Background paper on the risk assessment of Bisphenol A

Transcription

1 Background paper on the risk assessment of Bisphenol A Overview of previous risk assessments Prepared for the joint FAO/WHO expert meeting to review toxicological and health aspects of bisphenol A Ottawa, Canada 2 5 November 2010 Prepared by: Anna Beronius Institute of Environmental Medicine, Karolinska Institutet, Stockholm, Sweden Professor Annika Hanberg Institute of Environmental Medicine, Karolinska Institutet, Stockholm, Sweden Professor Christina Rudén Royal Institute of Technology, Department of Philosophy, Stockholm, Sweden

2 Table of Contents 1. Introduction Existing BPA risk assessment documents scope and conclusions regarding human health risks Risk assessments available from Europe Risk assessments available from the US and Canada Risk assessment available from Japan Estimations of human exposure levels Identification of critical study, critical effect and NOAEL Use of assessment factors and calculations of TDI and MOS/E Significance of low-dose data for human health risk assessment References Tables

3 1. Introduction This background document was written in order to summarize the health risk assessments of bisphenol A (BPA) conducted to date. It includes a summary of the main conclusions from existing risk assessment documents regarding health risks in the general population, as well as a review of how human exposure levels were estimated, the identification of critical effect(s) and corresponding NOAELs, the use of assessment, or uncertainty, factors and how data on so-called low-dose effects were handled. Several health risk assessments of BPA have been conducted recently by regulatory authorities as well as expert groups in Europe (SCF, 2002; ECB, 2003 and 2008; EFSA 2006,2008 and 2010), the United States (vom Saal et al., 2007; NTP-CERHR, 2008; US FDA, 2008a), Canada (Health Canada, 2008) and Japan (AIST, 2005). The format and structure vary depending on what type of authority conducted the assessment and the intended end-use. Some were carried out by regulatory agencies for the purpose of evaluation of the margin of safety/exposure (MOS/E) or proposal for a tolerable daily intake (TDI). Others were conducted by government-funded expert groups in order to answer specific questions regarding potential human health risks. In all cases potential adverse health effects of BPA were identified and evaluated and human exposure levels were estimated in order to draw conclusions about health risks at current exposure levels but not all risk assessments have set healthbased guidance values. However, conclusions regarding whether or not BPA poses a risk to human health vary between these assessments. BPA is an industrial chemical used to manufacture polycarbonate plastics and epoxy resins. It is produced in large volumes worldwide and measured concentrations of BPA in human blood, urine and other tissues confirm that exposure is widespread in the human population (Vandenberg et al., 2007; Calafat et al., 2008). In most of the risk assessments of BPA conducted to date food has been considered to be the main source of exposure. However, it has recently been suggested that other exposure routes, e.g. via the skin, may contribute significantly to the total exposure of BPA (Vandenberg et al., 2007; Stahlhut et al., 2009). The no observed adverse effect level (NOAEL) for BPA established for regulatory purposes in the US and Europe is currently 5 mg/kg body weight (bw) and day (EFSA, 2006; US FDA, 2008a). This NOAEL is based on two multigeneration studies in rats and mice, respectively (Tyl et al., 2002 and 2008). Above this NOAEL effects on adult and offspring body and organ weights in rats as well as liver effects in adult mice were observed. However, over the last decade a large number of studies have reported effects of BPA exposure at doses in the µg/kg bw-range (so called low-dose effects), leading scientists and others to question the NOAEL, which is three orders of magnitude higher. The first low-dose effects to be reported were effects on the development of the male reproductive tract in mice after exposure in utero (Nagel et al., 1997; vom Saal et al., 1998). Since then a significant number of animal studies have reported additional effects after low dose exposure, primarily developmental effects on male and female reproductive organs and function (e.g. Gupta, 2000; Aikawa et al., 2004; Nikaido et al., 2004; Markey et al., 2005; Timms et al., 2005; Ho et al., 2006; Durando et al., 2007) and neurobehavioral developmental effects (e.g. Adriani et al., 2003; Carr et al., 2003; Negishi et al., 2004; Ryan and Vandenberg, 2006; Nakagami, 2009, Cox et al., 2010). 3

