Binding and Activity of Tetrabromobisphenol A Mono-Ether Structural Analogs to Thyroid Hormone Transport Proteins and Receptors

Background: Tetrabromobisphenol A (TBBPA) mono-ether structural analogs, identified as the by-products or transformation products of commercial TBBPA bis-ether derivatives, have been identified as emerging widespread pollutants. However, there is very little information regarding their toxicological effects. Objective: We aimed to explore the potential thyroid hormone (TH) system–disrupting effect of TBBPA mono-ether structural analogs. Methods: The binding potencies of chemicals toward human TH transport proteins [transthyretin (TTR) and thyroxine-binding globulin (TBG)] and receptors [TRα ligand-binding domain (LBD) and TRβ-LBD] were determined by fluorescence competitive binding assays. Molecular docking was used to simulate the binding modes of the chemicals with the proteins. The cellular TR-disrupting potencies of chemicals were assessed by a GH3 cell proliferation assay. The intracellular concentrations of the chemicals were measured by high-performance liquid chromatography and mass spectrometry. Results: TBBPA mono-ether structural analogs bound to TTR with half maximal inhibitory concentrations ranging from 0.1μM to 1.0μM but did not bind to TBG. They also bound to both subtypes of TR-LBDs with 20% maximal inhibitory concentrations ranging from 4.0μM to 50.0μM. The docking results showed that the analogs fit into the ligand-binding pockets of TTR and TR-LBDs with binding modes similar to that of TBBPA. These compounds likely induced GH3 cell proliferation via TR [with the lowest effective concentrations (LOECs) ranging from 0.3μM to 2.5μM] and further enhanced TH-induced GH3 cell proliferation (with LOECs ranging from 0.3μM to 1.2μM). Compared with TBBPA, TBBPA-mono(2,3-dibromopropyl ether) showed a 4.18-fold higher GH3 cell proliferation effect and 105-fold higher cell membrane transportation ability. Conclusion: This study provided a possible mechanism underlying the difference in TTR or TR binding by novel TBBPA structural analogs. These compounds might exert TH system–disrupting effects by disrupting TH transport in circulation and TR activity in TH-responsive cells. https://doi.org/10.1289/EHP6498


Introduction
Tetrabromobisphenol A (TBBPA) and its derivatives are applied as brominated flame retardants (BFRs) in the production of commercial products such as plastics, textiles, electronics and electrical equipment, light sockets, water pipes, and thermal insulation foam (Alaee et al. 2003;). In addition to being used as an additive BFR, TBBPA ( ∼ 18% of the production volume) is also used as a reactive BFR to synthesize TBBPA derivatives (Covaci et al. 2009;Liu et al. 2019;Qu et al. 2016). Commercial TBBPA bis-ether derivatives-such as TBBPA-bis(glycidyl ether) (TBBPA-BGE), TBBPA-bis(allyl ether) (TBBPA-BAE), and TBBPA-bis(2,3-dibromopropyl ether) (TBBPA-BDBPE)are produced via modification of the hydroxyl groups (OH) on both sides of TBBPA with glycidyl ether, allyl ether, and 2,3dibromopropyl ether groups, respectively ( Figure 1) (European Commission 2006). Similar to TBBPA, TBBPA bis-ether derivatives have also been commonly used as additive or reactive BFRs to ultimately produce acrylonitrile butadiene styrene and highimpact polystyrene resins . Currently, TBBPA and its derivatives account for 60% of the total BFR market and have become the most widely applied BFRs in commercial products (Covaci et al. 2011).
In contrast to TBBPA bis-ether derivatives, TBBPA monoether structural analogs have only one OH group on the side of TBBPA replaced by a derivatized group. TBBPA mono-ether structural analogs-including TBBPA-mono(glycidyl ether) (TBBPA-MGE), TBBPA-mono(allyl ether) (TBBPA-MAE), and TBBPA-mono(2,3-dibromopropyl ether) (TBBPA-MDBPE)are frequently detected. These chemicals could be synthesized as by-products during commercial TBBPA bis-ether derivative production (Qu et al. 2011). Some mono-ether structural analogs were detected in brominated epoxy resin samples at levels of 46-160 mg=kg (Liu et al. 2015). Furthermore, TBBPA monoether structural analogs could be generated from TBBPA bis-ether derivatives during transformation processes (Liu et al. , 2019. Ether bond breakage was found to be an important transformation pathway for TBBPA bis-ether derivatives (Liu et al. 2017). TBBPA mono-ether structural analogs-including TBBPA-MGE, TBBPA-MAE, and TBBPA-MDBPE-could be transformed from commercial TBBPA bis-ether derivatives in the environment due to the breaking of ether bonds Qu et al. 2013). Moreover, the substantially increased ratio of TBBPA-MAE/ TBBPA-BAE (3.4:1) in the soil samples compared with the original ratio in the commercial products (9:2 × 10 −3 :1) suggested that certain portions of TBBPA derivatives could transform to TBBPA mono-ether structural analogs in the environment (Liu et al. 2017). In addition, the biotransformation from TBBPA to some TBBPA mono-ether structural analogs has been reported in various studies Gu et al. 2017), suggesting that TBBPA could also contribute to the occurrence of environmental TBBPA monoether structural analogs.
