NTP-CERHR expert panel report on the reproductive and developmental toxicity of bisphenol A

Authors


Preface

The National Toxicology Program (NTP)1 established the NTP Center for the Evaluation of Risks to Human Reproduction (CERHR) in June 1998. The purpose of the CERHR is to provide timely, unbiased, scientifically sound evaluations of the potential for adverse effects on reproduction or development resulting from human exposures to substances in the environment. The NTP-CERHR is headquartered at NIEHS, Research Triangle Park, NC, and is staffed and administered by scientists and support personnel at NIEHS.

Bisphenol A is a high-production volume chemical used in the production of epoxy resins, polyester resins, polysulfone resins, polyacrylate resins, polycarbonate plastics, and flame retardants. Polycarbonate plastics are used in food and drink packaging; resins are used as lacquers to coat metal products such as food cans, bottle tops, and water supply pipes. Some polymers used in dental sealants and tooth coatings contain bisphenol A. Exposure to the general population can occur through direct contact with bisphenol A or by exposure to food or drink that has been in contact with a material containing bisphenol A. CERHR selected bisphenol A for evaluation because of (1) high production volume; (2) widespread human exposure; (3) evidence of reproductive toxicity in laboratory animal studies; and (4) public concern for possible health effects from human exposures.

Relevant literature on bisphenol A was identified from searches of the PubMed (Medline) and Toxline databases through February 2007 using the term “bisphenol” and the bisphenol A CAS RN (80-05-7). References were also identified from databases such as REPROTOX, HSDB, IRIS, and DART, from the bibliographies of the literature reviewed, by members of the expert panel, and in public comments.

CERHR convened a 12-member, independent panel of government and non-government scientists to evaluate the scientific studies on the potential reproductive and developmental hazards of bisphenol A. The expert panel met publicly on March 5–7, 2007 and August 6–8, 2007. The Expert Panel Report on Bisphenol A is intended to (1) interpret the strength of scientific evidence that bisphenol A is a reproductive or developmental toxicant based on data from in vitro, animal, or human studies; (2) assess the extent of human exposures to include the general public, occupational groups, and other sub-populations; (3) provide objective and scientifically thorough assessments of the scientific evidence that adverse reproductive and developmental health effects may be associated with such exposures; and (4) identify knowledge gaps to help establish research and testing priorities to reduce uncertainties and increase confidence in future evaluations. This report has been reviewed by members of the expert panel and by CERHR staff scientists. Copies of this report have been provided to the CERHR Core Committee2 and will be made available to the public for comment.

Following the public comment period, CERHR will prepare the NTP-CERHR Monograph on the Potential Human Reproductive and Developmental Effects of Bisphenol A. This monograph will include the NTP Brief, the Expert Panel Report, and all public comments received on the Expert Panel Report. The NTP-CERHR Monograph will be made publicly available and transmitted to appropriate health and regulatory agencies.

Reports can be obtained from the web site (http://cerhr.niehs.nih.gov) or from: Michael D. Shelby, PhD, NIEHS EC-32, PO Box 12233, Research Triangle Park, NC 27709. E-mail: shelby@niehs.nih.gov

1.0 CHEMISTRY, USE AND HUMAN EXPOSURE

1.1 Chemistry

Section 1 is based initially on secondary review sources. Primary study reports are addressed by the Expert Panel if they contain information that is highly relevant for determining the effect of exposure on developmental or reproductive toxicity or if the studies were released subsequent to the reviews.

1.1.1 Nomenclature.

The CAS RN for bisphenol A is 80-05-7. Synonyms for bisphenol A listed in Chem IDplus (ChemIDplus, 2006) include: 2-(4,4′-Dihydroxydiphenyl)propane; 2,2-Bis(4-hydroxyphenyl)propane; 2,2-Bis(hydroxyphenyl)propane; 2,2-Bis(p-hydroxyphenyl)propane; 2,2-Bis-4′-hydroxyfenylpropan [Czech]; 2,2-Di(4-hydroxyphenyl)propane; 2,2-Di(4-phenylol)propane; 4,4′-(1-Methylethylidene)bisphenol; 4,4′-Bisphenol A; 4,4′-Dihydroxydiphenyl-2,2-propane; 4,4′-Dihydroxydiphenyldimethylmethane; 4,4′-Dihydroxydiphenylpropane; 4,4′-Isopropylidene diphenol; 4,4′-Isopropylidenebisphenol; 4,4′-Isopropylidene diphenol; Biphenol A; Bis(4-hydroxyphenyl) dimethylmethane; Bis(4-hydroxyphenyl)dimethylmethane; Bis(4-hydroxyphenyl)propane; Bisferol A [Czech]; Bisphenol. Bisphenol A; DIAN; Diano; Dimethyl bis(p-hydroxyphenyl)methane; Dimethylbis(p-hydroxyphenyl)methane; Dimethylmethylene-p,p′-diphenol; Diphenylolpropane; Ipognox 88; Isopropylidenebis(4-hydroxybenzene); Parabis A, Phenol; (1-methylethylidene)bis-, Phenol; 4,4′-(1-methylethylidene)bis-; Phenol, 4,4′-dimethylmethylenedi-; Phenol, 4,4′-isopropylidenedi-; Pluracol 245, Propane; 2,2-bis(p-hydroxyphenyl)-; Rikabanol; Ucar bisphenol A; Ucar bisphenol HP; beta,beta′-Bis(p-hydroxyphenyl)propane; beta-Di-p-hydroxyphenylpropane; p,p′-Bisphenol A; p,p′-Dihydroxydiphenyldimethylmethane; p,p′-Dihydroxydiphenylpropane; p,p′-Isopropylidenebisphenol; and p,p′-Isopropylidenediphenol.

1.1.2 Formula and molecular mass.

Bisphenol A has a molecular mass of 228.29 g/mol and a molecular formula of C15H1602 (European-Union, 2003). The structure for bisphenol A is shown in Figure 1.

Figure 1.

Structure for bisphenol A.

1.1.3 Chemical and physical properties.

Bisphenol A is a white solid with a mild phenolic odor (European-Union, 2003). Physicochemical properties are listed in Table 1.

Table 1. Physicochemical Properties of Bisphenol Aa
PropertyValue
  • a

    aStaples et al. (1998).

Odor thresholdNo data found
Boiling point220°C at 4 mm Hg; 398°C at 760 mm Hg
Melting point150–157°C
Specific gravity1.060–1.195 g/mL at 20–25°C
Solubility in water120–300 mg/L at 20–25°C
Vapor pressure8.7×10−10–3.96×10−7 mm Hg at 20–2 5°C
Stability/reactivityNo data found
Log Kow2.20–3.82
Henry constant1.0×10−10 atm m3/mol

1.1.4 Technical products and impurities.

Purity of bisphenol A was reported at 99–99.8%, and common impurities observed were phenol and ortho and para isomers of bisphenol A [reviewed in (European-Union, 2003)]. Terasaki et al. (2004) used reversed phase chromatography and nuclear magnetic resonance spectroscopy to characterize the composition of 5 commercial bisphenol A samples. The nominal purity of the samples was 97 or 98%. Actual purities were 95.3 to >99%. Up to 15 contaminants were identified among which were: 4-hydroxyacetophenone; 4,4′-(1,3-dimethylbutylidene) bisphenol; p-cumylphenol; 4-hydroxyphenyl isobutyl methyl ketone; 2,4*-dibhydroxy-2,2-diphenylpropane; 2,4′-dibhydroxy-2,2-diphenylpropane; 2,4-bis(4-hydroxycumyl)phenol; 2,3-dihydro-3-(4′-hydroxyphenyl)-1,1,3-trimethyl-1H-inden-5-ol; 2-(4′-hydroxyphenol)-2,2,4-trimethylchroman; and 4-(4′-hydroxyphenol)-2,2,4-trimethylchroman (Terasaki et al., 2005).

No information on trade names for bisphenol A was located.

1.1.5 Analytical considerations.

Measurement of bisphenol A in environmental and biologic samples can be affected by contamination with bisphenol A in plastic laboratory ware and in reagents (Tsukioka et al., 2004; Völkel et al., 2005). Accuracy is also affected by measurement technique, particularly at the very low concentrations that can now be measured. Enzyme-linked immunosorbent assay (ELISA) has poor correlation with the LC-ECD method and also the different ELISA kits correlate poorly with each other. ELISA methods may overestimate bisphenol A in biologic samples due to lack of specificity of the antibody and effects of the biologic matrix (Inoue et al., 2002; Fukata et al., 2006). Although high performance liquid chromatography (HPLC) with ultraviolet, fluorescence, or electrochemical detection can be sensitive to concentrations <0.5 ng/ml (Sajiki et al., 1999; Inoue et al., 2000; Kuroda et al., 2003; Sun et al., 2004), these methods are unable to make definitive identification of bisphenol A or bisphenol A glucuronides, because similar retention times may occur for the metabolites of other endogenous and exogenous compounds (Völkel et al., 2005). Use of LC-mass spectrometry (MS) with and without hydrolysis of bisphenol A glucuronide permits determination of free and total bisphenol A with a limit of quantification of 0.1 for MS (Sajiki et al., 1999) and 1 μg/L for MS/MS (Völkel et al., 2005). Gas chromatography (GC)/MS has been used with solid phase extraction after treatment with glucuronidase and derivatization to measure total bisphenol A with a limit of detection of 0.05 μg/L for MS (Tan and Mohd, 2003) and 0.1 μg/L for MS/MS (Calafat et al., 2005). Some of the variability in studies cited in this and subsequent sections may be due to differences in measurement techniques and to contamination. Bisphenol A glucuronidate can be an unstable product that can be degraded in acidic and basic pH solutions and can be hydrolyzed to free bisphenol A at neutral pH and room temperature in diluted rodent urine, placental and fetal tissue homogenates at room temperature. However, conjugates in urine are stable for at least 7 days when stored at −4°C and at least 180 days when stored at −70°C (Waechter et al., 2007; Ye et al., 2007).

1.2 Use and Human Exposure

1.2.1 Production information.

Bisphenol A is manufactured by the acid catalyzed condensation of phenol and acetone (SRI, 2004).

In 1998, members of the Society of the Plastics Industry Bisphenol A Task Group [assumed manufacturers of bisphenol A] included Aristech Chemical Corporation, Bayer Corporation, Dow Chemical Company, and Shell Chemical Company (Staples et al., 1998). Current manufacturers of bisphenol A in the U.S. are Bayer MaterialScience, Dow Chemical Company, General Electric, Hexion Specialty Chemicals, and Sunoco Chemicals (SRI, 2004) (S. Hentges, public comments, February 2, 2007). There are currently six bisphenol A and four polycarbonate plants in the U.S. (S. Hentges, personal communication, October 30, 2006); three of four polycarbonate plants are located within bisphenol A plants. In 2000, there were 13 epoxy plants in the U.S., but was not clear if all of the plants manufactured bisphenol A-containing epoxy resins.

In mid-2004, U.S. bisphenol A production volume was reported at 1.024 million metric tons [∼2.3 billion pounds] (SRI, 2004). A production volume of 7.26 billion g [16 million pounds] was reported for bisphenol A in 1991 (reviewed in HSDB, 2003). United States bisphenol A consumption was reported at 856,000 metric tons [∼1.9 billion pounds] in 2003 (SRI, 2004); 2003 consumption patterns included 619,000 metric tons [∼1.4 billion pounds] used in polycarbonate resins, 184,000 metric tons [∼406 million pounds] used in epoxy resins, and 53,000 metric tons [∼117 million pounds] used in other applications.

1.2.2 Use.

In 1999 and 2003, it was reported that most bisphenol A produced in the U.S. was used in the manufacture of polycarbonate and epoxy resins and other products [reviewed in (Staples et al., 1998; SRI, 2004)]. Polycarbonate plastics may be used in the manufacture of compact discs, “solid and multi wall sheet in glazing applications and film,” food containers (e.g., milk, water, and infant bottles), and medical devices [reviewed in (European-Union, 2003)]. Bisphenol A may have been used at one time in Europe in polyvinyl chloride cling film and plastic bags, but that use is believed to have been discontinued (European Food Safety Authority, 2006). Contact with drinking water may occur through the use of polycarbonate for water pipes and epoxy-phenolic resins in surface coatings of drinking water storage tanks [reviewed by (European Food Safety Authority, 2006)].

Polycarbonate blends have been used to manufacture injected molded parts utilized in alarms, mobile phone housings, coil cores, displays, computer parts, household electrical equipment, lamp fittings, and power plugs. Automotive and related uses for polycarbonate blends include light reflectors and coverings, bumpers, radiator and ventilation grills, safety glazing, inside lights, and motorcycle shields and helmets. Epoxy resins are used in protective coatings, structural composites, electrical laminates, electrical applications, and adhesives. The European Union (2003) reported that smaller volumes of bisphenol A are used in production of phenoplast, phenolic, and unsaturated polyester resins, epoxy can coatings, polyvinyl chloride (PVC) plastic, alkoxylated bisphenol A, thermal paper, and polyols/polyurethane. Other uses reported for products manufactured from bisphenol A included protective window glazing, building materials, optical lenses, and development of dyes [reviewed in (Staples et al., 1998)]. A search of the National Library of Medicine Household Products Database (NLM, 2006) revealed that bisphenol A-based polymers are used in coatings, adhesives, and putties available to the general pubic for use in automobiles, home maintenance and repair, and hobbies, but only 3 epoxy products, used for crafts and hobbies, contain bisphenol A itself.

Some polymers manufactured with bisphenol A are Food and Drug Administration (FDA)-approved for use in direct and indirect food additives and in dental materials, as reported in the Code of Federal Regulations (CFR) (FDA, 2006). In the CFR, bisphenol A is often referred to as 4,4′-isopropylidnediphenol. Polymers manufactured with bisphenol A are FDA-approved for use as anoxomers and in coatings, adhesives, single and repeated food contact surfaces, and tooth shade resin materials.

The European Union (2003) noted that resins, polycarbonate plastics, and other products manufactured from bisphenol A can contain trace amounts of residual monomer and additional monomer may be generated during breakdown of polymer. The American Plastics Council reports that residual bisphenol A concentrations in polycarbonate plastics and epoxy resins are generally <50 ppm (S. Hentges, personal communication, October 30, 2006). Polymer hydrolysis can occur at elevated temperature or extreme pH. An example of potential human exposure is migration of bisphenol A from a food container into the food. Exposure to bisphenol A through food is discussed in detail in Section 1.2.3.2.

1.2.3 Occurrence

1.2.3.1 Environmental fate and bisphenol A levels in environment:

Bisphenol A may be present in the environment as a result of direct releases from manufacturing or processing facilities, fugitive emission during processing and handling, or release of unreacted monomer from products (European-Union, 2003). According to the Toxics Release Inventory database, total environmental release of bisphenol A in 2004 was 181,768 pounds, with releases of 132,256 pounds to air, 3533 pounds to water, 172 pounds to underground injection, and 45,807 pounds to land (TRI, 2004).

Bisphenol A released to the atmosphere is likely degraded by hydroxy radicals (European-Union, 2003). Half-life for the reaction between bisphenol A and hydroxy radicals was estimated at 0.2 days. It was also noted that photolysis and photodegradation of bisphenol A in the atmosphere is possible and photo-oxidation half-lives of 0.74–7.4 hr were estimated [reviewed in (Staples et al., 1998; European-Union, 2003)]. The European Union (2003) noted that because of its low volatility and relatively short half-life in the atmosphere, bisphenol A is not likely to enter the atmosphere in large amounts. Removal by precipitation and occurrence in rain water were thought likely to be negligible. Because of its short half-life in the atmosphere, bisphenol A is unlikely to be transported far from emission points.

Based on vapor pressure and Henry constant (Table 1), the European Union (2003) and Staples et al. (1998) concluded that bisphenol A is of low volatility and not likely to be removed from water through volatilization. Both groups concluded that hydrolysis of bisphenol A in water is unlikely. However, there was disagreement on potential for photo-oxidation of bisphenol A in water. Based on physical and chemical properties, the European Union concluded that photolysis of bisphenol A in water is unlikely. Staples et al. (1998) noted that bisphenol A is able to absorb ultraviolet light, especially in a basic solution. Therefore, it was concluded that photolysis from surface water is possible, depending on conditions such as pH, turbidity, turbulence, and sunlight. Photo-oxidation half-life of bisphenol A in water was estimated at 66 hr to 160 days [reviewed in (Staples et al., 1998)]. Rapid biodegradation of bisphenol A from water was reported in the majority of studies reviewed by the European Union (2003) and Staples et al. (1998). A biodegradation half-life of 2.5–4 days was reported in a study measuring bisphenol A concentrations in surface waters near the receiving stream of a bisphenol A manufacturer [reviewed in (Staples et al., 1998)].

When the Staples et al. (1998) review was published, soil sorption constants had not been measured but were estimated at 314–1524. Based on such data, the European Union (2003) and Staples et al. (1998) concluded that bisphenol A adsorption to soils or sediments would be “modest” or “moderate.” Based on data for degradation of bisphenol A in water, the European Union (2003) predicted that bisphenol A would be degraded in soil and estimated a half-life of 30 days for degradation of bisphenol A in soil. Subsequent to the Staples et al. (1998) and European Union (2003) reviews, a study examining fate of 14C-bisphenol A in soils through laboratory soil degradation and batch adsorption tests was released by Fent et al. (2003). In that study, 14C-bisphenol A was dissipated and not detectable in 4 different soil types within 3 days. Soil distribution coefficients were determined at 636–931, and based on those values, the study authors concluded that bisphenol A has low mobility in soil. The study authors concluded that bisphenol A is not expected to be stable, mobile, or bioavailable from soils.

In studies reviewed by the European Union (2003) and Staples et al. (1998), bioconcentration factors for fish were measured at 3.5–68 and were found to be lower than values estimated from the Kow. Both groups concluded that potential for bioconcentration of bisphenol A is low in fish. Higher bioconcentration factors (134–144) were determined for clams [reviewed in (European-Union, 2003)].

Two studies examining aggregate exposures in preschool age children in the U.S. used GC/MS to measure bisphenol A concentrations in environmental media (Wilson et al., 2003, 2006). In the first study (Wilson et al., 2003), bisphenol A concentrations were measured in air outside 2 day care centers and the homes of 9 children. Bisphenol A was detected in 9 of 13 outdoor air samples at <0.100–4.72 ng/m3 (mean concentration=2.53 ng/m3 at day care centers; 1.26 ng/m3 at home). In indoor air from day care centers and homes, bisphenol A was detected in 12 of 13 samples at <0.100–29 ng/m3 (mean concentration=6.38 ng/m3 at day care centers; 11.8 ng/m3 at home). At those same locations, bisphenol A was detected in all of 13 samples of floor dust at means (range) of 1.52–1.95 (0.567–3.26) ppm (μg/g) and play area soils at means (range) of 0.006–0.007 (0.004–0.014) ppm (μg/g). In the second study (Wilson et al., 2006), bisphenol A concentrations were measured inside and outside at least 222 homes and 29 daycare centers. Bisphenol A was detected in 31–44% of outdoor air samples from each location; concentrations ranged from <LOD (0.9) to 51.5 ng/m3. Medians were <limit of detection (LOD). Indoor air samples (45–73%) contained detectable concentrations of bisphenol A; concentrations were reported at <LOD (0.9)–193 ng/m3. Median values were <LOD–1.82 ng/m3. Bisphenol A was detected in 25–70% of dust samples; concentrations were reported at <LOD (20) to 707 ng/g. Median values were <LOD–30.8 ng/g.

A second U.S. study used a GC/MS method to measure bisphenol A concentrations in dust from 1 office building and 3 homes and in air from 1 office building and 1 home (Rudel et al., 2001). Bisphenol A was detected in 3 of 6 dust samples (reporting limit >0.01 μg/extract) at concentrations of 0.25–0.48 μg/g dust. In indoor air samples collected from offices and residences, bisphenol A was detected in 3 of 6 samples (detection limit=∼0.5 ng/m3) at concentrations of 0.002–0.003 μg/m3. In another study using a GC/MS technique, bisphenol A concentrations in indoor air from 120 U.S. homes were below reporting limits (0.018 μg/m3) (Rudel et al., 2003). Median (range) bisphenol A concentration in dust in this study was 0.821 (<0.2–17.6) μg/g, with 86% of samples above the reporting limit.

Limited information is available for bisphenol A concentrations in U.S. water (Table 2). In 1996 and/or 1997, mean bisphenol A concentrations were reported at 4–8 μg/L in surface water samples near 1 bisphenol A production site but bisphenol A was not detected (<1 μg/L) in surface water near 6 of 7 bisphenol A production sites in the U.S. (Staples et al., 2000). Bisphenol A was detected at a median concentration (in samples with detectable bisphenol A above the reporting limit of 0.09 μg/L) of 0.14 μg/L and a maximum concentration of 12 μg/L in 41.2% of 85 samples collected from U.S. streams in 1999 and 2000 (Kolpin, 2002). In 2001 and 2002, bisphenol A was not detected (<0.001 μg/L) in effluent from a wastewater treatment plant in Louisiana, and concentrations were not quantifiable [quantification limit not defined] in samples collected from surface waters in Louisiana and in drinking water at various stages of treatment at plants in Louisiana and Ontario, Canada (Boyd et al., 2003). In water samples collected in Europe and Japan from the 1970 s through 1989, bisphenol A concentrations were ≤1.9 μg/L and in most cases were ≤0.12 μg/L [reviewed in (European-Union, 2003)].

Table 2. Concentrations of Bisphenol A Detected in Water
Sample typeDetection methodDetection rate (%)Concentration (μg/L) range [median]Reference
Surface water
 German riversGC-MS1000.005–0.014 [3.8]Kuch and Ballschmiter (2001)
 Louisiana, U.S.GC-MS0<MDL 0.1Boyd et al. (2003)
 U.S. streamsGC-MS41.2[0.14] max 12Kolpin et al. (2002)
 NetherlandsGC-MS78–93Max marine 0.33 Max fresh 21Belfroid et al. (2002)
Drinking water
 Louisiana, U.S.GC-MS0<MDL 0.1Boyd et al. (2003)
 Ontario, CanadaGC-MS0<MDL 0.1Boyd et al. (2003)
 GermanyGC-MS1000.005–0.002 [1.1]Kuch and Ballschmiter (2001)
Landfill leachate
 JapanGC-MS100740Kawagoshi et al. (2003)
 JapanGC-MS70% sites1.3–17, 200 [269]Yamamoto et al. (2001)
Sewage treatment works
 GermanyGC-MS940.005–0.047 [10]Kuch and Ballschmiter (2001)
 Louisiana, U.S.GC-MS0<MDL 0.1Boyd et al. (2003)
1.2.3.2 Potential exposures from food and water:

The European Union (2003) noted that the highest potential for human exposure to bisphenol A is through products that directly contact food. Examples of food contact materials that can contain bisphenol A include food and beverage containers with internal epoxy resin coatings and polycarbonate tableware and bottles, such as those used to feed infants.

In addition to commercial food sources, infants consume breast milk. Calafat et al. (2006) reported a median bisphenol A concentration of ∼1.4 μg/L [as estimated from a graph] in milk from 32 women (Table 3). Bisphenol A was measured after enzymatic hydrolysis of conjugates. Ye et al. (2006) found measurable concentrations of bisphenol A in milk samples from 18 of 20 lactating women. Free bisphenol A was found in samples from 12 women. The median total bisphenol concentration in milk was 1.1 μg/L (range: undetectable to 7.3 μg/L). The median free bisphenol A concentration was 0.4 μg/L (range: undetectable to 6.3 μg/L). Sun et al. (2004) used an HPLC method to measure bisphenol A concentrations in milk from 23 healthy lactating Japanese women. Bisphenol A concentrations ranged from 0.28–0.97 μg/L, and the mean±SD concentration was reported at 0.61±0.20 μg/L. No correlations were observed between bisphenol A and triglyceride concentrations in milk. Values from 6 milk samples were compared to maternal and umbilical blood samples reported previously in a study by Kuroda et al. (2003). Bisphenol A values were higher in milk, and the milk/serum ratio was reported at 1.3. Bisphenol A values in milk were comparable to those in umbilical cord serum. [It was not clear whether milk and serum samples were obtained from the same volunteers in the two studies.]

Table 3. Bisphenol A Concentrations in Human Breast Milk
Source (n)MethodLODFree (ng/ml) mean±SD (range)Total (ng/ml) mean±SD (range)Detection rate (%)Reference
  • a

    aEstimated from a graph.

Japanese (23)HPLC-Fl0.11 ng/ml0.61±0.20 (0.28–0.97) 100Sun et al. (2004)
Japanese (101) (colostrum 3 days after delivery)ELISANA 3.41±0.13 (1–7)100Kuruto-Niwa et al. (2007)
United States (20)HPLC-MS/MS0.3 ng/ml1.3 (<0.3–6.3)1.9 (<0.3–7.3)60 free 90 totalYe et al. (2006)
Japanese (3)GC-MS0.09 ng/g 0.46 (<0.09–0.65)67Otaka et al. (2003)
U.S. (32)NANANA1.4aNACalafat et al. (2006)

Studies have measured migration of bisphenol A from polycarbonate infant bottles or containers into foods or food simulants. Results of those studies are summarized in Table 4. Analyses for bisphenol A were conducted by GC/MS or HPLC. The European Union (2003) group noted that in many cases bisphenol A concentrations were below the detection limit in food simulants. When bisphenol A was detected, concentrations were typically ≤50 μg/L in simulants exposed to infant bottles and ≤5 μg/kg in simulants exposed to polycarbonate tableware. An exception is one study that reported bisphenol A concentrations at up to ∼192 μg/L in a 10% ethanol food simulant and 654 μg/L in a corn oil simulant (Onn Wong et al., 2005). In the study, cut pieces of bottles were incubated, and the study authors acknowledged that bisphenol A could have migrated from the cut edges. [The Expert Panel notes that incubations were at 70 or 100°C for 240 hr, representing conditions not anticipated for normal use of baby bottles.] One study conducted with actual infant food (formula and fruit juice) reported no detectable bisphenol A (Mountfort et al., 1997). Some studies examining the effects of repeated use of polycarbonate items noted increased leaching of bisphenol A with repeated use (Earls, 2000; Brede et al., 2003; CSL, 2004). It was suggested that the increase in bisphenol A migration was caused by damage to the polymer during use. Results from other reports suggested that leaching of bisphenol A decreased with repeated use, and it was speculated that available bisphenol A was present at the surface of the product and therefore removed by washing (Biles et al., 1997b; Kawamura et al., 1999; Haighton et al., 2002; European Union, 2003). One study (Kawamura et al., 1999) showed higher concentrations of bisphenol A in simulants exposed to products that had been recalled because of unacceptable residual concentrations of bisphenol A and other compounds. The study by Biles et al. (1997b) demonstrated that infant bottles exposed to 50 or 95% ethanol at 65°C for 240 hr leached bisphenol A at concentrations exceeding residual monomer concentrations, and it was suggested that hydrolysis of the polymer had occurred.

Table 4. Examination of Bisphenol A in Polycarbonate Food Contact Surfaces
Sample (location)ProcedureBisphenol A concentration in simulantReference
  • a

    aReviewed by European Union (2003).

  • b

    bReviewed by Haighton et al. (2002).

  • c

    cThe authors of this study identified an error in the units reported in their study and that the correct concentrations are 1000-fold higher than indicated in the article, the correct values are indicated in table above (T. Begley, email communication, August 6, 2007).