4 Canadian authorities have banned the use of BPA in feeding bottles following the risk assessment conducted by Health Canada in Furthermore, while the European Food Safety Authority (EFSA) has stated that BPA poses no risk to the general population at current exposure levels (EFSA, 2006, 2008 and 2010), France and Denmark have banned the use of BPA in products intended for young children, such as feeding bottles, cups and other food contact materials. Similar bans have been introduced in several US states and counties. 2. Existing BPA risk assessment documents scope and conclusions regarding human health risks Risk assessment scopes and conclusions regarding risk to human health have been summarized in Table 1. Conclusions relating to risk have been expressed differently in the different risk assessment documents, primarily because of different intended purposes of the evaluations, i.e. whether the purpose was to calculate a TDI or a MOS or to answer specific questions regarding potential human health risks. 2.1 Risk assessments available from Europe A European opinion on consumer exposure to BPA via the diet is available from the former Scientific Committee on Food (SCF, 2002). This brief opinion assessed the exposure of the general population to BPA via food, focusing mainly on carcinogenicity and reproductive toxicity. The SCF set a temporary TDI of 10 µg/kg bw and concluded that the estimated worst case exposure via foodstuffs, ranging between 0.48 and 1.6 µg/kg bw/day, was below the TDI. In 2006 EFSA published an opinion on BPA up-dating the SCF document (EFSA, 2006). EFSA assessed exposure to the general population via food, and particularly the exposure of infants, and focused on carcinogenicity and reproductive toxicity. It also paid specific attention to the low-dose studies of BPA. EFSA established a new TDI of 50 µg/kg bw and concluded that conservative estimates of exposure were less than 30% of the TDI for all population groups. EFSA also conducted a review in 2008 specifically addressing new data indicating that fetuses and newborns may not be able to metabolize BPA as efficiently as adults (EFSA, 2008). However, it was concluded that the risk assessment conducted in 2006 could be considered conservative for humans, including fetuses and infants, and no changes in the TDI for BPA were made. EFSA adopted a third opinion on BPA in September 2010 in which they 1) evaluated a dietary developmental neurotoxicity study in rats (Stump et al., 2010), 2) evaluated scientific literature published between 2007 and July 2010 in terms of relevance for the risk assessment of BPA, and 3) provided advice on the Danish evaluation of the Stump et al. study underlying the Danish ban on BPA in food contact materials for children aged 0-3 years. EFSA concluded that the study by Stump et al. was inconclusive and of limited value in the risk assessment of BPA. Further, EFSA stated that data published after 2006 did not warrant a revision of the TDI for BPA, which remains at 50 µg/kg bw. However, a minority opinion from one of the panel members was included stating that there are significant uncertainties about the validity of the NOAEL on which the TDI was based and that the TDI should be considered as temporary. Uncertainties were particularly expressed in regard to effects on brain receptor programming, immune modulation and enhancement of susceptibility to breast at doses below the current NOAEL. 4

5 Another European risk assessment of BPA was conducted within the EU Existing Substances Program for the former European Chemicals Bureau (ECB, 2003). The ECB assessment had a broader scope than the assessments conducted by the SCF and EFSA, considering effects on the environment as well as human health. The assessment was carried out in accordance with the guidelines set up for risk assessment within the Existing Substances Program and conclusions regarding risk were made using a MOS approach. No clear comment was made regarding the sufficiency of the MOS but the ECB concluded that there was currently no concern for human health risks for the general population. However, the EU member states involved in the assessment could not come to full agreement on the interpretation and relevance of studies reporting effects on development at low doses, in the µgrange. This disagreement was noted and further testing to investigate the potential of BPA to cause adverse effects on development at low doses was requested. An update of the ECB human health risk assessment was published in 2008, which considered new data generated since 2003 (ECB, 2008). Its purpose was to address the uncertainties expressed in the previous assessment concerning a NOAEL for developmental toxicity and report on the findings of additional developmental studies requested in that assessment. Thus the focus was specifically on the exposure of infants via food. The conclusion was that MOS values were sufficiently high and that there was no concern for human health risks in relation to any endpoint included in the assessment for any part of the general population, including infants. However, Denmark, Sweden and Norway disagreed with this conclusion in regard to developmental neurotoxicity. 2.2 Risk assessments available from the US and Canada In November 2006 the meeting Bisphenol A: An examination of the relevance of ecological, in vitro and laboratory animal studies for assessing risks to human health was held in Chapel Hill in the US. The meeting was sponsored by the NIEHS and US EPA and the purpose was to address the potential relationship between BPA exposure and negative trends in human health occurring in recent decades. Several panels of experts, mainly from the US but also from Japan and Europe, representing different disciplines were put together to tackle this question. Exposure levels in the general population were also discussed and estimated. The results and conclusions from these panels have been presented in a number of published articles (Crain et al., 2007; Keri et al., 2007; Richter et al., 2007; Vandenberg et al., 2007; vom Saal et al., 2007; Wetherill et al., 2007). In this paper these reports are collectively referred to as Chapel Hill, The Chapel Hill assessment focused on the exposure of the general population to BPA via food and the environment. Effects on development were specifically evaluated but effects after exposure in adulthood were also discussed, such as neurobehavioral effects, effects on fertility and the immune system and metabolic disorders. The Chapel Hill experts estimated much higher levels of human exposure to BPA than previous assessments (see section 3 on Estimations of human exposure levels ). In contrast to other assessments they also judged that studies reporting effects at low doses, in the µg-range, were relevant for human risk assessment (see section 6 on Significance of low-dose data for risk assessment ). The conclusion of the Chapel Hill assessment was thus that there is great cause for concern, for all population groups, in regard to health risks from BPA. Another assessment of BPA was conducted in the US by an expert panel at the National Toxicology Program Center for the Evaluation of Risks to Human Reproduction (NTP-CERHR, 2008). The risks from exposure via food and the environment were assessed and the focus was specifically to evaluate the reproductive toxicity of BPA at low doses. The NTP expressed some concern for effects on the brain, behavior, and prostate gland in fetuses, infants, and children and minimal concern for 5

