DBDPE

Article Cite This: Environ. Sci. Technol. 2019, 53, 8437−8446

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Bioconcentration, Biotransformation, and Thyroid Endocrine Disruption of Decabromodiphenyl Ethane (Dbdpe), A Novel Brominated Flame Retardant, in Zebrafish Larvae Xiaochen Wang,†,‡,⊥ Siyuan Ling,§,⊥ Kelan Guan,∥ Xiaojun Luo,∥ Lianguo Chen,† Jian Han,† Wei Zhang,§ Bixian Mai,∥ and Bingsheng Zhou*,†

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†

State Key Laboratory of Freshwater Ecology and Biotechnology, Institute of Hydrobiology, Chinese Academy of Sciences, Wuhan 430072, China ‡ University of Chinese Academy of Sciences, Beijing 100049, China § State Environmental Protection Key Laboratory of Environmental Risk Assessment and Control on Chemical Process, School of Resource and Environmental Engineering, East China University of Science and Technology, Shanghai 200237, China ∥ State Key Laboratory of Organic Geochemistry, Guangzhou Institute of Geochemistry, Chinese Academy of Sciences, Guangzhou 510640, P.R. China S Supporting Information *

ABSTRACT: The brominated flame retardant decabromodiphenyl ethane (DBDPE), an alternative to decabrominated diphenyl ether (BDE209), has become a widespread environmental contaminant, but its possible toxic effects to wildlife remain unknown. Using zebrafish as a model, we investigated the bioconcentration and impact of DBDPE on thyroid endocrine function after water-borne exposure, compared to BDE209. Zebrafish embryos were exposed to DBDPE or BDE209 (0, 3, 10, 30, 100, 300 nM) for 6 or 14 days. Chemical analysis revealed that DBDPE and BDE209 were bioconcentrated in zebrafish larvae, with similar magnitudes of accumulated concentrations. Based on screened by chromatograms, at least seven unknown compounds were observed in DBDPE-treated larvae, indicating biotransformation of the chemical. Significant increases in whole body content of triiodothyronine (T3) and thyroxine (T4) were detected in DBDPE-treated larvae, but decreased in BDE209-treated groups. Alterations in gene transcription along the related hypothalamic-pituitary-thyroid (HPT) axis were observed. Furthermore, the binding and transport protein transthyretin (TTR) was significantly increased in DBDPE exposure groups. Histological examination and stereological analysis showed no obvious pathological changes in the thyroid gland. The present study demonstrates for the first time the bioavailability, biotransformation and thyroid endocrine disruption associated with DBDPE exposure in fish. Further studies are warranted to identify the metabolites of DBDPE and to define its environmental risks to aquatic organisms.

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INTRODUCTION

Such restricted usage of PBDEs has led to an increase in the demand for alternative or novel BFRs (NBFRs). Among the most common NBFRs is decabromodiphenyl ethane (DBDPE), which has been used as an alternative for the legacy commercial deca-BDE (BDE209), and it has been produced and used in increasing amounts.4 In China, DBDPE is the most popular NBFRs and has been produced since 2005, with production increasing at 80% per year.5 Recent studies reported that the estimated production of DBDPE was 11 000

Brominated flame retardants (BFRs) are classified as high production volume chemicals used in the plastics and textile industries.1 Polybrominated diphenyl ethers (PBDEs) were widely added to consumer products such as plastics, textiles, furniture, and electric/electronic devices. In the past 15 years, increasing evidence has demonstrated persistent, bioaccumulative, and toxic effects of PBDEs to wildlife and humans in both aquatic and terrestrial environments.2,3 Due to their high ecological and health risks, commercial PBDEs have been globally listed as persistent organic pollutants (POPs) by the Stockholm Convention (http://chm.pops.int/ TheConvention/ThePOPs/). © 2019 American Chemical Society

Received: Revised: Accepted: Published: 8437

May 11, 2019 June 9, 2019 June 12, 2019 June 12, 2019 DOI: 10.1021/acs.est.9b02831 Environ. Sci. Technol. 2019, 53, 8437−8446

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Environmental Science & Technology t in 2006, and increased to 31 000 t in 2016.6 The domestic production of DBDPE from 2006 to 2016 was estimated at approximately 230 000 t.6 Furthermore, the annual production of DBDPE has exceeded that of deca-BDE since 2012, and may have exceeded the volume of PBDEs.6,7 The large-scale use of DBDPE has led to rapid growth of ubiquitous contamination in environmental medium, and among wild life and humans worldwide.4,5,8 Environmental monitoring indicated that DBDPE was the predominant NBFR, while the concentrations of DBDPE overwhelmed replacement of BDE209 in many samples.9 DBDPE is a highly hydrophobic chemical (LogKow = 11.1) and has similar properties to BDE209.4 Hence, environmental concentrations in surface water are generally low (96% purity) was purchased from Tokyo Chemical Industry Co., Ltd. (Tokyo, Japan); Decabrominated diphenyl ethers (BDE209, CAS #1163−19−5, >98% purity) was purchased from Wellington Laboratories (Ontario, Canada). The dimethyl sulfoxide (DMSO) used to make the stock solutions of DBDPE and BDE209 (0.1% v/v) was purchased from Sigma-Aldrich (St. Louis, MO). The methanesulfonate (MS-222) was purchased from SigmaAldrich (St. Louis, MO). DBDPE standard and BDE209 standard used in chemical analysis were purchased from AccuStandard, Inc. (New Haven, CT). All solvents were of HPLC grade and other chemicals used were of analytical grade. Zebrafish Maintenance and Embryo Exposure. The culture of adult zebrafish (wild type, AB strain, four-monthold) and embryo exposure was carried out following the method described previously.22 Briefly, embryos that had developed normally and reached the blastula stage (2 h post fertilization, hpf) were selected for chemical exposure. Approximately 500 normal embryos were randomly distributed into glass beakers containing 500 mL of DBDPE and BDE209 exposure solutions. There were four replicates for control and DBDPE (0, 3, 10, 30, 100, and 300 nM, or equal to 0, 2.91, 9.71, 29.10, 97.10, and 291.36 μg/L) or BDE209 (0, 3, 10, 30, 100, and 300 nM, or equal to 0, 2.88, 9.59, 28.80, 95.92, and 287.76 μg/L) exposure groups. The selected highest exposure concentration was based on a previous study, where zebrafish embryos exposed to 80 μg/L BDE209 demonstrated significant alteration of thyroid hormone and HPT axis gene transcription.23 Both the control and chemical treatment groups received 0.1% (v/v) DMSO. The exposure solution was renewed daily throughout the experimental period, and the temperature was kept at 28 ± 0.5 °C under a 14 h light:10 h dark cycle. The zebrafish larvae were fed with cultured live paramecia and Artemia twice daily. The larvae were randomly sampled at 6 and 14 dpf, anesthetized with 0.03% MS-222 (300 mg/L), immediately frozen in liquid nitrogen, and stored at −80 °C for subsequent gene, protein, and TH assays. A subset of larvae were analyzed for bioconcentration and metabolites analysis at 6 and 14 days. At 14 dpf, a subset of larvae were fixed in paraformaldehyde for histological examination. Details related to hatching, malformation, growth, and survival were also recorded. All studies were conducted in accordance with the Guidelines for the Care and Use of Laboratory Animals of the National Institute for Food and Drug Control of China. 8438