Currently, because of their high production volume and wide application in consumer products, TBBPA and its derivatives are emerging as a group of ubiquitous pollutants that have been frequently detected in various environmental media, including river water, seepage water, sewage water, sediment (Nyholm et al. 2013;Qu et al. 2011), air samples , and wild animals (Letcher and Chu 2010;. For example, TBBPA bis-ether derivatives and TBBPA mono-ether structural analogs were detected in the soil samples collected from sites around BFR production plants with levels ranging from below the method limit of detection (LOD) to 13 mg=g dry weight (Liu et al. 2017), and they were detected in some marine biological samples collected from the Chinese Bohai Sea with levels ranging from below the method LOD to 2:7 lg=g lipid weight . Some TBBPA bis-ether derivatives (such as TBBPA-BDBPE) and some TBBPA mono-ether structural analogs (such as TBBPA-MAE) were found at similar or higher levels than TBBPA in the samples collected from the same area (Liu et al. , 2017. Evidence has shown that the contamination of TBBPA-BDBPE in herring gulls collected from Great Lake locations could be dated back to at least the early 2000s (Gauthier et al. 2019), suggesting that this chemical might have existed in the environment for a long time. The above studies suggest that humans might be exposed to these chemicals due to their release from commercial products containing TBBPA bis-ether derivatives added as BFRs or via exposure from environmental media contaminated by the pollution of BFR production plants. More importantly, TBBPA bis-ether derivatives and TBBPA monoether structural analogs have been identified in various types of seafood, including mollusks, shrimp, crab, and fish samples collected from coastal areas of the Chinese Bohai Sea . Consumption of contaminated seafood could lead to direct exposure to these chemicals through oral exposure and represents an important health concern.
Thyroid hormone (TH) plays an important role in maintaining a normal physiological state and is essential for central nervous system development, metabolism, and growth in mammals (Bernal 2005). Impaired TH function has been associated with many adverse health effects, such as deficient neurological development, reduced immune activity, obesity, and metabolic disorders (Yen 2001). When TH is produced and released from the thyroid gland into circulation, most of it binds to the TH transport proteins transthyretin (TTR) and thyroxine-binding globulin (TBG) (Brent 2012). In target tissues, TH exerts its biological actions by interacting with TH receptors (TRs) and regulating their activity in TH-responsive cells (Aagaard et al. 2011). To date, a large number of epidemiological studies have provided extensive data on the high prevalence and incidence of thyroid disease (Bjoro et al. 2000;Kwon et al. 2018). The prevalence of diagnosed thyroid dysfunction in the form of hypothyroidism was found to be 3.82% in Europe (Garmendia Madariaga et al. 2014). The incidence of thyroid cancer has tripled or more in many developed countries over the past 30 y (Roman et al. 2017). Iodine nutrition, aging, smoking status, and genetic susceptibility could be key determinants of thyroid disease risk (Taylor et al. 2018). Increasing evidence has shown that various environmental endocrine disruptors might also be an important driving factor (Boas et al. 2012;Hofmann et al. 2009;Jugan et al. 2010;Patrick 2009;Tang et al. 2020). Therefore, unveiling the molecular mechanisms responsible for pollutant exposure-associated thyroid disorders is critical to understanding the role of dominant environmental exposure factors in the increased incidence of thyroid disease.
One of the most concerning adverse effects of TBBPA is its endocrine-disrupting potency as a TH system-disrupting chemical Yu et al. 2018;Zhang et al. 2014). Kim and Oh (2014) showed that TBBPA was weakly correlated with a disturbance in TH levels in the general population in Korea. Yang et al. (2016) demonstrated that THs presented a nonmonotonous doseeffect relationship in male Sprague-Dawley rats exposed to 0-1,000 mg=kg body weight (BW) per day and showed that the derived-reference dose was 0:6 mg=kg BW per day. By using zebrafish (Danio rerio) embryos, Zhu et al. (2018) demonstrated that TBBPA evoked TH system-disrupting effects by showing changes in the levels of THs and in the expression of some TH system-related genes after exposure to 200 lg=L TBBPA. TBBPA was reported to exert a TH system-disrupting effect via two possible mechanisms: a) through binding to TH transport proteins and disrupting TH homeostasis in blood (Meerts et al. 2000), or b) through binding to TRs and disturbing their activity and associated cellular functions (Chan and Chan 2012;Kitamura et al. 2002Kitamura et al. , 2005Lévy-Bimbot 2012). In contrast to the extensive toxicity investigations already conducted on TBBPA, information concerning the toxicological effects of TBBPA bis-ether derivatives or TBBPA mono-ether structural analogs is limited. Understanding their structure-dependent activity on TH transport proteins or receptors is important to unveil the mechanisms underlying their effects on the TH system. Together with environmental exposure evaluation, investigation of the toxicity of TBBPA derivatives and TBBPA structural analogs will help to clarify their potential health risks.
