Biomonitoring Equivalents (BEs) for tetrabromobisphenol A

Regulatory Toxicology and Pharmacology 102 (2019) 108–114

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Biomonitoring Equivalents (BEs) for tetrabromobisphenol A Sean M. Hays , Christopher R. Kirman

T

∗

Summit Toxicology, L.L.P., Bozeman, MT, USA

ARTICLE INFO

ABSTRACT

Keywords: Biomonitoring Equivalents Risk assessment Pharmacokinetics Tetrabromobisphenol A

Tetrabromobisphenol A (TBBPA) is a flame retardant used in a variety of products, including epoxy and polycarbonate resins. Relevant exposure to TBBPA has been assessed by measuring TBBPA in the blood of humans. Here, we derive Biomonitoring Equivalents (BEs) for TBBPA to interpret these, and future biomonitoring results for TBBPA in humans. The available toxicity risk values (TRVs) for TBBPA were all based on toxicology studies in rats. Several studies have been conducted in which TBBPA in blood of rats were measured following controlled oral doses of TBBPA. These data provide a robust relationship from which to derive BEs. BEs of 5.6 and 13.0 μg total TBBPA/L plasma were calculated for available cancer and non-cancer TRVs, respectively. Several studies have measured TBBPA in serum, with median concentrations less than 0.1 μg/L, indicating considerable margins of safety (MOS) for TBBPA based on the currently available biomonitoring studies.

1. Introduction Large biomonitoring studies, such as those conducted as part of the National Health and Nutrition Examination Survey (NHANES) in the United States (US), the Canadian Health Measures Survey (CHMS) and other national biomonitoring efforts such as those conducted in France and in Germany, are providing valuable data on the prevalence and concentrations of chemicals in biological matrices in the general population. Measured concentrations of chemicals in biological media such as blood or urine provide an integrated reflection of exposures that may occur via multiple routes and pathways (Sexton et al., 2004). However, due to the lack of guidance values for interpreting biomonitoring data for most environmental chemicals, these data are typically presented without any interpretation in the context of potential health risks. Biomonitoring Equivalents (BEs) provide one tool for evaluating such data in the context of existing risk assessments (Hays et al., 2007, 2008; LaKind et al., 2008). A BE is defined as the concentration or range of concentrations of a chemical or its metabolites in a biological medium (blood, urine, or other medium) that is consistent with an existing health-based exposure guidance value such as a reference dose (RfD) or Tolerable Daily Intake (TDI). Existing chemical-specific pharmacokinetic data are used to estimate biomarker concentrations that are consistent with the Point of Departure (POD) used in the derivation of an exposure guidance value (such as the RfD or TDI), and with the exposure guidance value itself. BEs can also be estimated corresponding to other types of exposure guidance factors, including recommended intakes of nutritionally

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essential elements (Hays et al., 2014). BEs can be estimated using available human or animal pharmacokinetic data (Hays et al., 2008), and BEs have been derived for over 100 compounds to date. BE values for multiple chemicals have been used to evaluate CHMS (St. Amand et al., 2014) and NHANES (Aylward et al., 2013) biomonitoring data across chemicals to examine relative levels of exposure in the context of risk assessment-derived exposure guidance values. In continuation of the effort to improve the interpretation of biomonitoring data, this work presents the background and derivation of BE values for tetrabromobisphenol A (TBBPA; CAS #79-94-7, mw = 543.9 g/mol). TBBPA is used as a reactive flame retardant in a variety of products, including epoxy and polycarbonate resins. Thus, for these uses, it is chemically bonded to chemical constituents, which diminishes the ability of TBBPA to migrate from products (Selleström and Jansson, 1995). TBBPA is also used as an additive flame retardant, but at much lower production amounts, in acrylonitrile butadiene styrene (ABS) (WHO, 1995). Assessing exposures to chemicals in the environment and in consumer products can be difficult, especially when exposures may occur via multiple sources and routes. Flame retardants have the added complexity of exposure sources; for example, dust emissions from consumer products in the home and workplace (Hays et al., 2006). As a result of this difficulty, more researchers and government agencies are relying more on biomonitoring instead of external dose characterization in order to assess exposures amongst humans, either as part of epidemiology studies or as regional and/or national exposure and health monitoring programs. The BEs created here can be used to interpret