DOI: 10.1021/acs.est.9b02831 Environ. Sci. Technol. 2019, 53, 8437−8446

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Environmental Science & Technology

Statistical Analysis. All data were initially verified for normality and homogeneity of variance using the Kolmogorov−Smirnov and Levene’s tests, respectively. The differences between the control and each exposure group were evaluated by one-way analysis of variance (ANOVA) followed by Tukey’s test. All analyses were performed with SPSS 16.0 (SPSS, Chicago, IL). The data are expressed as means ± standard error (SEM). P < 0.05 was considered statistically significant.

Quantification of DBDPE and BDE209 in Exposure Solutions and in Zebrafish Larvae. DBDPE and BDE209 concentrations in the exposure solutions and in larvae (6 dpf and 14 dpf, n = 3 replicates) were extracted and analyzed as described previously.23,24 The identification and quantification of parent compound and metabolites of both DBDPE and BDE209 were performed using an Agilent 7890A-5975C gas chromatograph−mass spectrometer (GC−MS) with an electron-capture negative ionization mode (ECNI) in the selected ion monitoring mode. Concentrations of analytes are expressed as nM in exposure solution or μg/g d.w. (dry weight) in larvae. The detailed protocols for extraction, clean up, analysis, and quality assurance and quality control (QA/ QC) are provided in the Supporting Information (SI) Text S1. The limit of detection (LOD) was 1 ng/g for BDE-209 and 10 ng/g for DBDPE. The limit of quantification (LOQ) was 10 ng/g for BDE-209 and 15 ng/g for DBDPE, respectively. RNA Isolation and Quantitative Real-Time Polymerase Chain Reaction (qRT-PCR) Assay. Larval samples (14 dpf; 20 larvae; n = 4 replicates) were collected and preserved with RNAiso Plus reagent according to the manufacturer’s instructions (Invitrogen, Carlsbad, CA). The total RNA extraction and purification, first-strand cDNA synthesis, and quantitative real-time PCR were carried out following the protocols as previously described.22 The primer sequences of the target genes were obtained from the literature or identified using the online Primer3 software (http://frodo.wi.mit.edu/; see SI Table S1). The housekeeping gene β-actin was selected as a reference and was determined to be stable in response to DBDPE and BDE209 exposure. The gene transcription levels were normalized to that of β-actin using the 2−ΔΔCt method. Protein Extraction and Western Blot Analysis. In the study, the expression of transthyretin (TTR) protein was selected for analysis. A Western blot was carried out following previously described methods23 using approximately 100 larvae (6 dpf and 14 dpf, 50 μg protein/sample) for each sample (n = 4 replicates). A detailed procedure is provided in SI Text S2. A quantitative measure of protein expression was obtained by densitometry, with the results normalized to GADPH expression. The sheep transthyretin antibody (Abcam, Cambridge, UK) has previously been verified to be reactive and suitable for zebrafish studies.23 Thyroid Hormone Extraction and Measurement. The methods for extraction of whole body THs content in larvae (6 dpf and 14 dpf, 200 larvae/sample, n = 4 replicates) were as previously described22 and a detailed procedure is provided in SI Text S3. The total T4 and T3 levels were measured using commercial enzyme-linked immunosorbent assay (ELISA) test kits purchased from Wuhan EIAab Science Co., Ltd. (Wuhan, China) following the manufacturer’s instructions. The intraassay variations reported by the manufacturer are 4.3% for total T4 and 4.5% for total T3, and the interassay variations for total T4 and total T3 are 7.5% and 7.2%, respectively. Thyroid Histology and Stereological Analysis. The 14 dpf larvae (six larvae per treatment, n = 4 replicates) were fixed in 4% paraformaldehyde and embedded in paraffin for subsequent serial sections of whole fish at 3 μm using a rotary microtome (Leica RM2125, Solms, Germany). The sections were stained with hematoxylin and eosin (H&E), and imaged under an Olympus CX31 light microscope (Olympus, Tokyo, Japan) for examination of morphological alterations. The surface area of each thyroid follicle was estimated using ImageJ software (Media Cybernetics, Bethesda, MD).