In the present study, we aimed to explore the characteristics of the disrupting potential of TBBPA derivatives and TBBPA structural analogs on the TH system. We chose three TBBPA derivatives (TBBPA-BGE, TBBPA-BAE, and TBBPA-BDBPE) and three TBBPA mono-ether structural analogs (TBBPA-MGE, TBBPA-MAE, and TBBPA-MDBPE) ( Figure 1) that are most frequently detected in the environment. To study their potential TH transport-disrupting effects, we evaluated their binding potencies with two major human TH transport proteins (TTR and TBG). We also determined the binding potencies of these chemicals to two subtypes of human TR LBDs (TRa-LBD and TRb-LBD). By using the GH3 cell proliferation assay, we evaluated their TR-disrupting potencies in cells. Molecular docking was used to simulate the binding modes of these chemicals with TH transport proteins and receptors. The cell membrane transportation ability between a TBBPA mono-ether structural analog and TBBPA was also compared by measuring their intracellular amounts in GH3 cells after exposure to investigate their effects and mechanisms.

Fluorescence Competitive Binding Assays
Fluorescence polarization-based competitive binding assays were used to determine the binding potency of the chemicals to TH transport proteins and receptors. Two fluorescence probes, fluoresceintriiodothyronine (F-T 3 ) and fluorescein-thyroxine (F-T 4 ), were synthesized in our previous works (Ren and Guo 2012;Ren et al. 2013). F-T 3 and F-T 4 were prepared by reacting fluorescein isothiocyanate with T 3 or T 4 , which resulted in derivatization of the amine groups of T 3 or T 4 with fluorescein. The competitive binding assays were carried out in a manner similar to a previous study (Qin et al. 2019). Briefly, 200 nM TTR and 50 nM F-T 4 were used in the TTR competitive binding assay, 100 nM TBG and 50 nM F-T 4 were used in the TBG competitive binding assay, and 200 nM TRa-LBD or 200 nM TRb-LBD and 50 nM F-T 3 were used in the TR-LBD competitive binding assays. These assays were performed in potassium phosphate buffer [50 mM potassium phosphate, 150 mM potassium chloride (KCl), 0:5 mM ethylenediaminetetraacetic acid (EDTA), pH 7.5]. For the TTR and TBG competitive binding assays, five different concentrations of competitors were tested with the highest concentration of 50 lM and 10-fold dilutions between each of the other concentrations. For the TR-LBDs competitive binding assays, seven different concentrations of competitors were tested with the highest concentration of 50 lM and 2-fold dilutions between each of the other concentrations. The protein, probe, and different concentrations of competitor were mixed in a total volume of 100 lL and incubated for 5 min at room temperature (controlled at 25°C), and then the fluorescence polarization was measured. For the competitive binding assays, we prepared three replicate wells per experiment for each group in a 96-well plate. The dimethyl sulfoxide (DMSO) content was maintained below 1%. Fluorescence polarization was detected by a SpectraMax i3x multimode detection platform (Molecular Devices). The excitation and emission wavelengths were 490 nm and 520 nm, respectively. The fluorescence polarization value was plotted as a function of competitor concentration to obtain the competition curve. The values of half maximal inhibitory concentration (IC 50 ) or 20% maximal inhibitory concentration (IC 20 ), the concentration of a ligand required to displace 50% or 20%, respectively, of a probe from the protein) were obtained from the linear interpolation between two concentrations located at approximately the 50% or 20% inhibition level. The relative binding potency (RBP) was calculated by dividing the IC 50 value of TBBPA (for TTR) or the IC 20 value of TBBPA (for TRa-LBD and TRb-LBD) by that of a competitor. The IC 50 , IC 20 , and RBP values (compared with TBBPA, which was set as 1) are listed in Table 1.