Corresponding author. E-mail address: [email protected] (S.M. Hays).

https://doi.org/10.1016/j.yrtph.2018.12.014 Received 4 May 2018; Received in revised form 29 October 2018; Accepted 19 December 2018 Available online 26 December 2018 0273-2300/ © 2019 The Authors. Published by Elsevier Inc. This is an open access article under the CC BY-NC-ND license (http://creativecommons.org/licenses/BY-NC-ND/4.0/).

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existing and future biomonitoring data for TBBPA in a public health risk context.

available risk assessments conducted for TBBPA and are organized by those completed before and after the completion of the NTP study. Pre-NTP study, The European Chemical Bureau (ECB) published in 2006 a risk evaluation for TBBPA. They chose a NOAEL (No Observed Adverse Effect Level) of 40 mg/kg-d for polycystic lesions associated with kidney tubule dilation in offspring from a rat developmental toxicity study (Fukuda et al., 2004). The ECB did not derive a TRV. The European Food Safety Authority (EFSA) in 2011 selected a Point of Departure (POD) for conducting a margin of exposure (MOE) analysis for TBBPA. The EFSA calculated a benchmark dose lower limit a the 10% response level (BMDL10) of 16 mg/kg-d for the endpoint of thyroid hormone homeostasis in rats. In 2012, Health Canada completed a screening evaluation for TBBPA. They chose a LOAEL (Lowest Observed Adverse Effect Level) of 140 mg/kg-d as a POD for an MOE evaluation. The 140 mg/kg-d LOAEL was associated with liver toxicity among offspring from a developmental study (Tada et al., 2006). Post-NTP study, the US EPA OPPT has concluded that TBBPA is likely to be carcinogenic to humans (EPA, 2014) but have not published a cancer or non-cancer TRV for TBBPA to date. Wikoff et al. (2015) has conducted the most thorough risk assessment to date and it was published after the release of the NTP study and thus included the results from the NTP study. Wikoff et al. (2015) calculated a non-cancer reference dose (RfD) (based on uterine hyperplasia) based on the National Toxicology Program (NTP, 2013) oral gavage dosing study in rats (Table 1). Pecquet et al. (2018) have calculated a RfDcancer for TBBPA based on incidence of uterine tumors from the NTP study. Pecquet and coauthors decided that the mode of action (MOA) for uterine tumors warranted a threshold based risk assessment. They utilized a rat to human body weight scaling factor to scale the rat doses to human equivalent doses and calculated a BMDL10 in rats to derive a human equivalent (HED) POD of 25 mg/kg-day. They then chose a composite uncertainty factor (UF) of 100 to derive a RfDcancer. These various risk assessments (Table 1) are used in the derivation of the BEs.

2. Methods Deriving a BE requires converting a regulatory guidance value in terms of an external dose (oral, inhalation or dermal) into a corresponding internal (e.g., blood) or urinary biomarker concentration (Hays et al., 2007). When deriving BEs, there are three critical evaluations as dictated by the BE Derivation Guidelines (Hays et al., 2008).

• Identify

•

•

available exposure guidance values established for the compound. These may include robust toxicity reference values such as Reference Dose (RfD), Tolerable Daily Intake (TDI), cancer potency factor (CPF) or robust governmental evaluations of the available data which conclude with selection of or calculation of a point of departure (POD) to serve as the basis of a margin of exposure (MOE) risk characterization. Identify available biomarker(s) for the compound and choose the best available biomarker for the compound that is most easily interpreted in a risk assessment context (Hays et al., 2008). While numerous biomarkers may be available, several factors should be weighed when choosing the best biomarker for support of a BE (Hays et al., 2008). These include an assessment of the specificity of the biomarker to exposures to the compound of interest, and, when multiple biomarkers are available, selection of the best biomarker should take into consideration the half-life of elimination and its relevance towards the mechanism of action for the toxic effects (Hays et al., 2008). Identify pharmacokinetic data/model for converting external doses into corresponding biomarker concentrations. In some cases, robust physiologically based pharmacokinetic models are available, but such models may not be required. Other useful data and models include simple non-physiological pharmacokinetic models, data from controlled dosing studies evaluating urinary excretion fractions, and direct measurement of biomarkers in appropriate animal studies.