  • ND, not detected.

Commercially available infant bottles containing residual bisphenol A concentrations of 7–46 ppm (U.S.)Common use: bottles were boiled for 5 min, filled with water or 10% ethanol, and stored at room temperature for up to 72 hr Worst case use: bottles were boiled for 5 min, filled with water or 10% ethanol, heated to 100°C for 0.5 hr, cooled to room temperature, and refrigerated for 72 hrND (LOD 5 ppb [μg/L]; corresponding to a food concentration of 1.7 ppb) following either procedureFDA (1996)
21 new and 12 used (1–2-year-old) infant bottles (U.K.)Bottles were pre-washed, steam sterilized, filled with boiling water or 3% glacial acetic acid, refrigerated at 1–5°C for 24 hr, and heated to 40°C before samplingND (LOD 10 μg/L) [ppb] from new bottles; ND (<10 μg/L ) to 50 μg/L from used bottles exposed to either simulant [mean not given]Earls et al. (2000)
Infant bottles with residual bisphenol A concentrations of 26 mg/kg [number tested not indicated] (U.K.)Bottles were sterilized with hypochlorite, in dishwasher, or by steam; filled with infant formula, fruit juice, or distilled water; microwaved for 30 sec and left to stand for 20 min (1 cycle); samples were analyzed after 3, 10, 20, or 50 cycles; other bottles were filled with distilled water and left to stand for 10 days at 40°CND (LOD 0.03 mg/kg) [<30μg/kg or ppb] under any conditionMountfort et al. (1997)
6 infant feeding bottles (country of purchase not known)Bottles were filled with water at 26°C and left to stand for 5 hr or filled with water at 95°C and left to stand overnightND (LOD 2 ppb [μg/L]) in bottles filled with water at 26°C and 3.1–55 ppb [μg/L] in bottles filled with water at 95°C.Hanai (1997)a
14 samples of new infant feeding bottles and tableware including a bowl, mug, cup, and dish recalled because residual bisphenol A and other phenol concentrations exceeded 500 ppm [mg/kg] (Japan)Products were exposed to n-heptane, water, 4% acetic acid, or 20% ethanol; in some cases simulant was heated to 60 or 95°C; in other cases, the object was boiled for 5 min; analyses were usually conducted after a 30-min contact periodUp to 40 ppb [μg/kg] from recalled products and ND (LOD 0.2) to 5 μg/kg from commercially available products.Kawamura et al. (1999)ab
Discs prepared from commercial food-grade polycarbonate resins (residual bisphenol A at 8800 to 11,200 μg/kg) from U.S. manufacturersMaterials exposed to water, 10% ethanol, or Miglyol (fractionated coconut oil) at 100°C for 6 hr or water, 3% acetic acid, 10% ethanol, or Migloyl at 49°C for 6–240 hrND (LOD 5 ppb [μg/L]) under all conditions.Howe and Borodinsky (1998)
2 infant bottles from JapanIn three repeated tests, boiling water was added to bottles; bottles were incubated at 95°C for 30 min and cooled to room temperature; before repeating the test a fourth time, the bottles were scrubbed with a brushBelow quantification limit (LOD 0.57 ppb [μg/L]) to mean concentrations of 0.75 ppb before brushing and <0.57 to 0.18 ppb after brushing.Sun et al. (2000)
4 new different brands of infant bottles (Argentina)Bottles were exposed to distilled water, 3% acetic acid, or 15% ethanol at 80°C for 2 min or distilled water at 100°C for 0.5 min1.1–2.5 ppb [μg/L].D'Antuono et al. (2001)
12 infant bottles (Norway)Bottles were tested before washing and following 51 and 169 dish washings; bottles were occasionally brushed (13 times by second test and 23 times by third test) and boiled (12 times by second testing and 25 times by third testing); unwashed bottles were rinsed with boiling water before testing; for testing, bottles were filled with hot water and incubated at 100°C for 1 hrMean (range) μg/L [ppb]: 0 washes: 0.23 (0.11–0.43) 51 washes: 8.4 (3.7–17) 169 washes: 6.7 (2.5–15)Brede et al. (2003)
18 infant bottles (12 tested) (U.K.)Bottles were tested before and after 20 and 50 dish washings; bottles were brushed after every 2 wash cycles; bottles were sterilized with boiling water, filled with 3% acetic acid, or 10% ethanol, and incubated at 70°C for 1 hrBefore washing: ND (LOD 1.1 ppb or μg/L) in 10% ethanol and ND (LOD 0.34 ppb or μg/L) in 3% acetic acid; 20 washes: ND to 4.5 ppb in 10% ethanol and ND to 0.51 ppb in 3% acetic acid; 50 washes: ND to 3.1 ppb in 10% ethanol and ND to 0.7 ppb in 3% acetic acidCSL (2004)
28 brands of new infant bottles (residual bisphenol A concentrations of <3 to 141 mg/kg) manufactured in Europe or Asia (Singapore)Bottles were cut and pieces were exposed to 10% ethanol at 70°C or corn oil at 100°C for 8–240 hrND (LOD 0.05) to 1.92 μg/in2[<5–192μg/L orppb] in 10% ethanol and ND (LOD 0.05) to 6.54 μg/in2[<5–654μg/L] in corn oil over the 240-hr exposure periodOnn Wong et al. (2005)
22 new infant bottles and 20 used (3–36 months) bottles (Netherlands)Bottles were immersed in boiling water for 10 min before testing and filled with distilled water or 3% acetic acid and incubated at 40 °C for 24 hrND in new bottles (<2.5 μg/L (LOD) [ppb] in distilled water and <3.9 μg/L (LOD) in 3% acetic acid) or in used bottles exposed to 3% acetic acid; not detected to non-quantifiable (<5 μg/L) in distilled water from used bottles.FCPSA (2005)
New unwashed infant bottles (number not indicated) (Japan)Bottles were exposed to water at 95°C for 30 minND (LOD 0.05 μg/L [ppb]) to 3.9 μg/L.Japanese studies reviewed in Miyamoto and Kotake (2006)
5-gallon water carboysWater was stored in the carboys for 3, 12, or 39 weeks, temperature not indicated0.1–0.5 μg/L [ppb] at 3 and 12 weeks and. 4.6–4.7 μg/L at 39 weekscBiles et al. (1997b)

High molecular weight, heat-cured bisphenol A-based epoxy resins are used as protective linings in cans for food and beverages and may be used in wine storage vats (European-Union, 2003). Residual bisphenol A monomer can migrate from the coatings to foods or beverages contained within cans. Studies were conducted to measure actual concentrations of bisphenol A in commercially available foods or to measure concentrations of bisphenol A leaching from can linings into food simulants. Because the actual measurement of bisphenol A concentrations in canned foods represents the most realistic situation, the CERHR review will focus on those data. Studies conducted with simulants will not be reviewed, with the exception of one study by Howe et al. (1998) that was considered by the FDA (1996) in their estimates of bisphenol A intake.

Bisphenol A concentrations detected in infant foods are summarized in Table 5, and bisphenol A concentrations detected in non-infant foods are summarized in Table 6. With the exception of isolated cases in which bisphenol A concentrations were measured at up to ∼0.8 mg/kg food, most measurements were below 0.1 mg/kg. The European Union also noted an extraction study conducted with an epoxy resin that is occasionally used to line wine vats. Based on that study, a worst-case scenario of 0.65 mg/L bisphenol A in wine was used. The European Union noted that the value represents a very worst-case exposure scenario but decided to use that number in risk estimates because no other value was available. [The Expert Panel notes that a study of bisphenol A in wine (Brenn-Struckhofova and Cichna-Markl,2006) identified a maximum concentration of 2.1μg/L (Table6).]

Table 5. Surveys of Bisphenol A Concentrations in Canned Infant Formulas or Food
Food (no. sampled)aBisphenol A concentration μg/kg or μg/LCountryReference
  • a

    aValues before and after heating in can and from non-dented and dented cans; values did not differ under the various conditions and were presented together.

  • ND, not detected.

Infant formula (14)Mean 5 (0.1–13.2 ppb [μg/L]); when diluted with water to make prepared formula mean concentrations would be 2.5 (0.05–6.6)U.S.Biles et al. (1997a) FDA (1996)
Infant formula (4)ND (LOD 2 μg/kg)U.K.Goodson et al. (2002) UKFSA (2001)
Infant formula (5)44–113 μg/kgTaiwanKuo and Ding (2004)
Infant dessert (3)18.9–77.3 μg/kgU.K.Goodson et al. (2002)
Infant vegetable food (4)<LOQ (LOQ 10 μg/kg)New ZealandThomson and Grounds (2005)
Infant dessert (3)<LOQ (LOQ 10 μg/kg)  
Table 6. Surveys of Bisphenol A Concentrations in Canned or Bottled Foods or Food Simulants
Food (no. sampled)Bisphenol A concentration (range in μg/kg unless specified)Country of purchaseaReference
  • a

    aAlthough cans were purchased in one or two countries for each study, most studies reported that cans were packaged in various locations throughout North America, Europe, or Asia.

  • b

    bThe UKFSA noted that the higher concentrations of bisphenol A detected in one meat product likely resulted from the use of bisphenol A as a cross-linking agent in the resin at that time.

  • c

    cValues were obtained from heated and non-heated cans but presented together because it could not be determined if heating resulted in differing extraction rates.

  • d

    dValues were determined before and after heating in can and from non-dented and dented cans; because the values did not differ under the various conditions, they were presented together.

  • e

    eTotal number of samples analyzed was not reported.

  • f

    fAs reported by study authors; detection limits not specified.

  • g

    gA maximum concentration of 121 ppb reported in the first phase of the study was determined to have resulted from analytical interference.

  • ND, not detected.

Vegetables with liquid (6)Mean (range) 16 (4–39)U.S.FDA (1996)
Liquids from canned vegetables or mushrooms (10)4.2±4.1 (SD) −22.9±8.8 μg/can [12±12–76±29 μg/kg]Spain, U.S.Brotons et al. (1995)
Coffee (13)ND–213 [median=11] (LOD 2)JapanKawamura et al. (1999) (reviewed in European-Union, 2003; English abstract available)
Black tea (9)ND–90 [median=<2] (LOD 2)  
Other tea (8)ND–22 [median=5.7] (LOD 2)  
Alcoholic beverages (10)ND except for 1 sample with 13 (LOD 2)  
Soft drinks (7)ND (LOD 2)  
Vegetables (10)9–48 [median=21]U.K.Goodson et al. (2002) UKFSA (2001)
Desserts (5)ND (LOD 2)–14 [median=10]  
Fruits (2)19 and 38  
Pastas (5)ND–41 [median=11] (LOD 7)  
Meats (5)16–422b [median=52]  
Fish (10)ND–44 [median=16.8] (LOD 2)  
Non-alcoholic or alcoholic beverages (11)ND except for 1 sample above LOD (LOD 2) but below LOQ (7)  
Soups (10)ND–21 [median=<2] (LOD 2)  
Vegetables, fruits, or mushrooms (14)ND (LOD 10)–95.3 in solid portion; ND (LOD 0.005 ug/mL)–0.004 μg/mL in liquid portion; ND–11.1 μg/can [85 μg/kg] total Yoshida et al. (2001)
Meat productsd (2)8.6–25.7U.K.Goodson et al. (2004)
Pastad (1)67.3–129.5  
Vegetables or beansc(2)11.3–14.4  
Soupc (1)18.5–39.1  
Puddingc (3)3.8–53.2  
Puddingd (1)18.5–28.1  
Grains and potatoese0f –75 [mean not given]JapanMiyamoto and Kotake (2006)
Sugar, sweets, snackse0f–4 [mean not given]  
Fatse0f  
Fruits (including canned drinks), vegetables, mushrooms, seaweedse0f–450 [mean not given]  
Seasoning and beveragese0f–213 [mean not given]  
Fish9–480 [mean not given]  
Meat and eggse12.5–602 [mean not given]  
Milk and dairy productse0c–6 [mean not given]  
“Other” [not specified further]e36–310 [mean not given]  
Canned fish (7)1–23 [median=6]JapanSajiki et al. (2007)
Canned meat (5)4–20 [median=10]  
Canned fruit (3)ND (LOD 0.2)  
Canned vegetables (13)3–78 [median=15]  
Canned soup (12)1–156 [median=15]  
Canned sauce (6)ND (LOD 0.2)–842 [median=220]  
Canned coconut milk56–247  
Drinks in plastic containers (3)ND (LOD 0.2) to 1 [median=0.3]JapanSajiki et al. (2007)
Cookies in plastic containers (4)1–14 [median=3.5]  
Soup in plastic containers (2)ND (LOD 0.2) and 3  
Fast food sandwiches (3)3 (all values)  
Food in paper containers (16)ND (LOD 0.2)–1 [median=<0.2]  
Fruits and vegetables (38)ND (LOQ 10)–24 [median=<10]New ZealandThomson and Grounds (2005)
Fish (8)ND (LOQ 20)–109 [median=<20–24]  
Soup (4)ND (LOQ 10)–16 [median=<20]  
Sauces (4)ND (LOQ 10) –21 [median=16]  
Meat (6)ND (LOQ 20)–98 [median=<20]  
Pasta (4)ND (LOQ 10)  
Dessert (2)ND (LOQ 20)  
Coconut cream (3)ND (LOQ 20)–192 [median=29]  
Soft drinks (4)ND (LOQ 10)  
Beverages (7)ND (LOD 0.9)–3.4 [median=0.4]AustriaBraunrath et al. (2005)
Vegetables (6) (only solid portion was analyzed, with the exception of tomatoes)8.5–35 [median=26]  
Fruits (4)5–24 [median=6.6]  
Canned fat-containing products such as soups, meats, and cream (9)2.1–37.6 [median=20.7]  
Tuna (9)<ND (LOQ 7.1)–102.7 [median=11.2]MexicoMunguía-López et al. (2005)
Beverage/beer cans exposed to 10% ethanol at 150°F [65.6°C] for 30 min and then 120°F [48.9°C] for 10 daysND (LOD 5)U.S.Howe et al. (1998) FDA (1996)
Food cans exposed to 10 or 95% ethanol at 250°F [121°C] for 2 hr and then 120°F [48.9°C] for 10 days or at 212°F [100°C] for 30 min and then 120°F [48.9°C] for 10 daysND (LOD 5)–95 (mean=37)g  
Honey (107 samples; ∼90% imported in epoxy-lined drums)ND (LOD 2)–33.3 [median=<2]JapanInoue et al. (2003b)
Wine stored in steel, wood, or plastic vats, filled into glass bottles, or purchased in local markets (59)<LOQ (0.2 ng/mL) to 2.1 μg/L; mean 0.58 in samples above the LOQAustriaBrenn-Struckhofova and Chichna-Markl (2006)
Solid food (309)ND (<∼0.8 )–192 [median=3.52–4.32]U.S.Wilson et al. (2006)
Liquid food (287)ND (<∼ 0.3)–17.0 [median=0.45–0.79]U.S.Wilson et al. (2006)

In one study, empty cans were filled with soup, beef, evaporated milk, carrots, or 10% ethanol (Goodson et al., 2004). The cans were then sealed, processed at 5, 20, or 40°C, and sampled at 1 or 10 days or 1, 3, or 9 months. Half the cans processed according to each condition were dented. It was determined that 80–100% of the bisphenol A migrated to food immediately after processing, and that bisphenol A concentrations did not change during storage or as a result of denting. The study authors concluded that most migration occurred during can processing. Boiling the cans or heating to 230°C did not increase migration of bisphenol A, but that finding appears to contrast with findings of others. Kang et al. (2003) examined the effects of temperature, duration of heating, glucose, sodium, and oil on migration of bisphenol A from cans. In cans filled with water, heating to 121°C compared to 105°C increased migration of bisphenol A but the duration of heating had no significant effect. Compared to cans filled with water, increased amounts of bisphenol A migrated from cans filled with 1–10% sodium chloride, 5–20% glucose, or vegetable oils and heated to 121°C. Takao et al. (2002) reported increased leaching of bisphenol A from cans into water when the cans were heated to ≥80°C.

A study examining aggregate exposures of U.S. preschool age children measured bisphenol A concentrations in liquid food and solid food served to the children at home and at child care centers (Wilson et al., 2003). Duplicate plates of food served to nine children were collected over a 48-hr period. GC/MS analyses were conducted on four liquid food samples and four solid food samples from the child care center and nine liquid food samples and nine solid food samples from home. Bisphenol A was detected in all solid food samples, three liquid food samples from the child care center, and two liquid food samples from the home. Concentrations of bisphenol A ranged from <0.100–1.16 ng/g [μg/kg] in liquid foods and from 0.172–4.19 ng/g [μg/kg] in solid food.

The study examining aggregate exposures of U.S. preschool age children was repeated with a larger sample and again measured bisphenol A concentrations in liquid food and solid food served to the children at home and at child care centers (Wilson et al., 2006). Bisphenol A concentrations were measured by GC/MS in food served over a 48-hr period to at least 238 children at home and 49 children at daycare centers. Bisphenol A was detected in 83–100% of solid food samples; concentrations were reported at <LOD (0.8) to 192 ng/g [μg/kg]. Sixty-nine to 80% of liquid food contained detectable concentrations of bisphenol A; concentrations were reported at <LOD (0.3)–17.0 ng/mL in liquid food. Data were also collected for hand wipes of 193 children at daycare centers and 60 children at home. Bisphenol A was detected in 94–100% of handwipe samples; concentrations ranged from <LOD [not defined] to 46.6 ng/cm2. Bisphenol A was detected in 85–89% of food preparation surface wipes from homes; concentrations were reported at <LOD [not defined] to 0.357 ng/cm2.

A review by Miyamoto and Kotake (2006) reported bisphenol A concentrations of 0.011–0.086 mg/kg in non-canned foods such as fats, fruits, fish, meat, and eggs. However, one study used GC-MS to examine bisphenol A in 14 types of produce purchased in southern Italy (Vivacqua et al., 2003). Bisphenol A concentrations were below the detection limit [not reported] in 5 produce samples. In the remaining samples, bisphenol A was detected at concentrations of 0.25±0.02 (SD) to 1.11±0.09 mg/kg. [These concentrations are equal to or higher than those found in canned foods, where the presumption is that the source is the epoxy liner of the container.]

Bisphenol A has been found in recycled paper products used for food processing at 10 or more times the concentrations found in non-recycled paper products [reviewed by the (European Food Safety Authority, 2006)]. Bisphenol A concentrations were up to 26 μg/g paper. Migration to food was not discussed.

Epoxy paints are used to coat the insides of residential drinking water storage tanks. Bisphenol A has been shown to migrate from painted concrete and stainless metallic plates; however, a water sample from a recently painted reservoir showed no detectable bisphenol A (Romero et al., 2002). When exposed to chlorine disinfectant, bisphenol A disappears within 4 hr, but the chlorinated bisphenol A congeners that are formed can remain in solution up to 20 hr when low chlorine doses are used (Gallard et al., 2004). The toxicity of these chlorinated bisphenol A congeners is unknown; however, there is some evidence that estrogenic activity and receptor binding remains after chlorination (Hu et al., 2002).

1.2.3.3 Potential migration from dental material:

Bisphenol A is used in the manufacture of materials found in dental sealants or composites (i.e., fillings) (European-Union, 2003). Examples of bisphenol A-derived materials used in dental sealants include bis-glycidyldimethacrylate and bisphenol A-dimethyl acrylate. Bisphenol A could potentially be present as an impurity or be released during degradation of the dental materials. Sealants are comprised of an organic matrix, while composites contain inorganic filler in addition to the organic matrix. According to the British Dental Association, filled composites would possibly produce lower exposure to bisphenol A than sealants, because they contain proportionately less resin than sealants [reviewed in (European-Union, 2003)]. During dental procedures, resin mixtures are applied as fluid monomers and polymerized in situ by ultraviolet or visible light. According to the European Union (2003), patients can be exposed to bisphenol A during the polymerization stage.

In a review of in vitro studies examining bisphenol A migration from dental sealants, the European Union (2003) concluded that release of bisphenol A is likely to occur only with degradation of the parent monomer. The data suggested that bis-glycidyldimethacrylate does not degrade; therefore, release of bisphenol A is only likely to occur with bisphenol A-dimethyl acrylate use. In vivo studies measuring bisphenol A in saliva following sealant application were reviewed in detail by CERHR because they provide the most relevant human exposure information.

Olea et al. (1996) measured saliva concentrations of bisphenol A for 1 hr before and 1 hr after application of 50 mg bis-glycidyldimethacrylate- and bisphenol A-dimethyl acrylate-based sealant across 12 molars of 18 patients. Concentrations of bisphenol A in saliva were measured by GC/MS and HPLC. Following treatment, saliva contained ∼90–931 μg bisphenol A. Based on an assumed saliva production rate of 0.5 mL/min, a saliva concentration of 3–30 μg/mL was estimated by the study authors. With the exception of 1 patient who was excluded from the study, bisphenol A was not detected in saliva before sealant application.

Arenholt-Binslev (1999) measured bisphenol A in saliva of 8 adult patients who each had four molars treated with 38 mg of 1 of 2 sealants, Delton LC or Visio-seal. Saliva was collected before, immediately after, and at 1 or 24 hr following treatment for measurement of bisphenol A concentrations by HPLC. Bisphenol A was detected at 0.3–2.8 ppm immediately after application of Delton LC sealant [bisphenol A-dimethyl acrylate sealant according to the European Union (2003)] but was not detected 24 hr later (detection limit=0.1 ppm [mg/L]). Bisphenol A was not detected in saliva of patients who received the Visio-seal sealant (bis-glycidyldimethacrylate sealant, according to the European Union). It was noted that saliva bisphenol A concentrations were much lower than those reported by Olea et al. (1996). Possible reasons for the inconsistencies in results between the 2 studies were stated to be differences in the amount of sealant used and co-elution of compounds that could have confounded bisphenol A analysis.

Fung et al. (2000) measured salivary bisphenol A concentrations in 40 patients treated with a dental sealant (Delton Opaque Light-cure Pit and Fissure Sealant) that was understood to contain bisphenol A-dimethyl acrylate, according to the European Union (2003). Eighteen patients in the low-dose group received 8 mg dental sealant on 1 tooth, and 22 patients in the high-dose group received 32 mg sealant on 4 teeth. Saliva and blood were collected for HPLC analysis before the procedure and at 1 and 3 hr and 1, 3, and 5 days after the procedure. More details of this study are included in Section 2.1.1.1. Analysis of the dental sealant revealed that bisphenol A concentrations were below the detection limit of 5 ppb. At 1 hr following treatment, Bisphenol A was detected only in saliva samples from 3 of 18 volunteers in the low-dose group and 13 of 22 samples from volunteers in the high-dose group. At 3 hr post-treatment, bisphenol A was detected in samples from 1 of 18 volunteers in the low-dose group and 7 of 22 volunteers from the high-dose group. Concentrations of bisphenol A in saliva at 1 and 3 hr following exposure were reported at 5.8–105.6 ppb [μg/L]. No bisphenol A was detected in saliva samples at 24 hr after treatment or in serum samples at any time point. Differences in bisphenol A concentrations and the presence of bisphenol A in saliva of the low-dose compared to the high-dose group at 1 and 3 hr achieved statistical significance. The European Union (2003) noted that the concentrations of saliva bisphenol A reported by Fung et al. (2000) were >250 times lower than those reported by Olea et al. (1996).

Sasaki et al. (2005) used ELISA to examine salivary bisphenol A concentrations in 21 patients before and after 1 cavity was filled with 0.1 g of composite resin. The resins consisted of bisphenol A diglycidylether methacrylate (i.e., bis-glycidyldimethacrylate), triethylene glycol dimethacrylate, and/or urethane dimethacrylate. Saliva was collected before treatment, during the 5 min following treatment, and then immediately after gargling with water. Following treatment, saliva bisphenol A increased [from ≤2 to ∼15–100μg/L]. Gargling reduced bisphenol A to near pretreatment concentrations [≤5μg/L] in most patients, with the exception of 1 patient with the highest bisphenol A concentration [reduced from ∼100 to 18μg/L]. [An increase in saliva bisphenol A concentrations was noted in 1 of 2 patients receiving a composite consisting solely of urethane dimethacrylate.] The study authors noted that cross-reactivity is possible with the ELISA technique, but that cross reactivity between bisphenol A diglycidylether methacrylate and triethylene glycol dimethacrylate is low. Therefore, the study authors thought it possible that they were measuring only bisphenol A. [As discussed inSection 1.1.5, ELISA may overestimate bisphenol A.]

Joskow et al. (2006) examined bisphenol A in urine and saliva of 14 adults treated with dental sealants. The volunteers received either Helioseal F (n=5) or Delton LC (n=9) sealant. Only the Helioseal F sealant was noted to carry the American Dental Association Seal of Acceptance. Sealant was weighed before and after application to determine the amount applied, and the numbers of treated teeth were recorded. The mean number of teeth treated was 6/person and the mean total weight of sealant applied was 40.35 mg/person. In a comparison of the two different sealants, no differences were reported for the number of teeth treated or amount of sealant applied. Saliva samples were collected before, immediately after, and 1 hr after sealant application. Urine samples were collected before and at 1 and 24 hr after sealant placement. A total of 14–15 saliva samples and 12–14 urine samples were collected at each time point. Samples were treated with β-glucuronidase and analyzed for bisphenol A concentrations using selective and sensitive isotope-dilution-MS-based methods. Saliva concentrations were highest immediately following treatment; mean concentrations were reported at 42.8 ng/mL in patients treated with Delton LC and 0.54 ng/mL in patients treated with Helioseal F. The highest mean urinary concentrations of bisphenol A were measured at 1 hr following exposure and were reported at 27.3 ng/mL in patients treated with Delton LC and 7.26 ng/mL in patients receiving the Helioseal F sealant. The study authors noted that saliva and urine bisphenol A concentrations after application of Helioseal F were comparable to baseline concentrations. More information on bisphenol A concentrations in saliva and urine is included in Section 2, and exposure estimates are provided in Section 1.2.4.1.2. The study authors noted that saliva concentrations detected in their study were ∼1000 times lower than those reported by Olea et al. (1996) but were within the ranges reported by Fung et al. (2000) and Sasaki et al. (2005). Analytical procedures and use of a large amount of sealant were noted as possible reasons for the higher values reported by Olea et al. (1996).

The European Union noted a study by Lewis et al. (1999) that characterized materials in 28 commercial resin-based composites and sealants, including those examined by Olea et al. (1996). HPLC and infrared analysis could not verify the presence of bisphenol A in any sealant product. Lewis et al. (1999) noted that in the study by Olea et al. (1996) another component in the resin may have been misidentified as bisphenol A because of difficulties with resolution.

In their review of studies examining bisphenol A concentrations in saliva of patients treated with dental sealants, the European Union (2003) noted that the higher concentrations reported may have resulted from interference during analysis and thus may overestimate bisphenol A exposures from dental treatments. It was concluded that dental treatment would likely result in saliva bisphenol A concentrations of 0.3–3 ppm. Because bisphenol A was generally not detected in saliva at time points beyond 1 hr after treatment, it was concluded that bisphenol A exposure resulting from dental treatments is likely to be an acute event. In their 2002 position statement, the American Dental Association stated that none of the 12 dental sealants that carry the American Dental Association Seal release bisphenol A (American Dental Association, 1998). On initial analysis, one of the sealants was found to leach trace concentrations of bisphenol A, but following implementation of quality controls by the manufacturer, bisphenol A could no longer be detected in the final product.