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RESULTS Developmental Toxicity. After 6 and 14 days of exposure, either DBDPE or BDE-209 did not significantly affect hatching, malformation, or survival rates compared with those of controls (SI Table S2). The 96 h hatching rates were over 90% in all groups, and the survival rates were over 82% and over 69% at 6 dpf and 14 dpf, respectively, in all groups. There was no significant difference in growth (weight) after 6 days exposure, but a significant decrease in growth was observed in the 300 nM DBDPE treatment group after 14 days. Concentrations in Exposure Solution and Bioconcentration in Larvae. The concentrations of DBDPE and BDE209 were measured just after (T0) and before (T24) renewal of exposure solutions. The measured mean concentrations of DBDPE in the exposure water just after renewing the exposure solutions (T0) were 2.95, 8.38, 20.93, 49.71, and 145.08 nM in the 3, 10, 30, 100, and 300 nM exposure groups, respectively. After 24 h (before renewal, T24), the measured mean concentrations of DBDPE were 1.03, 4.63, 11.07, 42.63, and 136.32 nM. In the BDE209 exposure groups, the measured mean concentrations (T0) were 2.85, 9.07, 27.39, 85.3, and 278.39 nM, respectively, while a small decrease in the concentrations was measured before renewal of the exposure concentrations (SI Figure S1). The concentration of DBDPE and BDE209 in the control group was below the detection limit. The DBDPE and BDE209 concentrations in zebrafish larvae were analyzed at 6 and 14 days (Figure 1). After 6 days of exposure, the total body burden of DBDPE and BDE209 in the larvae showed a concentration-dependent bioconcentration, with 0.56, 2.17, 5.27, 11.52, and 31.42 μg/g d.w. for DBDPE, and 0.61, 0.78, 1.39, 5.41, and 17.35 μg/g d.w. for BDE209 (Figure 1A). After 14 days exposure, DBDPE concentrations were 0.82, 2.66, 5.46, 13.54, and 33.03 μg/g d.w., respectively. The measured concentrations at 6 days were comparable to those at 14 days for DBDPE. However, BDE209 concentrations in all exposure groups increased significantly compared to the concentrations measured at 6 dpf, and measured as 1.64, 2.80, 14.95, 25.49, and 57.21 μg/g d.w. (Figure 1B). The concentration of DBDPE and BDE209 in zebrafish larvae in the control group was below the detection limit. Biotransformation of DBDPE and BDE209 in Zebrafish Larvae. Several metabolites in DBDPE-exposed larvae (300 nM) were screened at 6 dpf and 14 dpf, respectively (Figure 2). Based on the GC-ECNI−MS chromatograms (m/z 79 and 81), at least seven peaks of unknown, brominecontaining compounds eluting before DBDPE were observed, indicating that DBDPE biotransformation does occur in zebrafish larvae. Furthermore, the metabolic rate (ratio of parent compound) of DBDPE in larvae changed based on the exposure time. At 6 days exposure, seven peaks appeared before DBDPE, meanwhile DBDPE comprised 45.9% of the 8439

DOI: 10.1021/acs.est.9b02831 Environ. Sci. Technol. 2019, 53, 8437−8446

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Environmental Science & Technology

eight bromine-containing compounds present in the larvae (Figure 2A). Likewise, at least seven peaks were observed after 14 days exposure, of which DBDPE comprised 11.8% (Figure 2B). In the BDE209 treatment group, at least nine peaks appeared before BDE209, of which the proportion of BDE209 comprised 31.7% at 6 dpf (Figure 2C) and 14.7% at 14 dpf (Figure 2D). The detailed percentage of DBDPE, BDE209 and their potential metabolites of each peak in 6 dpf and 14 dpf larvae are shown in SI Table S3. Whole-Body Thyroid Hormone Contents. The total body THs were measured in the larvae at 6 and 14 dpf. In the DBDPE exposure groups, a concentration-dependent increase in T4 (52.4%, 76.2%, and 161.9%) and T3 levels (102.3%, 113.3%, and 132.6%) was observed after 6 days exposure, and showed a significant difference among the 30, 100, and 300 nM treatment groups (Figure 3A, B). In contrast, treatment with BDE209 did not cause significant alterations in T4 or T3 content (Figure 3A,B). After 14 days treatment, a concentration-dependent increase in both T4 (168.3%, 253.9%, and 489.5%) and T3 (392.4%, 540.2%, and 868.5%) was also measured in the DBDPE exposure groups, and showed a significant difference among the 30, 100, and 300 nM groups (Figure 3C,D). In the BDE209 treatment groups, a significant decrease in the T4 levels was observed, but T3 levels were significantly increased in the 100 and 300 nM groups, respectively (Figure 3C,D). HPT Axis Gene Expression. The gene expression patterns related to the HPT axis and selected TH-specific transporters were examined (Figure 4). In the 14 dpf larvae, crh gene expression was significantly increased following treatment with concentrations of 30, 100, and 300 nM DBDPE, while tshβ gene expression was increased in the 10, 30,100, and 300 nM treatment groups. In contrast, exposure to all concentrations of

Figure 1. DBDPE and BDE209 concentrations in zebrafish larvae. Zebrafish embryos/larvae after exposure to 0, 3, 10, 30, 100, and 300 nM of DBDPE or BDE209 for (A) 6 days and (B) 14 days. DBDPE and BDE209 levels were measured in 100 larvae, with three replicate samples. The data are expressed as the mean ± SEM.

Figure 2. Comparison of GC/ECNI−MS of chromatograms (m/z 79 and 81) after larvae exposure for 6 and 14 days. (A) extracts of 300 nM DBDPE-exposed larvae for 6 days; (B) extracts of 300 nM DBDPE-exposed larvae for 14 days; (C) extracts of 300 nM BDE209-exposed larvae for 6 days; (D) extracts of 300 nM BDE209-exposed larvae for 14 days. 8440

DOI: 10.1021/acs.est.9b02831 Environ. Sci. Technol. 2019, 53, 8437−8446

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Environmental Science & Technology

Figure 3. Measured concentrations of total T4 and T3 in zebrafish larvae after exposure to DBDPE (0, 3, 10, 30, 100, and 300 nM) or BDE209 (0, 3, 10, 30, 100, and 300 nM). (A) T4 in 6 days larvae; (B) T3 in 6 days larvae; (C) T4 in 14 days larvae; (D) T3 in 14 days larvae. Values are expressed as means ± SEM of four replicate samples. *P < 0.05, **P < 0.01, and ***P < 0.001 indicates a significant difference between exposure groups and the control group.

hormone-binding and transport proteins, ttr was decreased and showed a significant difference in the 100 and 300 nM groups. Among the nuclear TH receptor and TH transporter genes, upregulation of trα, trβ, and mct8 was observed in all DBDPE exposure groups, and showed a significant difference in the 300 nM treatment group. Among TH metabolism genes, ugt1ab expression was decreased in all DBDPE treatment groups, and showed a significant difference in the 100 and 300 nM groups (Figure 4A). Among the BDE209 exposure groups, upregulation of crh, tshβ, and tshr as well as dio1, dio2, dio3a, and dio3b were observed in the 10, 30, 100, and 300 nM groups, respectively. Ttr was increased and showed a significant difference in the 100 and 300 nM treatment groups. Transcription of trα, trβ, and mct8 was also upregulated in all exposure groups, and showed a significant difference at 300 nM. The gene transcription of ugt1ab was increased, and showed a significant difference in the 100 and 300 nM treatment groups (Figure 4B). Protein Abundance. TTR protein expression was measured after 6 and 14 days exposure. After exposure for 6 days, a concentration-dependent upregulation of TTR protein expression was observed with DBDPE treatment (543.1.6%, 628.3%, and 640.1% in the 30, 100, and 300 nM groups, respectively) (Figure 5). Similarly, upregulation of TTR was also observed with DBDPE treatment (371.4%, 476.1%, and 500.4% in the 30, 100, and 300 nM groups, respectively) at 14 dpf, compared with those in the control (Figure 5). Histological and Stereological Analysis. No obvious pathological thyroid changes were observed in fish exposed to 300 nM DBDPE (SI Figure S3). Stereological analysis also showed no difference in terms of follicle areas (SI Figure S3).

