Temporal Changes and Stereoisomeric Compositions of 1,2,5,6,9,10

Subscriber access provided by UNIVERSITY OF TOLEDO LIBRARIES

Article

Temporal changes and stereoisomeric compositions of 1,2,5,6,9,10hexabromocyclododecane (HBCD) and 1,2-dibromo-4-(1,2-dibromoethyl) cyclohexane (TBECH) in marine mammals from the South China Sea Yuefei Ruan, James C.W. Lam, Xiaohua Zhang, and Paul K.S. Lam Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.7b05387 • Publication Date (Web): 04 Feb 2018 Downloaded from http://pubs.acs.org on February 4, 2018

Just Accepted “Just Accepted” manuscripts have been peer-reviewed and accepted for publication. They are posted online prior to technical editing, formatting for publication and author proofing. The American Chemical Society provides “Just Accepted” as a service to the research community to expedite the dissemination of scientific material as soon as possible after acceptance. “Just Accepted” manuscripts appear in full in PDF format accompanied by an HTML abstract. “Just Accepted” manuscripts have been fully peer reviewed, but should not be considered the official version of record. They are citable by the Digital Object Identifier (DOI®). “Just Accepted” is an optional service offered to authors. Therefore, the “Just Accepted” Web site may not include all articles that will be published in the journal. After a manuscript is technically edited and formatted, it will be removed from the “Just Accepted” Web site and published as an ASAP article. Note that technical editing may introduce minor changes to the manuscript text and/or graphics which could affect content, and all legal disclaimers and ethical guidelines that apply to the journal pertain. ACS cannot be held responsible for errors or consequences arising from the use of information contained in these “Just Accepted” manuscripts.

Environmental Science & Technology is published by the American Chemical Society. 1155 Sixteenth Street N.W., Washington, DC 20036 Published by American Chemical Society. Copyright © American Chemical Society. However, no copyright claim is made to original U.S. Government works, or works produced by employees of any Commonwealth realm Crown government in the course of their duties.

Page 1 of 37

Environmental Science & Technology

1

Temporal changes and stereoisomeric compositions of

2

1,2,5,6,9,10-hexabromocyclododecane

3

1,2-dibromo-4-(1,2-dibromoethyl) cyclohexane (TBECH)

4

in marine mammals from the South China Sea

5

Yuefei Ruan†, James C.W. Lam*,†,‡, Xiaohua Zhang‡, Paul K.S. Lam*,†,§

6

†

(HBCD)

and

State Key Laboratory in Marine Pollution (SKLMP), Research Centre for the Oceans

7

and Human Health, Shenzhen Key Laboratory for Sustainable Use of Marine

8

Biodiversity, City University of Hong Kong, Hong Kong SAR, China

9

‡

Hong of Kong, Hong Kong SAR, China

10 11 12

Department of Science and Environmental Studies, The Education University of

§

Department of Biology and Chemistry, City University of Hong Kong, Kowloon, Hong Kong, China

James C. W. Lam*

Paul K. S. Lam*

Department of Science and

State Key Laboratory in Marine

Environmental Studies, The Education

Pollution, City University of Hong

University of Hong Kong, Hong Kong

Kong, Tat Chee Avenue, Kowloon,

SAR, China

Hong Kong SAR, China

Tel: +852-2948-8537

Tel: +852-3442-6828

Fax: +852-2948-7676

Fax: +852-3442-0303

E-mail: [email protected];

E-mail: [email protected]

[email protected]

13

1

ACS Paragon Plus Environment


14

ABSTRACT

15

Stereoisomeric compositions of 1,2,5,6,9,10-hexabromocyclododecane (HBCD) and

16

1,2-dibromo-4-(1,2-dibromoethyl)cyclohexane (TBECH) were investigated in the

17

blubber of two species of marine mammals, finless porpoises (Neophocaena

18

phocaenoides) and Indo-Pacific humpback dolphins (Sousa chinensis), from the South

19

China Sea between 2005 and 2015. The concentrations of ΣHBCD in samples of

20

porpoise (n = 59) and dolphin (n = 32) ranged from 97.2 to 6,260 ng/g lipid weight

21

(lw) and from 447 to 45,800 ng/g lw, respectively, while those of ΣTBECH were both

22

roughly two orders of magnitude lower. A significant increasing trend of ΣHBCD was

23

found in dolphin blubber over the past decade. The diastereomeric profiles exhibited

24

an absolute predominance of α-HBCD (mostly > 90%), while the proportions of four

25

TBECH diastereomers in the samples appeared similar. A preferential enrichment of

26

the (−)-enantiomers of α-, β-, and γ-HBCD was found in most blubber samples.

27

Interestingly, the body lengths of porpoises showed a significant negative correlation

28

with the enantiomer fractions of α-HBCD. Significant racemic deviations were also

29

observed for α-, γ-, and δ-TBECH enantiomeric pairs. This is the first report of the

30

presence of TBECH enantiomers in the environment. The estimated hazard quotient

31

indicates that there is a potential risk to dolphins due to HBCD exposure.

32

2


Page 2 of 37

Page 3 of 37


33

INTRODUCTION

34

Approximately one third of organic flame retardants on the market are categorized as

35

brominated flame retardants (BFRs).1 One of the most extensively used BFRs in

36

consumer products — polybrominated diphenyl ethers (PBDEs) — have been

37

identified as global pollutants, and technical mixtures containing these compounds

38

have been restricted worldwide.2,3 Thus, there has been a shift to the use of

39

alternatives. Two types of alternatives, namely, 1,2,5,6,9,10-hexabromocyclododecane

40

(HBCD) and 1,2-dibromo-4-(1,2-dibromoethyl) cyclohexane (TBECH), have been

41

regarded as possible replacements for PBDEs.4 Both of these compounds have similar

42

applications in expanded and extruded polystyrene.5–7 Nevertheless, HBCD has been

43

ubiquitously found in the environment, with very high concentrations (up to ppm

44

levels) frequently observed in marine mammals and seabirds.8,9 Since 2013, HBCD

45

has been listed under the Stockholm Convention on Persistent Organic Pollutants

46

(POPs) under Annex A for global elimination;10 however, its use is still permitted, and

47

it has a global annual production of more than 30,000 metric tons.11 TBECH has been

48

produced since 1970s, but information on its production is very limited and was,

49

formerly reported between 4.5 and 226 metric tons/year in the USA.12,13 Frequently

50

detected in environmental matrices, TBECH has been an emerging concern due to its

51

toxicity as the first potent environmental androgen receptor (AR) agonist, which was

52

found to be as potent as dihydrotestosterone in human AR activation.14 As indicated

53

by studies on indoor dust from European regions, TBECH could be a possible HBCD

54

replacement in the investigated area.15–17

55

These two BFRs can exhibit optical activities as chiral compounds. HBCD can

56

generate up to 6 pairs of enantiomers and 4 meso-forms, while TBECH can yield up

57

to 4 pairs of enantiomers.18–20 Technical products of HBCD (tHBCD) usually consist 3



58

of three diastereomeric pairs of enantiomers, α- (~12%), β- (~9%), and γ- (~79%)

59

racemates, with trace amounts of others such as δ- and ε-meso-isomers (~0.5%).21–24

60

Only one manufacturer (Albemarle Corporation. Charlotte, NC, USA) reported the

61

commercial formulation of TBECH (tTBECH), and it contains equimolar amounts of

62

α- and β-racemates with trace amounts of γ- and δ-diastereomers generated in a

63

high-temperature operation.20 With respect to the environmental relevance of their

64

stereoisomerism, a shift toward enrichment of α- over β- and γ-HBCD along trophic

65

levels has consistently occurred in food webs.25 Few studies have confirmed the

66

presence of HBCD meso-isomers in the environment.26 The prevalence of the

67

preferential bioaccumulation of (−)-α-HBCD can be observed more often in certain

68

animal species, but this seems to be independent of other biological factors such as

69

body composition, age/sex variation, nutritive condition, and tissue distribution.8,25,27

70

There is still very limited knowledge of the fate of these chiral contaminants in the

71

environment, particularly TBECH, which has been primarily reported as the sum of

72

all stereoisomers (ΣTBECH).28,29 Information about the environmental occurrence of

73

TBECH enantiomers has not yet been reported.

74

Although stereoisomers can exhibit different environmental fates and toxicities, the

75

stereoisomerism of environmental contaminants has not yet received sufficient

76

attention.30 The understanding of the toxicities of individual HBCD or TBECH

77

stereoisomers is far from adequate. In general, in vitro studies have found that all

78

(+)-HBCD appeared more toxic than their corresponding (−) forms, while γ- and

79

δ-TBECH were more potent in AR agonism than α- and β-TBECH.14,31,32 In vivo

80

studies on juvenile rainbow trout (Oncorhynchus mykiss) and adult zebrafish (Danio

81

rerio) have evidenced a stronger potency of γ-HBCD,33,34 while in fertilized eggs of D.