A study on orthodontic adhesives found no bisphenol A release from these materials after simulated aging (Eliades et al., 2007). Another study found plastic orthodontic brackets in water to release bisphenol A at 0.01–0.40 mg/kg material and denture base resin in water to release bisphenol A at 0.01–0.09 mg/kg material (Suzuki et al., 2000).

1.2.3.4 Bisphenol A concentrations measured in biological samples:

Bisphenol A concentrations detected in human blood are summarized in Table 7. Goodman et al. (2006) noted that although blood concentration may provide information on internal dose, it does not allow for estimates of daily intake. It was also noted that in many studies in which blood concentration of bisphenol A was measured, sample preparation and analysis methods were poorly reported. Many study groups used an ELISA method to measure blood bisphenol A concentration. As discussed in Section 1.1.5, the ELISA technique is likely to overestimate bisphenol A concentrations as a result of cross-reactivity with other substances and due to effects of biologic matrices (Inoue et al., 2002; Fukata et al., 2006; Goodman et al., 2006).

Table 7. Blood Concentrations of Bisphenol A in Adults
Population (n)Bisphenol A μg/LacMethodReference
  • a

    aMean±SD or median (range).

  • b

    bAs discussed in Section 1.1.5, ELISA may overestimate bisphenol A.

  • c

    cIt is uncertain whether parent, conjugated, or total bisphenol A was measured.

  • d

    dEstimated from a graph.

Germany
 Men (7)<0.5HPLC-MS/MSVölkel et al. (2005)
 Women (12)<0.5HPLC-MS/MSVölkel et al. (2005)
 Pregnant Caucasian women (37; 32–41 weeks gestation)4.4±3.9GC-MSSchönfelder et al. (2002a)
Japan
 Men (21; age 22–51)“almost all” <0.2 ng/mlHPLC-ECDFukata et al. (2006)
 Men (9; age 30–50)0.59±0.21 (0.38–1.0)HPLC-MSSajiki et al.(1999)
 Men (11)1.49±0.11 (SEM)ELISAbTakeuchi and Tsutsumi (2002)
 Women (31; age 22–51)“almost all” <0.2 ng/mlHPLC-ECDFukata et al. (2006)
 Women (12; age 30–50)0.33±0.54 (0–1.6)HPLC-MSSajiki et al.(1999)
 Women (14)0.64±0.10(SEM)ELISAbTakeuchi and Tsutsumi (2002)
 Pregnant women (37; late pregnancy)1.4±0.9ELISAbIkezuki et al.(2002)
 Pregnant women with normal karyotype early 2nd trimester (200)2.24 (0.63–14.36)ELISAbYamada et al. (2002)
 Pregnant women with abnormal karyotype early 2nd trimester (48)2.97 (∼0.0.7–18.5)dELISAbYamada et al. (2002)
 Pregnant women (9)0.43 (0.21–0.79)HPLC-FlKuroda et al. (2003)
 Infertile women (21)0.46 (0.22–0.87)HPLC-FlKuroda et al. (2003)
 Women with multiple miscarriages (45; mean age 31.6 years)2.59±5.23ELISAbSugiura-Ogasawara et al. (2005)
 Healthy woman (32; mean age 32 years)0.77±0.38ELISAbSugiura-Ogasawara et al. (2005)
 Women with polycystic ovary syndrome (16)1.04±0.10 (SEM)ELISAbTakeuchi and Tsutsumi (2002)
 Non-obese women with polycystic ovarian syndrome (13; average age 26.5 years)1.05±0.10 (SEM)ELISAbTakeuchi et al. (2004a)
 Obese women with polycystic ovarian syndrome (6; average age 24.7 years)1.17±0.16 (SEM)ELISAbTakeuchi et al. (2004a)
 Non-obese women (19; average age 27.5 years)0.71±0.09 (SEM)ELISAbTakeuchi and Tsutsumi (2002)
 Obese women (7; average age 28.8 years)1.04±0.09 (SEM)ELISAbTakeuchi et al. (2004a)
 Hyperprolactinemic women (7; average age 27.7 years)0.83±0.12 (SEM)ELISAbTakeuchi et al. (2004a)
 Amenorrheic women (7; average age 25.1 years)0.84±0.10 (SEM)ELISAbTakeuchi et al. (2004a)
 Women with normal uterine endometrium (11; mean age 48.9 years2.5±1.5ELISAbHiroi et al. (2004)
 Women with simple endometrium hyperplasia (10; mean age 48.4 years)2.9±2.0ELISAbHiroi et al. (2004)
 Women with complex endometrium hyperplasia (9; mean age 48.4 years)1.4±0.4ELISAbHiroi et al. (2004)
 Women with endometrial carcinoma (7; mean age 63.1 years)1.4±0.5ELISAbHiroi et al. (2004)

Several studies reported concentrations of bisphenol A in human urine; those studies are summarized in Table 8. As discussed in greater detail in Section 2, the majority of ingested bisphenol A is excreted in urine as bisphenol A glucuronide after acute exposure. Smaller amounts of bisphenol A are metabolized to and excreted as bisphenol A sulfate. Some of the studies determined concentrations of parent bisphenol A before and after digestion with glucuronidases. With the exception of Fujimaki et al. (2004) who used an ELISA technique to measure urinary bisphenol A, other study authors used HPLC, GC/MS, or LC/MS. Results from 394 participants of the National Health and Nutrition Examination Survey (NHANES) III survey are included in Table 8 (Calafat et al., 2005). Bisphenol A was detected in 95% of the participants, which indicated widespread exposure to bisphenol A in the U.S. Consistent with those findings, bisphenol A was detected in urine from 85 of 90 (94.4%) 6–8-year-old girls from the U.S. (Wolff et al., 2006). In a review of urinary bisphenol A data, Goodman et al. (2006) noted that in most cases, median total urinary bisphenol A concentration (the sum of parent and conjugated bisphenol A) were ∼1–2 μg/L. Two studies (Yang et al., 2003; Mao et al., 2004) reported urinary bisphenol A concentrations that were orders of magnitude higher than commonly observed concentrations, despite the use of apparently reliable analytical techniques. Goodman et al. (2006) has suggested that reported hormone concentrations for the study volunteers were also higher than expected, indicating the possibility of laboratory or reporting error. The use of urinary bisphenol A concentration to estimate daily exposures appears in Section 1.2.4.1.2.

Table 8. Urinary Concentrations of Bisphenol A and Metabolites in Adults or Children
   Urinary bisphenol A or metabolite concentrations as median (range) or mean±SEM, μg/La [detectable fraction, % >LOD] 
CountryStudy populationLOD (μg/L)FreeTotalGlucuronideSulfateReference
  • a

    aWith the exception of the study by Fujimaki et al. (2004), which used the potentially unreliable ELISA, the studies used analytical techniques based on HPLC, GC/MS, and LC/MS.

  • b

    bLimit of detection (LOD) for bisphenol A following digestion of conjugate was 0.3 μg/L.

  • c

    cSamples were only digested with β-glucuronidase and do not account for bisphenol A conjugated to sulfate.

  • d

    dVariance not indicated.

  • e

    eMinimum detection limit based upon free bisphenol A.

U.S.30 urine samples from demographically diverse, anonymous adult volunteers0.3< 0.3 (<0.3–0.6) [10%]2.12 (<LODb –19.8) [97%]1.4 (<LODb –19.0) [90%]0.3 (<LODb –1.8) [47%]Ye et al. (2005)
U.S.394 adult volunteers (men and women; 20–59 years old) fromthe NHANES III survey0.1 1.28 (10–95th percentile: 0.22–5.18) [95%]c  Calafat et al. (2005)
U.S.23 adults0.5 0.47 (<1–2.24) [52%]  Liu et al. (2005)
U.S.9 girls (9 years of age)0.5 2.4 (0.04–16) [89%]  Liu et al. (2005)
U.S.90 women (6–8 years of age; White, Black, Asian, or Hispanic ethnicity)8.36 1.8 (<0.3–54.3) [85%]  Wolff et al. (2006)
Germany7 men, 12 women1.14 (BPA) 10.1 (BPA monoglucuronide)<1.14 [0%] <26.26 [LOQ] Völkel et al. (2005)
Korea15 men (42.6±2.4 years of age)d10.28–2.36; 0.58±0.140.85–9.83 2.82±0.730.16–11.67 2.34±0.85<MDL –1.03; 0.49±0.27eKim et al. (2003b)
Korea15 women (43.0±2.7 years of age)d0.280.068–1.65; 0.56±0.101.00–7.64 2.76±0.54<MDLc –4.34 1.00±0.34<MDL –3.40; 1.20±0.32eKim et al. (2003b)
Korea34 men, 39 women (mean=48.5 years of age)0.012 Geometric mean: 9.54 (<0.012–586.14b) [75%]  Yang et al. (2003)
Korea81 men not occupationally exposed to bisphenol A  Geometric mean±SD 6.88±3.72  Yang et al. (2006)
Korea79 women not occupationally exposed to bisphenol A0.026 Geometric mean±SD 5.01±3.16 [97.5%]  Yang et al. (2006)
Japan48 woman college students0.2<0.2 [2%] 1.2 (0.2–19.1) [100%] Ouchi and Watanabe (2002)
JapanPooled urine samples from at least 5 people0.12<0.120.11–0.51  Brock et al. (2001)
Japan23 women, 46 men; in each volunteer, 2 samples per volunteer were combined 0.01–0.27Mean=0.81 (range: 0.14–5.47)  Tsukioka et al. (2004)
JapanWhole-day urine samples collected from 11 men and 11 women  Mean=0.81 (range 0.24–2.03)  Tsukioka et al. (2004)
JapanUrine collected from 3 volunteers0.02<0.10.22, 0.41, and 0.45[100% after deconjugation]  Kawaguchi et al. (2004)
JapanSpot urine samples collected from 56 women who were 1–9 months pregnant; 21–43 years of age1.1 <1.1 (<1.1–5.4)c (ELISA) [30%]  Fujimaki et al. (2004)
Japan21 men, 31 women; 22–51 years of age0.249/51 had <0.2 mean 0.34 (n=2) [4%]1.92±0.27 [98%]  Fukata et al. (2006)
China10 healthy man volunteers; 21–29 years of age2.8 <2.7 –3950 1220±1380d [60%]  Mao et al. (2004)
China10 healthy woman volunteers; 21–29 years of age2.8 30–3740 1290±1220d [100%]  Mao et al. (2004)

In humans, bisphenol A was measured in cord blood and amniotic fluid, demonstrating distribution to the embryo or fetus. Detailed descriptions of those studies are also presented below.

Engel et al. (2006) reported concentrations of bisphenol A in human amniotic fluid. Twenty-one samples were obtained during amniocentesis conducted before 20 weeks gestation in women who were referred to a U.S. medical center for advanced maternal age. Bisphenol A concentrations in amniotic fluid were measured using LC with electrochemical detection. Bisphenol A was detected in 10% of samples at concentrations exceeding the LOD (0.5 μg/L). Bisphenol A concentration ranges of 0.5–1.96 μg/L were reported.

Schönfelder et al. (2002b) examined bisphenol A concentrations in maternal and fetal blood and compared bisphenol A concentrations in blood of male and female fetuses. In a study conducted at a German medical center, blood samples were obtained from 37 Caucasian women between 32 and 41 weeks gestation. At parturition, blood was collected from the umbilical vein after expulsion of the placenta. Bisphenol A concentrations in plasma were measured by GC/MS. Control experiments were conducted to verify that bisphenol A did not leach from collection, storage, or testing equipment. Bisphenol A was detected in all samples tested, and concentrations measured in maternal and fetal blood are summarized in Table 9. Mean bisphenol A concentrations were higher in maternal (4.4±3.9 [SD] μg/L) than fetal blood (2.9±2.5 μg/L). Study authors noted that in 14 cases fetal bisphenol A plasma concentrations exceeded those detected in maternal plasma. Among those 14 cases, 12 fetuses were male. Analysis by paired t-test revealed significantly higher mean bisphenol A concentrations in the blood of male than female fetuses (3.5±2.7 vs. 1.7±1.5 ng/mL, P=0.016). Bisphenol A concentrations were measured in placental samples at 1.0–104.9 μg/kg.

Table 9. Concentrations of Bisphenol A in Maternal and Fetal Samples
 Bisphenol A concentrations, μg/L median (range) or mean±SD 
 Serum or plasma  
Study description (analytical method)MaternalFetalAmniotic fluidReference
  • a

    aAs discussed in Section 1.1.5, ELISA may overestimate bisphenol A. Some samples were verified by HPLC.

  • a, b

    bEstimated from a graph.

21 samples collected in women in the U.S. before 20 weeks gestation (LC with electrochemical detection)  0.5 (Non-detectable <0.5–1.96) 10% of samples detectableEngel et al. (2006)
37 German women, 32–41 weeks gestation (GC/MS)3.1 (0.3–18.9) 4.4±3.92.3 (0.2–9.2) 2.9±2.5 Schönfelder et al. (2002b)
37 Japanese women in early pregnancy (ELISA)a1.5±1.2  Ikezuki et al. (2002)
37 Japanese women in late pregnancy (ELISA)a1.4±0.9  Ikezuki et al. (2002)
32 Japanese infants at delivery (ELISA)a 2.2±1.8 Ikezuki et al. (2002)
32 Japanese amniocentesis samples at 15–18 weeks gestation (ELISA)a  8.3±8.9Ikezuki et al. (2002)
38 samples obtained at full-term cesarean section (ELISA)a  1.1±1.0Ikezuki et al. (2002)
200 Japanese women carrying fetuses with normal karyotype at 16 weeks mean gestation (ELISA)2.24 (0.63–14.36) 0.26 (0–5.62)Yamada et al. (2002)
48 Japanese women carrying fetuses with abnormal karyotypes at a 16 weeks mean gestation (ELISA)2.97 [0.718.5]b 0 [07.5]bYamada et al. (2002)
9 sets of maternal and umbilical cord blood samples obtained at birth in Japanese patients (HPLC)0.43 (0.21–0.79) 0.46±0.20.64 (0.45–0.76) 0.62±0.13 Kuroda et al. (2003)
180 Malaysian newborns (GC/MS) Non-detectable (<0.05) to 4.05 88% of samples detectable Tan and Mohd (2003)

Ikezuki et al. (2002) measured concentrations of bisphenol A in serum from 30 healthy premenopausal women, 37 women in early pregnancy, 37 women in late pregnancy, and 32 umbilical cord blood samples. Concentrations of bisphenol A were also measured in 32 samples of amniotic fluid obtained during weeks 15–18 of gestation, 38 samples of amniotic fluid obtained at full-term cesarean section, and 36 samples of ovarian follicular fluid collected during in vitro fertilization procedures. [It was not stated if different sample types were obtained from the same subjects.] An ELISA method was used to measure bisphenol A concentrations and results were verified by HPLC. The mean±SD concentration of bisphenol A in follicular fluid was reported at 2.4±0.8 μg/L. As summarized in Table 9 for maternal and fetal samples, concentrations of bisphenol A in follicular fluid were similar to those detected in the serum of fetuses and pregnant and non-pregnant women and in amniotic fluid collected in late pregnancy (∼1–2 μg/L). Bisphenol A concentrations in amniotic fluid samples collected in early pregnancy were ∼5-fold higher than in other samples, and the difference achieved statistical significance (P<0.0001). Study authors postulated that the higher concentrations of bisphenol A in amniotic fluid collected during gestation weeks 15–18 may have resulted from immature fetal liver function. They noted that according to unpublished data from their laboratory, the percentage of glucuronidated bisphenol A in mid-term amniotic fluid was ∼34%, which is much lower than reported values for other human fluids (>90%).

Yamada et al. (2002) measured bisphenol A concentrations in maternal serum and amniotic fluid from Japanese women. Samples were collected between 1989 and 1998 in women undergoing amniocentesis around gestation week 16. One group of samples was obtained from 200 women carrying fetuses with normal karyotypes, and a second group of samples was obtained from 48 women carrying fetuses with abnormal karyotypes. An ELISA method was used to measure bisphenol A concentrations. [As discussed inSection 1.1.5, ELISA may overestimate bisphenol A.] Concentrations of bisphenol A measured in maternal plasma and amniotic fluid are summarized in Table 9. Median concentrations of bisphenol A in maternal serum (∼2–3 μg/L) were significantly higher [∼10-fold] than concentrations in amniotic fluid (∼0–0.26 μg/L) in the groups carrying fetuses with normal and abnormal karyotypes. However, in 8 samples from women carrying fetuses with normal karyotypes, high concentrations (2.80–5.62 μg/L) of bisphenol A were measured in amniotic fluid. The study authors interpreted the data as indicating that bisphenol A does not accumulate in amniotic fluid in most cases but that accumulation is possible in some individuals. Bisphenol A concentrations in maternal blood were significantly higher [by ∼33%] in woman carrying fetuses with abnormal versus normal karyotypes. However, the study authors noted that the effect may not be related to bisphenol A exposure because there was no adjustment for maternal age, and concentrations in amniotic fluid did not differ between groups. In the group carrying fetuses with normal karyotypes, data obtained from 1989–1998 were summarized by year. Median bisphenol A concentrations in serum significantly decreased over that time from a concentration of 5.62 μg/L detected in 1989 to 0.99 μg/L in 1998.

Kuroda et al. (2003) used an HPLC method to measure bisphenol A concentrations in 9 sets of maternal and cord blood samples obtained from Japanese patients at the time of delivery. Bisphenol A concentrations were also measured in 21 sets of serum and ascitic fluid samples collected from sterile Japanese patients of unspecified sexes and ages. Results for pregnant women are summarized in Table 9. Mean±SD concentrations of bisphenol A were lower in maternal (0.46±0.20 ppb [μg/L]) than cord blood (0.62±0.13 ppb [μg/L]). There was a weak positive correlation (r=0.626) between bisphenol A concentrations in maternal and cord blood. There were no differences between pregnant and non-pregnant blood levels (Kuroda et al., 2003). Mean±SD concentrations of bisphenol A were higher in ascitic fluid (0.56±0.19 ppb [μg/L]) than in serum (0.46±0.20 ppb [μg/L]). The correlation between bisphenol A concentration in serum and ascitic fluid was relatively strong (r=0.785).

Tan and Mohd (2003) used a GC/MS method to measure bisphenol A concentrations in cord blood at delivery in 180 patients at a Malaysian medical center. Bisphenol A was detected in 88% of samples. As noted in Table 9, concentrations ranged from <0.10–4.05 μg/L.

Schaefer et al. (2000) measured concentrations of bisphenol A and other compounds in uterine endometrium of women undergoing hysterectomy for uterine myoma at a German medical center. Endometrial and fat samples were obtained between 1995–1998 from 23 women (34–51 years old) with no occupational exposure to bisphenol A. Samples were handled with plastic-free materials and stored in glass containers. Concentrations of environmental chemicals were measured in samples by GC/MS. None of 21 fat samples had detectable concentrations of bisphenol A. Bisphenol A was detected in 1 of 23 endometrial samples; the median concentration was reported at <1 μg/kg wet weight, and the range was reported at 0–13 μg/kg. [It is not known why a median value and range were reported when bisphenol A was only detected in 1 sample.]

As part of a study to compare an ELISA and an LC/MS method for biological monitoring of bisphenol A, Inoue et al. (2002) measured concentrations of bisphenol A in semen samples obtained from 41 healthy Japanese volunteers (18–38 years old). Analysis by the ELISA method indicated bisphenol A concentrations ranging from concentrations below the detection limit (2.0 μg/L) to 12.0 μg/L. The LC/MS method indicated that the bisphenol A concentration in all samples was <0.5 μg/L, the LOQ. The study authors concluded that the LC/MS method was more accurate and sensitive and that the ELISA method overestimated bisphenol A concentrations, possibly due in part to nonspecific antibody interactions.

1.2.4 Human exposure

1.2.4.1 General population exposure
1.2.4.1.1 Estimates based on bisphenol A concentrations in food or environment:

Wilson et al. (2003) estimated aggregate exposures to bisphenol A in preschool aged children (2–5 years) from the U.S. In 1997, numerous chemicals were surveyed, but only bisphenol A results are reported here. Ten child care centers were surveyed and the 2 centers with the highest and lowest overall concentrations of target pollutants were selected for the study. Both centers were located in North Carolina. Nine children who attended one of the child care centers participated in the study. Over a 48-hr period, bisphenol A concentrations were measured in indoor and outdoor air, dust, soil, and food; the ranges detected are summarized in Sections 1.2.3.1 and 1.2.3.2. In estimating exposures, absorption was considered to be 100%. Calculations considered ventilation rates, time spent indoors and outdoors, time spent at home and in day care, the measured weight of each child, assumed ingestion of dust and soil, and total weight of foods consumed. Mean (range) bisphenol A intake was estimated at 0.042981 (0.018466–0.071124) μg/kg bw/day.

Wilson et al. (2006) conducted a second study to estimate aggregate exposures in 257 U.S. children aged 1.5–5 years. Bisphenol A was one of the compounds assessed in this study of homes and daycare centers in 6 North Carolina and 6 Ohio counties in 2000–2001. Over a 48-hr period, bisphenol A concentrations were measured in indoor and outdoor air, dust, soil, food, and surface and hand wipes; the ranges detected are summarized in Sections 1.2.3.1 and 1.2.3.2. In estimating exposures, absorption was considered to be 50%. Calculations considered ventilation rates, time spent indoors and outdoors, time spent at home and in day care, the measured weight of each child, assumed ingestion of dust and soil, and total weight of foods consumed. Median (25th percentile to maximum) bisphenol A aggregate exposures were estimated at 2.56 (1.5–57.2) μg/day for children from North Carolina and 1.88 (1.27–48.6) μg/day in children from Ohio. Median (25th percentile to maximum) potential aggregate dose, assuming 50% absorption, was estimated at 0.0714 (0.0424–1.57) μg/kg bw/day in children from North Carolina and 0.0608 (0.0341–0.775) μg/kg bw/day in children from Ohio. The study authors noted that 99% of exposure occurred through dietary ingestion.

The European Union (2003) conducted a comprehensive exposure estimate that considered exposures resulting from food and environmental sources. Oral exposure estimates for children and adults were reported and are summarized in Table 10. Estimates were based on migration studies conducted with polycarbonate and concentrations of bisphenol A measured in foods packaged in epoxy-lined cans. Assumptions used in exposure estimates included 100% oral absorption and body weights of 70 kg for adults, 14.5 kg for 1.5–4.5-year-old children, 4.5 kg for 1–2-month-old infants, 7 kg for 4–6-month-old infants, and 8.7 kg for 6–12-month-old infants. Estimated exposures for children were said to represent realistic worst-case scenarios for food and drink intake relative to body weight.

Table 10. Bisphenol A Oral Exposure Estimates by the European Uniona
   Bisphenol A intake
Exposure source (exposed population)Daily food intakeBisphenol A concentration in foodμg/dayμg/kg bw/day
  • a

    aEuropean Union (2003).

  • b

    bThe European Union acknowledged that exposure through wine represents a very worst-case scenario.

Infant bottles0.699 L/day milk50 μg/L358
(1–2 month-old infant)    
Infant bottles0.983 L/day milk50 μg/L507
(4–6-month-old infant)    
Polycarbonate tableware (1.5–4.5-year-old child)2 kg food/day5 μg/kg100.7
Canned food0.375 kg canned food/day100 μg/kg405
(6–12- month-old infant)    
Canned food2 kg canned food/day100 μg/kg20014
(1.5–4.5-year-old child)    
Canned food1.0 kg canned food/day100 μg/kg1001.4
(adult)    
Wine0.75 L/day650 μg/L5007b
(adult)    
Canned food and wine0.75 L/day wine650 μg/L in wine6009b
(adult)1.0 kg canned food/day100 μg/kg food  

The European Union (2003) also estimated human environmental exposure to bisphenol A from sources such as drinking water, fish, plants, milk, meat, and air. The values were apparently obtained using the European Union System for the Evaluation of Substances (EUSES) model. Total regional exposure to bisphenol A was estimated at 0.0178 μg/kg bw/day. The highest local exposure was thought to occur in the vicinity of PVC-producing plants and was estimated at 59 μg/kg bw/day. Aggregate exposures in adults involving food, wine, and environmental sources were estimated at 9 μg/kg bw/day for regional scenarios and 69 μg/kg bw/day for worst-case local scenarios occurring near a PVC-manufacturing plant. However, it was noted in the European Union report that use of bisphenol A in PVC manufacture was being phased out.

The European Union (2003) noted that exposures to bisphenol A through dental sealant are single and rare events and do not lead to repeated exposure. Therefore, the issue was not considered further.

Exposures to bisphenol A from some consumer products were identified and characterized by the European Union (2003). Products included: marine antifouling agents used on boats, wood varnish, wood fillers, and adhesives. With the exception of adhesives for which frequent use was thought possible, exposure to the other products was considered to be relatively rare. Exposures were estimated based on factors such as epoxy and residual bisphenol A concentrations, exposure time, area of skin exposed, and possible generation of mists during processes such as brushing. Inhalation exposures by product were estimated at 3 × 10−4 μg for antifouling agents and 0.02 μg for wood varnish. Dermal exposure by product without protective clothing was estimated at 29 μg for antifouling agents, 3.6 μg for wood varnish, 9 μg for wood filler, and 14 μg for adhesives. [Dermal exposure to adhesives appears to be incorrectly reported as 1μg in Table 4.20 of the European Union review.] Exposure was estimated to be 1–2 orders of magnitude lower when protective clothing such as gloves was used. Assuming an absorption rate of 10%, dermal exposure to bisphenol A through adhesives was estimated at 0.02 μg/kg bw/day.

The European Commission, 2002) reviewed the report by the European Union (2003) in draft and suggested alternate exposure estimates. Those estimates and the assumptions used to support those estimates are summarized in Table 11.

Table 11. Bisphenol A Exposure Estimates by the European Commissiona
Age (body weight)Type of food and amount consumedConcentration of bisphenol A in food (μg/kg)Exposure estimate (μg/kg bw/day)
  • a

    aEuropean Commission (2002).

0–4-month old infant (4.5 kg)0.7 L of formula/day101.6
6–12-month old infant (8.8 kg)0.7 L of formula/day100.8
6–12-month old infant (8.8 kg)0.38 kg canned food/day200.85
4–6-year-old child (18 kg)1.05 kg canned food/day201.2
Adult (60 kg)1.05 kg canned food/day200.37
Adult (60 kg)0.75 L wine/day90.11

Miyamoto and Kotake (2006) estimated aggregate oral and inhalation exposure to bisphenol A in Japanese male children and adults. The estimates were based on unpublished Japanese data. This report is the only known study investigating potential exposure to children through mouthing of toys. Mouthing times were estimated by surveying the mothers of 50 infants and recording 25 infants on video camera. Mean±SD daily mouthing times were reported at 41.7±13.7 min for infants 0–5 months of age and 73.9±32.9 min for infants 6–11 months of age. Migration rates were estimated from 0 μg/cm2/min for toys that do not contain bisphenol A to 0.0162 μg/cm2/min, the highest value reported in the Japanese literature. It was assumed that most toys were not manufactured with polycarbonate, epoxy resins, or grades of PVC that contain bisphenol A. Surface area of toys was assumed to be 10 cm2. In estimating oral exposures to bisphenol A, intake from food was also considered. Bisphenol A concentrations measured in migration testing of polycarbonate bottles and food surveys are summarized in Section 1.2.3.2. Volume of food consumption and frequency of article use were considered in estimates of bisphenol intake through food. Bisphenol A concentrations in drinking water were considered to be 0–0.17 μg/L, and water intake was assumed to be 2 L/day. In estimating inhalation exposures, concentrations of bisphenol A were considered to range from 0–8.1 ng/m3 in indoor air and 0–28 ng/m3 in outdoor air. Time spent indoors and outdoors and breathing rates were considered. Absorption from lungs was assumed at 100%. Estimated exposures from mouthing of toys, food and water intake, and inhaled air are summarized in Table 12.