6 effects on the mammary gland and an earlier age for puberty for females in fetuses, infants, and children. The Chapel Hill and NTP-CERHR assessments are not quantitative risk assessments of BPA, i.e. no MOS/E calculations or quantitative comparisons between a TDI and exposure levels were conducted. However, both assessments compared levels of estimated human exposure to doses where adverse effects of BPA occur and stated qualitative conclusions regarding risk. In 2008 the US FDA issued a draft health risk assessment of BPA considering exposure to the general population via food and focusing primarily on infants and specifically developmental toxicity of BPA (US FDA, 2008a). Exposure levels and MOS were calculated for males and females separately. The FDA concluded that MOS values were sufficient and that there was no risk to any part of the population. This draft was then reviewed by a Subcommittee of FDA s Science Board, which released its report in October 2008 (US FDA, 2008b). The Subcommittee identified several significant concerns with the draft assessment. For example, the Subcommittee did not agree with the FDA s exclusion of the large number of non-glp studies that indicate effects at low doses (see section 6 on Significance of low-dose data for risk assessment ) or the NOAEL used as point of departure for calculating MOS. The FDA published an update in January 2010 stating that studies on low-dose effects of BPA cited by the Subcommittee and in the NTP-CERHR assessment from 2008 had been reviewed and that the FDA now shares the opinion of the NTP, i.e. that these studies provide reason for some concern regarding potential effects on the brain, behavior, and prostate gland in fetuses, infants, and children (US FDA, 2010). The FDA is currently conducting further studies on the safety of low doses of BPA, including studies carried out in collaboration with the National Institute for Environmental Health Sciences. The results from those studies are expected to be available by 2012 and will then be considered in further assessment of BPA by the FDA. A risk assessment of BPA is also available from the Canadian Health authority (Health Canada, 2008). This assessment evaluated the risk posed by BPA to the environment and to the general population, covering exposure of humans via food as well as via the environment. Focus was on the evaluation of the carcinogenicity and reproductive toxicity. A MOE approach was used. Although the derived MOE values were considered to be sufficient the uncertainties surrounding potential developmental neurotoxicity at low doses of BPA (below the established NOAEL) led to the conclusion that BPA may be present at levels constituting a risk to human health. 2.3 Risk assessment available from Japan A Japanese risk assessment of BPA conducted in 2005 is, since 2007, available in English from the Japanese National Institute of Advanced Industrial Science and Technology (AIST, 2005). The risk posed by BPA to the environment and to the general population, covering exposure of humans via food as well as via the environment, was evaluated. Exposure levels and MOE were calculated for males and females separately. MOE values were considered to be sufficiently high and the AIST concluded that there was no concern for human health risks in regard to any of the endpoints investigated or any part of the population. 3. Estimations of human exposure levels 6