82

rerio, a more pronounced embryo toxicity of γ/δ-TBECH than α/β-TBECH was

4


Page 4 of 37

Page 5 of 37


83

observed.35

84

Marine mammals, particularly those from coastal regions, have been suggested to

85

serve as important sentinel species for tracking spatio-temporal dynamics of

86

environmental contamination.36 Several BFRs have been linked with detrimental

87

health effects in marine mammals, such as the decrease in offspring survivorship.37

88

According to our previous studies, two resident marine mammal species in the South

89

China Sea, finless porpoises (Neophocaena phocaenoides) and Indo-Pacific

90

humpback dolphins (Sousa chinensis), were susceptible to exposure to high levels of

91

PBDEs and other BFRs.38 However, the stereoisomerism of these chemicals has not

92

been considered, especially regarding their environmental fate and potential risks to

93

these cetaceans.

94

To better understand the current status of the exposure and the risks posed by the

95

chiral BFRs, the present study aimed to (1) identify and quantify HBCD and TBECH

96

stereoisomers using novel methods for the investigation of the possible temporal

97

trends (2005‒2015) and physiology-related differences in the accumulation of these

98

contaminants in marine mammals from the South China Sea and (2) study the changes

99

in the stereoisomeric compositions of the chiral BFRs in marine mammals and

100

evaluate their potential risks to the cetaceans. To our knowledge, this is the first study

101

that demonstrates the environmental occurrence of TBECH as well as HBCD

102

meso-isomers in marine mammals in the Asian region, as well as the first report on the

103

identification of ε-HBCD in organisms and the detection of TBECH enantiomers in

104

the environment.

105

MATERIALS AND METHODS

106

Materials and Sample Collection. The standards used in this study were

5



107

purchased from Wellington Laboratories (Guelph, Ontario, Canada). The details of

108

standards, solvents, and reagents are described in the Supporting Information (SI).

109

The blubber samples of finless porpoises (n = 59) and Indo-Pacific humpback

110

dolphins (n = 32) were obtained from stranded cetaceans collected by the Agriculture,

111

Fisheries and Conservation Department in Hong Kong, China, between 2005 and

112

2015 (SI Table S1). Once in the laboratory, the samples were freeze-dried and stored

113

at −50 ºC until further processing.

114

Extraction and Cleanup. Analyses of blubber samples were accomplished by

115

using previously established methods with modifications.38 Briefly, ~1 g of blubber

116

sample was ground with anhydrous sodium sulfate and spiked with surrogates of 10

117

ng each, and then the sample was extracted via pressurized fluid extraction. Each

118

extract was purified for lipid removal by gel permeation chromatography and then

119

further purified by elution through anhydrous sodium sulfate, activated aluminum

120

oxide, and activated silica gel. The final reduced extract was spiked with internal

121

standards of 10 ng each before it was concentrated to dryness under a gentle nitrogen

122

stream and, finally, reconstituted to 100 µL in methanol or n-hexane.

123

Instrumental Analysis. Determination of HBCD stereoisomers was performed

124

using liquid chromatography‒tandem mass spectrometry (LC–MS/MS) consisting of

125

an Agilent (Palo Alto, CA, USA) 1290 Infinity LC system coupled to an AB SCIEX

126

(Woodlands, Singapore) QTRAP®3200 system with an electrospray ionization (ESI)

127

interface in negative multiple reaction monitoring (MRM) mode. An Agilent

128

(Lawrence, Kansas, USA) ZORBAX Eclipse Plus C18 column (4.6 mm i.d. × 100

129

mm, 3.5 µm) coupled to a Macherey-Nagel (Düren, Germany) NUCLEODEX β-PM

130

column (4 mm i.d. × 200 mm, 5 µm) was applied for chromatographic separation.

131

More details about LC–MS/MS determination can be found in the SI. 6


Page 6 of 37

Page 7 of 37


132

Determination of TBECH enantiomers was performed using gas chromatography–

133

mass spectrometry (GC–MS) with an Agilent (Shanghai, China) 7890A GC system

134

coupled to an Agilent (Santa Clara, CA, USA) 5975C inert XL MSD with a triple axis

135

detector in negative chemical ionization mode. A Supelco (Bellefonte, PA, USA)

136

CHIRALDEX B-TA capillary column (30 m × 0.25 µm i.d., 0.12 µm film thickness)

137

was used for chromatographic separation. The GC oven program is listed in SI Table

138

S2. More details about GC–MS determination can be found in the SI.

139

Quality Assurance and Quality Control (QA/QC). The QA/QC was

140

composed of procedure blanks and reference samples per batch of 10 samples. A

141

9-point standard calibration curve, with concentrations ranging from 5 to 2,000 ng/mL,

142

was used for the quantification of each stereoisomeric analyte. The regression

143

coefficients (r) were ≥ 0.995 for all calibration curves. Concentrations of analytes

144

were expressed as ng/g lipid weight (lw). Recoveries of spiked standards in the

145

reference samples, including eel muscles and corn oil, can be found in the SI.

146

Recoveries of spiked surrogates in the cetacean blubber were all in the range of 71%

147

to 97%. Recoveries of each native standard and surrogate are listed in SI Table S3.

148

The limit of detection (LOD) was defined as an instrument signal-to-noise ratio (SNR)

149

of ≥ 3:1. The method LODs of each HBCD and TBECH stereoisomer are listed in SI

150

Table S4.

151

Statistical Analysis. Statistical analysis was performed using Microsoft Excel

152

2016 to generate descriptive statistics, with all other statistical procedures conducted

153

using IBM SPSS Statistics 20.0 and OriginLab OriginPro 2015. For the calculation of

154

geometric mean and arithmetic mean, it was confirmed that different data treatments,

155

i.e. excluding values below the LOD, using ½ the LOD instead, or setting those

156

values to zero, had almost no effect on the conclusion and interpretation of the results. 7



157

Thus, values below the LOD were replaced with ½ LOD, in the same way as our

158

previous studies.38,39 Methods used for statistical hypothesis and correlation analysis

159

are described in the SI. The detection frequencies of HBCD and TBECH

160

stereoisomers in the samples are listed in SI Table S5. Enantiomer fractions (EFs)

161

were used to describe enantiomeric patterns,40 and the details of EF calculation can be

162

found in the SI.

163

RESULTS AND DISCUSSION

164

Levels of ΣHBCD and ΣTBECH in the Cetacean Blubber. ΣHBCD was

165

found in all blubber samples, whereas the detection frequencies of ΣTBECH (65.9%)

166

were much lower than those of ΣHBCD (SI Table S5). The concentrations of ΣHBCD

167

in the samples of porpoise and dolphin ranged from 97.2 to 6,260 ng/g lipid weight

168

(lw) and from 447 to 45,800 ng/g lw, respectively, while those of ΣTBECH ranged

169

from below the limit of detection (LOD) (< 0.8) to 125 ng/g lw and from < 0.8 to 362

170

ng/g lw, respectively (Table 1). Concentrations of both ΣHBCD and ΣTBECH in

171

porpoises were much lower than those in dolphins (p < 0.001), probably partially due

172

to the difference in the spatial distributions between these two species — dolphins

173

exploit prey species common in the estuary, while porpoises rely more on pelagic

174

habitats for food.41,42 As indicated by our previous studies, compared to the habitat of

175

porpoises, the habitat of dolphins is closer to the coastal regions of the Pearl River

176

Estuary, where the aquatic environment is more contaminated by different brominated

177

flame retardants.38,39 However, species-specific differences in contaminant uptake

178

from diets, accumulation, and metabolism should not be neglected and these deserve

179

for further investigation. Concentrations of ΣTBECH were roughly two orders of

180

magnitude lower than those of ΣHBCD in both porpoises and dolphins, in accordance

8


Page 8 of 37

Page 9 of 37


181

with previous findings of a lower bioaccumulation potential of TBECH than HBCD in

182

zebrafish (Danio rerio) after dietary exposure.43 This also suggests the much higher

183

production/use of tHBCD than tTBECH in the Pearl River Delta (PRD). Positive

184

correlations between the levels of ΣHBCD and ΣTBECH were observed in the two

185

species of marine mammals (Pearson’s r = 0.500, p < 0.0001 and Pearson’s r = 0.680,

186

p < 0.0001 for porpoises and dolphins, respectively), indicating that there may have

187

been a similar production/use pattern between tHBCD and tTBECH, a potential

188

HBCD replacement, in the PRD over the past decade. Notably, though HBCD has

189

been listed under the Stockholm Convention on Persistent Organic Pollutants (POPs)

190

for global elimination since 2013, the control on the production and use of HBCD in

191

China has not come into force until December 26, 2016 (SI Figure S3). In addition,

192

TBECH can be used as a substitute for HBCD, the regional market demand for

193

TBECH has kept increasing. Thus, it is not surprising that positive correlations

194

between the levels of ΣHBCD and ΣTBECH were observed.