Table 12. Average Estimated Exposure to Bisphenol A in Japanese Man Adults and Childrena
Exposure sourceBisphenol A concentration (other assumptions)Average estimated exposures in each age groupb (μg/kg bw/day)
  0–5 months6–11 months1–6 years7–14 years15–19 years19 years
  • a

    aMiyamoto and Kotake (2006).

  • b

    bAssumptions for bodyweights and most media intake levels were not provided.

Human milkNegligible00    
Formula (water)0–0.17 μg/L0.0120.0096    
Feeding bottle0–3.9 μg/L0.0150.014    
Infant food0–5.0 μg/kg 0.085    
Toys0–0.0162 μg/cm2/min (mean mouthing times of 41.7 min in 0–5-month-olds and 73.9 min in 6–11-month-olds)0.0260.069    
Air0–8.1 ng/m3 in indoor air and 0–28 ng/m3 in outdoor air (90% indoors/10% outdoors)0.00260.00240.00210.00170.00150.0015
Water0–0.17 μg/L (intake of 2 L/day)  0.0120.00530.00290.0027
Food and drink       
 Canned0–602 μg/kg  0.380.210.200.29
 Non-canned0–3 μg/kg  0.380.210.130.12
Tableware0–39.4 μg/meal/utensil (3 meals/day; 1–5 types of utensils used/meal)  0.400.120.0240.022
Total breast-fed: 0.028 formula-fed: 0.055breast-fed: 0.16 formula-fed: 0.181.20.550.360.43

Additional estimates of bisphenol A exposure through food are summarized in Table 5 and Table 6. Details of studies conducted by Earls et al. (2000) and Onn Wong et al. (2005) are presented in Section 1.2.3.2. Exposure estimates conducted by the FDA are described below. Limited details were available from the other studies that were presented in reviews.

The FDA (1996) estimated bisphenol A intake in infants and adults resulting from exposures to epoxy food-can linings and polycarbonate plastics. Exposure estimates occurring through contact of formula with polycarbonate bottles were based on results of a study conducted by the Chemistry Methods Branch of the FDA. The Chemistry Methods Branch also measured concentrations of bisphenol A in 5 brands of infant formula (14 samples total); the study is also published as Biles et al. (1997a). In estimating adult bisphenol A exposure through the consumption of canned foods, the FDA considered surveys conducted by the Chemistry Methods Branch, Brotons et al. (1995), and the Society of Plastics Industry Group. It appears that the study by the Society of Plastics Industry Group was later published by Howe et al. (1998) and included a re-analysis to correct some interferences observed in analytical methods. Exposure estimates and assumptions used to make the estimates are summarized in Table 13.

Table 13. Summaries of Studies Estimating Bisphenol A Exposures Solely from Foods
PopulationExposure sourceBasis and assumptions for estimatesExposure estimate μg/kg bw/dayReference
  • a

    aThe study authors acknowledged the use of aggressive migration testing conditions and conservative assumptions in calculations, thus leading to overestimated infant exposures.

InfantsPolycarbonate bottlesBisphenol A migration concentration of 15–20 μg/L; milk consumption of up to 550 mL/day; mean body weight of 11 kg0.75–1Earls et al. (2000)
Infants (0–3 months old)Polycarbonate bottlesMean upper-bound concentration of bisphenol A migration in 10% ethanol (0.64 μg/in2) and in corn oil (0.43 μg/in2); body weights reported by National Center for Health Statistics, and FDA Dietary Exposure Guidelines with modifications for properties of infant formula15–24aOnn Wong et al. (2005)
Not reportedFood from epoxy-lined cansBisphenol A concentrations of 5 ppb [μg/L] in beverages and 37 ppb [μg/kg] in other foods; FDA Dietary Exposure Guidelines: dietary intake of 3 kg/day, body weight of 60 kg0.105Howe et al. (1998) Haighton et al. (2002) NAS (1999)
AdultsCumulative exposures from food contacting cans and polycarbonate plastics22 ppb [μg/kg] bisphenol A in vegetables, consumption factor of 0.17 for food contacting polymer-coated metal, intake of 3 kg food/bw/day, 60 kg bw, and insignificant contribution from polycarbonate0.183FDA (1996)
InfantsCumulative exposures from food contacting cans and polycarbonate plasticsBisphenol A concentration of 6.6 μg/kg in prepared infant formula, <1.7 ppb [μg/L] in infant formula from polycarbonate bottles, consumption of 820 g food/day, and 4 kg infant weight1.75 
AdultsCanned foodsData from survey of canned foods and food intake patterns determined from surveysMean=0.0083 (0–0.29)Thomson and Grounds (2005)
AdultsCanned foods and canned fishData from survey of canned foods and food intake patterns determined from surveys0.0044 for men ≥25 0.0041 for women ≥25 0.0048 for men 19–24Thomson et al. (2003)
AdultsWineMaximum bisphenol A concentration of 2.1 ng/mL in wine, consumption of 0.75 L/day, and 60 kg body weight<0.026Brenn-Struckhofova and Cichna-Markel (2006)
Hospital patientsMeals served at 2 hospitalsMean intake from hospital diets was estimated at 1.3 (0.19–3.7) μg/day; [60 kg body weight was assumed][0.02 (0.003–0.06)]Miyamoto and Kotake (2006) Fujimaki et al. (2004)
Japanese adults and children∼200 food items were collected in a total diet studyNo details0.00475 for children 2–6 years 0.00195 for adultsMiyamoto and Kotake (2006)

Table 14 summarizes exposure estimates for aggregate or food exposures. Studies suggest that the majority of bisphenol A exposure occurs through food and that environmental exposures do not appear to substantially affect total exposure, with the possible exception of exposure near point sources. Table 14 includes estimates that CERHR believes to represent potentially realistic exposure scenarios and does not include data from extreme worst-case scenarios such as possible point–source exposures.

Table 14. Summary of Food or Aggregate Exposures to Bisphenol A
PopulationBasis of estimatesExposure estimate μg/kg bw/dayaReference
  • a

    aEstimates involving extreme worst case scenarios and Japanese data with very limited information were not included in this table.

1–2-month-old infantFood exposure (data from migration studies of polycarbonate bottles)8European Union (2003)
0–4-month-old infantFood exposure (data from migration studies of polycarbonate bottles)1.6European Commission (2002)
0–5-month-old infant (formula-fed)Aggregate exposure (based on formula, environmental, and toy exposures)0.055Miyamoto and Kotake (2006)
0–5-month-old infant (breast fed)Aggregate exposure (based on human milk, environmental, and toy exposures)0.028Miyamoto and Kotake (2006)
4–6-month-old infantFood exposure (data from migration studies of polycarbonate bottles)7European Union (2003)
6–11-month-old infant (formula-fed)Aggregate exposure (based on formula, food, environmental, and toy exposures)0.18Miyamoto and Kotake (2006)
6–11-month-old infant (breast-fed)Aggregate exposure (based on human milk, food, environmental, and toy exposures)0.16Miyamoto and Kotake (2006)
6–12-month-old infantFood exposure (data from survey of canned foods)5European Union (2003)
6–12-month-old infantFood exposure (data from migration studies with infant bottles and canned foods)1.65European Commission (2002)
InfantFood exposure (data from polycarbonate bottle leaching studies)0.75–1Earls (2000)
InfantFood exposures (contact with cans and polycarbonate plastics)1.75FDA (1996)
1.5–4.5-year-old childFood exposure (data from survey of canned foods and migration studies with polycarbonate tableware)14.7European Union (2003)
1–6-year-old childAggregate exposure (based on food, environmental, and tableware exposures)1.2Miyamoto and Kotake (2006)
1.5–5 year old childAggregate exposure (surveys of bisphenol in food, air, dust, soil and hand and surface wipes)0.06-0.07 (0.03–1.57)Wilson et al. (2006)
3–5-year-old childAggregate exposure (surveys of bisphenol in food, air, dust, and soil)0.04 (0.018–0.07)Wilson et al. (2003)
2–6-year-old childFood exposure (collection of 200 food items)0.004Miyamoto and Kotake (2006)
4–6-year-old childFood exposure (data from survey of canned foods)1.2European Commission (2002)
7–14-year-old childAggregate exposure (based on food, environmental, and tableware exposures)0.55Miyamoto and Kotake (2006)
15–19-year-old individualAggregate exposure (based on food, environmental, and tableware exposures)0.36Miyamoto and Kotake (2006)
Adult, ≥19 yearsAggregate exposure (based on food, environmental, and tableware exposures)0.43Miyamoto and Kotake (2006)
AdultFood exposure (data from survey of canned foods not including wine)1.4European Union (2003)
AdultFood exposure (data from surveys of canned food)0.37European Commission (2002)
AdultWine exposure (data from study of epoxy-lined wine drums)0.11European Commission (2002)
AdultWine exposure (data from wine samples)<0.026Brenn-Struckhofova and Cichna-Markel (2006)
AdultFood exposure (from contact with epoxy-lined cans and polycarbonate)0.183FDA (1996)
AdultsFood exposure (survey of canned foods)0.008Thomson and Grounds (2005)
AdultFood exposure (collection of 200 food items)0.002Miyamoto and Kotake (2006)
1.2.4.1.2 Estimates based on biological monitoring:

Goodman et al. (2006) noted that total urinary bisphenol A concentrations were useful for estimating bisphenol A intake. Because of extensive first-pass metabolism, little parent compound is systemically circulated, as discussed in more detail in Section 2. Because nearly 100% of an acute exposure to bisphenol A is excreted in urine within 24 hr (Völkel et al., 2002; Tsukioka et al., 2004), bisphenol A intake can be estimated by measuring bisphenol A in urine over a specified time interval. Arakawa et al. (2004) measured bisphenol A excretion over a 5-day period and reported intra- and inter-individual variability. As a result, caution was urged in using single time-point values to estimate long-term exposure. Typical daily intakes of bisphenol A estimated from urinary levels are <0.01–2.17ìg/kg bw/day (Table 15). A Monte Carlo simulation using the urine data of Tsukioka et al. (2004) and Arakawa et al. (2004) estimated mean exposures of 0.028–0.049 ug/kg bw/day for males and 0.034–0.059 ug/kg bw/day for females (Miyamoto and Kotake, 2006). Using the U.S. NHANES data and assumptions on excretion rates and body weight a median intake of 0.026 ug/kg bw/day is estimated. An estimated median exposure based on urinary bisphenol A concentrations in 6–8-year-old girls was 0.07 μg/kg bw/day (Wolff et al., 2006).

Table 15. Estimates of Bisphenol A Intakes Based on Urinary Excretion
PopulationBasis for estimatesMean or median (range) of estimated intake μg/kg bw/dayaReference
  • a

    aConsistent with estimates conducted by Goodman et al. (2006), body weights of 60 kg were assumed, unless otherwise indicated.

  • b

    bA 50-kg body weight was assumed.

22 Japanese adultsMean excretion of 1.68 μg/day (0.48–4.5 μg/day)0.028 (0.008–0.075)Tsukioka (2004)
36 Japanese male studentsMedian excretion of 1.2 μg/day (<0.21–14 μg/day)0.02 (<0.0035–0.23)Arakawa et al. (2004)
5 Japanese malesMedian excretion of 1.3 μg/day (<0.58–13 μg/day) over a 5-day period0.022 (<0.01–0.22)Arakawa et al. (2004)
Data from Tsukioka (2004) and Arakawa et al. (2004)Monte Carlo simulationsMean exposure: 0.028–0.049 in men and 0.034–0.059 in women; low exposures (5th percentile) 0.021–0.037 in men and 0.025–0.044 in women; high exposures (95th percentile): 0.037–0.064 in men and 0.043–0.075 in womenMiyamoto and Kotake (2006)
56 pregnant Japanese femalesBisphenol A concentration in one spot sample was normalized to creatinine and exposure was estimated using average creatinine and urine volume excretion rates, which resulted in a median intake of <2 μg/day (<0.3–7.9 μg/day)<0.04 (<0.006–0.16)bFujimaki et al. (2004)
48 Japanese female college studentsAuthors estimated bisphenol A intake of 0.6–71.4 μg/day, based on a median bisphenol A concentration of 0.77 ng/mg (0.1–11.9 ng/mg) creatinine in a spot urine sample, assumed creatinine excretion of 1200 mg/day and that 20% of the dose is excreted in urine. [CERHR recalculated values using a 100% urinary excretion rate that is consistent with human data]0.01–1.2 based on study author assumptions [0.015 (0.0020.24) based on a 100% urinary excretion rate]Ouchi and Watanabe (2002)
7 males and 12 females without intentional exposureAll measurements <LOD of 1.14 μg/LBased on 2 L urine excreted and 60 kg adult exposure <0.038Völkel et al. (2005)
394 participants in the NHANES III survey (U.S.)Median (10th–95th percentile) 1.32 (0.23–7.95) μg bisphenol A/g creatinine in a spot urine sample; [assumed 100% urinary excretion of bisphenol A in 24 hr and creatinine excretion of 1200 mg/day][median=0.026; 10th–95th percentile: 0.005-0.159]Calafat et al. (2005)
90 girls, 6–8-years-old (U.S.)Median (range) 1.8 μg/L (<0.3–54.3) [assumed 100% urinary excretion of bisphenol A in 24 hr; 1 L per day; 25 kg body weight ][0.07 (<0.0122.17)]Wolff et al. (2006)

Joskow et al. (2006) used values for total bisphenol A in urine to estimate exposure to bisphenol A following dental sealant application. Urinary concentrations of bisphenol A are reported in Table 8. Factors or assumptions used in the exposure estimates were recovery of bisphenol A in urine as its glucuronide conjugate within 24–34 hr following exposure, a 5.4-hr half-life of elimination for bisphenol A glucuronide, and a 1.5 L/day urinary excretion volume. Estimated doses of bisphenol A [based on a 60-kg bw] were 49–239 μg [0.82–4.0μg/kg bw] following application of Delton LC and 0–9.5 μg [0–0.16μg/kg bw] following application of Helioseal F. The study authors stated that the estimates were likely low because a substantial amount of bisphenol A was potentially eliminated by collection of saliva samples immediately following treatment.

1.2.4.2 Occupational exposure:

Occupational exposure to bisphenol A could potentially occur during its manufacture, in the production of polycarbonate plastics, and during the manufacture or use of epoxy resins, powder coatings paints, or lacquers (European-Union, 2003). Possible exposure to bisphenol A during PVC manufacture has been considered, but the European Union (2003) stated that the application was being phased out. According to the European Union, bisphenol A is generally available as granules, flakes, or pellets, thus reducing exposure potential. Bisphenol A is manufactured in closed systems, but exposure is possible during sampling, container filling, and plant maintenance. In the manufacture of polycarbonate, bisphenol A enters the plant and remains in a closed system before extrusion. Sampling is conducted by a closed loop system. Following extrusion, the polycarbonate is chopped into granules and bagged, and it is during that stage that exposure to residual bisphenol A (reported at ≤100 ppm) through dust is possible. However, it is noted that polycarbonate is stable and that residual bisphenol A is contained within the polymer matrix. The European Union stated that exposure to bisphenol A during the manufacture of polycarbonate items is not likely to exceed values observed during the manufacture of polycarbonate. In the production of epoxy resin, bisphenol A exposure is most likely during reactor charging, but exposure during maintenance is also possible. A residual bisphenol A concentration of 300 ppm was reported for epoxy resins, but it was noted that most bisphenol A was trapped within the resin matrix. Exposure to bisphenol A during production of epoxy paints is reported to be negligible. In the manufacture of powder epoxy coatings, exposure is thought possible during weighing and milling. Exposure to bisphenol A during the use of powder paints has been documented.

There are no known regulatory limits for occupational exposure to bisphenol A in the U.S. In 2004, the American Industrial Hygiene Association proposed a workplace environmental exposure level (WEEL) of 5 mg/m3 for bisphenol A. The draft WEEL was based on irritation observed in an inhalation toxicity study (American Industrial Hygiene Association, 2004). The value is consistent with the time weighted average (TWA) exposure limits established in Germany and the Netherlands (European-Union, 2003).

The European Union (2003) summarized occupational exposure data for bisphenol A in Europe and the U.S. Only measured data for bisphenol A are summarized in this report. The European Union stated that the values reported did not account for the effects of personal protective equipment in order to avoid difficulties in attempting to quantify protection provided. TWA bisphenol A concentrations measured in occupational settings are summarized in Table 16. The limited number of values reported indicated that bisphenol A concentrations were below 5 mg/m3. Bisphenol A exposures (>1 mg/m3) were observed in spraying of powdered bisphenol A-containing coatings, bisphenol A manufacture and manufacture of epoxy resins. The highest daily average exposures were observed in the manufacture of bisphenol A. There is limited information on short-term exposure to bisphenol A. In manufacture of bisphenol A one facility reported short-term task exposures from 0.13–9.5 mg/m3 (European-Union, 2003).

Table 16. TWA Measurements of Bisphenol A in the Workplace
Industry or activityLocation/yearNo. of samplesSample type8-hr TWA (mg/m3) mean (range)b
  • a

    aNIOSH (1979). Other data are from the European Union (2003).

  • b

    bRange given representing different occupational activities.

  • NS, not specified.

Bisphenol A manufacture
 VariousU.S./not specifiedNot specifiedBisphenol ANS (Not detected to 2.6)
 Filling big bagsEurope/19983Inhalable bisphenol A0.81 (0.21-1.79)
 Filling silo tankersEurope/19983Inhalable bisphenol A0.89 (<0.5-1.61)
 Various tasksEurope/19988Inhalable bisphenol A0.3 (0.13-0.62)
 Plant operatorEurope/not specified7Inhalable bisphenol ANS (0.021-1.04)
 MaintenanceEurope/not specified3Inhalable bisphenol ANS (0.52-1.35)
 MaintenanceEurope/1998–20008Bisphenol ANS (<0.05–0.62)
 Charging big bagsEurope/1996–19975Inhalable bisphenol A0.35 (0.02–0.93)
 Plant operatorEurope/not specified13Bisphenol A0.61 (0.02–2.13)
 Maintenance operatorEurope/not specified2Bisphenol A1.06 (0.4–2.08)
Epoxy resin manufacture
 Loading/unloadingU.S./1970-mid 1990s26Bisphenol A0.18 (<0.1-0.99)
 Bagging/palletizingU.S./1970-mid 1990s37Bisphenol A0.25 (<0.1-2.8)
 Process operatorsU.S./1970-mid 1990s25Bisphenol A0.26 (<0.1-1.1)
 Equipment  technicianU.S./1970-mid 1990s6Bisphenol A<0.1
 MaintenanceU.S./1970-mid 1990s2Bisphenol A0.8 (0.37-1.2)
Bisphenol A Use
 Powder paint useaU.S./∼19797 (3 personal and 4 area samples)Bisphenol A (plant 1)0.005 (0.004–0.006)
 21 (15 personal and 6 area samples)Bisphenol A (plant 2)0.175 (0.001–1.063)

Data for powder paint use summarized in Table 16 were obtained from a NIOSH Health Hazard Evaluation conducted at a company that manufactured fan and ventilation equipment (NIOSH, 1979). In Plant 1 of the company, parts were coated with an epoxy-based powder paint by dipping. At Plant 2, an epoxy-based powder was applied to parts via electrostatic spraying. As evident in the data in Table 16, exposures were higher at the plant utilizing electrostatic spraying. Monitoring for bisphenol A was discussed in 2 other NIOSH Health Hazard Evaluation reports. In those reports, bisphenol A was not detected in a plant where an epoxy resin coating was used in the manufacture of electronic resistors (NIOSH, 1984) or in a plant where an epoxy resin coating was applied to steam turbine generators (NIOSH, 1985). Rudel et al. (2001) used a GC/MS technique to measure bisphenol A concentrations at one United States workplace where plastics were melted and glued; a concentration of 0.208 μg/m3 was reported.

[Bisphenol A exposures in U.S. powder paint workers were estimated at ∼0.1–100μg/kg bw/day based on TWA exposures of 0.001–1.063 mg/m3, an inhalation factor of 0.29 m3/kg day (USEPA,1988), 100% absorption from the respiratory system, and 8 hr worked per day.]

No information was located for dermal exposure to bisphenol A in occupational settings. Using their Estimation and Assessment of Substance Exposure model, the European Union (2003) estimated that dermal exposure of workers to bisphenol A was unlikely to exceed 5 mg/cm2/day. It was noted that the highest potential exposure to bisphenol A would occur during bag filling and maintenance work.

One study provided information on biological monitoring of bisphenol A in workers exposed to an epoxy compound. In 3 Japanese plants, exposed workers included 42 men who sprayed an epoxy hardening agent consisting of a mixture of bisphenol A diglycidyl ether (10–30%), toluene (0–30%), xylene (0–20%), 2-ethoxyethanol (0–20%), 2-butoxyethanol (0–20%), and methyl isobutyl ketone (0–30%) (Hanaoka et al., 2002). The workers wore “protection devices” during spraying. Controls consisted of 42 male assembly workers from the same plants who did not use bisphenol A diglycidyl ether. In 1999, urine samples were collected periodically, treated with β-glucuronidase, and examined for bisphenol A by HPLC. Urinary bisphenol A concentrations were significantly higher in exposed workers (median: 1.06 μmol/mol creatinine [2.14μg/g creatinine]; range: <0.05 pmol to 11.2 μmol/mol creatinine [<0.1 pg to 22.6μg/g creatinine]) compared to controls (median: 0.52 μmol/mol creatinine [1.05μg/g creatinine]; range: <0.05 pmol to 11.0 μmol/mol creatinine [<0.1 pg to 22.2μg/g creatinine]). The difference of the averages was reported as 2.5 μmol/mol creatinine [5.05μg/g creatinine] (95% confidence interval [CI] 1.4–4.7 μmol/mol creatinine [2.8–9.5]). Bisphenol A was not detected in three exposed workers and one control. [Assuming excretion of 1200 mg/day creatinine (Ouchi and Watanabe,2002), mean (ranges) of bisphenol excretion in urine were 2.57μg/day (<0.12 pg to 27.1μg/day) in exposed workers and 1.26μg/day (<0.12 pg to 26.6μg/day) in unexposed workers. With an assumed body weight of 60 kg, bisphenol A occupational intake was estimated at 0.043μg/kg bw/day (<0.002 pg to 0.45μg/kg bw/day) in exposed workers and 0.021μg/kg bw/day (<0.002 pg to 0.44μg/kg bw/day) in unexposed workers.]

1.3 Utility of Data

Numerous studies reported bisphenol A concentrations in canned foods and infant formula. Experiments examined potential concentrations of bisphenol A resulting from leaching of bisphenol A from polycarbonate bottles under a variety of conditions. There minimal data available for bisphenol A concentrations in drinking water but these show concentrations below the limit of detection. Bisphenol A has been detected in surface waters and solid waste landfill leachates. Bisphenol A has been detected in indoor dust samples and indoor and outdoor air samples. Data for occupational exposure to bisphenol A in the U.S. are very limited. Only 2 studies reported TWA exposures to bisphenol A in U.S. workers. Several estimates of human bisphenol A exposure were developed using bisphenol A concentrations measured in food and the environment. Although very limited for U.S. populations, there are data reporting bisphenol A concentrations in urine, breast milk, and amniotic fluid, but none for blood or fetal blood. Exposure estimates have been derived from urinary bisphenol A concentrations in multiple studies.

1.4 Summary of Human Exposure

In 1999 and 2003, it was reported that most bisphenol A produced in the U.S. was used in the manufacture of polycarbonate and epoxy resins and other products [reviewed in (Staples et al., 1998; SRI, 2004)]. Polycarbonate plastics are used in various consumer products and the products most likely to contribute to human exposure are polycarbonate food containers (e.g., milk, water, and infant bottles). Epoxy resins are used in protective coatings. Food cans lined with epoxy resin are a potential source of human exposure. Some polymers manufactured with bisphenol A are FDA-approved for use in direct and indirect food additives and in dental materials (FDA, 2006). Resins, polycarbonate plastics, and other products manufactured from bisphenol A can contain trace amounts of residual monomer and additional monomer may be generated during breakdown of the polymer (European-Union, 2003).

Bisphenol A may be present in the environment as a result of direct releases from manufacturing or processing facilities, fugitive emissions during processing and handling, or release of unreacted monomer from products (European-Union, 2003). Because of its low volatility and relatively short half-life in the atmosphere, bisphenol A is unlikely to be present in the atmosphere in high concentrations (European-Union, 2003). A study of 222 homes and 29 day care centers found bisphenol A in 31–44% of outdoor air samples with concentrations of <LOD (0.9) to 51.5 ng/m3 (Wilson et al., 2006). Rapid biodegradation of bisphenol A in water was reported in the majority of studies reviewed by the European Union (2003) and Staples et al. (1998). Drinking water concentrations of bisphenol A at Louisiana and Detroit Michigan water treatment plants were below the limit of detection (<0.1 ng/L). Chlorinated congeners of bisphenol A resulting from chlorination of water may be degraded less rapidly (Gallard et al., 2004). Bisphenol A is not expected to be stable, mobile, or bioavailable from soils (Fent et al., 2003). A study of 222 homes and 29 day care centers found bisphenol A in 25–70% of indoor dust samples with concentrations of <LOD (20) to 707 ng/g (Wilson et al., 2006). The potential for bioconcentration of bisphenol A in fish is low (Staples et al., 1998; European-Union, 2003). Table 17 summarizes concentrations of bisphenol A detected in environmental samples and drinking water.

Table 17. Maximum Reported Bisphenol A Concentrations in U.S. Ambient Air and Dust Samples
SampleBisphenol A concentrationReference
Outdoor air<52 ng/m3 Monthly average 0.12–1.2 ng/m3Wilson et al. (2003, 2006); Matsumoto et al. (2005)
Indoor air≤193 ng/m3Wilson et al. (2003, 2006); Rudel et al. (2001, 2003)
Indoor dust≤17.6 μg/gWilson et al. (2003, 2006); Rudel et al. (2001, 2003)
Drinking water<0.1 (MDL) <0.005Boyd et al. (2003); Kuch and Ballschmiter (2001)

The highest potential for human exposure to bisphenol A is through products that directly contact food such as food and beverage containers with internal epoxy resin coatings and polycarbonate tableware and bottles, such as those used to feed infants (European-Union, 2003). Studies examining the extraction of bisphenol A from polycarbonate bottles or tableware into food simulants are summarized in Table 4. Studies measuring bisphenol A concentrations in canned infant foods are summarized in Table 5 and studies measuring bisphenol A concentrations in canned food are summarized in Table 6. Table 18 summarizes the general findings from all the food contact–material studies. Bisphenol A concentrations were measured in canned foods produced and purchased from various countries.