Molecular Docking Simulations
AutoDock Vina was used to simulate the interactions between the chemicals and TH transport proteins (or receptors). . The cells were cultured in Dulbecco's modified Eagle's/F-12 medium (DMEM/F12; Life Technologies; catalog no. 11330032) supplemented with 10% (vol/vol) heatinactivated fetal bovine serum (FBS; Life Technologies; catalog no. 26400044), and antibiotics (100 U=mL penicillin and 100 lg=mL streptomycin; Life Technologies; catalog no. 15140122). All cells were cultured in a humidified atmosphere composed of 95% air and 5% carbon dioxide at 37°C. The GH3 cell proliferation assay was carried out as described in a previous study (Qin et al. 2019). Briefly, the culture medium was replaced with the test medium [DMEM/F12 without FBS but with bovine insulin (10 mg=mL), ethanolamine (10 mM), sodium selenite (10 ng=mL), bovine serum albumin (500 mg=mL), and human apotransferrin (10 mg=mL)] and incubated for 24 h. The GH3 cells were then seeded in 96-well plates at a density of 5 × 10 3 cells per well in the test medium. In the GH3 cell proliferation assay, we used three replicate wells per experiment for each group. After 24 h, the cells were exposed to different concentrations of chemicals (0, 0.3, 0.6, 1.2, 2.5, 5, or 10 lM) with or without 2 lM amiodarone (a specific antagonist of TR) to study their effect on cell proliferation. After exposure for 96 h, the proliferative effect was assessed using a Cell Counting Kit-8 (CCK-8) assay (Dojindo; catalog no. CK04). The relative cell proliferation values were calculated by dividing the absorbance at 450 nm of the control group (0.1% DMSO) by that of a cell sample based on the CCK-8 assay. In the present study, we aimed to compare the GH3 cell proliferation activity of TBBPA bis-ether derivatives and TBBPA mono-ether structural analogs with TBBPA by comparing their effects at the same concentration. The relative activity (RA) was calculated by dividing the relative cell proliferation value of TBBPA by that of a TBBPA bis-ether derivative or TBBPA monoether structural analog at the concentration of 2:5 lM, which showed no cytotoxicity for all tested chemicals ( Figure S1; Table  S1). The cytotoxicity was also determined based on the results of the CCK-8 assay. TBBPA and its analogs at concentrations that were not different or had higher absorbance at 450 nm than the control group were considered to have no cytotoxicity. The relative cell proliferation values (compared with the 0.1% DMSO-control group) and RA values (compared with TBBPA, set as 1) are listed in Table  1. To study the antagonistic activity of TBBPA bis-ether derivatives and TBBPA mono-ether structural analogs against TR, cells were exposed to different concentrations of chemicals in the presence of 10 nm T 3 . The relative cell proliferation values were calculated by dividing the absorbance of the 10 nM T 3 group at 450 nm by that of a cell sample based on the CCK-8 assay.
Detection of Intracellular TBBPA-MDBPE and TBBPA GH3 cells ( ∼ 5 × 10 5 cells) were seeded in 25-cm 2 flasks in test medium. After 12 h, TBBPA-MDBPE or TBBPA was added to the medium to obtain a final concentration of 2:5 lM, which showed no cytotoxicity for these two chemicals ( Figure S1C,D). After 24 h, the cells were harvested by trypsin-EDTA (Life Technologies). The cells were washed four times with phosphate-buffered saline buffer [137 mM sodium chloride, 2:7 mM KCl, 10 mM disodium phosphate, and 1:8 mM monopotassium phosphate, pH 7.4] by centrifuging the cells at 4,000 rpm for 5 min at 4°C. Then, the cell density was counted by a Countess cell counter (Life Technologies). Intracellular TBBPA-MDBPE and TBBPA were extracted, and their amounts were detected according to the methods below.
Sample pretreatment. The sample pretreatment method was performed as described in a previous study with slight modifications . In brief, the cell pellet (approximately 5 × 10 5 cells) of each sample was mixed with 2 mL of hydrochloric acid (6 mol=L) and 6 mL of dimethylcarbinol. Next, 15 mL of a mixture of hexane and methyl tertiary butyl ether (1:1, vol/vol) was added to extract the chemicals from the cells by applying an ultrasonic cell disrupter for 5 min. The supernatant was collected after centrifugation at 4,000 rpm for 15 min. The above ultrasonic extraction process was repeated twice, and the supernatants were collected and mixed together. Next, 20 mL of 1% KCl and 20 mL of water were added to the collected extraction solution to precipitate the proteins and impurities. Anhydrous sodium sulfate was added to remove water. The extraction solution was then removed gently with a rotary evaporator and nitrogen blowing. The target compounds were then reconstituted in 1:5 mL of methanol and introduced to a HybridSPE-Phospholipid cartridge (30 mg, 1 mL; Supelco) for further purification. Finally, 50 ng of Deuterated-labeled TBBPA (D 10 -TBBPA) was spiked as an internal standard and redissolved in 1 mL of methanol before analysis.

TTR
TRa-LBD TRb-LBD GH3 The addition of the compound led to displacement of the fluorescence probe from TR-LBDs or TTR and did not provide an IC 20 or IC 50 value).
e The T 4 and T 3 results were obtained from our previous study (Qin et al. 2019).
retention times and selected ion pairs with the corresponding standard compounds.
Quantification and method validation. The concentrations of TBBPA-MDBPE and TBBPA were quantified using internal standards. Calibration curves were established based on 10 points (0.1, 0.5, 1, 5, 10, 25, 50, 100, 250, and 500 ng=mL), and the correlation coefficients (R 2 ) were >0:9. Three different levels of TBBPA-MDBPE and TBBPA (10, 100, and 1,000 ng=mL, n = 3) were added to the blank cell matrix to examine the recovery rate of the sample pretreatment method. The LOD and limit of quantification (LOQ) were determined using signal-to-noise ratios of 3 and 10, respectively. The LOD values for TBBPA-MDBPE and TBBPA were 0:10 ng=mL and 0:13 ng=mL, respectively. The LOQ values for TBBPA-MDBPE and TBBPA were 0:32 ng=mL and 0:43 ng=mL, respectively. All cell samples were analyzed in triplicate, and the same unspiked samples were prepared in the same batch as the blank controls.