2.2. Potential biomarkers Biomonitoring for exposures to TBBPA has primarily involved measuring total TBBPA (parent compound plus conjugates) in whole blood and/or serum (Kim et al., 2014; Dufour et al., 2016; Dirtu et al., 2008). A review of TBBPA biomonitoring data has been compiled by EPA, and includes measured levels of total TBBPA in milk, blood, adipose tissue and hair (https://www.epa.gov/assessing-and-managingchemicals-under-tsca/supplemental-file-1-tetrabromobisphenol-tbbpacasrn-79). TBBPA-glucoronide could be detected in urine of human volunteers experimentally dosed with TBBPA orally at 0.1 mg/kg body weight (Schauer et al., 2006), but no TBBPA or conjugates has been found in urine of the general population, despite one attempt to date (Yang et al., 2014). Schauer et al. (2006) only reported that less than 0.1% of the dose (0.1 mg/kg) was recovered in urine, thus not providing sufficient data to support derivation of a BE. Two studies have

These evaluation steps are described below for TBBPA. 2.1. Exposure guidance values and critical effects TBBPA has been the subject of numerous toxicity studies in laboratory animals. Despite this, very few toxicity risk values (TRVs – e.g., Reference Dose, Tolerable Daily Limit, etc.) have been derived by agencies for TBBPA (Table 1). The National Toxicology Program (NTP) finished a two-year bioassay in 2013, which included previously unobserved findings of tumor responses in animals. As a result, risk assessments conducted prior to the completion of the NTP study were based on noncancer endpoints. The following is a summary of the Table 1 Available toxicity reference values for TBBPA. Reference

Year

Type of TRV

Critical Study

Critical Effect

POD type

POD (mg/ kg-d)

HED (mg/ kg-d)

UF

TRV (mg/ kg-d)

Wikoff et al. (2015) Pecquet et al. (2018) EFSA (2011) ECB (2006)

2014 2017 2011 2006

RfD RfDcancer POD POD

72.8 103 16 40

18.2 25.6 NA NA

30 100 NA NA

0.6 0.26 NA NA

2012

POD

Uterine hyperplasia in rats Uterine tumors in rats Thyroid hormone homeostasis Polycystic lesions associated with kidney tubule dilation among offspring liver toxicity among female offspring

BMDL10 BMDL10 BMDL10 NOAEL

Health Canada (HC, 2012)

NTP (2013) NTP (2013) NR Fukuda et al. (2004) Tada et al. (2006)

LOAEL

140.5

NA

NA

NA

1 - NSRL reported as 20 mg/day (0.26 * 70 kg body weight). Body weight adjusted NSRL reported here. NR - Not Reported. NA - Not Applicable. 109

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detected TBBPA in maternal milk (Shi et al., 2009, 2013). BEs are not typically derived for chemicals in milk. The presence of a chemical in milk poses a hazard to the nursing infant. However, because a risk assessment can be done more easily based on external doses (intake) using the concentration of chemical in milk and estimates of milk consumption rates, a BE will not be calculated for TBBPA in maternal milk. When multiple biomarkers exist for exposures to a compound, the preferred biomarker to support derivation of a BE is informed by three primary factors; 1) specificity of the biomarker to exposures to the compound of interest (i.e., the biomarker cannot be formed by exposures to a different parent compound), 2) the mechanism of action of toxicity (e.g., it is more appropriate if the biomarker is the proximate toxicant and not a deactivated metabolite), and 3) the biomarker with the longest half-life so that it is more stable and can be detected longest after an exposure event. TBBPA and the conjugates are not likely to be formed following exposures to any other compound, therefore all three should be specific to TBBPA exposures. Lastly, TBBPA is moderately short-lived and may not be detectable at low environmental exposures, whereas the conjugates may be more likely to be present (Schauer et al., 2006). Given these issues and the fact that total TBBPA (parent plus conjugates) has been used in human biomonitoring studies, total TBBPA in blood will be used for the BE derivations. The following pharmacokinetic data informs these calculations.