Figure 4. Heatmap of the thyroid gene transcriptional alterations in (A) DBDPE (0, 3, 10, 30, 100, and 300 nM) and (B) BDE209 (0, 3, 10, 30, 100, and 300 nM) exposed-zebrafish larvae for 14 days. Each concentration is represented in columns, and each gene in lines. Blue indicates down-regulation, whereas red indicates up-regulation. Color intensity represents the extent of the fold changes. Data are expressed as means of four replicate samples.

DBDPE decreased expression of tshr in a dose-dependent manner, with significant differences detected in the 30, 100, and 300 nM treatment groups. Of the deiodinase genes, dio1 gene expression was increased in the 100 and 300 nM treatment groups, whereas dio2 expression was downregulated in all groups and showed a significant difference in the 30 nM treatment group. Dio3a and dio3b were decreased following DBDPE treatment in the 30, 100, and 300 nM groups and in the 100 and 300 nM groups, respectively. Among the thyroid 8441

DOI: 10.1021/acs.est.9b02831 Environ. Sci. Technol. 2019, 53, 8437−8446

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rat tissues.18 This suggests species-specific differences in uptake kinetics, metabolism, and deposition. The reasons for the difference among fish and mammals in terms of bioaccumulation remain to be elucidated. On the other hand, higher body burdens of DBDPE in larvae could also be due to a lower capacity to excrete the chemical. Indeed, a previous study on PBDE bioaccumulation in developing mice showed that young animals have a reduced ability to excrete PBDEs during development.25 It is also necessary to investigate whether DBDPE preferentially distributes to lipophilic tissues, as larval stages possess higher lipid contents. The bioavailability of DBDPE in larvae also suggests that developing larvae could be used for other toxicological studies. Additional studies regarding the bioaccumulation kinetics of DBDPE in developing fish and also potential hazardous effects to early developmental stages are needed. In addition to bioaccumulation of BDE209 and DBDPE, we also identified possible metabolites in exposed larvae based on chromatograms. Previous studies showed that BDE209 could be readily metabolized to other lower molecular weight metabolites (e.g., BDE-206, BDE-207, BDE-196, BDE-197, -183, -154, -153, -100, -99, and -47) in zebrafish larvae,23,26 or to penta- to octa-BDEs PBDE congeners in juvenile fathead minnows,27 indicating that larvae may have high biotransformation potential for BDE209. In this study, we did not further establish the identify of these BDE209 metabolites. DBDPE is considered to be more stable than BDE-209 because the alkyl group has a lower probability of debromination than an ether hydrocarbon.28 In addition to DBDPE, at least seven possible metabolites were identified in DBDPE-exposed larvae, suggesting that zebrafish larvae have the capability to readily metabolize DBDPE. The ratios of parental compound to amount of parent and potential metabolites for BDE209 and DBDPE were 14.7% and 11.8% at 14 dpf, respectively. It is suspected that DBDPE is also more readily metabolized compared with BDE209. The present results are consistent with a previous study in rats treated with DBDPE, where at least seven metabolites were found.18 Although the structure of these metabolites remained to be identified, several previous studies showed that exposure to DBDPE could induce upregulation of genes encoding for cytochrome P450 monooxygenase (CYP) (e.g., CYP3A2, CYP2B1/2) in rat liver,18 and CYP1A4/5 in cultured chicken embryo hepatocytes.29 These studies suggest that DBDPE could be biotransformed. As lower molecular weight PBDE metabolites have been shown to elicit a greater potential for more persistent bioaccumulation and toxicity than the parent BDE congeners,30 it is of great interest and importance to investigate the metabolic pathway(s) involved and to establish the identify of metabolites in order to more completely understand the toxicity of DBDPE. In the present study, a significant decrease in T4 and increased T3 levels was observed in BDE209 exposure groups at 14 days, but not at 6 days. This observation is consistent with our previous study.23 A significant increase in both T4 and T3 was observed in DBDPE-treated zebrafish larvae at 6 and 14 days, indicating disruption of thyroid endocrine function. Compare to the BDE209 treatment groups, a marked increase in both T4 and T3 were observed with DBDPE treatment, suggesting a distinct pattern in terms of altered TH levels a likely more potent effect of DBDPE on THs. It should also be mentioned that studies have indicated that in general the effects of PBDEs on T3 levels were much smaller than

Figure 5. Protein abundance of transthyretin (TTR) in zebrafish larvae after exposure to DBDPE (0, 3, 10, 30, 100, and 300 nM) for 6 and 14 days. (A) Representative Western blot of TTR; (B) Quantification of the relative expression of TTR. Results are expressed as mean ± SEM of four replicate samples. *P < 0.05, **P < 0.01, and ***P < 0.001 indicates a significant difference between exposure groups and the control group.

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DISCUSSION We employed developing zebrafish larvae for assessment of thyroid endocrine disruption resulting from DBDPE exposure, an alternative to BDE209. Bioconcentration of DBDPE in the larvae was evident, and DBDPE could be metabolized. Both T4 and T3 levels were changed by DBDPE exposure in the larvae. To the best of our knowledge, this is the first study to provide evidence of the bioavailability and biotransformation of DBDPE, and its potential for thyroid endocrine disruption in fish. Exposure to all concentrations of DBDPE and BDE209 did not cause any obvious developmental toxicity (e.g., hatching, malformation, and survival) to zebrafish larvae, within the range examined. This observation is consistent with previous studies, where exposure to 380 μg/L of BDE209 did not cause obvious developmental toxicity,23 further confirming lower acute toxicity of BDE209. The observed absence of morphological deformities with DBDPE treatment is consistent with the low acute toxicity of BDE209 to zebrafish embryo/larval development. Although DBDPE has a high molecular weight, high dosedependent measured body burdens of DBDPE in zebrafish larvae indicate bioavailability and bioaccumulation of this hydrophobic chemical from water under laboratory exposure conditions, suggesting that DBDPE can be deposited in larval tissues. Furthermore, the concentrations of BDE209 and DBDPE measured in larvae are similar in magnitude, suggesting that the deposition of DBDPE is similar to that of BDE209. However, a previous study reported that oral exposure to DBDPE and BDE209 in rats showed 3−5 orders of magnitude lower concentrations of DBDPE than BDE209 in 8442