195

The sex, life history and biological factors of marine mammals have been reported

196

to affect the accumulation pattern of some legacy organochlorines and PBDEs, where

197

the contaminant burden in adult males and juveniles has been generally higher than

198

that in adult females

199

placental transfer and/or lactation and metabolism capacity with age and/or dilution

200

with growth. In the present study, the effects of the marine mammal’s sex and age on

201

contaminant levels were also examined (Table 1). The levels of both ΣHBCD and

202

ΣTBECH in porpoises varied greatly with developmental stages. Significantly higher

203

levels of ΣHBCD were observed in adult males and adult females than in juveniles (p

204

< 0.01), and significantly higher levels of ΣTBECH were observed in adult males than

205

in adult females and juveniles (p < 0.05). Compared to age classes, sexual differences

44–47

. The main reasons for this finding are suggested to be

9



206

appear to have little influence on ΣHBCD accumulation in porpoises. Compared with

207

adult females, adult males had higher mean concentrations, but this difference was not

208

significant (p > 0.05). The accumulation pattern in porpoises indicates a combined

209

consequence of maternal offloading and increased bioaccumulation capacity with age.

210

In dolphins, however, no significant difference in the levels of either ΣHBCD or

211

ΣTBECH was found among juveniles, adult males, and adult females (p > 0.05), and

212

similar finding of POPs was also reported in cetaceans from Korean waters.47 The

213

reason for this finding is not entirely clear, but the finding suggests that HBCD loss

214

through maternal transfer may be lower in dolphins. However, the compensation via

215

continuous bioaccumulation long after reproduction for the loss of HBCD through

216

maternal transfer should not be neglected.

217

Additionally, the current status of HBCD and TBECH was examined on a global

218

scale. The highest ΣHBCD concentrations detected in marine cetaceans collected

219

from different regions are summarized in SI Table S6. The highest concentration in

220

marine cetaceans reported to date was observed in blubber samples of dolphin

221

investigated by the present study, suggesting a heavier use of HBCD and a greater

222

contamination burden in the PRD than in other regions. Likewise, because TBECH is

223

a novel BFR and due to its limited records in environmental matrices, the biological

224

tissues of concern were extended to those of any types of organisms ever reported (SI

225

Table S7), and the highest ΣTBECH concentration observed to date was also found in

226

dolphin blubber from the PRD. However, the environmental fate of TBECH is still

227

poorly understood and further study is needed.

228

Temporal Trends of ΣHBCD and ΣTBECH in Porpoises and Dolphins.

229

Time-series analysis was conducted on the present data set of blubber samples, and

230

the concentrations were used to perform a log-linear regression. To attenuate the 10


Page 10 of 37

Page 11 of 37


231

differences between the sexes caused by maternal transfer and possible age-related

232

dissimilarities, adult males with body lengths greater than 120 and 200 cm for

233

porpoises and dolphins, respectively, were selected for this temporal-trend analysis

234

(SI Figure S1). Individuals with body lengths smaller than 120 and 200 cm were

235

categorized as juveniles for porpoises and dolphins, respectively.38,39 Consequently,

236

the dataset of adult males shows a clear upward trend between 2005 and 2015.

237

Specifically, the concentrations of ΣHBCD (p < 0.05) and ΣTBECH (p < 0.01) were

238

both found to exhibit a significant increasing trend in dolphin blubber. Although the

239

size of the dataset was limited, the log-linear regression slope of ΣTBECH in adult

240

male dolphins (Pearson’s r = 0.745) was higher than that of ΣHBCD (Pearson’s r =

241

0.816), suggesting a possible rapid growth in tTBECH consumption in the PRD over

242

the past decade.

243

The ΣHBCD concentrations are not statistically different between the juveniles and

244

adult dolphins and between males and females; thus, pooling of more individual

245

samples is possible. The complete dataset (SI Figure S2) shows that the

246

concentrations of ΣHBCD exhibited a significant increasing trend in dolphin blubber

247

between 2005 and 2015 (Pearson’s r = 0.567, p < 0.001), with an estimated annual

248

rate of 8%; an increasing trend in ΣHBCD levels was observed in porpoise blubber,

249

but this trend was not significant (p > 0.05). In comparison, the concentrations of

250

ΣTBECH appeared to remain steady in the blubber of both porpoise and dolphin

251

between 2005 and 2015, which possibly reflected a relatively low exposure of these

252

animals to TBECH within this period. However, the levels were higher in recent years

253

(as indicated by the log-linear regression slope), which may suggest the increasing

254

usage of tTBECH.

255

Based on one of our previous studies on PBDEs and several PBDE alternatives (but 11



256

excluding HBCD and TBECH) in blubber samples of the same mammal species

257

collected between 2003 and 2012, we found a similar contamination level between

258

PBDEs and ΣHBCD (within the same order of magnitude) over the past decades.38

259

Note that the large-scale production and uses of tHBCD in China began around 2000

260

and proliferated after 2005.11 However, HBCD has been banned for production, use,

261

import and export in China since December 26, 2016 with a five-year exemption for

262

the use in EPS/XPS foam in building insulation; after that period, its use is planned to

263

cease by 2021. Between 2005 and 2015, the production of tHBCD in China over time

264

increased dramatically (SI Figure S3). It is interesting to observe that the production

265

of tHBCD in China had an increasing trend that was similar to the ΣHBCD

266

concentrations found in dolphin blubber. In this case, dolphins may act as a potent

267

biological indicator to reflect the regional contamination status and trends.

268

Diastereomer Distribution Patterns in the Cetacean Blubber. The

269

stereoisomerism of HBCD has recently gained much interest. Because TBECH has

270

been indicated as a potential replacement for HBCD in expanded polystyrene for

271

building insulation, it is possible that this contaminant will be of global concern in the

272

future.4,16 TBECH also contains stereoisomers. However, knowledge on the

273

stereoisomerism of TBECH in the environment is still very limited. In this study, α-,

274

β-, γ-, δ-, and ε-HBCD, as well as α-, β-, γ-, and δ-TBECH, were considered for

275

diastereomeric identification.

276

The α-diastereomer of HBCD was detected in all blubber samples, accounting for

277

95.3% and 98.2% (mean values) of ΣHBCD in porpoises and dolphins, respectively

278

(Figure 1A). The results shown in SI Figure S4 further demonstrated that α-HBCD

279

dominated the HBCD profiles in marine mammals among the studied physiological

280

traits. For β-, γ-, δ-, and ε-HBCD, in general, their detection frequencies in porpoise 12


Page 12 of 37

Page 13 of 37


281

blubber were lower than those in dolphin blubber (SI Table S5), while the

282

ε-meso-isomer, on average, possessed the least proportions in both marine mammal

283

species.

284

In porpoise blubber, significantly higher proportions of δ-HBCD were found in

285

adult males than in adult females (p < 0.01), but when juveniles and adults were

286

pooled, the HBCD diastereomeric compositions did not show any sexual differences

287

(p > 0.05) (SI Figure S4A); in comparison, in terms of developmental stage, the

288

proportions of α-HBCD in juveniles were significantly lower than those in adults (p
0.05). However, the ratios of α-HBCD to

294

γ-HBCD were much higher in dolphins than in porpoises (p < 0.01), suggesting a

295

possibly higher bioaccumulation potential of α-HBCD in dolphin blubber.

296

For TBECH diastereomeric compositions, α-TBECH had the highest mean values,

297

accounting for 35.3% and 35.8% of ΣTBECH in porpoises and dolphins, respectively;

298

δ-TBECH exhibited the lowest mean values, accounting for 15.5% and 17.3% of

299

ΣTBECH in porpoises and dolphins, respectively (Figure 1B). TBECH diastereomeric

300

compositions in blubber showed certain species differences between porpoises and

301

dolphins. The proportions of β-TBECH in porpoise blubber were significantly higher

302

than those in dolphin blubber (p < 0.05). Although the difference was not significant,

303

γ-TBECH in porpoises had a lower mean value than in dolphins (p > 0.05). Hence, a

304

higher potential for the preferential bioaccumulation of γ-TBECH and/or the

305

biotransformation from β-TBECH may have occurred in dolphins compared to 13



306

porpoises. For both marine mammal species, their sex and life history did not appear

307

to affect the compositions of TBECH diastereomers (p > 0.05).