7 The exposure assessments for the general population conducted in the different risk assessment documents have been summarized in Table 2. The type of exposure considered as well as conclusions concerning the most highly exposed subgroups and their highest exposure levels are presented. The three opinions from EFSA are based on the same exposure assessment and have therefore been combined in Table 2. In many of the assessments exposure both from dietary sources and from the environment were considered. Different approaches to estimate human exposure, i.e. basing calculations on environmental levels of BPA combined with food intake data and/or on biomonitoring data, were used in the different assessments. In the assessments by the SCF, EFSA and US FDA only dietary exposure was considered. However, the diet was generally assumed in most of the assessments to be the major source for BPA exposure in the general population. Most risk assessments of BPA have based estimates of exposure via the diet on studies or assumptions regarding food intake and BPA-concentrations in certain food commodities. The use of maximum and/or average values for BPA concentrations in food and food intake have varied between assessments. A probabilistic approach to exposure assessment was used only in the AIST risk assessment. The AIST considered ranges of BPA concentrations in food and food consumption and used Monte Carlo simulations to express exposure in the general population as probability distribution functions. In addition to estimating exposure based on levels in the environment and food the AIST and NTP- CERHR also calculated total human exposure to BPA based on urinary concentrations from biomonitoring data. These calculations of exposure generated lower estimations of exposure for adults and children older than 6 years than those based on environmental levels and food consumption data. No calculations based on urinary concentrations were made for small children (< 6 years old) and comparisons cannot be made for this age group. Estimated exposure levels based on urinary concentration were discussed but not included in the risk characterization in several of the other assessments (EFSA, 2006; ECB, 2008; US FDA, 2008a; Health Canada 2008). Only the Chapel Hill assessment includes estimations of exposure based on circulating blood levels of BPA. In most of the investigated risk assessments estimated exposure levels varied between µg/kg bw/day. However, the Chapel Hill experts argued that available data indicate that an oral intake of 100 mg BPA/day, i.e. 1.5 mg/kg bw/day considering a 65 kg human, would be required to explain the circulating levels of unconjugated BPA in blood (Vandenberg et al., 2007). In contrast to the other assessments investigated here, which based exposure estimations on food intake or urinary concentrations, the Chapel Hill experts based their estimation on reported human blood concentrations of BPA and physiology-based pharmacokinetic (PBPK) modelling. Human median blood levels were compared to expected blood levels after an oral dose of 50 µg/kg bw/day from rodent metabolism studies. Since no metabolism studies investigating the kinetics of BPA after oral administration of low doses (at or below 50 µg/kg bw/day) were available blood levels from 11 studies in rodents given oral doses between 500 µg/kg bw 1 g/kg bw were scaled down to represent an oral dose of 50 µg/kg bw/day. It was concluded by the Chapel Hill expert group that human exposure would have to greatly exceed 500 µg/kg bw/day in order to account for the current human circulating levels of BPA. To further support this conclusion the Chapel Hill experts referred to the observations of Shin et al. (2004) who found in their rodent PBPK model that an oral intake of 7

8 100 mg BPA/day was needed to explain the mean human circulating level of 1.49 ng BPA/ml reported in a Japanese study (Takeuchi et al., 2002). It was discussed by the Chapel Hill experts that the high BPA blood concentrations could indicate other important routes of exposure than orally via food or that human metabolism of BPA is not as efficient as generally believed. All assessments concluded that infants and/or young children probably have the highest BPAexposure in the general population. This can be explained by a relatively high dietary intake via polycarbonate feeding bottles and tableware as well as canned foods. Children also consume a large amount of food per kg bw compared to adults. In the ECB assessments modeled worst case environmental exposure scenarios for people living close to polyvinylchloride (PVC) or BPA production sites 1 resulted in exposures of 60 and 40 µg/kg bw/day, respectively, which is relatively high compared to the estimated exposure via food. In the AIST and US FDA assessments different exposure estimations were conducted for males and females. Female exposure levels were calculated to be slightly higher than exposure in males in both assessments. In the AIST assessment the average daily intake estimate for 1-6 year olds was the same (1.2 µg/kg bw/day) for boys and girls while the 95 th percentile value was marginally higher for girls (4.1 vs. 3.9 µg/kg bw/day). It was, however, stated that this difference was not due to the presence of sex-specific exposure pathways but rather gaps in food consumption data and average body weight (AIST, 2005). Estimations for exposure in adults based on urinary concentrations also indicated slightly higher average and 95 th percentile intake in women ( and µg/kg bw/day, respectively) than in men ( and µg/kg bw/day, respectively). These values were based on urinary concentration data from men used to calculate a daily intake per person, which was then divided by the mean weight for women and men, respectively. In other words, differences in estimated exposure between the genders reflected differences in body weight rather than differences in exposure or kinetics. Similarly, in the assessment by the US FDA BPA-intake per day was the same for infant boys and girls and the differences in daily BPA-intake per kg bw (2.25 vs µg/kg bw/day in 1-2 months old boys and girls, respectively) were due to lower average body weight of girls Identification of critical study, critical effect and NOAEL The critical study, critical effect and NOAEL identified for the general population in each risk assessment have been summarized in Table 3. Notably, one or both of two multi-generation studies conducted by Tyl et al. (2002 and 2008) carried out in rats and mice, respectively, were identified as the critical study or studies for risk characterization in almost all assessments, except by the Chapel Hill expert group. Unpublished reports of these studies were available to regulatory bodies, i.e. the SCF, ECB and EFSA, a couple of 1 In the 2008 ECB update it was concluded that use of BPA in PVC production had been phased out and therefore no longer constitute a relevant source of BPA exposure. 2 It seems unlikely that females would have a higher food intake per kg body weight. Further, the difference in estimated exposure levels between the genders was only very slight in both the AIST and US FDA assessments and it is doubtful that this would represent an important gender-difference. 8