Table 18. Maximum Reported Bisphenol A Concentrations Measured in Foods or Food Simulants
Exposure sourceBisphenol A concentrationTable reference
Polycarbonate infant bottles≤55 μg/L (<5 μg/L in U.S. study)Table 4
Polycarbonate tableware≤5 μg/kgTable 4
Canned infant formulas≤113 μg/L (<6.6 μg μg/kg in U.S. study of water mixed formula; <13 μg/kg in U.S. formula concentrate)Table 5
Canned infant foods≤77.3 μg/kg 
Canned foods≤842 μg/kg (≤39 μg/kg in U.S. studies)Table 6

Table 19 summarizes BPA concentrations reported in human body fluids. Measurement of bisphenol A concentrations are affected by measurement technique, particularly at the very low concentrations that can now be measured. ELISA has poor correlation with the LC-ECD method and also the different ELISA kits correlate poorly with each other. ELISA methods may overestimate bisphenol A in biologic samples due to lack of specificity of the antibody and effects of the biologic matrix (Inoue et al., 2002; Fukata et al., 2006). In addition, contamination from labware and reagents or sample degradation during storage can impact the accuracy of measurements. [The panel therefore finds the greatest utility in studies that use sensitive and specific analytical methods for biological samples (LC-MS or GC-MS) and report quality control measures for sample handling and analysis.]

Table 19. Maximum Reported Biological Measures of Bisphenol A Concentrations in Humans
Biological mediumPopulationConcentration free BPAa (μg/L)Total BPAa (μg/L)Reference
  • a

    aMeasurements by HPLC, GC/MS, and LC/MS only.

UrineAdult≤2.36 (<0.6 in U.S. study)≤3950 (<19.8 U.S. studies)Table 8
 Children <54 (2 U.S. studies) 
BloodGeneral<LOD (0.5)<LOD (0.5)Table 7
 Infertile women<0.87 Table 7
 Women<1.6 Table 7
 Men<1 Table 7
 Fetal<9.2 Table 9
Breast milkWomen<6.3 (U.S.)<7.3Table 3
Amniotic fluidFetus<1.96 (U.S) Table 9
SemenAdult<0.5 Inoue et al.(2002)
Saliva after dental sealantAdult<2800 Arenholt-Bindslev et al. (1999)

Table 20 summarizes food and/or aggregate exposure estimates calculated from bisphenol A concentrations in food, environmental and toy exposures along with estimates of consumption and body weights. It was noted that dietary sources account for 99% of exposure (Wilson et al., 2006). Metabolite-based estimates of bisphenol A used urinary concentrations along with estimates of urinary and/or creatinine excretion, and body weight.

Table 20. Summary of Reported Human Dose Estimates
Exposure sourcePopulationBPA μg/kg bw/dayNotesSource
Estimates based on intake
 FormulaInfant1.6–88 assumes 700 ml formula with 50 μg/LTable 14
 FormulaInfant1.0Assumes 4.5 kg, 700 ml formula with 6.6 μg/L from U.S. canned formulaExpert Panel
 Breast milkInfant1.0Assumes 4.5 kg, 700 ml with 6.3 μg/L from breast milkExpert Panel
 FoodInfant1.65–55 assumes 0.375 kg canned food at 100 μg/kgTable 14
 Child0.00475–1.21.2 assumes 1 kg canned food at 20 μg/kgTable 14
 Adult0.00195–1.41.4 assumes 1 kg canned food at 100 μg/kgTable 14
 AggregateInfant (formula)0.055–0.18Assumes 0–0.17 μg/L in formulaTable 14
 Infant (breast milk)0.028–0.16Assumes 0 exposure from breast milkTable 14
 Child0.042981–14.714.7 assumes 2 kg canned food at 100 μg/kgTable 14
 Adult0.36–0.43Assumes 0–602 μg/kg in canned foodTable 14
 OccupationalAdult0.043–100 EPA and Expert Panel
Estimates based on urinary metabolites
 AggregateChild0.07 (2.17)Median (max) U.S. 6–8-year-old girlsTable 15
 Adult0.026Median NHANESTable 15
 Adult0.66Assume max 19.8 μg/L from U.S., 2 L urine/day, 60 kgYe et al. (2005)

Dental sealant exposure to bisphenol A occurs primarily with use of dental sealants bisphenol A dimethylacylate. This exposure is considered an acute and infrequent event with little relevance to estimating general population exposures.

Very limited information is available for bisphenol A exposure in the U.S. workplace. Data obtained from the U.S. and Europe indicate highest potential exposures during spraying of powdered bisphenol A-containing coatings and during tank filling, plant operation activities, and maintenance work in plants where bisphenol A is manufactured. (European-Union, 2003). One study measured total urinary bisphenol A in Japanese workers who sprayed an epoxy compound (Hanaoka et al., 2002).

2.0 GENERAL TOXICOLOGY AND BIOLOGICAL EFFECTS

As discussed in Section 1.4, the quantified amount of free bisphenol A present in biological samples may be affected by contamination with bisphenol A in plastic laboratory ware and in reagents (Tsukioka et al., 2004; Völkel et al., 2005). In addition, the accuracy may also be affected by measurement technique, particularly at the very low concentrations that can now be measured. ELISA have the potential to overestimate bisphenol A in biologic samples due to lack of specificity of the antibody and effects of the biologic matrix (Inoue et al., 2002; Fukata et al., 2006). High performance liquid chromatography (HPLC) with ultraviolet, fluorescence, or electrochemical detection is unable to make definitive identification of bisphenol A or bisphenol A glucuronides, because similar retention times may occur for the metabolites of other endogenous and exogenous compounds (Völkel et al., 2005). Use of LC-tandem mass spectrometry (MS/MS) with and without hydrolysis of bisphenol A glucuronide permits determination of free and total bisphenol A with a limit of quantification of 1 μg/L (Völkel et al., 2005). Gas chromatography (GC)/MS/MS has been used with solid phase extraction after treatment with glucuronidase and derivatization to measure total bisphenol A with a limit of detection of 0.1 μg/L (Calafat et al., 2005). Bisphenol A glucuronidate has been shown to be unstable and can be hydrolyzed to free bisphenol A at neutral pH and room temperature in diluted urine of rats and in rat placental and fetal tissue homogenates at room temperature. Bisphenol A glucuronide can also be hydrolyzed and in some cases degraded to unknown components either in acidic or basic pH solutions of diluted urine, adding another potential source of error in the measurement of sample levels of bisphenol A and its conjugates (Waechter et al., 2007). These considerations taken together, suggest that it is possible that free bisphenol A concentrations measured in biological samples may be overestimated.

2.1 Toxicokinetics and Metabolism

The studies presented in this section demonstrate that bisphenol A is absorbed in humans and experimental animals following oral exposure. In humans and experimental animals, most of the dose is present in blood as the main metabolite, bisphenol A glucuronide, and smaller percentages are present as the parent compound. Bisphenol A and its metabolites are widely distributed in humans and animals. More than 90% of unmetabolized bisphenol A is reportedly bound to plasma protein. Bisphenol A is distributed to fetal fluids in humans and experimental animals, and a limited number of studies in humans demonstrate fetal concentrations of bisphenol A within an order of magnitude of concentrations in maternal blood. None of the studies detected bisphenol A glucuronide in fetal fluids. Transfer of bisphenol A to milk was demonstrated in humans and experimental animals. One study in humans reported bisphenol A in milk at concentrations exceeding maternal blood concentrations. In humans and experimental animals, most of a bisphenol A dose is metabolized to bisphenol A glucuronide before absorption. Studies in humans and experimental animals demonstrated that glucuronidation of bisphenol A can occur in the liver, and one study in rats demonstrated that bisphenol A is glucuronidated upon passage through the intestine. Bisphenol A glucuronide is excreted in the bile of rats, and enterohepatic cycling is thought to occur in rats but not humans. In humans, most of a bisphenol A dose is eliminated through urine as bisphenol A glucuronide. In rats, bisphenol A is eliminated through feces as bisphenol A and in urine as bisphenol A glucuronide.

2.1.1 Humans.

Human toxicokinetics studies that were judged potentially important to interpret developmental and reproductive toxicity were reviewed in full. These studies include reports of potential exposure of fetuses during pregnancy or of infants through human milk and reports of toxicokinetics or metabolism following low-dose exposure of humans. Information from secondary sources was included if the information was not considered to be critical to the interpretation of developmental and reproductive toxicity data.

2.1.1.1 Absorption:

Two studies described here examined oral absorption of bisphenol A from dental sealants, and one study examined in vitro dermal absorption. Bisphenol A (as parent or the monoglucuronide) is absorbed in humans as indicated by the detection of bisphenol A (and metabolites) in blood from the general population (Section 1) and in maternal and fetal fluids (Table 9).

Fung et al. (2000) examined the toxicokinetics of bisphenol A leaching from dental sealant. Volunteers included 18 men and 22 non-pregnant women (20–55 years of age) who did not have dental disease, existing composite resin restorations or pit and fissure sealants, or a history of resin exposure. Volunteers were treated with a widely used commercial dental sealant (Delton Opaque Light-cure Pit and Fissure Sealant). Components of the sealant were analyzed by HPLC. The low-dose group (n=7 men, 11 women) received 8 mg dental sealant on 1 tooth, and the high-dose group (11 men, 11 women) received 32 mg sealant on 4 teeth. Saliva and blood samples were collected before the procedure and at 1 and 3 hr and 1, 3, and 5 days after the procedure. Blood and saliva were analyzed by HPLC. Statistical analyses of data were conducted by nonparametric test, Wilcoxon signed rank test, and χ2 test. Analysis of the dental sealant revealed that bisphenol A concentrations were below the detection limit of 5 ppb. At 1 hr following treatment, bisphenol A was detected in samples from 3 of 18 volunteers in the low-dose group and 13 of 22 samples from volunteers in the high-dose group. At 3-hr post-treatment, bisphenol A was detected in samples from 1 of 18 volunteers in the low-dose group and 7 of 22 volunteers in the high-dose group. Concentrations of bisphenol A in saliva at 1 and 3 hr following exposure were reported at 5.8–105.6 ppb [μg/L]. No bisphenol A was detected in saliva samples at 24 hr or in serum samples at any time point. Differences between the low-dose and high-dose groups in bisphenol A saliva concentrations and in the proportion of bisphenol A-positive saliva samples at 1 and 3 hr achieved statistical significance. In the high-dose group, a significant difference in “readings” was observed between 1 and 3 hr. [The data as presented did not illustrate possible quantitative differences in saliva bisphenol A concentrations from the 2 dose groups or at different sampling times.]

Joskow et al. (2006) examined bisphenol A in urine and saliva of 14 adults (19–42 years old) treated with dental sealants. Excluded from the study were individuals with resin-based materials on their teeth, smokers, users of antihistamines, and patients with Gilbert syndrome. The volunteers received either Helioseal F (n=5) or Delton LC (n=9) sealant. Sealant was weighed before and after application to determine the amount applied, and the number of treated teeth was recorded. The mean number of teeth treated was 6/person and the mean total weight of sealant applied was 40.35 mg/person. In a comparison of the 2 sealants, no differences were reported for number of teeth treated or amount of sealant applied. Saliva samples were collected before treatment, immediately after, and at 1 hr following sealant application. Urine samples were collected before treatment and at 1 and 24 hr following sealant placement. A total of 14–15 saliva samples and 12–14 urine samples were collected at each time point. Samples were treated with β-glucuronidase and analyzed for bisphenol A concentrations using selective and sensitive isotope-dilution-MS-based methods. Table 21 summarizes changes in saliva and bisphenol A concentrations. Immediately and at 1 hr after sealant application, salivary concentrations of bisphenol A compared to baseline were significantly higher in the patients who received the Delton LC sealant. Bisphenol A concentrations in saliva increased >84-fold following application of the Delton LC sealant. Urinary concentrations of bisphenol A were increased 1 hr following application of the Delton LC sealant. Concentrations of bisphenol A in saliva and urine following application of Helioseal F were reported to be similar to baseline.

Table 21. Saliva and Urinary Concentrations of Total Bisphenol A in Adults Receiving Dental Sealantsa
 Mean±SD bisphenol A concentration (ng/mL)b
Collection timeBoth sealantsDelton LCHelioseal F
  • a

    aJoskow et al. (2006).

  • b

    bSamples were treated with β-glucuronidase.

Saliva
 Pretreatment0.30±0.170.34±0.190.22±0.03
 Immediately after treatment26.5±30.742.8±28.90.54±0.45
 1 hr post-treatment5.12±10.77.86±12.730.21±0.03
Urine (creatinine-adjusted)
 Pretreatment2.41±1.242.6±1.42.12±0.93
 1 hr post-treatment20.1±33.127.3±39.17.26±13.5
 24 hr post-treatment5.14±3.967.34±3.812.06±1.04

The European Union (2003) reviewed unpublished preliminary data from a human dermal absorption study. Skin samples obtained from 3 human donors (6 samples/donor/dose) were exposed to 5 or 50 mg/cm2 (3.18 or 31.8 mg/mL) 14C-bisphenol A in ethanol vehicle. Following evaporation of the vehicle, bisphenol A was resuspended in artificial sweat. Radioactivity was measured in receptor fluid at various time intervals over a 24-hr period. Radioactivity was measured in the stratum corneum and “lower” skin layer at 24 hr. Authors of the European Union report noted that tritiated water was not used as a marker for skin integrity. However, based on the patterns of results, they concluded that skin integrity was likely lost after 4–8 hr. The European Union authors therefore concluded that the only reliable data from the study were those for the cumulative percentage of the dose in receptor fluid at 8 hr, which was reported at 0.57–1.22% at 5 mg/cm2 and 0.491–0.835% at 50 mg/cm2. Because radioactivity in skin was not measured at 8 hr, the percentage of the applied dose remaining on skin and available for future absorption could not be determined. Based on ratios of receptor fluid concentrations and lower skin levels (1:2 to 1:8) at 24 hr, and assuming that the higher ratio applies to skin at 8 hr, the authors of the European Union report predicted that 10% of the dose would be present in “lower” skin layers. Therefore, dermal absorption of bisphenol A was estimated at 10%.

2.1.1.2 Distribution:

In humans, bisphenol A was measured in cord blood and amniotic fluid, demonstrating distribution to the embryo or fetus. Studies reporting bisphenol A concentrations in fetal and/or maternal compartments are summarized in Table 9. Detailed descriptions of those studies are also presented below.

Engel et al. (2006) reported concentrations of bisphenol A in human amniotic fluid. Twenty-one samples were obtained during amniocentesis conducted before 20 weeks gestation in women who were referred to a U.S. medical center for advanced maternal age. Bisphenol A concentrations in amniotic fluid were measured using LC with electrochemical detection. Bisphenol A was detected in 10% of samples at concentrations exceeding the LOD (0.5 μg/L). Bisphenol A concentration ranges of 0.5–1.96 μg/L were reported.

Schönfelder et al. (2002b) examined bisphenol A concentrations in maternal and fetal blood and compared bisphenol A concentrations in blood of male and female fetuses. In a study conducted at a German medical center, blood samples were obtained from 37 Caucasian women between 32–41 weeks gestation. At parturition, blood was collected from the umbilical vein after expulsion of the placenta. Bisphenol A concentrations in plasma were measured by GC/MS. Control experiments were conducted to verify that bisphenol A did not leach from collection, storage, or testing equipment. Bisphenol A was detected in all samples tested, and concentrations measured in maternal and fetal blood are summarized in Table 9. Mean bisphenol A concentrations were higher in maternal (4.4±3.9 [SD] μg/L) than fetal blood (2.9±2.5 μg/L). Study authors noted that in 14 cases fetal bisphenol A plasma concentrations exceeded those detected in maternal plasma. Among those 14 cases, 12 fetuses were male. Analysis by paired t-test showed significantly higher mean bisphenol A concentrations in the blood of male than female fetuses (3.5±2.7 vs. 1.7±1.5 ng/mL, P=0.016). Bisphenol A concentrations were measured in placenta samples at 1.0–104.9 μg/kg.

Ikezuki et al. (2002) measured concentrations of bisphenol A in serum from 30 healthy premenopausal women, 37 women in early pregnancy, 37 women in late pregnancy, and 32 umbilical cord blood samples. Concentrations of bisphenol A were also measured in 32 samples of amniotic fluid obtained during weeks 15–18 of gestation, 38 samples of amniotic fluid obtained at full-term cesarean section, and 36 samples of ovarian follicular fluid collected during in vitro fertilization procedures. [It was not stated if different sample types were obtained from the same subjects.] An ELISA method was used to measure bisphenol A concentrations and results were verified by HPLC. The mean±SD concentration of bisphenol A in follicular fluid was reported at 2.4±0.8 μg/L. As summarized in Table 7 for nonpregnant women and Table 9 for maternal and fetal samples, concentrations of bisphenol A in follicular fluid were similar to those detected in the serum of fetuses and pregnant and non-pregnant women and in amniotic fluid collected in late pregnancy (∼1–2 μg/L). Bisphenol A concentrations in amniotic fluid samples collected in early pregnancy were ∼5-fold higher than in other samples, and the difference achieved statistical significance (P<0.0001). Study authors postulated that the higher concentrations of bisphenol A in amniotic fluid collected during gestation weeks 15–18 may have resulted from immature fetal liver function. They noted that according to unpublished data from their laboratory, the percentage of glucuronidated bisphenol A in mid-term amniotic fluid was ∼34%, which is much lower than reported values for other human fluids (>90%).

Yamada et al. (2002) measured bisphenol A concentrations in maternal serum and amniotic fluid from Japanese women. Samples were collected between 1989–1998 in women undergoing amniocentesis around gestation week 16. One group of samples was obtained from 200 women carrying fetuses with normal karyotypes, and a second group of samples was obtained from 48 women carrying fetuses with abnormal karyotypes. An ELISA method was used to measure bisphenol A concentrations. [As discussed inSection 1.1.5, ELISA may overestimate bisphenol A.] Concentrations of bisphenol A measured in maternal plasma and amniotic fluid are summarized in Table 9. Median concentrations of bisphenol A in maternal serum (∼2–3 μg/L) were significantly higher [∼10-fold] than concentrations in amniotic fluid (∼0–0.26 μg/L) in the groups carrying fetuses with normal and abnormal karyotypes. However, in 8 samples from women carrying fetuses with normal karyotypes, high concentrations (2.80–5.62 μg/L) of bisphenol A were measured in amniotic fluid. The study authors interpreted the data as indicating that bisphenol A does not accumulate in amniotic fluid in most cases but accumulation is possible in some individuals. Bisphenol A concentrations in maternal blood were significantly higher [by ∼33%] in woman carrying fetuses with abnormal versus normal karyotypes. However, the study authors noted that the effect may not be related to bisphenol A exposure because there was no adjustment for maternal age, and concentrations in amniotic fluid did not differ between groups. In the group carrying fetuses with normal karyotypes, data obtained from 1989–1998 were summarized by year. Median bisphenol A concentrations in serum significantly decreased over that time from a concentration of 5.62 μg/L detected in 1989 to 0.99 μg/L in 1998.

Kuroda et al. (2003) used an HPLC method to measure bisphenol A concentrations in 9 sets of maternal and cord blood samples obtained from Japanese patients at the time of delivery. Bisphenol A concentrations were also measured in 21 sets of serum and ascitic fluid samples collected from sterile Japanese patients of unspecified sexes and ages. Results for pregnant women are summarized in Table 9. Mean±SD concentrations of bisphenol A were lower in maternal (0.46±0.20 ppb [μg/L]) than cord blood (0.62±0.13 ppb [μg/L]). There was a weak positive correlation (r=0.626) between bisphenol A concentrations in maternal and cord blood. Concentrations of bisphenol A in the blood of sterile patients are summarized Table 7. There were no differences between pregnant and non-pregnant blood levels (Kuroda et al., 2003). Mean±SD concentrations of bisphenol A were higher in ascitic fluid (0.56±0.19 ppb [μg/L]) than in serum (0.46±0.20 ppb [μg/L]). The correlation between bisphenol A concentration in serum and ascitic fluid was relatively strong (r=0.785).

Tan and Mohd (2003) used a GC/MS method to measure bisphenol A concentrations in cord blood at delivery in 180 patients at a Malaysian medical center. Bisphenol A was detected in 88% of samples. As noted in Table 9 concentrations ranged from <0.10–4.05 μg/L.

Calafat et al. (2006) reported a median bisphenol A concentration of ∼1.4 μg/L [as estimated from a graph] in milk from 32 women. Bisphenol A was measured after enzymatic hydrolysis of conjugates. Ye et al. (2006) found measurable milk concentrations of bisphenol A in samples from 18 of 20 lactating women. Free bisphenol A was found in samples from 12 women. The median total bisphenol concentration in milk was 1.1 μg/L (range: undetectable to 7.3 μg/L). The median free bisphenol A concentration was 0.4 μg/L (range: undetectable to 6.3 μg/L).

Sun et al. (2004) used an HPLC method to measure bisphenol A concentrations in milk from 23 healthy lactating Japanese women. Bisphenol A concentrations ranged from 0.28–0.97 μg/L, and the mean±SD concentration was reported at 0.61±0.20 μg/L. No correlations were observed between bisphenol A and triglyceride concentrations in milk. Values from six milk samples were compared to maternal and umbilical blood samples previously reported in a study by Kuroda et al. (2003). Bisphenol A values were higher in milk, and the milk/serum ratio was reported at 1.3. Bisphenol A values in milk were comparable to those in umbilical cord serum. [It was not clear whether milk and serum samples were obtained from the same volunteers in the two studies.]

Schaefer et al. (2000) measured concentrations of bisphenol A and other compounds in uterine endometrium of women undergoing hysterectomy for uterine myoma at a German medical center. Endometrial and fat samples were obtained between 1995–1998 from 23 women (34–51 years old) with no occupational exposure. Samples were handled with plastic-free materials and stored in glass containers. Concentrations of environmental chemicals were measured in samples by GC/MS. None of 21 fat samples had detectable concentrations of bisphenol A. Bisphenol A was detected in 1 of 23 endometrial samples; the median concentration was reported at <1 μg/kg wet weight, and the range was reported at 0–13 μg/kg. [It is not known why a median value and range were reported when bisphenol A was only detected in 1 sample.]

As part of a study to compare an ELISA and an LC/MS method for biological monitoring of bisphenol A, Inoue et al. (2002) measured concentrations of bisphenol A in semen samples obtained from 41 healthy Japanese volunteers (18–38 years old). Analysis by the ELISA method indicated bisphenol A concentrations ranging from concentrations below the detection limit (2.0 μg/L) to 12.0 μg/L. The LC/MS method indicated that the bisphenol A concentration in all samples was <0.5 μg/L, the LOQ. The study authors concluded that the LC/MS method was more accurate and sensitive and that the ELISA method overestimated bisphenol A concentrations, possibly due in part to nonspecific antibody interactions.

2.1.1.3 Metabolism:

Völkel et al. (2005) measured bisphenol A and metabolite concentrations in human urine following exposure to a low bisphenol A dose. The human volunteers consisted of 3 healthy females (25–32 years old) and 3 healthy males (37–49 years old) who were asked to refrain from alcohol and medicine intake for 2 days before and during the study. Volunteers received 25 μg D16-bisphenol A in drinking water [0.00028–0.00063 mg/kg bw based on reported body weights], a dose reported to represent a worst-case human exposure. Urine samples were collected at 0, 1, 3, 5, and 7 hr following exposure. Analyses for D16-bisphenol A and D16-bisphenol A-glucuronide were conducted by LC/MS and HPLC. Recovery of D16-bisphenol A-glucuronide in urine within 5 hr of dosing was 85% of dose in males and 75% of dose in females. Analysis following treatment of urine with glucuronidase resulted in recovery of 97% of the dose in males and 84% of the dose in females. The highest concentrations of bisphenol A glucuronide in urine were measured at 1 hr (221–611 pmol [50–139 ng bisphenol A eq]/mg creatinine) and 3 hr (117–345 pmol [27–79 ng bisphenol A eq]/mg creatinine) following exposure. Elimination half-life was estimated at 4 hr. Bisphenol A concentrations exceeding the detection limit were detected in only 2 urine samples at concentrations of ∼10 pmol [2 ng]/mg creatinine.

Völkel et al. (2002) examined toxicokinetics and metabolism of bisphenol A in humans administered a low dose. Volunteers in this study consisted of 3 healthy females (24–31 years of age) and 6 healthy males (28–54 years of age) who were non- or occasional smokers; volunteers were asked to refrain from alcohol and medicine intake for 2 days before and during the study. In two different studies, D16-bisphenol A was orally administered to volunteers via gelatin capsules at a dose of 5 mg (0.054–0.090 mg/kg bw). The dose was reported to be ∼10-fold higher than the estimated human exposure level of 0.6 mg/day. In the first study, urine samples were collected at 6-hr intervals until 42 hr following exposure and blood samples were collected at 4-hr intervals until 32 hr following exposure in 3 males and 3 females. In a second, more detailed study conducted in 4 of the male volunteers, blood samples were collected at 30–60-min intervals until 381 min following exposure. Samples were analyzed by GC/MS and LC/MS. In the first study, a terminal half-life of 5.3 hr was reported for D16-bisphenol A glucuronide clearance from blood. The half-life for urinary elimination was reported at 5.4 hr. D16-Bisphenol A glucuronide concentrations in plasma and urine fell below LOD at 24–34 hr post-dosing. Complete urinary recovery (100%) was reported for the D16-bisphenol A glucuronide. In the second study, maximum plasma concentration of D16-bisphenol A glucuronide (∼800 pmol [183 ng bisphenol A eq]/mL) was obtained 80 min after oral administration. The half-life for initial decline in plasma was reported at 89 min. Free D16-bisphenol A was not detected in plasma. According to study authors, the study demonstrated rapid absorption of bisphenol A from the gastrointestinal tract, conjugation with glucuronic acid in the liver, and rapid elimination of the glucuronide in urine. Study authors noted that the rapid and complete excretion of bisphenol A glucuronide in urine suggested that in contrast to rats, enterohepatic circulation did not occur in humans.

Table 8 in Section 1 provides information on bisphenol A and metabolites detected in human urine. A study conducted in the U.S. used an HPLC method to examine 30 urine samples collected from a demographically diverse adult population in 2000–2004 (Ye et al., 2005). Mean urinary compound composition was 9.5% bisphenol A, 69.5% bisphenol A glucuronide, and 21% bisphenol A sulfate conjugate. A study conducted in Korea used an HPLC method to examine urine collected from 15 men (mean age=42.6 years) and 15 women (mean age=43.0 years) (Kim et al., 2003b). Sex-related differences were observed for urinary metabolic profiles. Mean urinary compound composition in men was reported at 29.1% bisphenol A, 66.2% bisphenol A glucuronide, and 4.78% bisphenol sulfate conjugate. The urinary metabolite profile in females was 33.4% bisphenol A, 33.1% bisphenol A glucuronide, and 33.5% bisphenol A sulfate conjugate. The study authors concluded that women had a greater ability for sulfation than men.

2.1.1.4 Excretion:

As discussed in greater detail in Section 2.1.1.3, two studies in which human volunteers were administered low doses of D16-bisphenol A (∼0.00028–0.090 mg/kg bw) demonstrated that most of the dose (85–100%) was eliminated through urine (Völkel et al., 2002, 2005). In those studies, the half-lives for urinary elimination were reported at 4–5.4 hr. As discussed in more detail in Section 2.1.1.3, examination of human urine samples revealed that bisphenol A glucuronide and sulfate conjugates are present at higher concentrations than is the parent compound (Kim et al., 2003b; Ye et al., 2005).