Statistical Analysis
All experiments were conducted in triplicate. The data are expressed as the mean value plus or minus the standard deviation (n = 3). The p-values of the experimental data were analyzed using one-way analysis of variance tests, followed by a least significant difference multiple comparisons test (IBM SPSS Statistics 20). A p-value of <0:05 was considered to be statistically significant.
Based on the above competitive binding assays, we demonstrated that the three tested TBBPA bis-ether derivatives displayed very low or negligible binding potencies to TH transport proteins. In contrast, the TBBPA mono-ether structural analogs bound to TTR, and one chemical displayed a higher binding potency than TBBPA (TBBPA-MAE; 2-fold greater). We further used molecular docking analysis to simulate the binding modes of TBBPA mono-ether structural analogs with TTR. The identified binding sites of these chemicals on TTR are shown in Figure 3A-C, whereas the binding scores and interactions of these chemicals with TTR are shown in Table S4 and Figure S3, respectively. We found that the order of the docking scores (a higher score means higher potential binding potency) (Table S4) and the observed binding potencies (Table 1) of these chemicals were consistent with each other. As shown in Figure 3A-C, TBBPA-MGE and TBBPA-MDBPE fit into TTR with their OH groups oriented toward the entrance of the protein, whereas their derivatized groups were positioned toward the inner part of the binding pocket. TBBPA-MAE fit into TTR in the opposite orientation, with its OH group oriented toward the inner part of the binding pocket ( Figure  3B). TBBPA was also included in the docking analysis for comparison. We found that the binding sites of the TBBPA mono-ether structural analogs in TTR were similar to those of TBBPA ( Figure  3A-C). The identical moieties of these chemicals-including two benzene rings, four bromine atoms, two methyl groups, and one OH group-resided in similar sites, whereas the residual differently derivatized moieties of the TBBPA mono-ether structural analogs resided at the site where the other OH group of TBBPA was located ( Figure 3A-C). As shown in Table S4, the interactions, especially the hydrophobic interactions, between the TBBPA mono-ether structural analogs and TTR were also similar to those between TBBPA and TTR. Based on the above results, TBBPA mono-ether structural analogs and TBBPA likely bind to human TTR through a similar mechanism, and the derivatized groups were shown to modulate TTR-binding potency.

Interaction of TBBPA Mono-Ether Structural Analogs with Human TR-LBDs
We then studied the potential TH system-disrupting effects of the TBBPA mono-ether structural analogs via the mechanism of interaction with TR in TH-responsive cells. First, we explored the possibility of direct binding of these chemicals with TRs. We quantitatively determined the binding potencies of TBBPA, the three TBBPA mono-ether structural analogs and the three TBBPA bis-ether derivatives with two subtypes of human TR-LBDs, TRa-LBD and TRb-LBD. As shown in Figure 4A-C, Table S5, and Table 1 Table S6). The TR-LBD binding potencies of these TBBPA mono-ether structural analogs were approximately 60-to 500-fold lower than those of the natural ligand T 3 (Table 1). As shown in Figure 4D, Table 1 (Table 1). For TRb-LBD, the relative binding potencies were as follows: TBBPA-MAE >TBBPA >TBBPA-MGE >TBBPA-MDBPE > TBBPA-BGE TBBPA-BAE >TBBPA-BDBPE (Table 1).
We also used molecular docking analysis to simulate the binding modes of the TBBPA mono-ether structural analogs with the TR-LBDs. The identified binding sites of these chemicals in TRa-LBD and TRb-LBD are shown in Figure 3D-F and Figure  S5, respectively. The binding scores and interactions of these chemicals with TRa-LBD and TRb-LBD are shown in Tables S7  and S8 and Figures S6 and S7, respectively. The docking score order (Tables S7 and S8) and the observed binding potencies (Table 1) of these chemicals with the TR-LBDs were consistent with each other. As shown in Figure 3D,F, TBBPA-MGE and TBBPA-MDBPE fit into TRa-LBD with their OH groups oriented toward the entrance of the receptor, whereas their derivative groups were positioned toward the inner part of the binding pocket. TBBPA-MAE fit into TRa-LBD in an opposite orientation, with its OH group oriented toward the inner part of the binding pocket ( Figure 3E). Similarly, TBBPA-MGE and TBBPA-MDBPE fit into TRb-LBD with their OH groups oriented toward the entrance of the receptor, whereas TBBPA-MAE with its OH group oriented toward the inner part of the binding pocket ( Figure S5). TBBPA was also docked in the TR-LBDs for comparison. We found that the TBBPA mono-ether structural analogs had similar binding sites as TBBPA in the TR-LBDs (Figure 3D-F; Figure S5). Three replicate wells were included for each group in a 96-well plate. Error bars represent the standard deviation of three replicates. * p < 0:05, compared with the control group (1% dimethyl sulfoxide). The p-values of the experimental data were analyzed using a one-way analysis of variance, followed by a least significant difference multiple comparisons test (IBM SPSS Statistics 20). See the summary data in Table S3.