9 years before they were published in the open literature. Thus, for example, the Tyl-study published in 2008 was already evaluated and included in the 2006 assessment by EFSA. The large number of animals used, the wide range of doses investigated and the fact that they were conducted according to GLP and based on enhanced OECD guidelines specifically designed to pick up endocrine disrupting effects were often used as arguments to support the use of the studies by Tyl and co-workers as critical for risk characterization. It should be noted that these studies have also been criticized for errors in the reporting of the ages of animals, the techniques used for organ dissection, the high dose used for a positive control and the type of endpoints examined (Myers et al., 2009). A commentary has been published by Tyl in response to this criticism (Tyl, 2009). There are some variations in how the data by Tyl et al. (2002 and 2008) have been interpreted by different expert groups to establish the critical effect and the NOAEL for BPA. In many cases a NOAEL of 5 mg/kg bw/day for effects on adult body weight (F0 and offspring generations) and offspring organ weights in rats, as well as liver effects in mice, was established as the lowest NOAEL. The AIST, Health Canada and US FDA concluded that these effects were due to systemic toxicity and derived an additional NOAEL of 50 mg/kg bw/day for reproductive effects. These effects included reduced litter size in rats as well as effects on the development of the male reproductive tract and preputial separation in mice. The conclusions reached at the Chapel Hill expert meeting differ markedly from the conclusions of the other risk assessments. Primarily, the Chapel Hill experts argued that the studies reporting effects of BPA exposure at low doses should be considered relevant for human health risk assessment and that a NOAEL has not been clearly established. The NOAELs for the effects reported in the two Tyl-studies, as established in each risk assessment, are summarized in Table 4 and 5. The Chapel Hill assessment was not included since no critical effects from these studies were identified. There is some variation between risk assessments in regard to what effects and NOAELs from these studies were considered pivotal for risk characterization. In the ECB risk assessments the decreased adult body weights and changes in organ weights observed in rats at 50 mg/kg bw/day were not considered biologically relevant because effects were not consistent over generations. In addition, the observed decreases in body weights were judged to be slight (< 10%). Further, the liver changes observed in mice at 50 mg/kg bw/day were not considered toxicologically significant because they were not accompanied by changes in liver weight. Effects that were not consistent over generations or did not show a clear dose-response relationship were generally not considered treatment related in the ECB assessments, even though results were statistically significant. It was therefore concluded in these assessments that 50 mg/kg bw/day was the lowest NOAEL to be carried forward to the risk characterization 3. The NTP-CERHR identified a reproductive toxicity NOAEL of mg/kg bw/day for delayed preputial separation in rat offspring from the 2002 study by Tyl and co-workers, which was not identified in 3 Denmark, Sweden and Norway disagreed that this NOAEL also covered developmental neurotoxicity but as the majority of the European Member States supported the NOAEL the position of the Nordic countries was only included as a footnote in the ECB 2008 assessment. 4 The reason the NTP-CERHR reporting this dose as 4.75 instead of 5 mg/kg bw/day, as stated in other risk assessments, is that an average between dietary doses given to adult males and females was used rather than 9