308

Diastereomer Proportions of HBCD and TBECH across Different

309

Studies. In the present study, the diastereomeric compositions of a specific tHBCD

310

were ascertained in a product of technical mixtures purchased from Toronto Research

311

Chemicals (Toronto, Ontario, Canada), and the mixtures were composed of 8.47%,

312

5.69%, 85.3%, 0.28%, and 0.22% of α-, β-, γ-, δ-, and ε-HBCD, respectively, similar

313

to the HBCD commercial/technical formulations previously reported. Dozens of

314

studies published elsewhere have demonstrated that in abiotic matrices, γ-HBCD was

315

the prevailing diastereomer, which was called a technical mixture-like diastereomeric

316

pattern, whereas in biotic matrices, especially for organisms at higher trophic levels

317

such as the apex predators, α-HBCD usually had an absolute predominance, or was

318

even the exclusive diastereomer detected. Studies indicated that this pattern possibly

319

resulted from diastereoselective uptake and/or metabolism.8,25 However, there are

320

very few records on δ- and ε-meso isomers in environmental matrices, which is

321

probably not attributed to their absence in the environment, but rather a result of the

322

very limited number of studies addressing this issue.8 In our study, the diastereomeric

323

patterns of α-, β-, and γ-HBCD in blubber of porpoise and dolphin coordinated well

324

with the diastereomer-specific enrichment pattern referred to above, and, δ- and

325

ε-HBCD found in our samples may reflect a combined consequence of

326

bioaccumulation and biotransformation (including bioisomerization).

327

For TBECH, specifically, a product of technical mixtures was also purchased from

328

Toronto Research Chemicals for diastereomeric identification. The proportion of

329

α-TBECH was ~15% higher than that of β-TBECH, and trace amounts of γ- and

330

δ-TBECH were detected with the resultant proportions of 0.13% and 0.12%, 14


Page 14 of 37

Page 15 of 37


331

respectively. This finding is different from the diastereomeric profiles of tTBECH

332

reported in other studies, with equimolar amounts of α- and β-racemates consistently

333

quoted from Arsenault et al. (2008).20 To date, complete TBECH diastereomeric

334

profiles in the environment have not been reported. Interestingly, in this study,

335

compared with the diastereomeric composition in TBECH technical mixtures, a

336

preferential enrichment of γ- and δ-TBECH, rather than α- and β-TBECH, could be

337

clearly observed in the blubber of the two marine mammal species. To our knowledge,

338

this was the first study to report a diastereoselective bioaccumulation of TBECH

339

stereoisomers in the environment.

340

Enantiomer Fractions (EFs) of HBCD in the Cetacean Blubber. The

341

elution order of (+)- and (−)-enantiomers of α-, β-, and γ-HBCD has been determined

342

in previous studies using the same NUCLEODEX chiral column as that used in the

343

present study.48,49 TBECH also contains enantiomeric pairs, and the enantioselective

344

degradation of α-TBECH has been examined as an indication of microbial processing

345

by Wong et al. (2012), who reported a shift in degradation preference from the second

346

elution enantiomer (on the BGB-176MS column) to its antipode.50 However, the

347

chromatographic elution order of any TBECH enantiomeric pair as (+)- and

348

(−)-enantiomers has not been confirmed. Here, as previously mentioned, the

349

concentration ratio of the first-eluting enantiomer (E1) on the CHIRALDEX B-TA

350

column and the total concentration of both enantiomers (E1 + E2) was used for

351

calculation. The resulting EFs of each pair of enantiomers in the blubber of porpoise

352

and dolphin are shown in Figure 2.

353

All blubber samples showed that the EFs of α-HBCD enantiomers negatively

354

deviated from the racemic standards with statistical significance (p < 0.01) (Figure

355

2A), with averages of 0.371 ± 0.044 and 0.387 ± 0.034 in porpoise blubber and 15



356

dolphin blubber, respectively, which evidenced a strongly preferential enrichment of

357

(−)-α-HBCD. The mean EFs of β- and γ-HBCD were also far below 0.5, with

358

averages of 0.418 ± 0.137 and 0.419 ± 0.063, respectively in porpoise blubber, and

359

0.357 ± 0.136 and 0.282 ± 0.107, respectively in dolphin blubber (p < 0.05),

360

suggesting a predominance of all (−)-HBCD enantiomers (in terms of α-, β-, and

361

γ-HBCD) in these marine mammals.

362

Physiological traits, including development stages and sexual differences, affected

363

the enantioselective pattern of α-HBCD and γ-HBCD in blubber (SI Figure S5). For

364

α-HBCD, a similar pattern of enantiomer-specific accumulation was observed within

365

the same categories of physiological traits in both marine mammal species.

366

Significant differences in the EFs of α-HBCD were found between adult males and

367

adult females and between juvenile males and adult males (p < 0.05), where the

368

lowest values of the EFs of α-HBCD occurred in adult males. For γ-HBCD, another

369

pair of enantiomers, the lowest mean EFs occurred in the dolphin blubber of juvenile

370

males. In accordance with the extremely high levels of α-HBCD found in blubber

371

(1,420 ± 1,320 ng/g lw and 7,580 ± 9,340 ng/g lw for porpoises and dolphins,

372

respectively), a possible explanation for the specific pattern observed above is the

373

maternal transfer of (−)-α-HBCD from mature females to their calves via gestation

374

and lactation, resulting in higher amounts of (−)-α-HBCD in adult males. The

375

stereoisomer-specific maternal transfer of persistent hydrophobic contaminants has

376

also been reported by several studies.51,52

377

Correlations between Enantiomer Fractions (EFs) of HBCD and Body

378

Lengths of Cetaceans. Studies have reported that a decreasing EF trend with

379

increasing age and/or animal body length was found for α-HBCD in the blubber of

380

Atlantic white-sided dolphins (Lagenorhynchus acutus) from the eastern coast of the 16


Page 16 of 37

Page 17 of 37


381

United States,51 for α-HCH in the blubber of ringed seals (Phoca hispida) from the

382

Northwest Polynya,53 and for PCB-95 in the blubber of bowhead whales (Balaena

383

mysticetus) from the northern coast of Alaska.54 For both porpoises and dolphins,

384

there was no difference in the EFs of α-HBCD among juvenile males, juvenile

385

females and adult females (p > 0.05); therefore, pooled samples of males and females

386

are possible for the assessment of the potential interaction between enantiomeric

387

distribution and body length as a surrogate for growth (SI Figure S6A). As a result,

388

the body length of porpoises (p < 0.0001), rather than that of dolphins (p > 0.05), was

389

significantly inversely correlated (Pearson’s r = ‒0.575) with the EFs of α-HBCD in

390

blubber. In other words, a steady excess of (‒)-α-HBCD in blubber was found as the

391

porpoise body length increased, suggesting that enantiomer-specific accumulation of

392

this enantiomer in porpoises gradually occurs over time.

393

Several studies have evidenced the dietary changes during early life stages of

394

certain marine mammals and the subsequent changes in the accumulation patterns of

395

persistent hydrophobic contaminants.45,55–57 With respect to the porpoises’ sex, the

396

correlation between the EFs of α-HBCD in blubber and body length was further

397

examined by using first-order linear regression, and interestingly, the EFs of α-HBCD

398

were not correlated with the body length of female porpoises (p > 0.05) but were

399

highly correlated with that of male porpoises [Pearson’s r = ‒0.706, R2 (linear) =

400

0.485, p < 0.0001]. This finding suggests that the EFs of α-HBCD may remain

401

constant in these female marine mammals during their life spans, which is probably a

402

cumulative consequence of a gradual dietary change from maternal dependency

403

before weaning to independent foraging for small prey in juveniles as they grow. In

404

the present study, because no correlation was found between the concentrations of

405

α-HBCD and animal body lengths in either porpoises or dolphins (p > 0.05), the

17



406

stereoselective metabolism of α-HBCD was not likely to account for the significant

407

decline in EFs as the porpoise body length increased, which further supports the

408

above-mentioned implications of the maternal transfer of (−)-α-HBCD from mature

409

females, possibly accompanied by the gradual change in diet from milk to prey in

410

these juvenile marine mammals.