Comparison of the Membrane Transportation Ability of TBBPA-MDBPE and TBBPA in GH3 Cells
Based on the above results, we found that TBBPA-MDBPE showed much higher (4.18-fold) GH3 cell proliferation activity than TBBPA ( Figure 5C,D and Table 1). Given that TBBPA-MDBPE has a distinctly different hydrophobicity (Log K ow ) from TBBPA (Table 1), we hypothesized that its higher GH3 cell proliferation activity might be due to its higher membrane transportation ability. To test this hypothesis, we measured the intracellular amounts of TBBPA-MDBPE and TBBPA in GH3 cells after exposure to the same concentration (2:5 lM) of these two compounds. We determined the intracellular amounts of TBBPA-MDBPE and TBBPA in GH3 cells at the same exposure concentration of 2:5 lM. At this concentration, TBBPA showed no difference from the control group ( Figure S1D), whereas TBBPA-MDBPE showed a higher absorbance at 450 nm than the control group ( Figure  S1C), suggesting that these two chemicals had no cytotoxicity to GH3 cells at a concentration of 2:5 lM. The cell membrane should remain intact when the cells are exposed to these chemicals at this concentration, and the amount of chemicals in GH3 cells was probably related to their membrane transportation ability. We used HPLC and MS/MS to detect the concentrations of TBBPA-MDBPE and TBBPA in the tested samples. The LOD values for TBBPA-MDBPE and TBBPA were 0:10 ng=mL and 0:13 ng=mL, respectively ( Table 2). The LOQ values for TBBPA-MDBPE and TBBPA were 0:32 ng=mL and 0:43 ng=mL, respectively ( Table  2). The recovery rates were in the range of 68% to 80% and 76% to 120% for TBBPA-MDBPE and TBBPA, respectively ( Table 2). The above results demonstrated the accuracy of the sample pretreatment and the chemical analysis methods. As shown in Table 2, after 24 h of exposure, the amounts of TBBPA-MDBPE and TBBPA in 5 × 10 5 GH3 cells were approximately 2,013:56 ng (2,733.5 pmol) and 14:10 ng (25.9 pmol), respectively. The results showed that the amount of TBBPA-MDBPE in the GH3 cells was approximately 105-fold higher than that of TBBPA when the cells were exposed to the same concentration (2:5 lM). This demonstrated that TBBPA-MDBPE had a much higher membrane transportation ability than TBBPA.

Discussion
In mammals, TH synthesis and secretion are finely modulated by the hypothalamic-pituitary-thyroid axis and regulate normal blood circulating TH homeostasis (Mendoza and Hollenberg 2017). Various endocrine-disrupting chemicals have the potential to disturb TH homeostasis via competitive binding with the TH transport proteins TTR and/or TBG in the blood (Meerts et al. 2000;Ren and Guo 2012). Previously, Meerts et al. (2000) used a 125 I-T 4 radio ligand displacement assay to demonstrate that TBBPA bound to human TTR. Here, by using a human TH transport protein competitive binding assay that was established in our previous study (Ren and Guo 2012), we also demonstrated the TTR-binding potential of TBBPA ( Figure 2D). Similar to . Three replicate wells were included for each group in a 96well plate. Error bars represent the standard deviation of three replicates. * p < 0:05, compared with the control group (1% dimethyl sulfoxide). The p-values of the experimental data were analyzed using a one-way analysis of variance, followed by a least significant difference multiple comparisons test (IBM SPSS Statistics 20). See the summary data in Table S5. TBBPA, TBBPA mono-ether structural analogs bound to human TTR (Figure 2A-C). We further revealed that TBBPA monoether structural analogs specifically bound to TTR at the TH binding site based on the site-specific fluorescence probe F-T 4 . Moreover, the TTR-binding potency of TBBPA-MAE was higher than that of TBBPA and even higher than that of the natural  Table S9. Note: DMSO, dimethyl sulfoxide. ligand T 4 (Table 1), indicating the strong unexpected TH transport-disrupting potential of this chemical. TTR is critical in the transport of TH across the blood-brain barrier and plays an important role in TH action in the brain (Zheng et al. 2001). Therefore, binding of these chemicals to TTR may not only affect blood circulating TH homeostasis but also decrease the availability of TH specifically in the brain. Moreover, it has also been demonstrated that TTR mediates the transport of environmental pollutants into the placenta (Meerts et al. 2002). The binding of TBBPA mono-ether structural analogs to TTR might also . Three replicate wells were included for each group in a 96-well plate. Error bars represent the standard deviation of three replicates. The relative cell proliferation values were calculated by dividing the absorbance value at 450 nm of the 10 nm T 3 group by that of each cell sample based on the Cell Counting Kit-8 assay. * p < 0:05, compared with the group treated with 10 nm T 3 . The p-values of the experimental data were analyzed using a one-way analysis of variance, followed by a least significant difference multiple comparisons test (IBM SPSS Statistics 20). See the summary data in Table S10. mediate the translocation of these chemicals across the placental barrier and lead to exposure to the fetus.