DOI: 10.1021/acs.est.9b02831 Environ. Sci. Technol. 2019, 53, 8437−8446

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Environmental Science & Technology those on T4 levels in rodent and fish.30,31 However, a high magnitude of change in T3 levels was observed with DBDPE treatment. These observations suggest that T3 is more sensitive to DBDPE stress, whereas the differential response of T3 may relate to different doses, exposure durations, and mode of administration. However, a distinct pattern of impact between BDE209 and DBDPE on THs in zebrafish larvae is evident, suggesting different mechanisms of action may disturb the TH system. THs have an important role in the regulation of early fish development. However, harmful effects of excessive THs have been reported in zebrafish larvae, including impaired growth of both pectoral and pelvic fins and severe developmental defects; in particular, they inhibit the development of scales and pigment pattern.32 Significant growth inhibition was also observed in the larvae after 14 days exposure to DBDPE, although both T4 and T3 levels were increased. Consistent with these observations, significant growth inhibition and higher TH levels were also observed in zebrafish larvae exposed to a PBDEs mixture (DE-71).33 To examine whether DBDPE exposure can impact thyroid structure and thereby influence thyroid function, we performed histological and morphometric examination of the thyroid. We did not observe obvious morphological alterations in thyroid tissues at the highest concentration delivered (300 nM). Hence it seems unlikely that exposure to DBDPE directly affects thyroid structure. We also examined gene expression profiles related to the HPT axis as well as specific transporters for TH following exposure to BDE209 and DBDPE for 14 days. The HPT axis is regulated through a negative feedback mechanism in which the level of thyroid-stimulating hormone (TSH) secreted by the pituitary controls the production and release of T4 by the thyroid follicles.34 Gene expression along the HPT axis is often examined in relation to chemical exposure and its impact on the thyroid system. Indeed, gene transcription is sensitive to altered TH levels. For example, in the present study, transcript levels of crh and tshβ were both significantly upregulated after exposure to various concentrations of BDE209, suggesting a negative feedback mechanism response to reduced T4 levels. Reduction of T4 with accompanying upregulation of crh and tshβ gene transcription have been observed in zebrafish larvae exposed to DE-7122 or BDE209,23 as well as in fathead minnow.26 We observed that the expression of crh and tshβ was also increased in response to DBDPE exposure, while T4 was also increased. However, an expected positive feedback response, i.e., increased T4 levels accompanying reduced regulatory gene transcription, was not observed in this study. Hence, we suggest that measurement of gene transcription may not always be suitable to predict physiological function. It is important to measure CRH and TSH concentrations and to evaluate whether the normal regulatory mechanism is in place or whether it may be impaired due to DBDPE treatment. In the present study, we also examined gene expression of deiodinases, which encode enzymes that regulate TH concentrations peripherally. In zebrafish, TH deiodinase genes include dio1/2/3, with dio3 having two paralogs (dio3a, 3b).35,36 Previous studies demonstrated that dio2 catalyzes the outer ring deiodination (ORD) of T4 to produce the bioactive T3, whereas dio3 (dio3a, b) catalyzes the inner ring deiodination (IRD) of T4 and T3 producing the inactive metabolites reverse T3 (rT3) and 3, 3′-diiodo-L-thyronine (T2). Dio1 is thought to play a minimal role in plasma TH homeostasis but has a considerable influence on iodine

recovery and TH degradation.37 In our study, the expression of dio2, dio3a, and dio3b was decreased while dio1 was increased in the DBDPE exposure groups. This result is consistent with previous reports showing that hyperthyroidism suppresses Dio2 activities and corresponding mRNA expression in fish, whereas hypothyroidism is associated with an increase.38 Increased dio1 may help to degrade the T3 leading to decreased levels.39,40 Likewise, Dio3 is the major inactivating pathway of THs and thyroidal status often parallels hepatic Dio3 activity, increasing during hyperthyroidism and decreasing during hypothyroidism.41,42 Hence, a decrease in dio3a and dio3b gene expression may also reflect a response to increased TH levels. However, it should also be noted that deiodinase gene responses are very sensitive to altered TH levels. A previous study suggested that hepatic mRNA levels of these enzymes do not correlate with observed changes in plasma TH levels, indicating that additional biomarkers are needed to assess thyroid disruption by environmental contaminants.43 TH performs its biological activity by binding to thyroid hormone receptors (TRs). As the major thyroid hormone receptor (TR) isoforms, trα and trβ can bind T3 and mediate TH-regulated gene expression.44 In our study, both trα and trβ were upregulated with DBDPE and BDE209 treatment. This result agrees with previous studies, where T3 increased and was accompanied by upregulation of TRα and TRβ expression in BDE209-treated zebrafish larvae.23 Similar findings were observed in T3-treated zebrafish larvae.42 Hence, upregulation of TRs may be a response to increased T3 levels upon DBDPE and BDE209 exposure. The observed upregulation of TRs in the present study may be due to the increased T3 levels and would influence the transcription of other genes involved in thyroid function. We observed that ugt1ab gene transcription was reduced and increased with DBEPE and BDE209 treatment, respectively. It has been suggested that UDPGTs play an important role in TH metabolism in the liver and increased ugt1ab expression could lead to increased elimination of THs to maintain TH homeostasis via the major pathway for T4 conjugation.39 An increase in ugt1ab gene expression upon BDE209 exposure is consistent with previous studies showing that zebrafish exposed to DE-71 (10 μg/L) and BDE209 have decreased T4 levels.22,23 In contrast, ugt1ab gene expression was decreased and T4 was increased with DBDPE treatment. A plausible explanation for the altered TH signaling observed throughout the HPT axis could be that increased T4 levels are accompanied by decreased UGT activities or gene transcription. In zebrafish, TH transporters include lat1, lat2, mct8, and oatp1c1.45 It has been suggested that MCT8 transports thyroxine into cells, ensuring its action on gene expression during neurodevelopment.46 In zebrafish, mct8 is increasingly expressed in the CNS during embryonic and larval development, and is strongly present in neurons and oligodendrocytes.45 Mct8 knockout zebrafish displayed severe neurological and behavioral deficits, confirming the crucial function of MCT8 in CNS development.46 Previous studies reported that higher concentrations of THs upregulate expression of mct8 mRNA expression, a response that appears to be more sensitive to T3 than T4.42 As a whole, these results provide evidence that TH status influences the transcriptional dynamics of mct8. Hence, the observed increased in mct8 8443