10 any of the other risk assessments. The NTP-CERHR assessment did not focus on general toxic effects, i.e. body or organ weight changes, and therefore the NOAEL of 5 mg/kg bw/day for effects on body and organ weights in rat and liver effects in mouse was not considered. Even though delayed preputial separation was only seen in the F1 generation at 47.5 mg/kg bw/day the NTP-CERHR expert panel identified 4.75 mg/kg bw/day as the pivotal NOAEL to carry forward to risk characterization. In the ECB, AIST and US FDA assessments a NOAEL of 50 mg/kg bw/day was established for this effect. It seems that EFSA (2006) also established that the NOAEL for preputial separation was 50 mg/kg bw/day based on the Tyl study, but this is not very clearly stated. The SCF and Health Canada assessments did not discuss a NOAEL for this effect. The NTP-CERHR and Health Canada expert groups expressed some uncertainty in setting the NOAEL. The NTP-CERHR expert group stated that 4.75 mg/kg bw/day was the NOAEL above which there was sufficient evidence of effects but that there was also limited evidence for other effects, mainly on the development of the brain and behavior, at lower doses in other studies. The same reasoning lies behind the uncertainties expressed by Health Canada concerning their NOAEL of 5 mg/kg bw/day. In response to uncertainties expressed in risk assessments concerning the potential of BPA to cause effects on the development of brain and behavior new studies in rodents investigating developmental neurotoxicity have been generated recently (e.g. Cox et al., 2010; Ryan et al., 2010; Stump et al., 2010; Xu et al., 2010). One study was conducted in compliance with OECD and US EPA developmental neurotoxicity guidelines and GLP (Stump et al., 2010). These studies are discussed in detail in the section on Neurobehavioral Effects in Animals. 5. Use of assessment factors and calculations of TDI and MOS/E A critical aspect of risk assessment is the use of assessment, or uncertainty, factors (Falk-Filipsson et al., 2007) to account for differences in sensitivity between species and between individuals or the lack of such knowledge. Historically, the default assessment factor for health risk assessment has been 100, consisting of factors 10 for inter-, and intra-species differences, respectively. The assessment factor may be adjusted depending on available knowledge about, for example, species differences in toxicokinetics, and confidence in the data material. Additional assessment factors may be added to account for other considerations, such as uncertainty in the NOAEL or lack of data, the nature of the effect(s), duration of exposure or route-to-route extrapolation (Falk-Filipsson et al., 2007). The assessment factor may be applied to derive an estimate of a dose-level considered safe for life-time exposure, e.g. an ADI or TDI, or to evaluate whether the margin between the NOAEL and estimated human exposure (MOS/E) is sufficient. Assessment factors as well as MOS/E and TDI values established in BPA risk assessments have been summarized in Table 3. Assessment factors were not stated in all assessments. In the cases where they were used, these factors often differed between assessments and thus may illustrate how different expert groups have judged the differences in sensitivity between species and individuals, as the dose in females. In the original studies by Tyl et al. [16, 17] actual dose was calculated from concentrations (ppm) in feed, generating slightly higher doses in female animals than in males. 10

11 well as the uncertainty in the data material, differently. However, the scientific justifications for determining the size of the assessment factors were seldom explained or motivated. In the assessments by SCF and EFSA assessment factors were derived in order to calculate a TDI for BPA. In these assessments the default factors of 10 for inter-species differences as well as for interindividual differences were applied, resulting in a total factor of 100. In addition, another factor of 5 was added in the SCF assessment for uncertainties in the database, i.e. the conflicting results regarding effects on development after exposure to low doses of BPA, based on the low-dose studies available at the time. The NOAEL of 5 mg/kg bw/day was thus divided by an uncertainty factor of 500 to yield a TDI of 10 µg/kg bw. The extra factor of 5 was not applied in the assessment by EFSA (2006) because the uncertainties surrounding the low-dose developmental studies were considered to have been resolved. This decision was primarily based on the results from the second multi-generation study by Tyl and co-workers in mice (Tyl et al., 2008). EFSA (2006) stated that new data, generated since the SCF assessment and including the Tyl mouse study, sufficiently supported the conclusion that BPA does not cause adverse effects on development at low doses. The European TDI for BPA was subsequently increased to 50 µg/kg bw. In the EFSA 2010 opinion the use of an assessment factor of 10 for inter-species differences was discussed and was concluded to be conservative based on the available information on toxicokinetics for BPA, i.e. that BPA is eliminated faster in humans than in rodents after oral exposure resulting in lower internal exposure in humans compared to rodents. MOS values were calculated in the ECB and US FDA assessment while MOE values were calculated by the AIST and Health Canada. However, the definitions of MOS and MOE are the same in these cases, i.e. the NOAEL divided by the estimated human exposure. In the AIST assessment a factor of 100 (10 for inter-species differences and 10 for inter-individual differences) was in the analysis of MOE values. Thus, an MOE 100 was considered as sufficient to protect human health. An extra factor for uncertainties in the data material was discussed but not included. In the up-dated ECB assessment (2008) separate MOS values were calculated for 1) increased liver weight and hepatocyte hypertrophy, 2) reduced body weight, increased kidney weight and nephropathy, as well as 3) effects on fertility and development. These MOS values were compared to different assessment factors of 175, 70 and 40, respectively. The factor of 175 for effects on the liver included factors 17.5 for inter-species differences and the default 10 for inter-individual differences in sensitivity. The factor 17.5 for inter-species differences was derived by multiplying the allometric scaling factor of 7, for extrapolating from mice to humans, with a default factor of 2.5 for remaining uncertainties concerning inter-species differences in sensitivity. The factor of 70 for effects on body weight and the kidney included the allometric scaling factor of 7 for inter-species differences in sensitivity and the default 10 for inter-individual differences. The factor of 40 for effects on reproduction included an allometric scaling factor of 4 for inter-species differences in sensitivity, for extrapolating from rats to humans, and the default 10 for inter-individual differences in sensitivity. The default factor of 2.5 for remaining differences in sensitivity between rodents and humans was reduced to 1 for effects on body weight and the kidney or reproduction. The reason given was that in humans BPA is readily excreted in urine after glucuronidation by the liver, while in rodents BPA is excreted via bile and undergoes enterohepatic recirculation resulting in higher levels of free 11