411

For γ-HBCD, considering that for both porpoises and dolphins, there was no

412

difference in the EFs of γ-HBCD among juvenile males, juvenile females and adult

413

females (p > 0.05), pooled samples of males and females are also possible for

414

evaluating the potential interaction between enantiomeric distribution and body length

415

(SI Figure S6B). In contrast to the decreasing trend of the EFs of α-HBCD with

416

increasing body length, the body length of dolphins (p < 0.05), rather than that of

417

porpoises (p > 0.05), was significantly positively correlated (Pearson’s r = ‒0.575)

418

with the EFs of γ-HBCD in blubber. Thus, a continuous preferential elimination of

419

(‒)-γ-HBCD in blubber was found as the dolphin body length increased, implying that

420

enantiomer-specific depuration of this enantiomer by dolphins occurs over time. With

421

respect to the sex of the dolphin, the correlation between the EFs of γ-HBCD in

422

blubber and the body length was further examined by using first-order linear

423

regression, and identical to the case of α-HBCD in porpoises, the EFs of γ-HBCD

424

were not correlated with the body length of female dolphins (p > 0.05), but were

425

highly correlated with that of male dolphins [Pearson’s r = 0.798, R2 (linear) = 0.604,

426

p < 0.01]. This result indicates that the EFs of γ-HBCD may remain constant in

427

female dolphins during their life spans. Sex-specific enantioselective accumulation

428

and/or depuration of either γ-HBCD enantiomer in dolphins may be a possible

429

explanation for this phenomenon.

430

Enantiomer Fractions (EFs) of TBECH in the Cetacean Blubber. For 18


Page 18 of 37

Page 19 of 37


431

TBECH enantiomers, in the present study, almost all blubber samples showed that the

432

EFs of α-TBECH positively deviated from the racemic standards with statistical

433

significance (p < 0.05), with averages of 0.587 ± 0.073 and 0.569 ± 0.053 in porpoises

434

and dolphins, respectively (Figure 2B). This finding indicates that a preferential

435

accumulation of E1-α-TBECH occurred in the blubber of these marine mammals.

436

Interestingly, for dolphins, although no difference was found in the EFs of α-TBECH

437

among juveniles, adult males and adult females (p > 0.05), the EFs of α-TBECH in

438

adult males (indicated by the mean and median values) appeared to have the greatest

439

deviation from the racemic standards (SI Figure S7). Therefore, similar to the case of

440

α-HBCD, the maternal transfer of E2-α-TBECH from mature females to their

441

offspring was likely to occur in dolphins.

442

In comparison, no enantioselective accumulation of β-TBECH was observed in

443

either porpoises or dolphins (p < 0.05). Note that EFs could be calculated only for

444

samples with the concentration of a specific TBECH diastereoisomer above the

445

method LOD. In this study, the detection frequencies of the γ- and δ-enantiomer pairs

446

in dolphins were both too low (~30-50%) to compare their EFs among different

447

physiological traits with sufficient precision. However, a preferential enrichment of

448

the E2-γ-enantiomer and the E1-δ-enantiomer seemed to be consistent for both

449

porpoises and dolphins (Figure 2B). Additionally, in porpoises, significantly higher

450

proportions of E1-δ-TBECH were observed in females than in males (p < 0.05). To

451

date, the toxicological significance of the enantioselective bioaccumulation of

452

TBECH stereoisomers is unknown, which warrants urgent investigation.

453

Preliminary Evaluation of the Hazards of Chiral BFRs to Marine

454

Mammals. In this screening assessment, hazard quotients (HQs) were determined

455

based on best-case and worst-case scenarios (HQbcs and HQwcs, respectively) to 19



456

provide a preliminary quantitative hazard evaluation of ΣHBCD and ΣTBECH to

457

porpoises and dolphins. HQs were calculated by dividing the measured environmental

458

concentration (MEC) [ng/g wet weight (ww)] by the lowest levels of estimated “body

459

burden” at which critical harmful effects were observed in mammalian toxicological

460

studies (hereafter referred to as tentative critical concentration, TCC).

461

The no-observed-adverse-effect levels (NOAELs) are available for several

462

mammals (rodents) with the investigated contaminants.58,59 In the absence of

463

dolphin/porpoise-speciﬁc toxicity data, by applying a body weight scaling factor

464

following the method of Lam & Lam (2015),60 these NOAEL values were converted

465

into TCC, which were used in the present risk assessment. In addition, an uncertainty

466

factor of 10 was applied to the calculation to account for the cross-species

467

extrapolation involved in the assessment process, and the whole-body concentration

468

was estimated to be 1/5 the blubber concentration.60–62 Because the extrapolation

469

between species and tissue types involves many uncertainties, the results of this

470

assessment will require further refinement when more data becomes available. The

471

corresponding HQs of ΣHBCD and ΣTBECH for porpoises and dolphins in this study

472

are summarized in SI Table S8. It is important to note that the hazard of HBCD to

473

finless porpoises and TBECH to aquatic organisms was evaluated for the first time.

474

The results of the assessment for ΣHBCD revealed that the HQbcs values for the

475

marine mammals were below unity, while the HQwcs values for both porpoises and

476

dolphins exceeded unity, suggesting that the whole-body concentrations of ΣHBCD

477

pose potential risks to these marine mammals. The HQs of one dolphin individual

478

exceeded 1, suggesting a moderate hazard to 3.13% of the investigated population of

479

dolphins from HBCD exposure, while not any porpoise individual had a HQ value

480

greater than 1, indicating that porpoises are much less threatened by exposure to 20


Page 20 of 37

Page 21 of 37


481

HBCD. Nearly 14% and 47% of the investigated populations of porpoises and

482

dolphins, respectively, were at risk (or had a low to moderate hazard) due to HBCD

483

exposure. HBCD may have induced potential detrimental effects from mild impaired

484

oligodendroglial development at the adult stage by long-term maternal exposure in

485

porpoises and dolphins in this studied region. In contrast, none of the HQs of

486

ΣTBECH detected exceeded a value of 0.01, thus posing no hazard (in terms of mild

487

renal lesions and inflammation) to porpoises and dolphins. Nonetheless, a study on

488

TBECH should be ongoing, as this chemical is a possible replacement for HBCD.

489

HBCD stereoisomers have been demonstrated to exhibit different biological

490

activities, thus resulting in different toxic effects on organisms. The stereoisomers

491

may interact additively, synergistically or antagonistically, even within the same

492

toxicity endpoint. Accordingly, the TCCs of tHBCD used in the preliminary hazard

493

evaluation may overestimate or underestimate the actual risks posed by ΣHBCD (a

494

mixture of varying proportions of HBCD stereoisomers) for porpoises and dolphins.

495

Different cytotoxicities for the three pairs of HBCD enantiomers were demonstrated

496

on human HepG2 cells (SI Table S9),32 and a set of stereoisomer-based threshold

497

values were derived based on the difference in toxic potency among the enantiomers

498

and were used for comparison with the MECmax values of individual HBCD

499

enantiomers in blubber.

500

Among the MECmax values of six HBCD enantiomers in porpoise blubber (SI Table

501

S10), only the MECmax value of (+)-α-HBCD (1,030 ng/g ww) exceeded the

502

significant toxic level of the identical enantiomer in terms of the increase of LDH

503

leakage. Similar to the result of porpoises, only the MECmax value of (+)-α-HBCD in

504

dolphin blubber (3,330 ng/g ww) exceeded the significant toxic level. Nevertheless,

505

the MECmax value of (−)-α-HBCD in dolphin blubber (4,750 ng/g ww) was very close 21



506

to the significant level of the identical enantiomer. As suggested by Zhang et al.

507

(2008), the γ-diastereomer was likely to exert the most significant toxicological

508

effects.32 Based on the above findings, the actual risks posed by ΣHBCD may be

509

overestimated, as indicated by using the HQ approach in the present study.

510

Likewise, the threshold values of the cytotoxicity for TBECH diastereomers to

511

human HepG2 cells (SI Table S11) could be used for comparison with the MECmax

512

values of individual TBECH diastereomers in blubber. The potency of TBECH

513

diastereomers in the activation of the human androgen receptor was indicated by

514

Khalaf et al. (2009) to follow the order of γ-, δ-, α-, and β-TBECH.14 Among the

515

MECmax values of four TBECH diastereomers in cetaceans (SI Table S10), the

516

MECmax value of γ-TBECH in porpoise blubber exceeded the EC50 value of 50:50

517

γ/δ-TBECH, while the MECmax values of γ- and δ-TBECH in dolphin blubber were

518

much higher than the EC50 values of 50:50 γ/δ-TBECH and 25:75 γ/δ-TBECH,

519

respectively. Given that γ-TBECH appears to be the most potent human androgen

520

receptor activator among the four TBECH diastereomers, its EC50 value should be

521

lower than that of 50:50 γ/δ-TBECH. For this reason, porpoises and dolphins may

522

suffer a certain hazard level in terms of androgen agonism from γ-TBECH exposure,

523

and this risk posed by exposure to ΣTBECH may increase for the marine mammals in

524

the studied region.