2.1.2 Experimental animal.

Original animal studies that were potentially important for the interpretation of developmental and reproductive toxicity were reviewed thoroughly. Examples included:

  • Studies examining toxicokinetics or metabolism in pregnant or lactating animals;

  • Studies examining toxicokinetic difference observed with different doses or exposure routes;

  • Studies looking at age-related differences in toxicokinetics or metabolism; and

  • Studies in non-rodent species such as primates.

Secondary sources were utilized for general information not considered critical to the interpretation of developmental and reproductive toxicity data.

2.1.2.1 Absorption:

In rats orally exposed to bisphenol A at doses ≤100 mg/kg bw, maximum bisphenol A concentrations (Cmax) were generally measured in plasma within 0.083–0.75 hr following exposure (Pottenger et al., 2000; Takahashi and Oishi, 2000; Yoo et al., 2001; Domoradzki et al., 2004; Negishi et al., 2004b). At doses of 1 or 10 mg/kg bw, time to maximum bisphenol A concentration (Tmax) in plasma was longer in postnatal day (PND) 21 rats (1.5–3 hr) than in PND 4 and 7 rats (0.25–0.75 hr) (Domoradzki et al., 2004). In a limited number of studies in which rats were subcutaneously (s.c.) dosed with up to 100 mg/kg bw bisphenol A, time (0.5–4 hours) to reach Cmax was longer than with oral dosing, although the findings were not always consistent (Pottenger et al., 2000; Negishi et al., 2004b). In one study, Tmax was comparable in oral and intraperitoneal (i.p.) dosing of rats (Pottenger et al., 2000). Another study reported that Cmax was attained at 0.7 hr in monkeys orally exposed to 10 or 100 mg/kg bw bisphenol A and at 0.5 hr in chimpanzees orally exposed to 10 mg/kg bw bisphenol A (Negishi et al., 2004b). In the same study, a longer Tmax (2 hr) was observed following exposure of monkeys and chimpanzees to the same doses by s.c. injection compared to oral intake. Additional details for these studies are presented below.

As discussed in greater detail in Section 2.1.2.3, bisphenol A is glucuronidated in the liver and intestine, and most of the dose is absorbed as bisphenol A glucuronide following oral exposure of rats (Domoradzki et al., 2004). In ovariectomized rats gavaged with bisphenol A, bioavailability of bisphenol A was reported at 16.4% at a 10 mg/kg bw dose and 5.6% at a 100 mg/kg bw dose (Upmeier et al., 2000). The findings are fairly consistent with a second study in which maximum plasma values of free bisphenol A represented low percentages [<2–8%] of the total radioactive dose in rats orally administered bisphenol A at 10 or 100 mg/kg bw (Pottenger et al., 2000); maximum values of free bisphenol A represented higher percentages of the radioactive dose in rats given 10 or 100 mg/kg bw s.c. [64–82% free bisphenol A] or i.p. [19–54%] (Pottenger et al., 2000). Percentages of parent bisphenol A in blood were also higher in monkeys exposed intravenously (i.v.; 5–29%) than orally (0–1%) (Kurebayashi et al., 2002). Similarly, HPLC analysis of plasma conducted 1 hr following s.c. or gavage dosing of 4 female 21-day-old Sprague–Dawley rats/group with bisphenol A revealed higher bisphenol A plasma concentrations with s.c. than with gavage dosing (Table 22) (Yamasaki et al., 2000). One study in male and female rats gavaged with 10 mg/kg bw bisphenol A demonstrated higher plasma concentrations of bisphenol A in immature animals than in adults (10.2–48.3 μg/g [mg/L] plasma at 4 days of age; 1.1–1.4 μg/g [mg/L] plasma at 7 days of age; 0.2 μg/g [mg/L] plasma at 21 days of age; and 0.024–0.063 μg/g [mg/L] plasma in adulthood) (Domoradzki et al., 2004).

Table 22. Plasma Bisphenol A Concentrations in 21-Day-Old Rats at 1 Hr Following Oral Gavage or S.C. Dosinga
 Plasma concentration, μg/L
Dose, mg/kg bwInjection (s.c.)Oral gavage
  • Values presented as mean±SD.

  • a

    aYamasaki et al. (2000).

0 (sesame oil vehicle)Not detectedNot detected
894.6±58.0Not examined
40886.3±56.4Not detected
1602948±768.8198.8±88.2
800Not examined2879.0±2328.3

A review by the European Union (2003) noted that in the study by Pottenger et al. (2000), fecal excretion represented the highest proportion of the eliminated dose (74–83% in males and 52–72% in females) following oral or parenteral exposure of rats to 10 or 100 mg/kg bw bisphenol A. The authors of the European Union report therefore concluded that absorption [assumed to be of the radioactive dose] is likely extensive following oral intake. Adding to the proof of extensive oral absorption is the observation that >50% of fecal elimination occurred at 24 hr post-dosing, a time period beyond the average gastrointestinal transit time of 12–18 hr for rats. Possible explanations provided for the detection of parent compound in feces were cleavage of conjugates within intestines and enterohepatic circulation.

2.1.2.2 Distribution
2.1.2.2.1 Pregnant or lactating animals:

Information on distribution in pregnant or lactating rats is presented first followed by other species. Studies including oral exposures are summarized before those with parenteral exposures.

Takahashi and Oishi (2000) examined disposition and placental transfer of bisphenol A in F344 rats. Rats were orally administered 1000 mg/kg bw bisphenol A (>95% purity) in propylene glycol on gestation day (GD) 18 (GD 0=day of vaginal plug). Rats were killed at various time points between 10 min and 48 hr after bisphenol A dosing. At each time point, 2–6 dams and 8–12 fetuses obtained from 2–3 dams were analyzed. Blood was collected from dams and kidneys, livers, and fetuses were removed for measurement of bisphenol A concentrations by HPLC. Results are summarized in Table 23. Study authors noted the rapid appearance of bisphenol A in maternal blood and organs and in fetuses. Concentrations of bisphenol A at 6 hr following dosing were 2% of peak concentrations in maternal blood and 5% of peak concentrations in fetuses. It was noted that in fetuses, area under the time-concentration curve (AUC) was higher and mean retention time, variance of retention time, and terminal half-life were longer than in maternal blood.

Table 23. Toxicokinetic Endpoints for Bisphenol A in Rats Dosed With 1000 mg/kg bw Bisphenol A on GD 18a
 Maternal tissue 
EndpointBloodLiverKidneyFetus
  • a

    aTakahashi and Oishi (2000).

Cmax, mg/L14.717136.29.22
Tmax, min20202020
AUC, mg·hr/L13.170084.022.6
Mean retention time, hr10.629.312.020.0
Variance in retention time, hr squared203657227419
Half-life, hr    
 From 20–40 min0.09520.1780.2450.55
 From 40 min–6 hr2.581.752.981.60
 From 6–48 hr4.65No dataNo data173

Dormoradzki et al. (2003) examined metabolism, toxicokinetics, and embryo-fetal distribution of bisphenol A in rats during 3 different gestation stages. Sprague–Dawley rats were gavaged with bisphenol A (99.7% purity)/radiolabeled 14C-bisphenol A (98.8% radiochemical purity) at 10 mg/kg bw. Bisphenol A was administered to 1 group of non-pregnant rats and 3 different groups of pregnant rats on GD 6 (early gestation), 14 (mid gestation), or 17 (late gestation). GD 0 was defined as the day that sperm or a vaginal plug were detected. Blood, urine, and feces were collected at multiple time points between 0.25 and 96 hr post-dosing. It appears that most and possibly all samples were pooled. Four rats in each group were killed at 96 hr post-dosing. Maternal organs, 6 embryos or fetuses/dam (when possible), and placentas were collected. Samples were analyzed for radioactivity and bisphenol A and/or bisphenol A glucuronide by HPLC/liquid scintillation spectrometry.

In all groups, 90–94% of radioactivity was recovered. Elimination of bisphenol A and its metabolites is discussed in Section 2.1.2.4. At 96 hr following dosing, low percentages of the dose were present in carcass (∼1–6%) and tissues such as brain, fat, liver, kidney, ovary, uterus, and skin. The only quantifiable data in placentas and fetuses at 96 hr were obtained in the GD 17 group, and those samples contained 0.01–0.07% of the bisphenol A dose. Standard deviations for maternal and fetal tissues generally exceeded 50% of the mean. Study authors concluded that disposition of radioactivity was similar in pregnant and non-pregnant rats.

Toxicokinetic data obtained from plasma profiles are summarized in Table 24. The authors stated that there was high inter-animal variability. The presence of two Cmax values was noted by the authors, and they stated that it was the result of enterohepatic circulation of radioactivity. Bisphenol A was not quantifiable in most plasma samples. Because bisphenol A glucuronide represented most (∼95–99%) of the radioactivity, plasma profiles for that metabolite were nearly identical to profiles for radioactivity.

Table 24. Toxicokinetic Data for Radioactivity in Pregnant and Non-Pregnant Rats Gavaged With 10 mg/kg bw 14C-bisphenol Aa
EndpointNon-pregnantGD 6–10GD 14–18GD 17–21
  • a

    aDormoradzki et al. (2003).

Cmax1, mg eq/L0.7160.3700.4821.006
Tmax1, hr0.250.250.250.25
Cmax2, mg eq/L0.1710.3360.2110.278
Tmax2, hr18122412
Time to non-quantifiable level, hr72Not determined7296
AUC
 14C, mg-eq·hr/L6.112.47.110.2
 Bisphenol A glucuronide, mg-eq·hr/L5.812.36.89.7
Percent as bisphenol A glucuronide95.199.295.895.1

A second study was conducted by Dormoradzki et al. (2003) to measure bisphenol A and bisphenol A glucuronide concentrations in maternal and fetal tissues. Rats were gavaged with radiolabeled bisphenol A at 10 mg/kg bw on GD 11, 13, or 16. Blood was collected over a 24-hr period. Five rats/group/time period were killed at 0.25, 12, and 96 hr post-dosing. Maternal blood and organs, yolk sacs/placentas, and embryos/fetuses were removed for measurement of bisphenol A and bisphenol A glucuronide. Yolk sacs/placentas and fetuses were pooled at most time periods. Results are summarized in Table 25.

Table 25. Maternal and Fetal Concentrations of Bisphenol A Following Gavage Dosing of Dams With 10 mg/kg bw Bisphenol Aa
 Bisphenol A concentration, mg/L or mg/kg
  Maternal plasmaYolk sac/placentaEmbryo/fetus
ExposureGlucuronideParentGlucuronideParentGlucuronideParent
  • Data expressed as mean±SD or single values for individual or pooled data.

  • a

    aDormoradzki et al. (2003).

  • a, b

    bLimit of detection (LOD) for bisphenol A reported at 0.005–0.029.

  • c

    cDetected only in two animals at the concentrations listed.

GD 11, 0.2 mCi
 0.25 hr1.060±0.2580.0410.062<LODb<LOD<LOD
 12 hr0.099±0.036<LOD<LOD<LOD<LOD<LOD
 96 hrNANA<LOD<LOD<LOD<LOD
GD 13, 0.2 mCi
 0.25 hr0.868±0.1890.0780.0360.019<LOD0.013
 12 hr0.117±0.0330.0080.0130.009<LOD<LOD
 96 hr Not analyzed due to insufficient radioactivity   
GD 16, 0.2 mCi
 0.25 hr1.768±0.7830.485, 0.129c0.223±0.1040.166±0.0690.031, 0.009c0.122, 0.020c
 12 hr0.174±0.045<LOD0.025±0.0050.034±0.002NANA
 96 hr Not analyzed due to insufficient radioactivity 0.0160.008
GD 16, 0.5 mCi
 0.25 hr1.699±0.5010.064±0.0250.342±0.1040.095±0.0310.013±0.0080.018±0.011

At 0.25 hr following dosing, bisphenol A glucuronide concentrations in maternal plasma were similar in groups dosed on GD 11 and 13 but concentrations were 1.7–2 times higher in the group dosed on GD 16. At 12 hr post-dosing in all exposure groups, bisphenol A glucuronide concentrations in maternal plasma were reduced 7- to 11-fold from values observed at 0.25 hr. Levels of radioactivity in plasma were not sufficient for analysis at 96 hr post-dosing. Bisphenol A was detected in maternal plasma at 0.25 hr post-dosing in rats that were exposed to a higher radioactive concentration (0.5 mCi compared to 0.2 mCi) on GD 16; bisphenol A concentrations were 26.5-fold lower than bisphenol A glucuronide concentrations.

In animals dosed on GD 11, bisphenol A glucuronide was only detected in yolk sac/placenta at 0.25 hr post-dosing and the concentration was ∼17 times lower than the concentration detected in maternal blood for the same time period. With dosing on GD 11, bisphenol A glucuronide was not detected in embryos and bisphenol A was not detected in yolk sac/placenta or embryos. In animals dosed on GD 13, bisphenol A glucuronide was detected in yolk sac/placenta at 0.25 and 12 hr post-dosing and concentrations were 9–24-fold lower than those detected in maternal plasma for the same time period. Bisphenol A was also detected in yolk sac/placenta at 0.25 and 12 hr after dosing on GD 13 and concentrations were similar to those detected in the blood of 2 dams. A lower concentration of bisphenol A was detected in embryos of dams at 0.25 hr following dosing on GD 13, and bisphenol A was the only moiety detected in embryos. Following dosing on GD 16, bisphenol A glucuronide and bisphenol A were detected in yolk sac/placenta at 0.25 and 12 hr post-dosing. Concentrations of bisphenol A glucuronide in yolk sac/placenta were 7- to 8-fold lower than concentrations detected in maternal plasma. From 0.25 to 12 hr, concentrations of bisphenol A decreased 4.9-fold and concentrations of bisphenol A glucuronide decreased 9-fold. Mean concentrations of bisphenol A in yolk/sac placenta following exposure on GD 16 were similar to the blood concentration detected in 1 of 2 dams.

In yolk sac/placenta and fetuses of dams dosed with a higher level of radioactivity (0.5 mCi) on GD 16, bisphenol A glucuronide and bisphenol A were detected at 0.25 hr following dosing. Compared to concentrations detected in placenta, fetal concentrations of bisphenol A glucuronide were ∼26-fold lower and bisphenol A concentrations were 5-fold lower. Bisphenol A concentrations were lower than bisphenol A glucuronide concentrations by 3.6-fold in yolk sac/placenta and by 0.7-fold in fetuses. Study authors concluded that there is no selective affinity for bisphenol A or bisphenol A glucuronide by the yolk sac/placenta or embryo/fetus.

Kurebayashi et al. (2005) examined distribution of radioactivity in pregnant and lactating rats dosed with 14C-bisphenol A. Pregnant rats were orally dosed with 0.5 mg/kg bw 14C-bisphenol A on GD 12, 15, or 18. The rats were killed at 30 min or 24 hr following dosing (n=1/time period) and examined by whole-body radioluminography. Study authors noted that the distribution of label was nearly identical in dams at each gestation time point. At 30 min following dosing, the concentration of radioactivity in dam blood was ∼31–43 μg bisphenol A eq/L. The highest concentration of radioactivity was detected in maternal liver (∼219–317 μg bisphenol A eq/kg) and kidney (∼138–270 μg bisphenol A eq/kg); concentrations in other tissues (lung, ovary, placenta, skin, and uterus) were ∼10-fold lower. Fetuses, fetal membranes, and yolk sacs did not contain quantifiable levels of radioactivity at 30 min following maternal exposure at any gestation time point. At 24 hr following exposure of dams, radioactivity concentrations in blood (∼4–11 μg bisphenol A eq/L) were nearly 3–10-fold lower than values obtained at 30 min following exposure. Levels of radioactivity remained highest in liver. At 24 hr following exposure, radioactivity was only detected in fetuses and fetal tissues from dams dosed on GD 18. Radioactivity levels in fetuses or fetal tissues compared to maternal blood were ∼30% in fetuses, nearly equal in fetal membranes, and ∼5-fold higher in yolk sacs. Study authors concluded that there was limited distribution of radiolabel to fetuses.

In another study by Kurebayashi et al. (2005), a lactating rat was orally dosed with 0.5 mg/kg bw 14C-bisphenol A on PND 11 and caged with 5 neonatal rats for 24 hours. One male and one female neonatal rat were killed at the end of the 24-hr period and examined by whole-body radioluminography. The 3 remaining neonates were caged for 24 hr with a dam that was not exposed to bisphenol A. One male and one female neonate were then killed and examined by whole-body radioluminography. In pups killed immediately after being nursed by the lactating dam exposed to 14C-bisphenol A, most of the radioactivity was detected in intestinal contents (∼30–46 μg bisphenol A eq/kg) and lower levels were found in gastric contents and urinary bladder (<10 μg bisphenol A eq/kg). After being nursed for 24 hr by a dam that was not exposed to bisphenol A, radioactivity was only detected in intestinal contents and the level was ∼20–40% of that measured in pups examined immediately after being nursed by dams receiving 14C-bisphenol A.

An additional 3 lactating dams were dosed with 0.5 mg/kg bw 14C-bisphenol A on PND 11 for examination of radioactivity in plasma and milk over a 48-hr period. Table 26 summarizes toxicokinetic endpoints for radioactivity in milk and plasma. Study authors concluded that there was significant secretion of 14C-associated radioactivity into milk.

Table 26. Toxicokinetic Endpoints for Radioactivity in Lactating Rats Orally Administered 0.5 mg/kg bw 14C-Bisphenol A on PND 11a
EndpointMilkMaternal plasma
  • a

    aKurebayashi et al. (2005).

Cmax, μg-eq/L4.4627.2
Tmax, hr84
Elimination half-life, hr2631
AUC (0–48 hr), μg-eq·hr/L)156689

Miyakoda et al. (1999) examined placental transfer of bisphenol A in rats. Wistar rats were administered an oral dose of bisphenol A (99% purity) at 10 mg/kg bw on GD 19. Blood was collected and fetuses were removed at 1, 3, and 24 hr following dosing. Bisphenol A concentrations were measured in plasma and fetuses by GC/MS. [A statement in Figure3of the study indicated that values were the means of 5 or 7 experiments; it is possible the authors meant that 5 or 7 dams were dosed.] Concentrations of bisphenol A peaked in maternal plasma and fetuses within 1 hr of dosing, with bisphenol A concentrations measured at ∼34 ppb [μg/L] in maternal plasma and 11 ppb [μg/kg] in fetuses. At 3 hr after dosing, bisphenol A concentrations were ∼10% of peak concentrations in maternal plasma and 40% of peak concentrations in fetuses. At 24 hr post-dosing, bisphenol A concentrations in fetuses were detected at 70% of peak value and concentrations in fetuses were more than twice the concentrations in maternal plasma. Study authors concluded that bisphenol A is rapidly transferred to the fetus and tends to remain longer in fetuses than in maternal blood.

Snyder et al. (2000) examined the toxicokinetics of bisphenol A in lactating rats. On PND 14, lactating CD rats were gavaged with 100 mg/kg bw 14C-bisphenol A. Milk, blood, and organs were collected from 2–4 dams/group at 1, 8, 24, or 26 hr after dosing. [While the text indicates collection of samples at 26 hr, Table3of the study indicates collection at 24 hr. The collection time reported in the study table was used when there were discrepancies between text and table.] Animals were injected with oxytocin before milk collection. Radioactivity in pup carcasses was measured at 2, 4, 6, and 24 hr following exposure of dams; 8–16 pups/time period were examined [pup data does not appear to be analyzed by litter]. Samples were analyzed by scintillation counting, HPLC, and/or nuclear magnetic resonance. At 1 and 8 hr following exposure, the highest percentage of the radioactive dose was detected in intestine with contents (75–83%). Among the other organs examined, the highest percentage of the radioactive dose was detected in liver (0.38–0.74%) and much lower percentages were detected in kidney and lung (≤0.02%). Low percentages of the radioactive dose were also detected in milk (≤0.0020%), blood (∼0.006%), plasma (∼0.01%), and fat (≤0.004%). Compared to earlier time periods, radioactivity levels were lower at 24 hr post-dosing (26% of the dose detected in intestine and contents), but distribution was similar. At all 3 sampling time points, radioactivity levels were highest in plasma > blood > milk. The major radioactivity peak in plasma was represented by bisphenol A glucuronide at 1, 8, and 26 hr following exposure. Bisphenol A glucuronide also represented the major radioactive peak detected in milk. Radioactivity levels in pups amounted to <0.01% of the maternal dose. Radioactivity levels in pups tended to increase over time. From 2–24 hr following exposure, mean±SD radioactivity levels rose from 44±24 to 78±11 μg bisphenol A eq/pup.

Yoshida et al. (2004) compared bisphenol A concentrations in rats and their offspring during the lactation period. The main focus of the study was developmental toxicity, which is discussed in Section 3.2.3.2. In the distribution study, Donryu rats (12–19/group) were gavaged with bisphenol A at 0 (carboxymethylcellulose solution), 0.006, or 6 mg/kg bw/day from GD 2 to the day before weaning (21 days post-delivery). Bisphenol A concentrations were measured in maternal and pup serum, milk, and pup liver by GC/MS on PND 10, 14, and/or 21. Milk samples were obtained from pup stomachs. Pup serum and liver samples were pooled. Two to six dams/litter were examined in each dose group and time period. Samples of tap water, drinking water from plastic containers, and feed were measured for bisphenol A content by HPLC. Bisphenol A was not detected in fresh tap water but was detected at ∼3 μg/L following storage of that water in plastic containers. Bisphenol A concentration in feed was ∼40 μg/kg. Results for maternal and fetal tissues are summarized in Table 27, 28. Bisphenol A concentrations in the serum of high-dose-dams were significantly elevated compared to the control group on PND 21. No other significant differences were observed in bisphenol A concentrations in samples between treated and control groups.

Table 27. Bisphenol A Concentrations in Maternal and Pup Samples During Lactation in Rats Gavaged With Bisphenol Aa
   Dose group, mg/kg bw/day
   00.0066
SampleTime of analysisSexBisphenol A concentration, ppb [μg/L or μg/kg]
  • a

    aYoshida et al. (2004).

  • b

    bValues are presented as mean±SD.

  • c

    cPup samples were pooled.

Damb
 SerumPND 21 3±04±011±4
 MilkPND 10 28±98±218±3
 PND 14 255±78205±7185±50
Pupc
 SerumPND 10Female41023
  Male1557
 PND 14Female543
  Male454
 PND 21Female939
  Male14920
 LiverPND 10Female131217
  Male9914
Pupc
 LiverPND 14Female2210018
  Male451416
 PND 21Female607037
  Male69960
Table 28. Toxicokinetic Endpoints for Bisphenol A in Pregnant Rats iv Dosed With 2 mg/kg bw Bisphenol Aa
EndpointCompartment
 Maternal serumPlacentaFetusAmniotic fluid
  • Values presented as mean±SD.

  • a

    aShin et al. (2002).

AUC, μg·hr/L905.5±275.84009±962.71964.7±678.5180.4±102.0
Elimination half-life, hr2.5±0.92.2±0.82.2±0.83.9±3.1
Mean residence time, hr3.0±1.12.0±0.53.0±0.95.6±4.7
Cmax, μg/L927.3±194.31399.2±323.7794±360.675.1±59.7
Tmax, hrNo data0.1±0.10.6±0.30.3±0.2

Kim and Huang (2003) used an HPLC method to measure bisphenol A concentrations in rat dams and their offspring. Dams were gavaged with bisphenol A (>99.7% purity) at doses of 0 (corn oil vehicle), 0.002, 0.020, 0.200, 2, or 20 mg/kg bw/day on GD 7–17. Dams and offspring were killed at 21 days following parturition, and serum was collected for measurement of bisphenol A. Development effects observed in this study are summarized in Section 3.2.1.1. Bisphenol A was not detected in the serum of dams at the two lowest doses. Respective concentrations of bisphenol A in the serum of dams at the 3 highest doses were 0.900, 0.987, and 1.00 mg/L. In offspring, bisphenol A was not detected in serum at the 3 lowest doses. At the 2 highest doses, the respective concentrations of bisphenol A in offspring were 0.69 and 0.74 mg/L in males and 0.71 and 0.82 mg/L in females.

Shin et al. (2002) examined elimination of bisphenol A from maternal–fetal compartments of rats. On 1 day between GD 17 and 19, four Sprague–Dawley rats were i.v. injected with 2 mg/kg bw bisphenol A. Amniotic fluid, placenta, and fetuses were collected at multiple intervals between 5 min and 8 hr following injection. Bisphenol A concentrations in samples were measured by HPLC. Transfer rate constants and clearance rates were determined using a five-compartment model consisting of maternal central, maternal tissue, placental, fetal, and amniotic fluid compartments. Toxicokinetic findings are summarized in Moors et al. (2006) evaluated the kinetics of bisphenol A in pregnant rats on GD 18 after a single i.v. dose of 10 mg/kg bw. Unconjugated bisphenol A represented almost 80% of total bisphenol A 5 min after injection, 50% of total bisphenol A 20 min after injection, and ∼10% of total bisphenol A 6 hr after the injection. The half-life of free bisphenol A in the dam's blood was 0.34 hr, and the half-life of total bisphenol A was 0.58 hr. Bisphenol A in fetal tissues peaked 20–30 min after maternal injection at 4.0 mg/kg in placenta, 3.4 mg/kg in fetal liver, and 2.4 mg/kg in remaining fetal tissues. Peak maternal blood bisphenol A had been 3.8 mg/L shortly after injection.

Rapid distribution of bisphenol A was observed in placenta, fetus, and amniotic fluid. Bisphenol A concentrations in placenta and fetus remained higher than those in maternal serum over most of the sampling period. Amniotic fluid contained the lowest concentration of bisphenol A. Decay curves in amniotic fluid, fetus, and placenta paralleled decay curves in maternal serum. Transfer rate constants and clearance rates are summarized in Table 29. Transfer rate constants were greater in the direction of amniotic fluid to fetus or placenta than in the opposite direction. The elimination rate constant and clearance rate from the fetal compartment were much lower than for the maternal central compartment. The clearance rate from placenta to fetus was higher than clearance rate from fetus to placenta. The authors calculated that 65.4% of the bisphenol A dose was delivered to the fetus, 33.2% to the maternal central compartment, and 1.4% to amniotic fluid. According to the study authors, the low transfer rate from the fetal to amniotic compartment suggested minimal fetal excretion of unchanged bisphenol A through urine and feces into the amniotic fluid. They also noted that the small fetal compartment transfer constant compared to the relative fetal–placental transfer constant indicated minimal metabolism by the fetus. Authors estimated that 100% of bisphenol A was eliminated from the fetus via the placental route and concluded that fetal elimination represents 0.05% of total elimination from the maternal–fetal unit.

Table 29. Intercompartmental Transfer and Clearances in Pregnant Rats Following Intravenous Bisphenol Aa
CompartmentTransfer rate constant, hr−1Clearance rate mL/min
  • Values presented as mean±SD.

  • a

    aShin et al. (2002).