oral doses. These include:

• Schauer et al. (2006) gavage dosed rats with a single oral dose of

•

•

•

2.3. Pharmacokinetic data Bioavailability estimates for TBBPA are impacted by the fact that TBBPA undergoes significant first-pass metabolism and elimination via bile (Kuester et al., 2007). Within 2 h of oral dosing in rats, 50% of the oral dose could be recovered in the bile (Kuester et al., 2007). The fraction of dose getting distributed systemically, thus is very low, with systemic bioavailability estimates in rats ranging from 1.6% (Kuester et al., 2007) to 4.8% (Knudsen et al., 2014). Once absorbed, TBBPA is conjugated with glucornide and/or sulfate in rats (Kang et al., 2009; Knudsen et al., 2014; Schauer et al., 2006). Fecal elimination is the primary route of elimination for TBBPA. Greater than 80% of a radioactive iv dose in rats is eliminated in the feces within 36 h post dosing, whereas less than 0.5% was eliminated in urine (Kuester et al., 2007). TBBPA is excreted in urine as TBBPA-sulfate and TBBPA-glucoronide in rats and only TBBPA-glucoronide in humans (Schauer et al., 2006). Total TBBPA in blood has a relatively short terminal half-life of 7–10 h in rats (Kang et al., 2009; Knudsen et al., 2014) and 2.2 days in humans (Hagmar et al., 2000). Very low concentrations (< 2%) are found in all tissues in rats 24 h following oral doses (250 mg/kg of 14C labeled material, 50 μCi/kg), with the greatest concentration of total radioactivity 24 h post dosing remaining in the intestinal lumen. Small but measurable radioactivity could be found in liver, muscle, adipose tissue and blood (Knudsen et al., 2014). TBBPA does not appear to be preferentially distributing to adipose tissue (Knudsen et al., 2014; Kuester et al., 2007). In humans orally dosed with 0.1 mg/kg TBBPA, parent compound was not detected in blood samples, however, TBBPA-glucuronide could be detected for up to 72 h post-dosing (Schauer et al., 2006), suggesting rapid conjugation in humans. TBBPA-glucuronide could be detected in urine samples for at least 96 h post-dosing (Schauer et al., 2006).

•

300 mg/kg TBBPA in corn oil and measured TBBPA, TBBPA-glucoronide and TBBPA-sulfate in plasma for 100 h post-dosing. The authors reported free TBBPA plasma area under the curve (AUC) of 1028 nmol-hr/mL, 161 nmol-hr/mL for TBBPA-glucoronide, and 9251 nmol-hr/mL for TBBPA-sulfate. Kuester at al. (2007) gavage dosed rats with single oral doses of 2, 20 or 200 mg/kg TBBPA in ethanol:cromophor:saline and measured free TBBPA in whole blood for 8 h post dosing. The authors fit a compartmental model to the free TBBPA blood-time profile data and calculated whole blood AUC for only the 20 mg/kg dosed animals (24 μg-min/mL; extrapolated to infinity). We contacted the authors to get access to the data for the other two dose groups, but the data were no longer available from the authors. Kang et al. (2009) gavage dosed rats with single oral doses of 200, 500 or 1000 mg/kg TBBPA in corn oil for either 1 day or for 14 days of repeated daily doses. The authors report serum AUC for the animals dosed for a single day; 107.3, 152.8 and 234.8 μg-hr/mL total TBBPA (free TBBPA plus conjugates) for the 200, 500 and 1000 mg/ kg TBBPA dosed animals, respectively. Total TBBPA was generated by treatment of 100 μL serum samples with 10 μL of acetic acid to cause deproteinization (Kang et al., 2009). Knudsen et al. (2014) gavage dosed rats with single oral doses 25, 250 or 1000 mg/kg TBBPA in ethanol, water and emulsifying agent. Blood samples were collected for 24-h post-dosing. The authors only determined free TBBPA plasma AUC (1729 nmol-min/mL) for the 250 mg/kg dosed animals. Borghoff et al. (2016) gavage dosed rats with 50, 100, 250, 500 and 1000 TBBPA mg/kg in corn oil either once daily for 28 days or for a single day. Blood samples were taken at 4 h post dosing in the single dose study and at 4 and 8 h post dosing following the 28th day of daily dosing. The authors report plasma concentration of TBBPA and glucuronide and sulfate conjugates of TBBPA for the time points sampled. Given that the authors did not collect more time points than the two time points on the days of sampling, daily AUC estimates cannot be generated. However, the data from the 4 and 8 h time points can be used to provide an estimate of average concentrations of each analyte on the day of sampling since the average blood concentration of total TBBPA occurs between 5 and 10 h postdosing (based on an analysis of Kang et al., 2009 and Schauer et al.,2006).