525

According to the above results, the influence of stereoisomerism on the

526

environmental fate and effects of chiral chemicals can introduce substantial

527

uncertainty into a risk assessment of their potential hazards. Further investigations of

528

the exposure of these marine mammals to chiral contaminants could focus on whether

529

a specific enantioselective trend matches that of their primary prey.

22


Page 22 of 37

Page 23 of 37


530

ASSOCIATED CONTENT

531

Supporting Information. Additional information on materials and methods.

532

Tables containing biological information for individual porpoises and dolphins,

533

recoveries of each native standard and surrogate, method LODs and detection

534

frequencies of each stereoisomer, concentrations of ΣHBCD and ΣTBECH in

535

biological tissues reported, and toxicological effects caused by HBCD and TBECH

536

stereoisomers. Figures showing data of the annual production of tHBCD in China,

537

temporal distributions in adult males, compositions of HBCD diastereomers in

538

porpoise blubber, and EFs of TBECH categorized by different physiological traits.

539

ACKNOWLEDGEMENTS

540

This project was supported by the General Research Fund (CityU 11100614 and

541

11338216) and Early Career Scheme (EdUHK 28300317) from the Hong Kong

542

Research Grants Council. The authors thank the Agriculture, Fisheries, and

543

Conservation Department, Hong Kong SAR Government and Ocean Park

544

Conservation Foundation, Hong Kong (OPCFHK) for their assistance with blubber

545

collection.

546

23



547

REFERENCES

548

(1)

Toms, L. M. L.; Hearn, L.; Kennedy, K.; Harden, F.; Bartkow, M.; Temme, C.;

549

Mueller, J. F. Concentrations of polybrominated diphenyl ethers (PBDEs) in

550

matched samples of human milk, dust and indoor air. Environ. Int. 2009, 35 (6),

551

864–869.

552

(2)

UNEP. Guidance for the inventory of polybrominated diphenyl ethers

553

( PBDEs ) listed under the Stockholm Convention on Persistent Organic

554

Pollutants. United Nations Environ. Program. 2012.

555

(3)

ethers in China: An overview. Chemosphere 2012, 88 (7), 769–778.

556 557

Ma, J.; Qiu, X.; Zhang, J.; Duan, X.; Zhu, T. State of polybrominated diphenyl

(4)

Verhoeven, J. K.; Bakker, J.; Bruinen de Bruin, Y.; Hogendoorn, E. A.; de

558

Knecht, J. A.; de Knecht, W. J. G. M.; Posthuma, L.; Struijs, J.; Vermeire, T.

559

G.; van Wijnen, H. J.; et al. From risk assessment to environmental impact

560

assessment of chemical substances

561

http://www.rivm.nl/bibliotheek/rapporten/601353002.pdf.

562

(5)

Andersson, P. L.; Oberg, K.; Orn, U. Chemical characterization of brominated

563

flame retardants and identification of structurally representative compounds.

564

Environ. Toxicol. Chem. 2006, 25 (5), 1275–1282.

565

(6)

Gemmill, B.; Pleskach, K.; Peters, L.; Palace, V.; Wautier, K.; Park, B.;

566

Darling, C.; Rosenberg, B.; McCrindle, R.; Tomy, G. T. Toxicokinetics of

567

tetrabromoethylcyclohexane (TBECH) in juvenile brown trout (Salmo trutta)

568

and effects on plasma sex hormones. Aquat. Toxicol. 2011, 101 (2), 309–317.

569 570

(7)

Rani, M.; Shim, W. J.; Han, G. M.; Jang, M.; Song, Y. K.; Hong, S. H. Hexabromocyclododecane in polystyrene based consumer products: An 24


Page 24 of 37

Page 25 of 37


evidence of unregulated use. Chemosphere 2014, 110, 111–119.

571 572

(8)

Koch, C.; Schmidt-Kötters, T.; Rupp, R.; Sures, B. Review of

573

hexabromocyclododecane (HBCD) with a focus on legislation and recent

574

publications concerning toxicokinetics and -dynamics. Environ. Pollut. 2015,

575

199, 26–34.

576

(9)

Bao, L. J.; Wei, Y. L.; Yao, Y.; Ruan, Q. Q.; Zeng, E. Y. Global trends of

577

research on emerging contaminants in the environment and humans: A

578

literature assimilation. Environ. Sci. Pollut. Res. 2015, 22 (3), 1635–1643.

579

(10)

UNEP. SC-6 / 13 : Listing of hexabromocyclododecane. Stockholm Convention

580

Organic on Persistent Pollutants: Persistent Organic Pollutants Review

581

Committee Seventh meeting. 2013.

582

(11)

Li, L.; Weber, R.; Liu, J.; Hu, J. Long-term emissions of

583

hexabromocyclododecane as a chemical of concern in products in China.

584

Environ. Int. 2016, 91, 291–300.

585

(12)

sediment of the Great Lakes. Environ. Sci. Technol. 2012, 46 (6), 3119–3126.

586 587

Yang, R.; Wei, H.; Guo, J.; Li, A. Emerging brominated flame retardants in the

(13)

Tomy, G. T.; Pleskach, K.; Arsenault, G.; Potter, D.; Mccrindle, R.; Marvin, C.

588

H.; Sverko, E.; Tittlemier, S. Identification of the novel cycloaliphatic

589

brominated flame retardant 1,2-dibromo4-(1,2-dibromoethyl)cyclohexane in

590

Canadian Arctic beluga (Delphinapterus leucas). Environ. Sci. Technol. 2008,

591

42 (2), 543–549.

592

(14)

Khalaf, H.; Larsson, A.; Berg, H.; McCrindle, R.; Arsenault, G.; Olsson, P. E.

593

Diastereomers of the brominated flame retardant 1,2-dibromo-4-(1,2

594

dibromoethyl)cyclohexane induce androgen receptor activation in the HepG2 25



595

hepatocellular carcinoma cell line and the LNCaP prostate cancer cell line.

596

Environ. Health Perspect. 2009, 117 (12), 1853–1859.

597

(15)

Cequier, E.; Ionas, A. C.; Covaci, A.; Marcé, R. M.; Becher, G.; Thomsen, C.

598

Occurrence of a broad range of legacy and emerging flame retardants in indoor

599

environments in Norway. Environ. Sci. Technol. 2014, 48 (12), 6827–6835.

600

(16)

Newton, S.; Sellström, U.; De Wit, C. A. Emerging flame retardants, PBDEs,

601

and HBCDDs in indoor and outdoor media in Stockholm, Sweden. Environ. Sci.

602

Technol. 2015, 49 (5), 2912–2920.

603

(17)

Tao, F.; Abdallah, M. A.-E.; Harrad, S. Emerging and legacy flame retardants

604

in UK indoor air and dust: Evidence for replacement of PBDEs by emerging

605

flame retardants? Environ. Sci. Technol. 2016, 50 (23), 13052–13061.

606

(18)

Wong, C. S. Environmental fate processes and biochemical transformations of

607

chiral emerging organic pollutants. Anal. Bioanal. Chem. 2006, 386 (3), 544–

608

558.

609

(19)

Arsenault, G.; Konstantinov, A.; Marvin, C. H.; MacInnis, G.; McAlees, A.;

610

McCrindle, R.; Riddell, N.; Tomy, G. T.; Yeo, B. Synthesis of the two minor

611

isomers, δ-and ε-1,2,5,6,9,10-hexabromocyclododecane, present in commercial

612

hexabromocyclododecane. Chemosphere 2007, 68 (5), 887–892.

613

(20)

Arsenault, G.; Lough, A.; Marvin, C.; McAlees, A.; McCrindle, R.; MacInnis,

614

G.; Pleskach, K.; Potter, D.; Riddell, N.; Sverko, E.; et al. Structure

615

characterization and thermal stabilities of the isomers of the brominated flame

616

retardant 1,2-dibromo-4-(1,2-dibromoethyl)cyclohexane. Chemosphere 2008,

617

72 (8), 1163–1170.

618

(21)

Peled, M.; Scharia, R.; Sondack, D. Thermal rearrangement of 26


Page 26 of 37

Page 27 of 37


hexabromo-cyclododecane (HBCD). Ind. Chem. Libr. 1995, 7 (C), 92–99.