Maternal central to maternal tissue3.4±2.638.2±26.5
Maternal tissue to maternal central1.7±1.350.2±36.7
Maternal central to placental0.7±0.58.3±5.4
Placental to maternal central23.6±14.72.2±1.3
Placental to fetal46.4±29.24.1±2.1
Fetal to placental22.8±28.07.6±6.0
Fetal to amniotic fluid0.00001±0.000020.00001±0.00001
Fetal0.0062±0.00440.0024±0.0015
Amniotic fluid to fetal14.0±21.00.8±1.1
Amniotic fluid to placental7.9±6.70.7±0.7
Placental to amniotic fluid1.0±1.30.1±0.1
Maternal central0.9±0.69.7±5.3

Moors et al. (2006) evaluated the kinetics of bisphenol A in pregnant rats on GD 18 after a single i.v. dose of 10 mg/kg bw. Unconjugated bisphenol A represented almost 80% of total bisphenol A 5 min after injection, 50% of total bisphenol A 20 min after injection, and ∼10% of total bisphenol A 6 hr after the injection. The half-life of free bisphenol A in the dam's blood was 0.34 hr, and the half-life of total bisphenol A was 0.58 hr. Bisphenol A in fetal tissues peaked 20–30 min after maternal injection at 4.0 mg/kg in placenta, 3.4 mg/kg in fetal liver, and 2.4 mg/kg in remaining fetal tissues. Peak maternal blood bisphenol A had been 3.8 mg/L shortly after injection.

Yoo et al. (2001) examined mammary excretion of bisphenol A in rats. At 4–6 days postpartum, 4–6 lactating female Sprague–Dawley rats/group were i.v. injected with bisphenol A at 0.47, 0.94, or 1.88 mg/kg bw and then infused with bisphenol A over a 4-hr period at rates of 0.13, 0.27, or 0.54 mg/hour. Blood samples were collected at 2, 3, and 4 hr, and milk was collected at 4 hr following initiation of infusion. Before collection of milk, rats were injected with oxytocin to increase milk production. HPLC was used to measure bisphenol A concentrations in serum. Differences in data for mean systemic clearance were analyzed by analysis of variance (ANOVA). Results are summarized in Table 30. The study authors noted extensive excretion of bisphenol A into milk, with milk concentrations exceeding serum concentrations. No significant differences were reported for systemic clearance rates between the 3 doses. Steady state concentrations of bisphenol A in maternal serum and milk increased linearly according to dose.

Table 30. Toxicokinetic Endpoints in Lactating Rats Infused With Bisphenol Aa
 Bisphenol A infusion rate, mg/hr
Endpoint0.130.270.54
  • Data presented as mean±SD.

  • a

    aYoo et al. (2001).

Systemic clearance, mL/min/kg119.2±23.8142.4±45.3154.1±44.6
Steady state serum bisphenol A concentration, ng/mL66.1±15.5120.0±34.7217.1±65.0
Steady state milk bisphenol A concentration, ng/mL173.1±43.3317.4±154.4493.9±142.2
Milk/serum ratio2.7±0.92.6±1.22.4±0.6

Kabuto et al. (2004) reported bisphenol A concentrations in mice indirectly exposed to bisphenol A during gestation and lactation. The focus of the study was oxidative stress; more details are presented in Section 3.2.7. Six ICR mouse dams were given drinking water containing 1% ethanol vehicle or bisphenol A at 5 or 10 μg/L. [Based on the reported water intake of 5 mL/day and an assumed body weight of 0.02 kg (USEPA,1988), it is estimated that bisphenol A intakes in mice at the start of pregnancy were 0.0013 and 0.0025 mg/kg bw/day.] Mice gave birth about 3 weeks following mating and pups were housed with dams for 4 weeks. [Based on an assumed body weight of 0.0085 kg and assumed water intake rate of 0.003 L/day (USEPA,1988), it is estimated that intake of bisphenol A in weanling males was 0.0018 and 0.0035 mg/kg bw/day.] At 4 weeks of age, male pups were killed and a GC/MS technique was used to measure bisphenol A concentrations in brain, kidney, liver, and testis in an unspecified number of control pups and in four pups from the 10 μg/L group. Study authors reported that they could not detect bisphenol A in control pups. In pups from the 10 μg/L group, the highest concentration of bisphenol A was detected in kidney (∼24 μg/kg wet weight), followed by testis (∼20 μg/kg wet weight), brain (∼18 μg/kg wet weight), and liver (∼11 μg/kg wet weight).

Zalko et al. (2003) examined metabolism and distribution of bisphenol A in pregnant CD-1 mice. A series of studies was conducted in which mice were treated with 3H-bisphenol A (>99.9% purity)/unlabeled bisphenol A (>99% purity). Mice were exposed to different regimens; biological samples examined included blood, liver, fat, gall bladder, uterus, ovaries, digestive tract and contents, urine, and feces. In the first exposure scenario, mice were s.c. injected with 0.025 mg/kg bw labeled/unlabeled bisphenol A on GD 17; three animals/time period were examined at 0.5, 2, and 24 hr following dosing. In the second exposure scenario, 2 mice/group were s.c. injected with 50 mg/kg bw bisphenol A on GD 17 and killed 24 hr following dosing. In the third scenario, 3 non-pregnant female mice/group were “force-fed” a single oral dose of 0.025 mg/kg bw bisphenol A; urine and feces were collected over 24 hr, and animals were killed at 24 hr. Biological samples were analyzed by scintillation analysis, HPLC, MS, and/or nuclear magnetic resonance.

In pregnant mice injected with 0.025 mg/kg bw/day bisphenol A and examined 24 hr later, 85.68% of the radioactivity was recovered. The highest percentages of radioactivity were detected in the digestive tract and contents (∼45%) and feces (∼21%). Less radioactivity was detected in the litter (∼4%), liver (∼2%), bile (∼2%), urine (∼6%), and carcass (∼3%). Blood, ovaries, uterus, placenta, amniotic fluid, fat, and cage washes each contained <1% of the radioactive dose. At 0.5 hr following dosing, levels of radioactivity were highest in uterus > liver > placenta > fetus > amniotic fluid > ovaries > carcass > blood. Radioactivity levels in tissues were lower by 24 hr following exposure. [Compared to radioactive levels detected in tissues at 24 hr, levels detected at 0.5 hr were ∼12-fold higher in uterus, 3-fold higher in liver, 8-fold higher in placenta, 3.5-fold higher in fetuses, 2-fold higher in amniotic fluid, and 3.5-fold higher in ovaries.] The only information provided for mice s.c. dosed with 50 mg/kg bw bisphenol A and examined 24 hr later was for radioactivity levels in organs; the highest levels (pg/g) were detected in uterus > blood > ovary > carcass > liver. Study authors stated that distribution of radioactivity was comparable in mice treated with 50 and 0.025 mg/kg bw bisphenol A. In the mice orally dosed with 0.025 mg/kg bw bisphenol A and examined 24 hr later, levels of radioactivity in blood, ovaries, and uterus were reported to be significantly lower [by ∼1–2 orders of magnitude] than levels in animals exposed by s.c. injection, but the level in the liver was not significantly different. There was significantly more residue in mouse carcass after oral than s.c. dosing (∼2.5 fold) (A. Soto, personal communication, March 2, 2007). No qualitative differences in metabolites were observed following oral or s.c. exposure. [Data were not shown by study authors.] Distribution of parent compound and metabolites detected in maternal and fetal tissues is summarized in Table 31. Further discussion on metabolites is included in Section 2.1.2.3.

Table 31. Qualitative Analysis of Maternal and Fetal Tissues Following Injection of Mice With 0.025 mg/kg bw Radiolabeled Bisphenol A on GD 17a
 Bisphenol A-associated compound detected
 Hydroxylated glucuronideDouble glucuronideMetabolite FbGlucuronideParentOthers
Hr after dose11.5 ng/g%17.5 ng/g%24.0 ng/g%25.0 ng/g%33.5 ng/g%%
  • Data presented as mean±SD

  • a

    aZalko et al. (2003).

  • b

    bMost likely bisphenol A glucuronide conjugated to acetylated galactosamine or glucosamine.

Maternal plasma
 0.50.07±0.0130.11±0.0240.11±0.0241.01±0.19391.06±0.19419
 20.02±0.0120.03±0.0140.03±0.0140.55±0.14630.15±0.041710
 240.04±0.0420 0 00.13±0.0565 015
Placenta
 0.5 0 00.46±0.4825.50±4.242515.98±12.02721
 20.03±0.0210.04±0.0310.37±0.0773.13±2.34621.32±0.95263
 240.05±0.0450.04±0.0240.64±0.19590.21±0.22190.06±0.0466
Fetus
 0.50.05±0.0310.04±0.0400.46±0.2753.83±2.65444.20±2.16491
 20.02±0.0210.01±0.0200.37±0.22131.93±0.45660.48±0.55163
 240.01±0.011 00.11±0.07130.51±0.12600.13±0.16152
Amniotic fluid
 0.50.10±0.1410.19±0.1420.09±0.1318.17±6.55830.90±0.8994
 20.06±0.0310.07±0.0310.26±0.1554.82±4.81880.10±0.0722
 240.13±0.0580.01±0.0210.37±0.09240.70±0.13440.03±0.03220
Maternal liver
 0.50.12±0.1200.18±0.2406.22±1.751812.90±2.813710.85±2.773112
 20.08±0.0810.77±0.2582.16±0.91204.95±1.82451.51±0.971313
 240.16±0.1420.35±0.1370.99±0.42162.56±1.62361.72±1.182317

Uchida et al. (2002) examined distribution of bisphenol A in pregnant mice and monkeys. On GD 17 (GD 0=day of vaginal plug), ICR mice were s.c. injected with bisphenol A 100 mg/kg bw in sesame oil vehicle. More than 3 mice/time point were killed at various points between 0.5–24 hr following injection. An untreated control group consisted of 6 mice. [Data were not presented for controls.] Maternal and fetal serum and organs were collected. Among organs collected were fetal uteri and testes, which were pooled. On GD 150, 2 Japanese monkeys (Macaca fuscata) were s.c. injected with 50 mg bisphenol A/kg bw and at 1 hr following injection, fetuses were removed by cesarean section. Two untreated fetuses were used as controls. Maternal and fetal serum and organs, not including reproductive organs, were collected from monkeys. Bisphenol A concentrations were measured by GC/MS in mouse and monkey samples.

In mice, bisphenol A was detected within 0.5 hr of exposure in all tissues examined, including placenta, maternal and fetal serum, liver, and brain, and fetal uterus, and testis. Bisphenol A concentrations were higher in fetal than maternal serum and liver. [Peak concentrations were observed within 0.5–1 hr in most tissues, with the exception of fetal brain (2 hr), and concentrations remained elevated for 1–6 hr, depending on tissue. More than one peak was observed in fetal serum, uterus, and testis.] In exposed monkeys, bisphenol A was found at the highest concentrations (15.6–72.50 mg/kg) in fetal heart, intestine, liver, spleen, kidney, thymus, muscle, cerebrum, pons, and cerebellum; bisphenol A concentrations in the same organs from control monkeys were measured at 3.70–22.80 mg/kg. Lower concentrations of bisphenol A were detected in umbilical cord and maternal and fetal serum of the exposed group (1.70–6.10 mg/kg) and control group (0.02–0.25 mg/kg). The study authors stated that the most likely source of bisphenol A in control monkeys was the feed, which was found to contain bisphenol A. The study authors concluded that the placental barrier does not protect the fetus from bisphenol A exposure.

Halldin et al. (2001) examined distribution of bisphenol A in quail eggs or hens. After injection of fertilized quail egg yolk sacs with 67 μg/g 14C-bisphenol A egg on incubation day 3, <1% of radioactivity was detected in embryos at incubation day 6 or 9. A similar finding was reported for diethylstilbestrol. At incubation day 6, no specific localization was observed in the embryo but in 10- and 15-day-old embryos a high amount of radioactivity was observed in liver and bile. [Low transfer of labeled bisphenol A to the egg was reported after oral or i.v. dosing of quail hens (with apparently 105μg bisphenol A), but concentrations in eggs were not quantified by study authors.]

2.1.2.2.2 Non-pregnant and non-lactating animals:

Domoradzki et al. (2004), examined the effects of dose and age on toxicokinetics and metabolism of bisphenol A in rats. Neonatal and adult male and female Sprague–Dawley rats were gavaged with 14C-bisphenol A (∼99% radiochemical purity)/non-radiolabeled bisphenol A (99.7% purity). Three neonatal rats/age/sex/time period were dosed on PND 4, 7, and 21 with 1 or 10 mg/kg bw bisphenol A. Adult rats (11 weeks old) [number treated not specified] were dosed with 10 mg/kg bw bisphenol A. Blood samples were collected at various time points from 0.25–24 hr post-dosing in neonatal rats and from 0.25–96 hr in adult rats. Plasma samples were pooled on PND 4. Immature rats were killed at 24 hr post-dosing, and adult rats were killed at 96 hr post-dosing. Brain, liver, kidneys, skin, and reproductive organs were collected from neonatal rats. Levels of radioactivity, bisphenol A, and/or metabolites were analyzed in blood and tissue samples using HPLC and liquid scintillation spectrometry.

In neonatal and adult rats, radioactivity levels in plasma generally peaked within 0.25–0.75 hr. With the exception of 0.25 hr post-dosing on PND 4, when plasma radioactivity levels were ∼4-fold higher in males than females, plasma radioactivity levels were generally similar in male and female rats. At 24 hr post-dosing, plasma radioactivity levels were 4–100 times lower in all groups of neonatal rats. Trends were noted for decreasing radioactivity levels with increasing age. Information related to dose- and age-related effects on metabolism is presented in Section 2.1.2.3.

Toxicokinetic values for bisphenol A are listed in Table 32. Cmax and AUC values for bisphenol A decreased with increasing age, especially following dosing with 10 mg/kg bw. Bisphenol A concentrations were lower in adults than neonates. No patterns were observed for half-lives, and the authors stated that values in neonates may not have been reliable because bisphenol A concentrations were near the LOD at the end of the 24-hr observation period. Ratios of Cmax and AUC values for the 10 and 1 mg/kg bw doses were different at each age and generally decreased with age. Plasma bisphenol A concentrations were very low in adults dosed with 10 mg/kg bw; therefore, few data were available.

Table 32. Toxicokinetic Values for Bisphenol A in Rats Following Gavage Dosing With 1 or 10 mg/kg bwa
 Age at exposure and sex
 PND 4PND 7PND 21Adult
EndpointMaleFemaleMaleFemaleMaleFemaleMaleFemale
  • Data missing from table cells were not determined.

  • a

    aDomoradzki et al. (2004).

Bisphenol A dose: 1 mg/kg bw
 Tmax, hr0.250.250.250.2533  
 Cmax, mg/L0.030.060.040.080.0050.006  
 Half-life, hr7.27.321.88.8    
 AUC, mg·hr/L0.20.10.10.1    
Bisphenol A dose: 10 mg/kg bw
 Tmax, hr0.250.250.250.251.51.50.250.75
 Cmax, mg/L48.310.21.11.40.20.20.0240.063
 Half-life, hr176.711.48.54.36.6“0”“0”
 AUC, mg·hr/L23.17.21.91.71.11“0”“0”
Ratio of value at 10 to 1 mg/kg bw/day
 Cmax161017027.517.5    
 AUC115.2721917    

Toxicokinetic values for bisphenol A glucuronide are listed in Table 33. Peak plasma concentrations of bisphenol A glucuronide were 9–22 times higher in neonates than adult rats dosed with 10 mg/kg bw bisphenol A. AUC values for bisphenol A glucuronide were also higher in neonates than adults [∼2–6 times higher]. In neonates dosed with 1 mg/kg bw, AUC values and elimination half-lives for bisphenol A glucuronide decreased with age. Ratios of Cmax and AUC values for the 10 and 1 mg/kg bw doses were nearly proportional. In adults dosed with 10 mg/kg bw, bisphenol A glucuronide concentrations peaked at 0.25 hr and secondary peaks were observed at 18 and 24 hr. In neonates dosed with 10 mg/kg bw, concentrations of bisphenol A glucuronide peaked at 0.75–1.5 hr and then bisphenol A glucuronide was eliminated in an apparently monophasic manner. Half-lives of elimination were shorter in neonates compared to adults. In neonatal rats, the bisphenol A glucuronide represented 94–100% of the 1 mg/kg bw dose and 71–97% of the 10 mg/kg bw/day dose. In adult rats, ∼100% of the dose was represented by bisphenol A glucuronide.

Table 33. Toxicokinetic Values for Bisphenol A Glucuronide in Rats Following Gavage Dosing With 1 or 10 mg/kg bw Bisphenol Aa
 Age at exposure and sex
 PND 4PND 7PND 21Adult
EndpointMaleFemaleMaleFemaleMaleFemaleMaleFemale
  • Data missing from table cells were not determined.

  • a

    aDomoradzki et al. (2004).

Bisphenol A dose: 1 mg/kg bw
 Tmax, hr0.750.750.750.250.250.25  
 Cmax, mg/L1.31.521.10.80.8  
 Half-life, hr26.124.26.66.44.24.1  
 AUC, mg·hr/L99.67.77.74.13.3  
 AUCBPA-glucuronide/AUCBPA45967777    
Bisphenol A dose: 10 mg/kg bw
 Tmax, hr1.51.51.50.750.750.750.250.25
 Cmax, mg/L13.16.36.610.310.47.80.60.7
 Half-life, hr7.39.89.18.44.44.422.510.8
 AUC, mg·hr/L8050.358.960.960.356.131.59.8
 AUCBPA-glucuronide/AUCBPA3.5731365556  
Ratio of value at 10 to 1 mg/kg bw/day
 Cmax10.14.23.39.4139.8  
 AUC8.95.27.67.914.717  

Half-life and AUC data for bisphenol A-derived radioactivity in organs of neonatal rats are summarized in Table 34. Radioactivity was distributed to all organs and dose-related increases were observed. The study authors noted lower concentrations in brain than in other tissues. [Levels of radioactivity in reproductive organs compared to those in plasma varied at each evaluation period but were usually within the same or one order of magnitude lower.] With the exception of males dosed with 10 mg/kg bw bisphenol A, half-lives decreased with age. There were some disproportionate increases in ratios of AUC at 10 and 1 mg/kg bw.

Table 34. Distribution of Radioactivity to Tissues at 24 Hr Following Dosing With Radiolabeled Bisphenol Aa
 PND 4PND 7PND 21
TissueHalf-life hrAUC mg · hr/kgAUC ratio of dosesHalf-life hrAUC mg · hr/kgAUC ratio of dosesHalf-life hrAUC mg · hr/kgAUC ratio of doses
  • a

    aDomoradzki et al. (2004).

Females, 1 mg/kg bw
Brain11.70.4 6.70.2 3.60.1 
Liver187.5 7.97.1 3.62.9 
Kidney18.19.4 7.39.5 5.03.0 
Ovary11.77.3 6.03.5 3.70.9 
Uterus7.48.3 6.23.0 3.41.0 
Carcass11.222.2 10.016.6 4.08.3 
Plasma19.59.4 6.47.8 3.63.5 
Females, 10 mg/kg bw
Brain7.23.38.38.02.512.54.91.717.0
Liver11.144.86.010.059.68.44.539.113.5
Kidney15.243.94.78.666.67.05.336.512.2
Ovary6.5136.218.75.069.719.93.621.123.4
Uterus15.2127.015.34.8108.536.23.430.630.6
Carcass6.6112.85.17.0130.77.94.8100.912.2
Plasma9.261.06.58.167.08.63.759.016.9
Males, 1 mg/kg bw
Brain14.10.3 6.00.3 3.40.1 
Liver19.76.1 6.67.3 3.73.2 
Kidney19.38.5 7.08.6 4.63.4 
Testis10.33.4 5.72.0 3.40.8 
Carcass11.122.2 9.017.3 4.19.0 
Plasma24.09.2 6.67.7 3.44.2 
Males, 10 mg/kg bw
Brain3.14.715.78.02.99.74.71.717.0
Liver11.648.47.911.862.08.55.140.912.8
Kidney5.468.98.19.859.66.96.930.48.9
Testes5.836.810.87.622.111.15.28.110.1
Carcass8.3111.75.08.6135.57.84.895.210.6
Plasma6.9113.012.39.969.09.04.062.014.8

The study authors concluded:

  • Metabolism of bisphenol A to its glucuronide conjugate occurred as early as PND 4 in rats;

  • Dose-dependent differences occurred in neonatal rats, as noted by a larger fraction of the lower dose being metabolized to the glucuronide; and

  • There were no major sex differences in metabolism or toxicokinetics of bisphenol A.

Pottenger et al. (2000) examined the effects of dose and route on metabolism and toxicokinetics of bisphenol A in rats. Information focusing on toxicokinetics is summarized primarily in this section, while metabolic data are summarized primarily in Section 2.1.2.3. Adult male and female F344 rats were dosed with 14C-bisphenol A (99.3% radiochemical purity)/non-radiolabeled bisphenol A (99.7% purity) at doses of 10 or 100 mg/kg bw by oral gavage or i.p. or s.c. injection. Blood was collected at multiple time points between 0.083 and 168 hr post-dosing, and excreta were collected for 7 days. Animals were killed 7 days post-dosing. Blood, brain, gonads, kidneys, liver, fat, skin, uterus, and carcass were analyzed by liquid scintillation counting and HPLC. Some samples were analyzed by HPLC/electrospray ionization/MS.

Toxicokinetic endpoints for bisphenol A in blood are summarized in Table 35. Study authors noted that concentration-time profiles of bisphenol were dependent on dose, exposure route, and sex. The longest Tmax was observed with s.c. dosing. Cmax and AUC values were lowest following oral administration. Time to non-quantifiable concentrations of bisphenol A was longest following s.c. exposure. The only sex-related difference was a higher Cmax value in females than males following oral dosing. In most cases, bisphenol A toxicokinetics were linear across doses within the same administration route, as noted by approximate proportionate increases in Cmax and AUC values from the low to the high-dose. Toxicokinetics data for radioactivity in plasma are summarized in Table 36. Concentrations of radioactivity were dependent on exposure route and to a lesser extent, dose and sex. AUC values for radioactivity were lowest following oral exposure. Time to non-quantifiable concentration was longest following s.c. dosing. For most groups, Cmax and AUC values were proportionate across doses within the same exposure route. A second part of the study examined metabolites and is summarized in Section 2.1.2.3.

Table 35. Toxicokinetic Endpoints for Bisphenol A in Blood Following Dosing of Rats by Gavage or Injectiona
 Exposure route and doses (mg/kg bw)
Endpoint10 oral100 oral10 i.p.100 i.p.10 s.c.100 s.c.
  • Missing values were not determined.

  • a

    aPottenger et al. (2000).

  • b

    bMean±SD.

  • c

    cNon-quantifiable (0.01 μg/g at 10 mg/kg bw and 0.1 μg/g at 100 mg/kg bw).

Males
 Tmax, hrN/A0.0830.50.250.750.5
 Cmax, mg/L, hrbc0.22±0.090.69±0.089.7±1.270.39±0.165.19±0.98
 Time to non-quantifiable concentration, hr0.0830.758121824
 AUC, mg · hr/L 0.11.116.42.624.5
Females
 Tmax, hr0.250.250.250.2540.75
 Cmax, mg/L, hrb0.04±0.032.29±1.820.87±0.1513.13±4.130.34±0.063.97±0.6
 Time to non-quantifiable concentration, hr1 24724872
 AUC, mg·hr/L0.424.41.426.23.131.5
Table 36. Toxicokinetics for Radioactivity Following Dosing of Rats with Bisphenol A Through Different Exposure Routesa
 Exposure route and doses (mg/kg bw)
Endpoint10 oral100 oral10 i.p.100 i.p.10 s.c.100 s.c.
  • a

    aPottenger et al. (2000).

Males
 Tmax, hr0.250.250.50.2510.75
 Cmax, mg eq/L, hr0.73±0.223.92±1.931.26±0.0929.3±11.70.61±0.246.33±0.43
 Time to non-quantifiable concentration, hr7272969696144
 AUC, mg-eq · hr/L8.166.516.917015.5218
Females
 Tmax, hr0.0830.250.250.50.750.75
 Cmax, mg eq/L, hr1.82±0.6628.33±8.642.27±0.1967.81±7.330.52±0.065.66±0.95
 Time to non-quantifiable concentration, hr727272120120168
 AUC, mg-eq·hr/L9.5494.915.324721.6297

Upmeier et al. (2000) examined toxicokinetics in rats exposed to bisphenol A through the oral or i.v. route. Ovariectomized DA/Han rats (130–150 g bw) were exposed to bisphenol A by i.v. injection with 10 mg/kg bw or oral gavage with 10 or 100 mg/kg bw. Blood was collected from treated rats at multiple time points until 2 hr following i.v. dosing and 3 hr following oral dosing. Three to five rats were sampled during each time period. To reduce stress, only some of the rats were sampled at each time point. In control animals, blood was collected 2 hr following dosing with vehicle. Bisphenol A concentrations in plasma were measured by GC/MS. Dosing with 10 mg/kg bw i.v. resulted in a maximum plasma concentration of 15,000 μg/L bisphenol A. Concentrations decreased to 700 μg/L within 1 hr, 100 μg/L within 2 hr, and non-detectable concentrations by 24 hr followingexposure. The apparent final elimination half-life was estimated at 38.5 hr. In rats gavaged with 10 mg/kg bw, an initial maximum blood concentration of 30 μg/L was obtained at 1.5 hr. A decrease in bisphenol A blood concentration at 2.5 hr was followed by a second peak of 40 μg/L at 6 hr, leading study authors to conclude that enterohepatic cycling was occurring. The same patterns of bisphenol A concentrations in blood were observed following gavage dosing with 100 mg/kg bw. Peak concentrations were observed at 30 min (150 μg/L) and 3 hr (134 μg/L) following exposure. According to the study authors, the differences in peak concentrations observed between the two doses suggested lower bioavailability at the high-dose than at the low dose. Oral bioavailability of bisphenol A was estimated at 16.4% at the low dose and 5.6% at the high-dose.

Yoo et al. (2001) examined toxicokinetics of a low i.v. dose and a higher gavage dose of bisphenol A in male rats. Five adult male Sprague–Dawley rats/group were administered bisphenol A by i.v. injection at a dose of 0.1 mg/kg bw or by gavage at a dose of 10 mg/kg bw. Multiple blood samples were collected until 3 hr following i.v. dosing and 24 hr following gavage dosing. HPLC was used to measure bisphenol A concentrations in serum. Route-specific differences in mean systemic clearance were analyzed by Student t-test. Results are summarized in Table 37. The study authors noted bi-exponential decay of serum bisphenol A concentrations following i.v. dosing, significantly longer elimination half-life with oral than i.v. exposure, and low oral bioavailability of bisphenol A.

Table 37. Toxicokinetic Values for Bisphenol A in Adult Rats Exposed to Bisphenol A Through the Intravenous or Oral Routea
 Bisphenol A dosing
Endpoint0.1 mg/kg bw, i.v.10 mg/kg bw, gavage
  • Data presented as mean±SD.

  • a

    aYoo et al. (2001).