The results of Borghoff et al. (2016) are selected to support derivation of BEs since the results provide a wide range of dose groups, animals were dosed for 28 days (as opposed to single oral doses administered by all the other studies), and the authors measured free and conjugated TBBPA in plasma, which supports derivation of BEs for free and total TBBPA in plasma, both of which are likely to be included in biomonitoring studies. All data from Borghoff et al. (2016) were obtained from the study authors. The 4 and 8 h time point concentrations were averaged. Results from all dose groups (Fig. 1) exhibits some nonlinearities at the higher dose groups (500 and 1000 mg/kg). Therefore, the data from the dose groups of 250 mg/kg and less were used in establishing linear relations between dose and plasma concentration (Figs. 2 and 3). Since the highest POD from the existing exposure guidance values for TBBPA is 140 mg/kg, dose groups of 250 mg/kg and less covers the range of doses of interest for deriving BEs. The linear relationship between parent TBBPA and dose is:

2.4. BE derivation approach All TRVs are based on toxicity studies in rats. Given that biomonitoring for TBBPA is via total TBBPA (parent plus conjugates) or parent TBBPA in blood, the most direct method for deriving the BEs (Hays et al., 2008) for TBBPA is to estimate concentrations of total TBBPA and parent TBPPA in blood in rats at the PODs used in derivation of the TRVs. Several PK studies have been conducted in rats that measured TBBPA and/or total TBBPA (TBBPA plus conjugates) in blood following

TBBPA in plasma (ug/L) = 18.02 * TBBPA Dose (mg/kg)

(1)

The sum of moles of TBBPA, TBBPA-GA and TBBPA-S for each dose group were converted to TBBPA molar equivalents and then to mass concentration (ug/L; see Fig. 3). The relationship between TBBPA dose 110

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Fig. 1. Plasma concentrations of TBBPA, TBBPA-GA and TBBPA-S in rats following 28 days of single daily oral doses of TBBPA. See text for description of data. Source Borghoff et al., 2016.

and total TBBPA equivalents (parent TBBPA and conjugates – consistent with analyzing for TBBPA after deconjugation methods) in plasma is: Total TBBPA in plasma (ug/L) = 21.7 * TBBPA dose (mg/kg)

were converted into corresponding BEPOD_Animal and BEPOD values, respectively, using equation (1) (parent TBBPA in plasma) and 2 (total TBBPA equivalents in plasma). The BEPOD values ranged from 330 to 460 μg TBBPA/L (Table 2) and 390–560 μg total TBBPA/L (Table 3). BEPOD_Animal values ranged from 290 to 2530 μg TBBPA/L (Table 2) and 350–3050 μg total TBBPA/L (Table 3). EFSA, ECB and Canada only chose PODs and conducted margin of exposure analyses. Thus, they only have BEPOD_Animals associated with them. Wikoff et al. (2015) applies UFs to the HED POD to derive a Reference Dose (RfD). The BE associated with the RfD is calculated by dividing the BEPOD by the composite UF of 30 used by Wikoff et al. (2015), yielding a BE of 11.0 μg TBBPA/L (Table 2) and 13.0 μg total TBBPA/L (Table 3). The BE associated with the NSRL derived by Pecquet et al. (2018) was calculated by dividing the respective BEPOD by the composite UF of 100 used by Pecquet et al., yielding a BE of 4.6 μg TBBPA/L in blood (Table 2) or 5.6 μg total TBBPA/L in blood (Table 3). The BEs derived for TBBPA were based on a robust toxicokinetic study in rats that measured parent TBBPA and conjugates in plasma following 28 days of controlled exposures to TBBPA over a fairly wide range of doses. This helps to yield high confidence in the BE estimates. The guidelines for the derivation of BEs indicate that the full