DOI: 10.1021/acs.est.9b02831 Environ. Sci. Technol. 2019, 53, 8437−8446

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Environmental Science & Technology Author Contributions

gene expression with both DBDPE and BDE209 treatment is likely a response to increased T3 concentrations. We found that TTR protein expression showed a dosedependent increase in the DBDPE exposure groups. This observation is consistent with our previous study in zebrafish larvae exposed to BDE209 for 14 days.23 TTR is known to be involved in TH transport and metabolic processes related to the thyroid axis in fish.47 In this study, we observed an increase in TTR protein expression while mRNA expression was reduced. Previous studies showed a significant increase in circulating levels of TTR in response to increased levels of either T3 or T4, whereas reduced levels of THs were also accompanied by a reduction in plasma TTR although hepatic transcription of TTR was unchanged.48 This observation is in agreement with the notion that gene expression patterns may be less likely to be useful as biomarkers of developmental TH disruption.42 Presumably increased TTR should help to maintain a similar bound/free hormone ratio as a compensatory response to the acute increase in free TH levels induced by treatment, and suggests a potential regulatory mechanism.49 In summary, we identified significant bioaccumulation of DBDPE in zebrafish larvae, accompanied by evidence of biotransformation of DBDPE and disruption of thyroid function. As a chemical with high-volume production and increased risk of contamination to wildlife and the environment, a clear understanding of the potential toxicity of DBDPE to organisms is very limited. Future studies are needed to examine the toxicokinetics and tissue distribution of DBDPE within an organism, in particular related to its metabolism and identification of metabolites. Future research should determine if exposure to environmentally relevant concentrations of DBDPE affects wildlife health. As thyroid hormones have crucial roles in early central nervous system development, an investigation of the effects of DBDPE on neurons is also needed.

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Notes

The authors declare no competing financial interest.

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ACKNOWLEDGMENTS This work was supported by the National Natural Science Foundation of China (grant number 21737005); the Strategic Priority Research Program of the Chinese Academy of Sciences (grant number XBD14040103); and the State Key Laboratory of Freshwater Ecology and Biotechnology (grant number 2019FBZ03).

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REFERENCES

(1) de Wit, C. A. An overview of brominated flame retardants in the environment. Chemosphere 2002, 46, 583−624. (2) Law, R. J.; Allchin, C. R.; de Boer, J.; Covaci, A.; Herzke, D.; Lepom, P.; Morris, S.; Tronczynski, J.; de Wi, C. A. Levels and trends of brominated flame retardants in the European environment. Chemosphere 2006, 64, 187−208. (3) Vuong, A. M.; Yolton, K.; Dietrich, K. N.; Braun, J. M.; Lanphear, B. P.; Chen, A. Exposure to polybrominated diphenyl ethers (PBDEs) and child behavior: Current findings and future directions. Horm. Behav. 2018, 101, 94−104. (4) Covaci, A.; Harrad, S.; Abdallah, M. A.; Ali, N.; Law, R. J.; Herzke, D.; de Wit, C. A. Novel brominated flame retardants: a review of their analysis, environmental fate and behaviour. Environ. Int. 2011, 37, 532−556. (5) McGrath, T. J.; Ball, A. S.; Clarke, B. O. Critical review of soil contamination by polybrominated diphenyl ethers (PBDEs) and novel brominated flame retardants (NBFRs); concentrations, sources and congener profiles. Environ. Pollut. 2017, 230, 741−757. (6) Shen, K.; Li, L.; Liu, J.; Chen, C.; Liu, J. Stocks, flows and emissions of DBDPE in China and its international distribution through products and waste. Environ. Pollut. 2019, 250, 79−86. (7) Qi, H.; Li, W.; Liu, L.; Zhang, Z.; Zhu, N.; Song, W.; Ma, W.; Li, Y. Levels, distribution and human exposure of new non-BDE brominated flame retardants in the indoor dust of China. Environ. Pollut. 2014, 195, 1−8. (8) Yu, G.; Bu, Q.; Cao, Z.; Du, X.; Xia, J.; Wu, M.; Huang, J. Brominated flame retardants (BFRs): a review on environmental contamination in China. Chemosphere 2016, 150, 479−490. (9) Zhen, X.; Tang, J.; Liu, L.; Wang, X.; Li, Y.; Xie, Z. From headwaters to estuary: Distribution and fate of halogenated flame retardants (HFRs) in a river basin near the largest HFR manufacturing base in China. Sci. Total Environ. 2018, 621, 1370− 1377. (10) Law, K.; Halldorson, T.; Danell, R.; Stern, G.; Gewurtz, S.; Alaee, M.; Marvin, C.; Whittle, M.; Tomy, G. Bioaccumulation and trophic transfer of some brominated flame retardants in a Lake Winnipeg (Canada) food web. Environ. Toxicol. Chem. 2006, 26, 190−190. (11) Wu, J.; Guan, Y.; Zhang, Y.; Luo, X.; Zhi, H.; Chen, S.; Mai, B. Trophodynamics of hexabromocyclododecanes and several other nonPBDE brominated flame retardants in a freshwater food web. Environ. Sci. Technol. 2010, 44, 5490−5495. (12) He, M.; Luo, X.; Chen, M.; Sun, Y.; Chen, S.; Mai, B. Bioaccumulation of polybrominated diphenyl ethers and decabromodiphenyl ethane in fish from a river system in a highly industrialized area, South China. Sci. Total Environ. 2012, 419, 109−115. (13) Tao, L.; Zhang, Y.; Wu, J.; Wu, S.; Liu, Y.; Zeng, Y.; Luo, X.; Mai, B. Biomagnification of PBDEs and alternative brominated flame retardants in a predatory fish: Using fatty acid signature as a primer. Environ. Int. 2019, 127, 226−232. (14) Luo, X.; Zhang, X.; Liu, J.; Wu, J.; Luo, Y.; Chen, S.; Mai, B.; Yang, Z. Persistent halogenated compounds in waterbirds from an e-

ASSOCIATED CONTENT

S Supporting Information *

The Supporting Information is available free of charge on the ACS Publications website at DOI: 10.1021/acs.est.9b02831. Text S1: Extraction and analysis of DBDPE and BDE209; Test S2: Western Blot; Test S3: Thyroid hormone extraction and assay; Table S1: Primer sequences for the quantitative reverse transcriptionpolymerase chain reaction; Table S2: Development index; Table S3: Percentage of DBDPE and potential metabolites; Figure S1: measured concentrations; Figure S2: Comparison of GC/ECNI−MS chromatogram (m/ z79 and 81); Figure S3: Histological and stereological analysis of zebrafish larvae (PDF)

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Co-first author.