Various TBBPA structural analogs, including TBBPA bisether derivatives, have been designed and synthesized as BFRs with the aim of enhancing some desired properties, such as safety to human health (Alaee et al. 2003). Here, we demonstrated that the TTR-binding potencies of the TBBPA mono-ether structural analogs were higher than those of the bis-ether derivatives (Table  1), which adds important information regarding the effects of the OH derivatized groups on the interaction between TBBPA structural analogs and TTR. Moreover, these results showed that the TBBPA mono-ether structural analogs, which occur as byproducts and transformation products in the environment, displayed greater potential TH-disruptive effects than TBBPA bisether derivatives.
Molecular docking showed that, like TBBPA, the TBBPA mono-ether structural analogs also fit into the ligand-binding pocket of TTR, and they had similar binding modes (Figure 3; Table S4). In view of the similar binding sites of the three tested TBBPA mono-ether structural analogs in the TTR, the different protein binding potencies could be explained by the variations in the interactions between the differently derivatized groups and the protein. For TTR, TBBPA-MAE (with an allyl group having a moderate size and Log K ow compared with the other two chemicals) displayed higher binding potency than the other two TBBPA mono-ether structural analogs. Compared with the other two chemicals, TBBPA-MAE might maintain a better balance between hydrophobic effects and steric hindrance in the TTR ligand-binding pocket, which in turn leads to its higher binding potency.
TR, as a nuclear hormone receptor, has been demonstrated to regulate the expression of various target genes responsible for development and energy homeostasis (Astapova et al. 2008;Morte et al. 2002). Disruption of the normal function of TR in THresponsive cells could lead to various adverse effects such as neurotoxicity and metabolic disorders (Heindel et al. 2017;Schreiber et al. 2010). Both apo-TRs (receptors without ligand) and ligandbound TRs could regulate target gene expression. In the absence of ligands, apo-TRs could heterodimerize with retinoid-X receptor (RXR) and recruit the binding of nuclear corepressors, which ultimately repress or silence basal target gene expression (Mendoza and Hollenberg 2017). Binding of T 3 to TRs could alter the conformations of TRs in such a way that the corepressor complex is replaced by a coactivator, which in turn activates gene expression (Harvey and Williams 2002). Therefore, for environmental pollutants, both direct binding to apo-TRs and competitive binding to T 3 -bound TRs could interfere with the normal function of TRs.
Previously, Kollitz et al. (2018) used a 125 I-T 3 radio ligand displacement assay to demonstrate that TBBPA bound to human TRb-LBD. Here, by using human TR-LBD competitive binding assays, we also demonstrated the binding of TBBPA to human TR-LBDs. We further demonstrated that, like TBBPA, TBBPA mono-ether structural analogs bound to both human TRa-LBD and TRb-LBD (Figure 4A-C; Figure S4A-C). The site-specific fluorescence probe F-T 3 demonstrated the specific binding of TBBPA mono-ether structural analogs to human TR-LBDs at the TH binding sites. Molecular docking showed that the mono-ether structural analogs fit into ligand-binding pockets of TR-LBDs (Figure 3; Tables S7 and S8), which supported the binding of these chemicals to TRs. Based on the results of the competitive binding assay and molecular docking, we demonstrated the binding ability of TBBPA mono-ether structural analogs to the TR-LBDs ( Figure 4A-C) and even had high potentials similar to that of the natural TH T 3 .
It has been demonstrated that activation of TR in rat pituitary tumor GH3 cells regulates cell proliferation activity (Gutleb et al. 2005). Therefore, the GH3 cell proliferation assay has been widely applied to study the cellular TR-disrupting potencies of pollutants (Kitamura et al. 2002;Qin et al. 2019). Our results showed that the TBBPA mono-ether structural analogs induced GH3 cell proliferation, and these effects were inhibited by the specific TR antagonist amiodarone ( Figure 5A-C). In addition to inducing GH3 cell proliferation after individual exposure, TBBPA and the TBBPA mono-ether structural analogs further enhanced T 3 -induced GH3 cell proliferation ( Figure 6A-D), which is in line with the results of a previous study on TBBPA (Kitamura et al. 2002). In some studies of other pollutants, such as tributyltin and triphenyltin, it was suggested that these chemicals might enhance T 3 -induced TR activation in GH3 cells through binding to RXR (Mengeling et al. 2016(Mengeling et al. , 2018. Therefore, the role of TBBPA and TBBPA mono-ether structural analogs binding to other receptors in the enhancement of T 3 -induced GH3 cell proliferation cannot be excluded, which warrants further detailed investigation. Notably, we found that one of the tested TBBPA mono-ether structural analogs, TBBPA-MDBPE, showed much higher GH3 cell proliferation activity than TBBPA (Figure 5C,D; Table 1). This did not appear to be correlated with the results obtained from the TR-LBDs binding assays and molecular docking assays, which showed that TBBPA-MDBPE had much weaker binding potencies to the TR-LBDs than TBBPA (Table 1). By using the developed LC-ESI-MS/MS method, we demonstrated that the intracellular amount of TBBPA-MDBPE was 105-fold higher than that of TBBPA, suggesting that TBBPA-MDBPE has a higher ability to transport across the cell membrane into GH3 cells than TBBPA ( Table 2). The proper hydrophobicity of TBBPA-MDBPE might lead to its higher membrane transportation ability, which could then lead to its higher potential to stimulate GH3 cell proliferation. The bioavailability of a pollutant, which normally refers to a complex in vivo outcome-including absorption, distribution, metabolism, and excretion-is relevant to its toxicity (Bradham et al. 2018). The membrane transportation ability of a pollutant, which is directly correlated to its hydrophobicity, is one of the factors affecting bioavailability (Semple et al. 2004). Therefore, the membrane transportation ability of TBBPA-MDBPE to cells is a critical factor that determines its higher potential to stimulate GH3 cell proliferation.