619 620

(22)

Heeb, N. V.; Schweizer, W. B.; Kohler, M.; Gerecke, A. C. Structure

621

elucidation of hexabromocyclododecanes - A class of compounds with a

622

complex stereochemistry. Chemosphere 2005, 61 (1), 65–73.

623

(23)

Koeppen, R.; Becker, R.; Emmerling, F.; Jung, C.; Nehls, I. Enantioselective

624

preparative HPLC separation of the HBCD-stereoisomers from the technical

625

product and their absolute structure elucidation using X-ray crystallography.

626

Chirality 2007, 19 (3), 214–222.

627

(24)

Heeb, N. V.; Bernd Schweizer, W.; Mattrel, P.; Haag, R.; Gerecke, A. C.;

628

Schmid, P.; Zennegg, M.; Vonmont, H. Regio- and stereoselective

629

isomerization of hexabromocyclododecanes (HBCDs): Kinetics and

630

mechanism of γ- to α-HBCD isomerization. Chemosphere 2008, 73 (8), 1201–

631

1210.

632

(25) Marvin, C. H.; Tomy, G. T.; Armitage, J. M.; Arnot, J. A.; McCarty, L.; Covaci,

633

A.; Palace, V. Hexabromocyclododecane: Current understanding of chemistry,

634

environmental fate and toxicology and implications for global management.

635

Environ. Sci. Technol. 2011, 45 (20), 8613–8623.

636

(26)

Han, Q.; Song, H.; Gao, S.; Zeng, X.; Yu, Z.; Yu, Y.; Sheng, G.; Fu, J.

637

Determination of ten hexabromocyclododecane diastereoisomers using two

638

coupled reversed-phase columns and liquid chromatography/tandem mass

639

spectrometry. Rapid Commun. Mass Spectrom. 2014, 28 (13), 1473–1478.

640

(27)

Sun, Y. X.; Luo, X. J.; Mo, L.; He, M. J.; Zhang, Q.; Chen, S. J.; Zou, F. S.;

641

Mai, B. X. Hexabromocyclododecane in terrestrial passerine birds from

642

e-waste, urban and rural locations in the Pearl River Delta, South China: Levels,

27



643

biomagnification, diastereoisomer- and enantiomer-specific accumulation.

644

Environ. Pollut. 2012, 171, 191–198.

645

(28)

Huber, S.; Warner, N. A.; Nygård, T.; Remberger, M.; Harju, M.; Uggerud, H.

646

T.; Kaj, L.; Hanssen, L. A broad cocktail of environmental pollutants found in

647

eggs of three seabird species from remote colonies in Norway. Environ.

648

Toxicol. Chem. 2015, 34 (6), 1296–1308.

649

(29)

Fan, X.; Kubwabo, C.; Rasmussen, P. E.; Wu, F. Non-PBDE halogenated

650

flame retardants in Canadian indoor house dust: Sampling, analysis, and

651

occurrence. Environ. Sci. Pollut. Res. 2016, 23 (8), 7998–8007.

652

(30)

Chen, D.; Hale, R. C.; Letcher, R. J. Photochemical and microbial

653

transformation of emerging flame retardants: Cause for concern? Environ.

654

Toxicol. Chem. 2015, 34 (4), 687–699.

655

(31)

Hamers, T.; Kamstra, J. H.; Sonneveld, E.; Murk, A. J.; Kester, M. H. A.;

656

Andersson, P. L.; Legler, J.; Brouwer, A. In vitro profiling of the

657

endocrine-disrupting potency of brominated flame retardants. Toxicol. Sci.

658

2006, 92 (1), 157–173.

659

(32)

Zhang, X.; Yang, F.; Xu, C.; Liu, W.; Wen, S.; Xu, Y. Cytotoxicity evaluation

660

of three pairs of hexabromocyclododecane (HBCD) enantiomers on Hep G2

661

cell. Toxicol. Vitr. 2008, 22 (6), 1520–1527.

662

(33)

Palace, V. P.; Pleskach, K.; Halldorson, T.; Danell, R.; Wautier, K.; Evans, B.;

663

Alaee, M.; Marvin, C.; Tomy, G. T. Biotransformation enzymes and thyroid

664

axis disruption in juvenile rainbow trout (Oncorhynchus mykiss) exposed to

665

hexabromocyclododecane diastereoisomers. Environ. Sci. Technol. 2008, 42

666

(6), 1967–1972.

28


Page 28 of 37

Page 29 of 37


667

(34)

Du, M.; Fang, C.; Qiu, L.; Dong, S.; Zhang, X.; Yan, C.

668

Diastereoisomer-specific effects of hexabromocyclododecanes on hepatic aryl

669

hydrocarbon receptors and cytochrome P450s in zebrafish (Danio rerio).

670

Chemosphere 2015, 132, 24–31.

671

(35)

Pradhan, A.; Kharlyngdoh, J. B.; Asnake, S.; Olsson, P. E. The brominated

672

flame retardant TBECH activates the zebrafish (Danio rerio) androgen receptor,

673

alters gene transcription and causes developmental disturbances. Aquat. Toxicol.

674

2013, 142–143, 63–72.

675

(36)

Fair, P. A.; Adams, J.; Mitchum, G.; Hulsey, T. C.; Reif, J. S.; Houde, M.;

676

Muir, D.; Wirth, E.; Wetzel, D.; Zolman, E.; et al. Contaminant blubber

677

burdens in Atlantic bottlenose dolphins (Tursiops truncatus) from two

678

southeastern US estuarine areas: Concentrations and patterns of PCBs,

679

pesticides, PBDEs, PFCs, and PAHs. Sci. Total Environ. 2010, 408 (7), 1577–

680

1597.

681

(37)

Hall, A. J.; Thomas, G. O.; Mcconnell, B. J. Exposure to persistent organic

682

pollutants and first-year survival probability in cray seal pups. Environ. Sci.

683

Technol. 2009, 43 (16), 6364–6369.

684

(38)

Zhu, B.; Lai, N. L. S.; Wai, T. C.; Chan, L. L.; Lam, J. C. W.; Lam, P. K. S.

685

Changes of accumulation profiles from PBDEs to brominated and chlorinated

686

alternatives in marine mammals from the South China Sea. Environ. Int. 2014,

687

66, 65–70.

688

(39)

Lam, J. C. W.; Lau, R. K. F.; Murphy, M. B.; Lam, P. K. S. Temporal trends of

689

hexabromocyclododecanes (HBCDs) and polybrominated diphenyl ethers

690

(PBDEs) and detection of two novel flame retardants in marine mammals from

29



Hong Kong, South China. Environ. Sci. Technol. 2009, 43 (18), 6944–6949.

691 692

(40)

Harner, T.; Wiberg, K.; Norstrom, R. Enantiomer fractions are preferred to

693

enantiomer ratios for describing chiral signatures in environmental analysis.

694

Environ. Sci. Technol. 2000, 34 (1), 218–220.

695

(41)

Barros, N. B.; Thomas A. Jefferson Parsons, and E. C. M. Feeding habits of

696

Indo-Pacific humpback dolphins (Sousa chinensis) stranded in Hong Kong.

697

Aquat. Mamm. 2004, 30 (1), 179–188.

698

(42)

Barros, N. B.; Jefferson, T. A.; Parsons, E. C. M. Food habits of finless

699

porpoises (Neophocaena Phocaenoides) in Hong Kong waters. Raffles Bull.

700

Zool. 2002, No. SUPPL. 10, 115–123.

701

(43)

Nyholm, J. R.; Norman, A.; Norrgren, L.; Haglund, P.; Andersson, P. L.

702

Uptake and biotransformation of structurally diverse brominated flame

703

retardants in zebrafish (Danio rerio) after dietary exposure. Environ. Toxicol.

704

Chem. 2009, 28 (5), 1035–1042.

705

(44)

Kajiwara, N.; Kamikawa, S.; Amano, M.; Hayano, A.; Yamada, T. K.;

706

Miyazaki, N.; Tanabe, S. Polybrominated diphenyl ethers (PBDEs) and

707

organochlorines in melon-headed whales, Peponocephala electra, mass

708

stranded along the Japanese coasts: Maternal transfer and temporal trend.

709

Environ. Pollut. 2008, 156 (1), 106–114.

710

(45)

Yordy, J. E.; Wells, R. S.; Balmer, B. C.; Schwacke, L. H.; Rowles, T. K.;

711

Kucklick, J. R. Life history as a source of variation for persistent organic

712

pollutant (POP) patterns in a community of common bottlenose dolphins

713

(Tursiops truncatus) resident to Sarasota Bay, FL. Sci. Total Environ. 2010,

714

408 (9), 2163–2172.