AUTHOR INFORMATION

Corresponding Author

*Phone: 86-27-68780042; fax: 86-27-68780123; e-mail: [email protected]. ORCID

Xiaojun Luo: 0000-0002-2572-8108 Lianguo Chen: 0000-0003-3730-7842 Bixian Mai: 0000-0001-6358-8698 Bingsheng Zhou: 0000-0003-0119-1868 8444

DOI: 10.1021/acs.est.9b02831 Environ. Sci. Technol. 2019, 53, 8437−8446

Article

Environmental Science & Technology waste recycling region in South China. Environ. Sci. Technol. 2009, 43, 306−311. (15) Zhu, B.; Lai, N.; Wai, T.; Chan, L.; Lam, J.; Lam, P. Changes of accumulation profiles from PBDEs to brominated and chlorinated alternatives in marine mammals from the South China Sea. Environ. Int. 2014, 66, 65−70. (16) Noyes, P. D.; Lema, S. C.; Macaulay, L. J.; Douglas, N. K.; Stapleton, H. M. Low Level Exposure to the Flame Retardant BDE209 Reduces Thyroid Hormone Levels and Disrupts Thyroid Signaling in Fathead Minnows. Environ. Sci. Technol. 2013, 47, 10012−21. (17) Costa, L. G.; de Laat, R.; Tagliaferri, S.; Pellacani, C. A mechanistic view of polybrominated diphenyl ether (PBDE) developmental neurotoxicity. Toxicol. Lett. 2014, 230, 282−294. (18) Wang, F.; Wang, J.; Dai, J.; Hu, G.; Wang, J.; Luo, X.; Mai, B. Comparative Tissue Distribution, Biotransformation and Associated Biological Effects by Decabromodiphenyl Ethane and Decabrominated Diphenyl Ether in Male Rats after a 90-Day Oral Exposure Study. Environ. Sci. Technol. 2010, 44, 5655−5660. (19) Sun, R.; Shang, S.; Zhang, W.; Lin, B.; Wang, Q.; Shi, Y.; Xi, Z. Endocrine Disruption Activity of 30-day Dietary Exposure to Decabromodiphenyl Ethane in Balb/C Mouse. Biomed. Environ. Sci. 2018, 31, 12−22. (20) Wang, Y.; Chen, T.; Sun, Y.; Zhao, X.; Zheng, D.; Jing, L.; Zhou, X.; Sun, Z.; Shi, Z. A comparison of the thyroid disruption induced by decabrominated diphenyl ethers (BDE-209) and decabromodiphenyl ethane (DBDPE) in rats. Ecotoxicol. Environ. Saf. 2019, 174, 224−235. (21) Smythe, T. A.; Butt, C. M.; Stapleton, H. M.; Pleskach, K.; Ratnayake, G.; Song, C. Y.; Riddell, N.; Konstantinov, A.; Tomy, G. T. Impacts of Unregulated Novel Brominated Flame Retardants on Human Liver Thyroid Deiodination and Sulfotransferation. Environ. Sci. Technol. 2017, 51, 7245−7253. (22) Yu, L.; Deng, J.; Shi, X.; Liu, C.; Yu, K.; Zhou, B. Exposure to DE-71 alters thyroid hormone levels and gene transcription in the hypothalamic-pituitary-thyroid axis of zebrafish larvae. Aquat. Toxicol. 2010, 97, 226−233. (23) Chen, Q.; Yu, L.; Yang, L.; Zhou, B. Bioconcentration and metabolism of decabromodiphenyl ether (BDE-209) result in thyroid endocrine disruption in zebrafish larvae. Aquat. Toxicol. 2012, 110− 111, 141−148. (24) Cristale, J.; Quintana, J.; Chaler, R.; Ventura, F.; Lacorte, S. Gas chromatography/mass spectrometry comprehensive analysis of organophosphorus, brominated flame retardants, by-products and formulation intermediates in water. J. Chromatog. A 2012, 1241, 1− 12. (25) Staskal, D. F.; Hakk, H.; Bauer, D.; Diliberto, J. J.; Birnbaum, L. S. Toxicokinetics of polybrominated diphenyl ether congeners 47, 99, 100, and 153 in mice. Toxicol. Sci. 2006, 94, 28−37. (26) Chen, L.; Wang, X.; Zhang, X.; Lam, P. K. S.; Guo, Y.; Lam, J. C. W.; Zhou, B. Transgenerational endocrine disruption and neurotoxicity in zebrafish larvae after parental exposure to binary mixtures of decabromodiphenyl ether (BDE-209) and lead. Environ. Pollut. 2017, 230, 96−106. (27) Noyes, P. D.; Hinton, D. E.; Stapleton, H. M. Accumulation and Debromination of Decabromodiphenyl Ether (BDE-209) in Juvenile Fathead Minnows (Pimephales promelas) Induces Thyroid Disruption and Liver Alterations. Toxicol. Sci. 2011, 122, 265−274. (28) Kierkegaard, A.; Sellstrom, U.; McLachlan, M. S. Environmental analysis of higher brominated diphenyl ethers and decabromodiphenyl ethane. J. Chromatog. A 2009, 1216, 364−375. (29) Egloff, C.; Crump, D.; Chiu, S.; Manning, G.; McLaren, K. K.; Cassone, C. G.; Letcher, R. J.; Gauthier, L. T.; Kennedy, S. W. In vitro and in ovo effects of four brominated flame retardants on toxicity and hepatic mRNA expression in chicken embryos. Toxicol. Lett. 2011, 207, 25−33. (30) Tomy, G. T.; Palace, V. P.; Halldorson, T.; Braekevelt, E.; Danell, R.; Wautier, K.; Evans, B.; Brinkworth, L.; Fisk, A. T. Bioaccumulation, biotransformation, and biochemical effects of