The presence of TBBPA derivatives and analogs in the environment has attracted a large amount of attention, especially regarding their potential toxicity due to their unique structures. Previously, a few studies have discussed the endocrine-disrupting potency, neurotoxicity, and preadipocytes of TBBPA and TBBPA analogs (Eng et al. 2019;Liu et al. , 2020Pullen et al. 2003). However, detailed molecular interactions have not been elucidated, leaving the mechanism underlying the different toxicities of TBBPA analogs remaining elusive. Here, in terms of TH disrupting potency, based on in vitro and in silico methods, we demonstrated exact interaction sites between the chemicals and associated receptors. The TBBPA mono-ether Table 2. Limit of detection (LOD), limit of quantification (LOQ), recovery rates at different concentrations, and the determined intracellular amount of TBBPA-mono(2,3-dibromopropyl ether) (TBBPA-MDBPE) and tetrabromobisphenol A (TBBPA).

Results
TBBPA-MDBPE TBBPA LOD (ng/mL) 0.10 0.13 LOQ (ng/mL) 0.32 0.43 Recovery at 10 ng=mL (%) 68 ± 27 76 ± 18 Recovery at 100 ng=mL (%) 70 ± 8 104 ± 19 Recovery at 1,000 ng=mL (%) 80 ± 10 120 ± 12 Intracellular amount (ng) 2,013:56 ± 91:86 14:10 ± 2:50 structural analogs also showed higher membrane transportation ability than TBBPA, which determined the higher potential to stimulate the proliferation of GH3 cells. Taken together, our results showed a possible molecular mechanism underlying the higher TH disrupting potency of TBBPA mono-ether structural analogs. The exposure sources of TBBPA and TBBPA analogs are contaminated environmental matrices, food, and commercial products containing TBBPA (Abou-Elwafa Abdallah 2016; Liu et al. 2015Qu et al. 2011). Recently, TBBPA in human breast milk in the general population in China was reported to be approximately at 1:57 ng=g lipid weight (Huang et al. 2020). The TBBPA concentration in human serum samples collected from the general population in Korea was approximately 45:6 ng=g lipid weight ( ∼ 0:5 nM) (Kim and Oh 2014). To date, because of the lack of commercial standards, there are no data about the exposure levels of TBBPA derivatives or TBBPA mono-ether structural analogs in human samples. Therefore, additional large-scale surveys on the burden to the human body of these chemicals are still needed in the future. The increasing production and usage of TBBPA bis-ether derivates and the continuous transformation into TBBPA mono-ether structural analogs could lead to their higher environmental and human exposure levels (Qu et al. 2013). Currently, the transformation of TBBPA bis-ether derivatives into TBBPA mono-ether structural analogs in the environment has been demonstrated. However, it is unclear whether the TBBPA derivatives could be transformed into TBBPA mono-ether structural analogs by metabolic enzymes in humans, which also needs to be studied in the future. Because some in vivo evidence has shown that TBBPA exposure impacted the TH axis, the effect of TBBPA mono-ether structural analogs on the TH system should be given great attention. Comprehensive in vivo animal toxicology experiments and human exposure studies are needed in the future and which are important to fully evaluate the risk of these TBBPA structural analogs.

Conclusion
Based on in vitro and in silico methods with clear mechanisms, we demonstrated the potential TH system-disrupting effects of TBBPA analogs by binding to TTR or TRs. The derivative groups also play a critical role in the cell membrane transportation ability, which in turn determines the effect on GH3 cells. Our study provides clear mechanisms regarding the higher toxicity of TBBPA mono-ether structural analogs than that of TBBPA and TBBPA bis-ether derivatives. To design safer TBBPA derivatives, proper derivative groups should be considered not only for the avoidance of high binding potency to TH receptors but also for the reduction of membrane transportation ability.