12 circulating BPA and longer half-life compared to humans. It was therefore concluded that humans would be less sensitive to these effects of BPA than rodents. It is noteworthy that this factor was excluded even though information on inter-species differences in toxicodynamics was lacking. Instead its exclusion was based on knowledge of kinetic differences between rodents and humans. Since the internal dose to the liver was considered to be comparable between rodents and humans the default factor was not reduced when considering effects on the liver. No additional factors for differences between experimental conditions and exposure, dose response, i.e. the distance between NOAEL and LOAEL, the type of effects or lack of confidence in the database were considered necessary. The US FDA considered an uncertainty factor of 1000 in their MOS analysis. The factor consisted of factors 10 for inter-species differences, 10 for inter-individual differences and an additional 10 to extrapolate from subchronic to chronic exposure in regard to systemic toxicity, i.e. effects on body and organ weights. However, the FDA Science Board Subcommittee pointed out in their review that the MOS analysis in the FDA assessment was a bit unclear (US FDA, 2008b). The analysis seems to incorporate additional factors of 10 (other than the default 10 each for inter-species and inter-individual differences) both for extrapolating from subchronic to chronic exposure and for irreversible reproductive or developmental effects. It thus appears that with four areas of uncertainty the total assessment factor should be greater than 1,000. Despite the controversies surrounding reported effects of BPA at low doses, and the uncertainties expressed in several risk assessments in regard to evaluating these data, no assessments for BPA have incorporated an assessment factor accounting for uncertainties in the database since the assessment by the SCF in However, in the 2010 EFSA opinion one panel member argued that due to the significant uncertainties in the NOAEL the present TDI should be considered as temporary and that this would usually mean also including an additional assessment factor of Significance of low-dose data for human health risk assessment The debate surrounding BPA is characterized by differing views concerning the significance of the reported low-dose effects for health risk assessment. The significance attributed to the low-dose data can be separated into two parts, i.e. the reliability of the data and their relevance for human health risk assessment. The term reliability can be explained as quality of data, e.g. the reproducibility of results and degree of certainty in these results. Here we are discussing the lowdose data material as a whole and not the significance of individual studies. Table 6 summarizes the conclusions from the BPA risk assessments concerning the reliability and relevance of the low-dose data as well as their overall significance for human health risk assessment. Discussions and conclusions on reliability and relevance in the BPA risk assessments were often intertwined, making it difficult to separate them. There was also little agreement between risk assessments in the use of terminology to express opinions on reliability, relevance and overall significance of data. As a result some interpretation of these discussions was needed to enable the comparisons in table 6. The term reliability was seldom used. However, the reproducibility or robustness of the low-dose study results was sometimes discussed and were interpreted here as indicating reliability. 12