(2)

The RfD established by Wikoff et al. (2015) and the RfDcancer established by Pecquet et al. (2018) in Table 1 calculated PODs and converted these to human equivalent dose (HED) PODs using allometric scaling (see Wikoff et al., 2015 and Pecquet et al., 2018 for additional details). Based on the BE derivation guidelines (Hays et al., 2008), a HED POD is used to calculate a BEPOD (Hays et al., 2008). For both of these TRVs, uncertainty factors (UFs) were applied to the HED PODs to derive the RfD and RfDcancer. The approach for deriving BEPODs and BEs for the RfD and RfDcancer is provided in Fig. 4. The remaining TRVs in Table 1 established PODs, but did not conduct any human equivalent conversions or apply an UFs. In this case, the BE derivation guidelines require calculation of a BEPOD_Animal (Hays et al., 2008) and no BEs (Fig. 5). 3. Results and discussion The PODs and HED PODs used in each of the TRVs listed in Table 1

Fig. 2. Plasma concentrations of TBBPA, TBBPA-GA, TBBPA-S in rats following 28 days of single daily oral doses of TBBPA. Data is the same as Fig. 1, but only for dose groups of 250 mg/kg and less. 111

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Fig. 3. Plasma concentration of TBBPA equivalents (parent TBBPA plus TBBPA molar equivalents of TBBPA-GA and TBBPA-S) in rats following 28 days of single daily oral doses of TBBPA.

complement of UFs should be utilized to convert the BEPOD to the BE, except for when the analyte measured in biomonitoring studies is the parent compound or the proximate toxicant in blood. In this case, the inter-species PK UF is replaced by the use of the exposure metric accounting for species differences in PK. The Wikoff et al. RfD utilized a composite UF of 30 and accounted for species differences by calculating a Human Equivalent Dose (HED) by adjusting the animal derived POD by a body weight scaling factor. Deriving the BEPOD based on the POD in rats, and using the remaining composite UF of 30 to convert the BEPOD to the BE is consistent with the BE derivation guidelines (Hays et al., 2008). Likewise, the NSRL derived by Pecquet et al. (2018) used a composite UF of 100, which included factors of UFA (animal to human) of 3, UFH (human variability) of 10, and UFD (database deficiency) of 3. Pecquet et al. (2018), likewise, calculated a HED POD. The BE associated with the RfDcancer was calculated by dividing the BEPOD by the composite UF (100) used by Pecquet et al.

Fig. 5. Schematic for approach to deriving BEs for TBBPA (parent and total) in plasma associated with the TRVs established by EFSA (2011), ECB (2006), and Health Canada (2012).

Fig. 4. Schematic for approach to deriving BEs for TBBPA (parent and total) in plasma associated with the RfD (Wikoff et al., 2015) and the RfDcancer (Pecquet et al., 2018). 112

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ECB (2006)

BE Derivation Step POD, mg/kg-day 72.8 103 16.0 40.0 HED POD, mg/kg- 18.2 25.6 NA NA day Conv. Factor (TBBPA in plasma (ug/L) = 21.7 * Dose (mg/kg)) BEPOD_Animal (ug/ NAb NAb 350 870 L) BEPOD (ug/L) 390 560 NA NA UF 30 100 NA NA BE (ug/L)a 13.0 5.6 NA NA

Health Canada (2012) 140.5 NA 3050 NA NA NA

NA - Not Applicable. a BE is calculated by dividing the BEPOD by the UF. b A BEPOD-Animal is not calculated since the TRV derived a HED POD.

