure 1. Spatial distribution of organophosphate triesters (a) and diesters (b) in indoor and outdoor dust samples collected across China (unit: ng/g dw).

29

Figure 2. Spearman’s rank correlation analysis of organophosphate di- and triester concentrations in indoor and outdoor dust samples collected across China (Spearman’s rho represents correlation coefficient and p value represents significance).

30

Figure 3. Correlation analysis of concentrations of organophosphate triesters (a) and diesters (b) between paired indoor and outdoor dust samples collected across China.

31

Figure 4. The molar concentration ratios of organophosphate diesters to their corresponding triester parent compounds (The black horizontal line inside each box represents median, the boxes represent 25th and 75th percentiles, whiskers represent a value of 1.5 standard deviation and the dots represent outliers).

32

Table 1. Concentrations (ng/g dw) of organophosphate (OP) triesters in indoor and outdoor dust samples collected across China OP triesters

Indoor dust (n=44) Average Median

Min.

Max.

TMP TEP TPP TNBP TIBP TBOEP TEHP TCEP TCIPP TDCIPP TPHP TMPP EHDPP CDPP IDDP BPDP TBPHP RDP BDP ∑Alkyl-

1.02 101 -* 111 13.3 135 231 441 1320 220 206 16.4 105 26.2 6.58 29.7 2.22 11.9 11.5 594



9.36 2550
∑Cl-

1980

1510

169

∑Aryl-

391

200 3.98 2380

∑Oligomeric- 23.3 ƩOP triesters 2990 *Dash means not available.

Outdoor dust (n=97) Average Median

Min.

Max.

0.939 34.0 35.2 9.22 35.7 107 108 233 68.2 69.5 36.0 23.1 5.75 7.08 3.17 0.19 2.64 8.06 222



15.2 397
7050

409

212

17.4

5320

1.20

3450

144

71.8

2.88

2040


226 9540

10.7 788

1.00 466


460 5960

DF %

41 75 0 100 95 98 98 100 100 80 98 98 100 82 77 82 11 39 93

33

MDL DF % 30 81 0 100 91 96 98 100 100 69 98 98 100 65 92 52 2 20 54

0.32 1.22 2.11 0.95 0.56 1.87 2.02 0.79 2.33 0.90 0.40 0.35 0.29 1.00 0.31 0.28 1.56 0.74 0.52

Table 2. Concentrations (ng/g dw) of organophosphate (OP) diesters in indoor and outdoor dust samples collected across China OP diesters DEP DPP DNBP DIBP BBOEP BEHP BCEP BCIPP BDCIPP DPHP BMPP ∑Alkyl∑Cl∑ArylƩOP diesters

Indoor dust (n=44) Average Median

80.5 0.30 15.8 16.4 5.67 75.0 18.5 6.62 2.51 144 2.17 194 27.6 146 367

38.9
Min.


Max.

667 2.90 72.8 59.8 69.4 472 131 94.0 23.6 2810 18.8 700 152 2810 3000

DF %

95 27 98 93 70 100 43 45 25 100 95

34

Outdoor dust (n=97) Average Median

73.1 0.75 11.6 11.3 2.66 31.9 10.1 3.12 2.22 39.7 0.74 131 15.4 40.4 187

28.9 0.43 5.41 7.09 0.71 13.9
MDL Min.


Max.

1960 7.36 291 144 37.7 280 293 38.6 36.4 1340 11.4 1980 298 1340 2000

DF %

95 63 98 97 78 99 36 22 20 100 70

0.20 0.07 0.08 0.18 0.24 0.03 0.56 0.68 0.10 0.12 0.01

Table 3. Estimated daily intake (ng/kg bw/d) of organophosphate (OP) triesters and diesters via indoor and outdoor dust exposure in China Indoor dust Adults Average OP triesters TEP* TIBP TPHP AlkylClArylOligomericƩOP triesters

3.15E-03 2.16E-03 2.07E-02 7.29E-02 3.53E-01 4.67E-02 9.30E-04 5.55E-01

Outdoor dust Adults

Children High 6.83E-02 1.24E-02 2.50E-01 6.21E-01 1.20E+00 3.07E-01 2.47E-02 1.93E+00

Average

High

Average

Children High

Average

High

3.91E-02 2.68E-02 2.56E-01 9.05E-01 4.38E+00 5.79E-01 1.15E-02 6.88E+00

8.47E-01 1.54E-01 3.11E+00 7.71E+00 1.50E+01 3.81E+00 3.07E-01 2.40E+01

8.94E-04 2.86E-04 1.07E-03 6.32E-03 1.20E-02 4.07E-03 5.65E-05 2.64E-02

6.90E-03 1.56E-03 1.37E-02 3.62E-02 7.34E-02 2.39E-02 1.74E-03 1.64E-01

4.86E-03 1.55E-03 5.82E-03 3.43E-02 6.52E-02 2.21E-02 3.07E-04 1.43E-01

3.75E-02 8.45E-03 7.46E-02 1.97E-01 3.99E-01 1.30E-01 9.47E-03 8.90E-01

OP diesters DEP 9.08E-03 1.10E-01 1.13E-01 DIBP 2.36E-03 1.32E-02 2.93E-02 DPHP 1.11E-02 1.16E-01 1.38E-01 Alkyl3.12E-02 1.33E-01 3.87E-01 Cl1.73E-03 3.18E-02 2.15E-02 Aryl1.14E-02 1.19E-01 1.41E-01 ƩOP diesters 6.07E-02 2.09E-01 7.53E-01 *Detailed EDIs of each OP ester were listed in Tables S13 and S14.

1.37E+00 1.64E-01 1.44E+00 1.65E+00 3.95E-01 1.48E+00 2.59E+00

1.64E-03 4.01E-04 6.44E-04 4.65E-03 6.79E-05 6.60E-04 5.48E-03

1.15E-02 2.05E-03 8.92E-03 2.23E-02 4.94E-03 9.13E-03 3.19E-02

8.89E-03 2.18E-03 3.50E-03 2.53E-02 3.69E-04 3.59E-03 2.98E-02

6.25E-02 1.11E-02 4.85E-02 1.21E-01 2.68E-02 4.96E-02 1.73E-01

35

Highlights Organophosphate tri- and di-esters are ubiquitous in dust samples from China TCIPP and DPHP are the predominant tri- and di-esters, respectively, in dust TEHP and TPHP are the sources of their corresponding diesters in dust Human exposure levels of DEP and DPHP through dust are notable

36