13 In several cases no clear comment on the overall significance of the low-dose data for human health risk assessment was made. However, if they were not considered reliable or relevant, or if they were not used as pivotal data in risk characterization, it must be implied that low-dose data were not considered significant. Conversely, as in the Chapel Hill case, if data were considered to be both reliable and relevant it should follow that they were significant for risk assessment. The Chapel Hill experts were alone in concluding that the low-dose data were both reliable and relevant for human health risk assessment. The low-dose data were not considered significant to the extent that it could be used for risk characterization in any of the other risk assessments. However, some uncertainties in regard to evaluating these data were explicitly expressed in the SCF, ECB (2003), Health Canada and NTP-CERHR assessments. In the 2003 ECB assessment, the EU member states had differing views concerning the reliability of reported low-dose effects on development, as well as their biological relevance to humans. An additional multi-generational study in mice, also covering low doses in the µg/kg bw-range, was requested to settle the uncertainties surrounding developmental toxicity of BPA. As a result the second study by Tyl et al. (2008) was carried out and was available for later risk assessment. This study was conducted according to regulatory test guidelines and GLP and did not show any significant effects at low doses. However, it did not include any investigations of developmental neurotoxicity. Effects on neurobehavioral development, as well as other reproductive toxicity endpoints, were however reported in several additional low-dose studies published after In the assessments from EFSA, ECB (2008) and US FDA the growing data material on low-dose effects was considered to be neither reliable nor relevant. Similar to the 2003 assessment from ECB, however, there was disagreement between the European countries concerning the relevance of certain effects reported in low dose studies in the ECB 2008 assessment. Denmark, Sweden and Norway disagreed with regard to the assessment of developmental neurotoxicity and stated that several of the low-dose studies investigating developmental effects on the brain and behavior could not be dismissed in the risk characterization. No uncertainties concerning the lack of significance of the low-dose data were expressed in the assessments from EFSA or US FDA. However, significant uncertainty about the validity of the NOAEL of 5 mg/kg bw/day was expressed by one of the panel members in the 2010 assessment by EFSA and was included as a minority opinion. Further, compliance with GLP principles and internationally validated and accepted test guidelines, such as the OECD guidelines, was considered a strength or even a quality criterion for evaluating toxicity data in the assessments by ECB, EFSA, NTP-CERHR and US FDA. These considerations thus also influenced the evaluation of data reliability in the BPA case. Requirements of compliance with GLP and test guidelines may result in the exclusion of a large amount of available toxicological data from the final step of risk assessment, i.e. risk characterization, since many studies are carried out in academic settings and not in accordance with test guidelines or standard operational procedures. 13

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16 Nikaido Y, Yoshizawa K, Danbara N, Tsujita-Kyutoku M, Yuri T, Uehara N, et al Effects of maternal xenoestrogen exposure on development of the reproductive tract and mammary gland in female CD-1 mouse offspring. Reprod Toxicol 18: Richter CA, Birnbaum LS, Farabollini F, Newbold RR, Rubin BS, Talsness CE, et al In vivo effects of bisphenol A in laboratory rodent studies. Reprod Toxicol 24: Ryan BC, Hotchkiss AK, Crofton KM, Gray LE, Jr In utero and lactational exposure to bisphenol A, in contrast to ethinyl estradiol, does not alter sexually dimorphic behavior, puberty, fertility, and anatomy of female LE rats. Toxicol Sci 114(1): Ryan BC, Vandenbergh JG Developmental exposure to environmental estrogens alters anxiety and spatial memory in female mice. Horm Behav 50: Shin BS, Kim CH, Jun YS, Kim DH, Lee BM, Yoon CH, et al Physiologically based pharmacokinetics of bisphenol A. J Toxicol Environ Health A 67: Stahlhut RW, Welshons WV, Swan SH Bisphenol A data in NHANES suggest longer than expected half-life, substantial nonfood exposure, or both. Environ Health Perspect 117: Stump DG, Beck MJ, Radovsky A, Garman RH, Freshwater LL, Sheets LP, et al Developmental neurotoxicity study of dietary bisphenol A in Sprague-Dawley rats. Toxicol Sci 115(1): Takeuchi T, Tsutsumi O Serum bisphenol A concentrations showed gender differences, possibly linked to androgen levels. Biochem Biophys Res Commun 291: Timms BG, Howdeshell KL, Barton L, Bradley S, Richter CA, vom Saal FS Estrogenic chemicals in plastic and oral contraceptives disrupt development of the fetal mouse prostate and urethra. PNAS 102: Tyl RW, Myers CB, Marr MC, Sloan CS, Castillo NP, Veselica MM, et al Two-generation reproductive toxicity study of dietary bisphenol A in CD-1 (Swiss) mice. Toxicol Sci 104: Tyl RW, Myers CB, Marr MC, Thomas BF, Keimowitz AR, Brine DR, et al Three-generation reproductive toxicity study of dietary bisphenol A in CD Sprague-Dawley rats. Toxicol Sci 68: Tyl RW Basic Exploratory Research Versus Guideline-Compliant Studies Used for Hazard Evaluation and Risk Assessment: Bisphenol A as a Case Study. Environ Health Perspect 117: US Food and Drug Administration (FDA). 2008a. Draft assessment of bisphenol A for use in food contact applications. Available on-line at: US Food and Drug Administration (FDA). 2008b. Science Board Subcommittee on Bisphenol A. Scientific Peer-Review of the Draft Assessment of Bisphenol A for Use in Food Contact Applications. Available on-line at 16

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