Table 4 Available biomonitoring data for TBBPA in blood.

Study Population

Biomonitoring of TBBPA exposures reported to date have measured TBBPA in serum and whole blood (Table 4). None of the studies explicitly specified if their analytical methods yielded parent TBBPA only or total TBBPA (parent plus conjugates). Generally, TBBPA serum concentrations are below 0.8 μg/L at the maximum, with central tendencies less than 0.1 μg/L. Some biomonitoring studies have reported lipid adjusted TBBPA blood concentrations (EPA, 2014), however, since TBBPA does not appear to be sequestered in lipid stores (Hakk et al., 2000), it is unlikely this is required or provides any additional insights. We were unable to locate any studies that have analyzed for the distribution of TBBPA within various fractions of blood. The BEs derived here are for TBBPA in the plasma fraction of blood. Future studies on this issue would help guide the biomonitoring community on the most appropriate way to report TBBPA blood concentrations. The BEs reported here are based on measures of total TBBPA (parent TBBPA plus conjugates) and parent TBBPA only. The appropriate BEs should be used to interpret TBBPA biomonitoring data (parent only or total TBBPA) in plasma. The BERfD_cancer of 4.6 μg parent TBBPA/L is the lowest BE calculated. Comparing this to the available biomonitoring studies suggests a margin of safety (MOS) > 40 at the central tendencies of the populations tested. Comparing the biomonitoring data to the BEPODs suggests Margins of Exposure (MOE) in excess of 2000. The BEs carry the same functional definition as the underlying TRV. As such, the TRVs available for TBBPA are to be used as risk management tools to help assist companies and regulatory agencies decide if 113

0.08 0.09 < 0.015 NR NR NR NR

NR – Not Reported. NS – Not Specified.

EFSA (2011)

2004 2004

Pecquet (2018) RfDcancer

TBBPA (parent or total)

Wikoff (2015) RfD

serum serum serum whole blood

Organization and Exposure Guidance Value

Blood Fraction

Organization and Guidance Value

Kicinski et al. (2012) Thomsen et al. (2007) Peters (2005) Peters (2004) WWF (2004)

Table 3 Derivation of Biomonitoring Equivalents for total TBBPA in plasma.

NR NR NR 100 21 ∼33 ∼50

Detection Frequency (%)

NA - Not Applicable. a BE is calculated by dividing the BEPOD by the UF. b A BEPOD-Animal is not calculated since the TRV derived a HED POD.

NS NS NS NS NS NS NS

NA NA NA

serum serum

2530

Individuals (n = 7) Pooled (n = 14) 515 adolescents, age 13.6–17 years Pooled (n = 20) Maternal individual (n = 42) Individuals (n = 91) Individuals (n = 47)

140.5 NA

2007 1999 2008–2011 1975–2002

BE Derivation Step POD, mg/kg-day 72.8 103 16.0 40.0 HED POD, mg/kg- 18.2 25.6 NA NA day Conv. Factor (TBBPA in plasma (ug/L) = 18.0 * Dose (mg/kg)) BEPOD_Animal (ug/ NAb NAb 290 720 L) BEPOD (ug/L) 330 460 NA NA UF 30 100 NA NA BE (ug/L)a 11.0 4.6 NA NA

Health Canada (2012)

Dirtu (2008)

ECB (2006)

Concentration central tendency (μg/L)

EFSA (2011)

Location

Pecquet (2018) RfDcancer

Year

Wikoff (2015) RfD

Belgium Belgium Belgium; Flanders Norway The Netherlands The Netherlands Europe

Organization and Exposure Guidance Value

Study

Organization and Guidance Value

Range (μg/L)

Table 2 Derivation of Biomonitoring Equivalents for parent TBBPA in plasma.

NR NR [ < 0.015–0.186] [0.002–0.005] [0.06–0.19] [0.056–0.787] [0.002–0.333]

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current and future potential exposures to TBBPA pose an unreasonable risk of harm to public health. The TRVs, and thus BEs and BEPODs, are not intended to be used as a diagnostic tool to assess whether individuals with TBBPA in their blood will or will not experience adverse health outcomes.

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