brominated diphenyl ethers in juvenile lake trout (Salvelinus namaycush). Environ. Sci. Technol. 2004, 38, 1496−1504. (31) Zhou, T.; Ross, D. G.; DeVito, M. J.; Crofton, K. M. Effects of short-term in vivo exposure to polybrominated diphenyl ethers on thyroid hormones and hepatic enzyme activities in weanling rats. Toxicol. Sci. 2001, 61, 76−82. (32) Brown, D. D. The role of thyroid hormone in zebrafish and axolotl development. Proc. Natl. Acad. Sci. U. S. A. 1997, 94, 13011− 13016. (33) Yu, L.; Lam, J. C.; Guo, Y.; Wu, R.; Lam, P. K.; Zhou, B. Parental Transfer of Polybrominated Diphenyl Ethers (PBDEs) and Thyroid Endocrine Disruption in Zebrafish. Environ. Sci. Technol. 2011, 45, 10652−10659. (34) Heijlen, M.; Houbrechts, A. M.; Darras, V. M. Zebrafish as a model to study peripheral thyroid hormone metabolism in vertebrate development. Gen. Comp. Endocrinol. 2013, 188, 289−296. (35) Guo, C.; Chen, X.; Song, H.; Maynard, M. A.; Zhou, Y.; Lobanov, A. V.; Gladyshev, V. N.; Ganis, J. J.; Wiley, D.; Jugo, R. H.; Lee, N. Y.; Castroneves, L. A.; Zon, L. I.; Scanlan, T. S.; Feldman, H. A.; Huang, S. A. Intrinsic expression of a multiexon type 3 deiodinase gene controls zebrafish embryo size. Endocrinology 2014, 155, 4069− 4080. (36) Dong, W.; Macaulay, L.; Kwok, K. W. H.; Hinton, D. E.; Stapleton, H. M. Using whole mount in situ hybridization to examine thyroid hormone deiodinase expression in embryonic and larval zebrafish: a tool for examining OH-BDE toxicity to early life stages. Aquat. Toxicol. 2013, 132−133, 190−199. (37) Bianco, A. C.; Kim, B. W. Deiodinases: implications of the local control of thyroid hormone action. J. Clin. Invest. 2006, 116, 2571− 2579. (38) Van der Geyten, S.; Byamungu, N.; Reyns, G. E.; Kuhn, E. R.; Darras, V. M. Iodothyronine deiodinases and the control of plasma and tissue thyroid hormone levels in hyperthyroid tilapia (Oreochromis niloticus). J. Endocrinol. 2005, 184, 467−479. (39) Lee, J.; Kim, S.; Park, Y.; Moon, H.; Choi, K. Thyroid Hormone-Disrupting Potentials of Major Benzophenones in Two Cell Lines (GH3 and FRTL-5) and Embryo-Larval Zebrafish. Environ. Sci. Technol. 2018, 52, 8858−8865. (40) Kim, S.; Sohn, J.; Ha, S.; Kang, H.; Yim, U.; Shim, W.; Khim, J.; Jung, D.; Choi, K. Thyroid Hormone Disruption by WaterAccommodated Fractions of Crude Oil and Sediments Affected by the Hebei Spirit Oil Spill in Zebrafish and GH3 Cells. Environ. Sci. Technol. 2016, 50, 5972−5980. (41) Heijlen, M.; Houbrechts, A. M.; Bagci, E.; Van Herck, S. L.; Kersseboom, S.; Esguerra, C. V.; Blust, R.; Visser, T. J.; Knapen, D.; Darras, V. M. Knockdown of type 3 iodothyronine deiodinase severely perturbs both embryonic and early larval development in zebrafish. Endocrinology 2014, 155, 1547−1559. (42) Walter, K. M.; Miller, G. W.; Chen, X.; Yaghoobi, B.; Puschner, B.; Lein, P. J. Effects of thyroid hormone disruption on the ontogenetic expression of thyroid hormone signaling genes in developing zebrafish (Danio rerio). Gen. Comp. Endocrinol. 2019, 272, 20−32. (43) Picard-Aitken, M.; Fournier, H.; Pariseau, R.; Marcogliese, D. J.; Cyr, D. G. Thyroid disruption in walleye (Sander vitreus) exposed to environmental contaminants: cloning and use of iodothyronine deiodinases as molecular biomarkers. Aquat. Toxicol. 2007, 83, 200− 211. (44) Power, D. M.; Llewellyn, L.; Faustino, M.; Nowell, M. A.; Bjornsson, B. Th.; Einarsdottir, I. E.; Canario, A. V. M.; Sweeney, G. E. Thyroid hormones in growth and development of fish. Comp. Biochem. Physiol., Part C: Toxicol. Pharmacol. 2001, 130C, 447−459. (45) Vancamp, P.; Darras, V. M. From zebrafish to human: A comparative approach to elucidate the role of the thyroid hormone transporter MCT8 during brain development. Gen. Comp. Endocrinol. 2018, 265, 219−229. (46) Sugiura, M.; Nagaoka, M.; Yabuuchi, H.; Akaike, T. Overexpression of MCT8 enhances the differentiation of ES cells 8445

DOI: 10.1021/acs.est.9b02831 Environ. Sci. Technol. 2019, 53, 8437−8446

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Environmental Science & Technology into neural progenitors. Biochem. Biophys. Res. Commun. 2007, 360, 741−745. (47) Power, D. M.; Elias, N. P.; Richardson, S. J.; Mendes, J.; Soares, C. M.; Santos, C. R. A. Evolution of the thyroid hormone-binding protein, transthyretin. Gen. Comp. Endocrinol. 2000, 119, 241−255. (48) Morgado, I.; Santos, C. R. A.; Jacinto, R.; Power, D. M. Regulation of transthyretin by thyroid hormones in fish. Gen. Comp. Endocrinol. 2007, 152, 189−197. (49) Morgado, I.; Campinho, M. A.; Costa, R.; Jacinto, R.; Power, D. M. Disruption of the thyroid system by diethystilbestrol and ioxynil in the sea bream (Sparus aurata). Aquat. Toxicol. 2009, 94, 87−93.

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DOI: 10.1021/acs.est.9b02831 Environ. Sci. Technol. 2019, 53, 8437−8446