30


Page 30 of 37

Page 31 of 37


715

(46)

Park, B. K.; Park, G. J.; An, Y. R.; Choi, H. G.; Kim, G. B.; Moon, H. B.

716

Organohalogen contaminants in finless porpoises (Neophocaena phocaenoides)

717

from Korean coastal waters: Contamination status, maternal transfer and

718

ecotoxicological implications. Mar. Pollut. Bull. 2010, 60 (5), 768–774.

719

(47) Moon, H. B.; Kannan, K.; Choi, M.; Yu, J.; Choi, H. G.; An, Y. R.; Choi, S. G.;

720

Park, J. Y.; Kim, Z. G. Chlorinated and brominated contaminants including

721

PCBs and PBDEs in minke whales and common dolphins from Korean coastal

722

waters. J. Hazard. Mater. 2010, 179 (1–3), 735–741.

723

(48)

Dodder, N. G.; Peck, A. M.; Kucklick, J. R.; Sander, L. C. Analysis of

724

hexabromocyclododecane diastereomers and enantiomers by liquid

725

chromatography/tandem mass spectrometry: Chromatographic selectivity and

726

ionization matrix effects. J. Chromatogr. A 2006, 1135 (1), 36–42.

727

(49)

Marvin, C. H.; Macinnis, G.; Alaee, M.; Arsenault, G.; Tomy, G. T. Factors

728

influencing enantiomeric fractions of hexabromocyclododecane measured

729

using liquid chromatography/tandem mass spectrometry. Rapid Commun. Mass

730

Spectrom. 2007, 21, 1925–1930.

731

(50)

Wong, F.; Kurt-Karakus, P.; Bidleman, T. F. Fate of brominated flame

732

retardants and organochlorine pesticides in urban soil: Volatility and

733

degradation. Environ. Sci. Technol. 2012, 46 (5), 2668–2674.

734

(51)

Peck, A. M.; Pugh, R. S.; Moors, A.; Ellisor, M. B.; Porter, B. J.; Becker, P. R.;

735

Kucklick, J. R. Hexabromocyclododecane in white-sided dolphins: Temporal

736

trend and stereoisomer distribution in tissues. Environ. Sci. Technol. 2008, 42

737

(7), 2650–2655.

738

(52)

Vorkamp, K.; Bester, K.; Rigét, F. F. Species-specific time trends and

31



739

enantiomer fractions of hexabromocyclododecane (HBCD) in biota from east

740

greenland. Environ. Sci. Technol. 2012, 46 (19), 10549–10555.

741

(53)

Fisk, A. T.; Holst, M.; Hobson, K. A.; Duffe, J.; Moisey, J.; Norstrom, R. J.

742

Persistent organochlorine contaminants and enantiomeric signatures of chiral

743

pollutants in ringed seals (Phoca hispida) collected on the east and west side of

744

the Northwater Polynya, Canadian Arctic. Arch. Environ. Contam. Toxicol.

745

2002, 42 (1), 118–126.

746

(54)

Hoekstra, P. F.; Wong, C. S.; O’Hara, T. M.; Solomon, K. R.; Mabury, S. A.;

747

Muir, D. C. G. Enantiomer-specific accumulation of PCB atropisomers in the

748

bowhead whale (Balaena mysticetus). Environ. Sci. Technol. 2002, 36 (7),

749

1419–1425.

750

(55)

Orr, A. J.; Newsome, S. D.; Laake, J. L.; Vanblaricom, G. R.; Delong, R. L.

751

Ontogenetic dietary information of the California sea lion (Zalophus

752

californianus) assessed using stable isotope analysis. Mar. Mammal Sci. 2012,

753

28 (4), 714–732.

754

(56)

Cadieux, M. A.; Muir, D. C. G.; Béland, P.; Hickie, B. E. Lactational transfer

755

of polychlorinated-biphenyls (PCBs) and other organochlorines in St.

756

Lawrence beluga whales (Delphinapterus leucas). Arch. Environ. Contam.

757

Toxicol. 2016, 70 (1), 169–179.

758

(57)

Peterson, S. H.; McHuron, E. A.; Kennedy, S. N.; Ackerman, J. T.; Rea, L. D.;

759

Castellini, J. M.; O’Hara, T. M.; Costa, D. P. Evaluating hair as a predictor of

760

blood mercury: The influence of ontogenetic phase and life history in pinnipeds.

761

Arch. Environ. Contam. Toxicol. 2016, 70 (1), 28–45.

762

(58)

Saegusa, Y.; Fujimoto, H.; Woo, G. H.; Inoue, K.; Takahashi, M.; Mitsumori,

32


Page 32 of 37

Page 33 of 37


763

K.; Hirose, M.; Nishikawa, A.; Shibutani, M. Developmental toxicity of

764

brominated flame retardants, tetrabromobisphenol A and

765

1,2,5,6,9,10-hexabromocyclododecane, in rat offspring after maternal exposure

766

from mid-gestation through lactation. Reprod. Toxicol. 2009, 28 (4), 456–467.

767

(59)

Curran, I. H. A.; Liston, V.; Nunnikhoven, A.; Caldwell, D.; Scuby, M. J. S.;

768

Pantazopoulos, P.; Rawn, D. F. K.; Coady, L.; Armstrong, C.; Lefebvre, D. E.;

769

et al. Toxicologic effects of 28-day dietary exposure to the flame retardant

770

1,2-dibromo-4-(1,2-dibromoethyl)-cyclohexane (TBECH) in F344 rats.

771

Toxicology 2017, 377, 1–13.

772

(60)

Lam, J. C. W.; Lam, P. K. S. Occurrence and ecological risk of halogenated

773

flame retardants (HFRs) in coastal zones. In Persistent Organic Pollutants

774

(POPs): Analytical Techniques, Environmental Fate and Biological Effects;

775

Zeng, E. Y., Ed.; Elsevier, 2015; Vol. 67, pp 389–409.

776

(61)

Yordy, J. E.; Pabst, D. A.; Mclellan, W. A.; Wells, R. S.; Rowles, T. K.;

777

Kucklickz, J. R. Tissue-specific distribution and whole-body burden estimates

778

of persistent organic pollutants in the bottlenose dolphin (Tursiops truncatus).

779

Environ. Toxicol. Chem. 2010, 29 (6), 1263–1273.

780

(62) Tanabe, S.; Tatsukawa, R.; Tanaka, H.; Maruyama, K.; Miyazaki, N.; Fujiyama,

781

T. Distribution and total burdens of chlorinated hydrocarbons in bodies of

782

striped dolphins (Stenella coeruleoalba). Agric. Biol. Chem. 1981, 45 (11),

783

2569–2578.

784 785

33



Page 34 of 37

786

Table 1. The concentrations (displayed with three significant digits) of ΣHBCD and ΣTBECH in the blubber of porpoises and dolphins

787

categorized by these marine mammals’ age and sex [shown as mass fractions (ng/g lipid weight)]. The structural formula of each analyte is

788

shown in the first column.

Analyte

ΣHBCD

Species (n)

Porpoise (59)

Dolphin (32)

ΣTBECH

Porpoise (59)

Dolphin (32)

Age Class

Minimum

Median

Maximum

Mean ± SD

Adult male

394

1,780

6,260

2,010 ± 1,460

Adult female

97.2

1,150

5,110

1,670 ± 1,590

Juvenile

129

509

1,720

657 ± 426

Adult male

447

2,650

12,600

4,590 ± 4,050

Adult female

997

5,410

45,800

11,500 ± 12,500

Juvenile

1,760

3,290

26,900

5,910 ± 6,640

Adult male

< 0.8

45.4

125

43.6 ± 31.5

Adult female

< 0.8

11.2

93.5

21.4 ± 28.7

Juvenile

< 0.8

19.5

60.2

21.2 ± 22.5

Adult male

< 0.8

33.5

193

56.2 ± 58.2

Adult female

8.08

101

341

110 ± 92.0

Juvenile

< 0.8

16.7

362

81.7 ± 110

789 34


Page 35 of 37


790 791

Figure 1. Stereoisomeric compositions of (A) the five HBCD diastereomers, (B) the

792

four TBECH diastereomers in the blubber of porpoises and dolphins (“*”: p < 0.05;

793

“**”: p < 0.01; “***”: p < 0.001).

794

35



795 796

Figure 2. EFs of (A) the three pairs of HBCD enantiomers, (B) the four pairs of

797

TBECH enantiomers in the blubber of porpoises and dolphins, respectively (“*”: p

Temporal Changes and Stereoisomeric Compositions of 1,2,5,6,9,10

Recommend Documents