Distribution half-life, min6.1±1.3 
Terminal elimination half-life, hr0.9±0.321.3±7.4
AUC, μg·hr/L16.1±3.285.6±33.7
Systemic clearance, mL/min/kg107.9±28.7 
Steady-state volume of distribution, L/kg5.6±2.4 
Cmax, μg/L 14.7±10.9
Tmax, hr 0.2±0.2
Apparent volume of distribution, L/kg 4273±2007.3
Oral clearance, mL/min/kg 2352.1±944.7
Absolute oral bioavailability, % 5.3±2.1

Kurebayashi et al. (2003) conducted a series of studies to examine toxicokinetics and metabolism of bisphenol A in adult F344N rats exposed through the oral or i.v. route. In these studies, radioactivity levels were measured by scintillation counting. Bisphenol A or its metabolites were quantified by HPLC, electrospray ionization/MS, or nuclear magnetic resonance. As discussed in greater detail in Section 2.1.2.4, fecal excretion was the main route of elimination for radioactivity following oral or i.v. dosing of rats with 0.1 mg/kg bw 14C-bisphenol A. A study describing biliary excretion and metabolites in bile is summarized in Section 2.1.2.3. Toxicokinetic endpoints were determined in a study in which blood was drawn from 3 male rats/group at various time points between 0.25–48 hr following oral gavage or i.v. dosing with 0.1 mg/kg bw bisphenol A. Results of the study are summarized in Table 38. Rapid absorption of radioactivity was observed following oral dosing. AUC values were significantly lower for oral than i.v. dosing. In a another study, rats were administered 14C-bisphenol A by i.v. injection and blood was collected 30 min later for determination of blood/plasma distribution and protein binding. At a blood radioactivity level of 80 nM [18 μg bisphenol A eq/L], preferential distribution to plasma was observed, with the blood/plasma ratio reported at 0.67. At radioactivity levels of 6–31 μg-eq/L (27–135 nM), plasma protein binding was reported at 95.4%. Additional studies reviewed by Teeguarden et al. (2005) reported plasma protein binding of bisphenol A at ∼90–95%. An additional study by Kurebayashi et al. (2003) compared metabolic patterns and excretion following exposure to a higher bisphenol A dose; that study is discussed inSection 2.1.2.3.

Table 38. Toxicokinetic Endpoints for 14C-Bisphenol A-Derived Radioactivity in Rats Exposed to 0.1 mg/kg bw 14C-Bisphenol A Through the Oral or I.V. Routea
EndpointI.V. exposureOral exposure
  • Data presented as mean±SD. Missing values are not applicable or were not reported.

  • a

    aKurebayashi et al. (2003).

  • b

    bP < 0.05 compared to i.v. exposure.

  • c

    cVariances not reported.

Tmax, hr 0.38±0.10
Cmax, μg-eq/L 5.5±0.3
Half-life-α, hr0.59±0.09No data
Half-life-β, hr39.5±2.144.5±4.1
Absorbance rate, hr−1 3.6±1.0
Volume of distribution, L/kg27.0±0.7No data
Total body clearance. L/hr/kg0.522±0.0110.544±0.049
Mean residence time, hr51.7±2.4No data
AUC, μg-eq·hr/L
  0–6 hr33.9±1.618.4±0.7b
  0–24 hr79.3±3.360.0±7.1b
  0–48 hr118±4102±13b
  0–∞192±4185±16
Oral bioavailabilityc
  0–6 hr 0.54
  0–24 hr 0.76
  0–48 hr 0.86
  0–∞ 0.97

Kurebayashi et al. (2005) administered 14C-bisphenol A to adult male and female F344 rats (3/dose/sex) at doses of 0.020, 0.1, or 0.5 mg/kg bw orally or 0.1 or 0.5 mg/kg bw by i.v. injection. Plasma samples were analyzed for radioactivity over a 72-hr period to determine toxicokinetic endpoints. Results are summarized in Table 39. Study authors noted that the AUC was almost linearly correlated with dose. Several peaks were observed with oral or i.v. exposure, indicating enterohepatic cycling, according to the study authors. Study authors noted that substantially lower AUC values in females than in males following oral exposure could have resulted from lower absorption and/or a higher elimination rate. Distribution of radioactivity was evaluated 0.5, 24, and 72 hr following oral administration of 0.1 mg/kg bw bisphenol A to adult male and female Wistar rats (3/sex/time point). At 0.5 hr following exposure, most of the radioactivity (∼12–51 μg bisphenol A eq/kg) was found in kidney and liver. [A large amount of radioactivity was also reported for intestinal contents, but those data were not shown by the study authors.]. Lower amounts of radioactivity (∼2–7 μg bisphenol A eq/kg or L) were detected in adrenal gland, blood, lung, pituitary gland, skin, and thyroid gland of both sexes; uterus; and bone marrow, brown fat, and mandibular gland of males. In males,<μg bisphenol A eq/kg was detected in skeletal muscle and testis. Radioactivity was non-quantifiable in brain and eye of both sexes; epididymis, prostate gland, and heart of males; and bone marrow, brown fat, skeletal muscle, and mandibular gland of females. At ≥24 hr following exposure, radioactivity was detected primarily in only kidney, liver, and intestinal contents, with the exception of ∼3 μg bisphenol A eq/L detected in blood of males at 24 hr following dosing. Study authors noted that elimination of radioactivity from some tissues appeared to occur more rapidly in females than in males. Distribution in pregnant animals was also examined and is described in Section 2.1.2.2.1.

Table 39. Toxicokinetic Endpoints for Plasma Radioactivity in Rats Dosed With 14C Bisphenol Aa
 Route and dose (mg/kg bw)
 OralI.V.
Endpoints20100500100500
  • Data presented as mean±SD.

  • a

    aKurebayashi et al. (2005).

Males
 Elimination half-life, hr78±5218±321±319±221±3
 AUC, μg-eq·hr/L36±6178±44663±164266±46865±97
 Apparent absorption, %828160  
Females
 Elimination half-life, hr20±722±1318±813±316±2
 AUC, μg-eq·hr/L14±599±19500±43190±451029±81
 Apparent absorption, %355050  

Kabuto et al. (2003) reported distribution of bisphenol A in mice. Male ICR mice were i.p. dosed with bisphenol A at 0, 25, or 50 mg/kg bw/day for 5 days and killed 6 hr following the last dose. Bisphenol A concentrations in tissues of animals from the high-dose group were determined by GC/MS. In mice of the high-dose group, the highest concentrations of bisphenol A were detected in kidney (∼2.02 mg/kg wet weight) and body fat (∼1.25 mg/kg wet weight). Lower concentrations of bisphenol A (≤0.42 mg/kg wet weight or mg/L) were detected in brain, lung, liver, testis, and plasma.

Kurebayashi et al. (2002) examined the toxicokinetics of a low bisphenol A dose in Cynomolgus monkeys following gavage or i.v. dosing. Three adult male and female monkeys were dosed with 0.1 mg/kg bw 14C-bisphenol A (99% radiochemical purity)/non-radiolabeled bisphenol A [purity not reported]. Monkeys were dosed by i.v. injection on Day 1 of the study and by gavage on Day 15 of the study. Urine and feces were collected for 7 days post-dosing. Blood samples were collected at various time points from 0.083–72 hr following i.v. dosing and for 0.25–71 hr after oral dosing. Binding to plasma protein was determined at some time points over 0.25–4 hr. Samples were analyzed by liquid scintillation counting and HPLC. Following oral or i.v. exposure, the percentage of radioactivity recovered in excreta and cage washes was 81–88% over a 1-week period. As discussed in greater detail in Section 2.1.2.4, most of the radioactivity was excreted in urine and very little was excreted in feces. Toxicokinetic endpoints are summarized in Table 40. Based on the toxicokinetic values, study authors concluded that absorption of bisphenol A following oral exposure was rapid and high, and terminal elimination half-lives of bisphenol A/metabolites were longer following i.v. than oral exposure. As discussed in more detail in Section 2.1.2.3, glucuronide compounds were the major metabolites detected in urine, and higher percentages of the radioactive dose in plasma were represented by bisphenol A following i.v. than oral dosing.

Table 40. Toxicokinetic Endpoints for Radioactivity in Male and Female Cynomolgus Monkeys Exposed to 14C-Bisphenol A Through IV Injection or by Gavagea
EndpointMaleFemale
  • [Mean±SD assumed based on data presentations elsewhere in this study.]

  • a

    aKurebayashi et al. (2002).

Intravenous exposure
 AUC, μg-eq·hr/L377±85382±96
 Volume of distribution, L/kg1.58±0.111.82±0.41
 Half-life, hr13.5±2.614.7±2.1
 Total body clearance, L/hr/kg0.27±0.050.28±0.08
 Mean residence time, hr5.93±0.916.68±0.72
Oral exposure
 AUC, μg-eq·hr/L265±74244±21
 Tmax, hr1.00±0.870.33±0.14
 Cmax. μg-eq/L104±85107±37
 Half-life. hr9.63±2.749.80±2.15
 Bioavailability0.70±0.160.66±0.13

Negishi et al. (2004b) compared toxicokinetics of bisphenol A in female F344/N rats, Cynomolgus monkeys, and Western chimpanzees. Bisphenol A was administered by oral gavage and s.c. injection at doses of 10 or 100 mg/kg bw/day to rats and monkeys and 10 mg/kg bw to chimpanzees. Three rats/dose/time point were killed before and at various times between 0.5 and 24 hr following bisphenol A administration. Three monkeys/group and 2 chimpanzees were first exposed orally and 1 week later by s.c. injection. In monkeys, blood samples were drawn before and at various times from 0.5–24 hr after dosing. In chimpanzees, blood was drawn before and at multiple time points between 0.25–24 hr following dosing. Bisphenol A was measured in serum by ELISA, and toxicokinetics endpoints were determined. Results are summarized in Table 41. The study authors noted that the bioavailability of bisphenol was lowest in rats<chimpanzees<monkeys following exposure through either route. In most cases, bisphenol A was not detected in rat serum following oral administration of the 10 mg/kg bw dose. In all species, higher bioavailability was observed with s.c. than oral dosing.

Table 41. Toxicokinetic Endpoints for Bisphenol A by ELISA in Rats, Monkeys, and Chimpanzeesa
 10 mg/kg bw100 mg/kg bw
EndpointsOralS.C.OralS.C.
  • Data were not reported in cases where table cells are empty.

  • a

    aNegishi et al. (2004b).

Rat (data presented as mean±SD)
 Cmax, μg/L 872±164580±3983439±679
 Tmax, hr 1.00.51.0
 AUC0–4 h, μg·hr/L 1912±262506±3139314±2634
 AUC0–24 h, μg·hr/L 3377±3341353±46223,001±6387
Monkey (data presented as mean±SD)
 Cmax, μg/L279±±92057,934±19025732±52510,851±3915
 Tmax, hr0.7±0.22.0±0.00.7±0.22.0±0.0
 AUC0–4 h, μg·hr/L3209±53615,316±585614,747±249548,010±11,641
 AUC0–24 h, μg·hr/L3247±58739,040±10,73852,595±8951189,627±21,790
Chimpanzee (data presented for 2 animals)
 Cmax, μg/L325; 962058; 1026Dose not administered
 Tmax, hr0.5; 0.52.0; 2.0  
 AUC0–4 h, μg·hr/L491; 2355658; 3109  
 AUC0–24 h, μg·hr/L1167; 81321,141; 12,492  

In a subsequent report (Tominaga et al., 2006), these authors noted that ELISA may overestimate bisphenol A concentrations due to non-specific binding. They reported measurements by LC-MS/MS in animals evaluated using the same study design [possibly the same specimens reported previously]. These results are summarized in Table 42. The authors proposed that primates, including humans, may completely glucuronidate orally-administered bisphenol A on its first pass through the liver and excrete it in the urine whereas bisphenol A remains in the rat for a more extended period due to enterohepatic recirculation. They suggested that the rat may not be a good model for human bisphenol A kinetics.

Table 42. Toxicokinetic Endpoints for Bisphenol A by LC-MS/MS in Rats, Monkeys, and Chimpanzeesa
 10 mg/kg bw100 mg/kg bw
EndpointsOralS.C.OralS.C.
  • a

    aTominaga et al. (2006).

  • b

    b1 or 2 animals.

Rat (data presented as mean±SD)
 Cmax, μg/L2.1±1.6746±8047.5±10.62631±439
 Tmax, hr0.7±0.30.8±0.30.5±0.01.2±0.8
 t1/2, hrnot calculated3.2±0.7not calculated4.5±0.7
 AUC0–4 h, μg·hr/L4.2b1542±20043.2±9.76926±1071
 AUC0–24 h, μg·hr/L7.2b1977±182350±29415,576±2263
Monkey (data presented as mean±SD)
 Cmax, μg/L11.5±2.24213±331928.6±3.97010±3045
 Tmax, hr1.0±0.91.7±0.63.3±1.22.7±1.2
 t1/2, hr8.9±3.03.8±0.84.5±0.712.9±3.6
 AUC0–4 h, μg·hr/L21.4±6.18828±430985.3±18.619,981±7567
 AUC0–24 h, μg·hr/L42.5±7.318,855±3870350±1379,796±21,750
Chimpanzee (data presented as mean for 2 animals)
 Cmax, μg/L5.5703Dose not administered
 Tmax, hr0.81.0  
 t1/2, hr6.84.2  
 AUC0–4 h, μg·hr/L13.32148  
 AUC0–24 h, μg·hr/L33.16000  
2.1.2.3 Metabolism:

Information is arranged in this section according to species. In rats, study summaries are arranged in order of those providing general or route-specific information on metabolites, specifics on organs or enzyme isoforms involved in metabolism, and pregnancy-, sex-, or age-related effects on metabolism.

Pottenger et al. (2000) examined the effects of dose and route on toxicokinetics of bisphenol A in rats. Disposition of bisphenol A and its metabolites in urine and feces is described in this section, while results of the toxicokinetics study are described in Section 2.1.2.2. Five adult F344 rats/sex/group were dosed with 14C-bisphenol A (99.3% radiochemical purity)/non-radiolabeled bisphenol A (99.7% purity) at doses of 10 or 100 mg/kg bw by oral gavage or i.p. or s.c. injection. Excreta were collected for 7 days. Samples were analyzed by HPLC or HPLC/electrospray ionization/MS. The percentage of radioactivity recovered from all groups was 84–98%. Fecal elimination represented the largest percentage of radioactivity in all exposure groups (52–83%). Eight peaks were identified in feces, and the largest peak (representing 86–93% of radioactivity) was for unchanged bisphenol A. Elimination of radioactivity through urine was ∼2-fold higher in females (21–34%) than males (13–16%) in all dose groups. Fourteen different peaks were identified in urine. It was estimated that radioactivity in urine was represented by bisphenol A monoglucuronide (57–87%), bisphenol A (3–12%), and bisphenol A sulfate (2–7%). Some differences were noted for retention of radioactivity following dosing by gavage (0.03–0.26%), i.p. injection (0.65–0.85%), and s.c. injection (1.03–1.29%).

Metabolites associated with bisphenol A exposure were examined in a second study by Pottenger et al. (2000). Three rats/sex/dose/route/time point were dosed with 14C-bisphenol A/non-radiolabeled bisphenol A at 10 or 100 mg/kg bw by oral gavage or i.p. or s.c. injection. Rats were killed at 2 different time points following dosing, Tmax, and the time when bisphenol A concentrations were no longer quantifiable. Times at which rats were killed were determined by data obtained during the first study. Plasma samples were pooled at each time period and examined by HPLC or HPLC/electrospray ionization/MS. Qualitative and quantitative differences were observed for parent compound and metabolites in plasma following exposure through different routes. Following oral exposure, bisphenol A glucuronide was the most abundant compound detected in plasma at both time periods (Cmax and time when parent compound was not quantifiable) and represented 68–100% of total radioactivity. Following i.p. or s.c. exposure, unmetabolized bisphenol A was the most abundant compound at Tmax; levels of radioactivity represented by unmetabolized bisphenol A were 27–51% following i.p. exposure and 65–76% following s.c. exposure. Only 2–8% of radioactivity was represented by bisphenol A following oral exposure. Some compounds observed following i.p. or s.c. exposure were not observed following oral exposure. A compound tentatively identified as a sulfate conjugate was observed following i.p. exposure and represented a small portion of radioactivity. An unresolved peak of 3 compounds was observed following i.p. or s.c. exposure, at the time when parent compound was not quantifiable and represented that major percent of radioactivity for that time point. Three additional unidentified, minor peaks were observed following i.p. or s.c. but not oral exposure. The major sex differences observed were higher Cmax values for bisphenol A and bisphenol A glucuronide in females than males, especially following i.p. administration. A review by the European Union (2003) noted that the substantially higher concentrations of parent compound with i.p. and s.c. compared to oral exposure indicated the occurrence of first-pass metabolism following oral intake.

Elsby et al. (2001) examined bisphenol A metabolism by rat hepatocytes. In the hepatocyte metabolism study, hepatocytes were isolated from livers of adult female Wistar rats and incubated in dimethyl sulfoxide (DMSO) vehicle or bisphenol A 100 or 500 μM [23 or 114 mg/L] for 2 hr. Metabolites were identified by HPLC or LC/MS. Data were obtained from 4 experiments conducted in duplicate. At both concentrations, the major metabolite was identified as bisphenol A glucuronide, which was the only metabolite identified following incubation with 100 μM bisphenol A. Two additional minor metabolites identified at the 500 μM concentration included 5-hydroxy-bisphenol A-sulfate and bisphenol A sulfate. Another part of the study comparing metabolism of bisphenol A by rat and human metabolites is discussed in Section 2.1.1.3. Another study (Pritchett et al., 2002) comparing metabolism of bisphenol A in humans, rats, and mice is also summarized in Section 2.1.1.3.

In neonatal rats gavaged with 1 or 10 mg/kg bw 14C-bisphenol A on PND 4, 7, and 21 and adult rats gavaged with 10 mg/kg bw bisphenol A, the major compounds detected in plasma were bisphenol A glucuronide and bisphenol A (Domoradzki et al., 2004). Up to 13 radioactive peaks were identified in neonatal rats dosed with 10 mg/kg bw and 2 were identified in neonates dosed with 1 mg/kg bw/day. At the 10 mg/kg bw dose, the concentration of bisphenol A glucuronide detected in plasma increased with age. Metabolic profiles were generally similar in males and females. The study authors noted that metabolism of bisphenol A to its glucuronide conjugate occurs as early as PND 4 in rats. However, age-dependent differences were observed in neonatal rats, as noted by a larger fraction of the lower dose being metabolized to the glucuronide. More details from this study are included in Section 2.1.2.2.

Kurebayashi et al. (2005) used a thin layer chromatography technique to examine metabolite profiles in blood, urine, and feces of 3 male rats orally dosed with 0.5 mg/kg bw 14C-bisphenol A. [The procedure did not identify metabolites.] Parent bisphenol A represented ∼2% of the dose in plasma at 0.25 and 6 hr post-dosing and ∼0.3% of the dose at 24 hr after exposure. Unmetabolized bisphenol A represented 1.6% of compounds in urine and 77.2% of compounds in feces collected over a 24-hr period. Free bisphenol A represented 47.1% of compounds in urine following β-glucuronidase hydrolysis of urine, and there was an almost equivalent decrease in a metabolite the study authors identified as “M2.” Therefore, the study authors stated that M2 was most likely bisphenol A glucuronide. M2 was the major metabolite identified in plasma (∼74–77%) and urine (∼40%).

The European Union (2003) reviewed studies by Atkinson and Roy (1995a,b) that reported two major and several minor adducts in DNA obtained from the liver of CD-1 rats dosed orally or i.p. with 200 mg/kg bw bisphenol A. Chromatographic mobility of the two major adducts was the same as that observed when bisphenol A was incubated with purified DNA and a peroxidase or microsomal P450 activation system. The profile closely matched that of adducts formed with the interaction between bisphenol O-quinone and purified rat DNA deoxyguanosine 3′-monophosphate. Formation of the adduct appeared to be inhibited by known inhibiters of cytochrome P (CYP) 450. It was concluded that bisphenol A is possibly metabolized to bisphenol O-quinone by CYP450.

Biliary excretion of bisphenol A and its metabolites following oral or i.v. dosing with bisphenol A was examined by Kurebayashi et al. (2003). Bile ducts of 3 rats/sex/group were cannulated, and the rats were dosed with 0.1 mg/kg bw 14C-bisphenol A (>99% radiochemical purity) in phosphate buffer vehicle by oral gavage or i.v. injection. Biliary fluid was collected every 2 hr over a 6-hr period to determine percent total biliary excretion and percent of dose represented by bisphenol A glucuronide. Results are summarized in Table 43. The study authors noted that the importance of biliary excretion following oral or i.v. dosing. 14C-bisphenol A-glucuronide was the predominant metabolite in bile.

Table 43. Biliary Excretion in Male and Female Rats Exposed to 0.1 mg/kg bw 14C-Bisphenol A Through the Oral or Intravenous Routea
 MaleFemale 
ParametersI.V.OralI.V.Oral
  • a

    aKurebayashi et al. (2003).

Biliary excretion, %
  0–2 hr48323528
  0–4 hr61445039
  0–6 hr66505845
Radioactivity in bile represented by glucuronide, %84868788
Dose excreted as glucuronide in bile, %55435040

In another study by Kurebayashi et al. (2003), biliary, fecal, and urinary metabolites were examined in male rats gavaged with 100 mg/kg bw bisphenol A or D16-bisphenol A in corn oil. Bile was collected over an 18-hr period, and urine and feces were collected over a 72-hr period. The primary metabolite detected in urine was bisphenol A glucuronide, which represented 6.5% of the dose. Lower percentages of the dose (≤1.1%) were present in urine as bisphenol A and bisphenol A sulfate. In feces, the primary compound detected was bisphenol A, which represented 61% of the dose. No glucuronide or sulfate conjugated metabolites of bisphenol A were detected in feces. Most of the dose in bile consisted of bisphenol A glucuronide (41% of the dose). Bisphenol A represented 0.3% of the dose in bile. The study authors noted that as with oral or i.v. exposure to a smaller dose, feces was the main route of elimination for bisphenol A and bile was the main elimination route for bisphenol A glucuronide.

A study by Yokota et al. (1999) examined the hepatic isoform of uridine diphosphate glucuronosyltransferase (UDPGT) involved in the metabolism of bisphenol A and distribution of the enzyme in organs of Wistar rats. Using yeast cells genetically engineered to express single rat UDPGT enzymes, it was determined that UGT2B1 was the only isoform capable of glucuronidating bisphenol A. Microsomal UDPGT activity toward bisphenol A was demonstrated in liver, kidney, and testis, but activity was minimal in lung and brain. [Minimal activity was also observed for intestine]. Northern blot analyses revealed high expression of UGTB1 only in liver. It was demonstrated that 65% of glucuronidation activity was absorbed by binding with anti-UGTB1, indicating that additional isoforms are likely involved in glucuronidation of bisphenol A.

The intestine was determined to play a role in the metabolism of bisphenol A in rats. Nine-week-old male Sprague–Dawley rats were orally administered 0.1 mL of a solution containing 50 g/L bisphenol A [5 mg total or17 mg/kg bw assuming a body weight of0.3 kg (USEPA,1988)] (Sakamoto et al., 2002). Rats were killed at multiple time intervals between 15 min and 12 hr following exposure. The small intestine was removed and separated into upper and lower portions. Intestinal contents were removed from each section. Bisphenol A and metabolite concentrations were measured by HPLC. Activities and expression of β-glucuronidase were determined. A large amount of bisphenol A glucuronide was detected in the upper and lower portions of the small intestine, and a large amount of free bisphenol A was detected in the cecum. Less bisphenol A was detected in colon and feces. The observations lead the study authors to conclude that free bisphenol A generated in the cecum as a result of deconjugation was reabsorbed in the colon. The presence of large amounts of bisphenol A glucuronide in the small intestine at 12 hr following exposure suggested that bisphenol A was reabsorbed in the colon and re-excreted as the glucuronide. As determined in an assay using p-nitrophenol-β-d-glucuronide as a substrate, ∼70% of total β-glucuronidase activity was present in the cecum and 30% in the colon. Western blot analysis revealed a large amount of bacterial β-glucuronidase protein in cecum and colon contents.

Glucuronidation and absorption of bisphenol A in rat intestine were studied by Inoue et al. (2003a). Intestines were obtained from 8-week-old male Sprague–Dawley rats, and the small intestine was divided into 4 sections. Small intestine and colon were everted and exposed to 40 mL of a solution containing bisphenol A at 10, 50, or 100 μM [2.3, 11, or 23 mg/L, resulting in delivery of 91, 456, or 913μg bisphenol A to the everted intestine]. Every 20 min during a 60-min time period, reaction products were collected from serosal and mucosal sides and analyzed by HPLC. Optimal glucuronidation was observed at 50 μM [11 mg/L]. At 60 min following exposure to 50 μM bisphenol A, ∼37% of bisphenol A was absorbed by the small intestine and ∼83% was glucuronidated. Approximately 74.7% of the glucuronide was excreted on the mucosal side and ∼25.3% transported to the serosal side of small intestine. Slightly greater absorption of bisphenol A in the colon (48.6%) compared to the proximal jejunum (37.5%) was observed at 60 min following exposure to the 50 μM solution. Transport of both bisphenol A and bisphenol A glucuronide to the serosal side of intestine increased distally and was greatest in the colon. Minimal mucosal excretion was observed in the colon.

Inoue et al. (2004) compared glucuronidation of bisphenol A in pregnant, non-pregnant, and male rats. Livers of 4 male and non-pregnant Sprague–Dawley rats/group were perfused via the portal vein for 1 hr with solutions containing bisphenol A at 10 or 50 μM [2.3 or 11 mg/L]. The total amount of bisphenol A infused into livers was 1.5 or 7.5 μmol [0.34 or 1.7 mg]. On GD 20 or 21, livers of 4 pregnant Sprague–Dawley rats were perfused for 1 hr with 10 μM [2.3 mg/L] bisphenol A. At the start of perfusion, excreted bile and perfusate in the vein were collected every 5 min for 1 hr. Samples were analyzed by HPLC. Statistical analyses were conducted by Student t-test and ANOVA. Bisphenol A glucuronidation in the liver was 59% in male rats and 84% in non-pregnant female rats perfused with the 10 μM solution. The glucuronide was excreted primarily through bile in both males and females, but a significantly higher amount was excreted through bile in non-pregnant females than in males. The total amount of glucuronide excreted into bile and vein was ∼1.4-fold higher in females than males following perfusion with the 10 μM [2.3 mg/L] solution. At the 50 μM [11 mg/L] concentration, bisphenol A glucuronidated within liver was 66% in males and 91% in females. In males the glucuronide was excreted mainly in bile, and in females, a higher amount of glucuronide was excreted in the vein. In livers of pregnant rats perfused with the 10 μM [2.3 mg/L] solution, 69% of bisphenol A was glucuronidated in the liver. Percentages of glucuronide excretion were 54.5% through bile and 45.5% through the vein in pregnant rats. In a comparison of pregnant rats and non-pregnant rats perfused with 10 μM [2.3 mg/L] bisphenol A, biliary excretion in pregnant rats was half that observed in non-pregnant rats, and venous excretion in pregnant rats was 3-fold higher than in non-pregnant rats. To determine the pathway of bisphenol A glucuronide excretion, livers of 4 male Eisai hyperbilirubinemic rats, a strain deficient in multidrug resistance-associated protein, were perfused with 50 μM [11 mg/L] bisphenol A. During and after perfusion, nearly all of the bisphenol A was excreted into the vein, thus indicating that multidrug resistance-associated protein mediates biliary excretion of bisphenol A glucuronide. The study authors concluded that bisphenol A is highly glucuronidated and excreted into bile using a multidrug resistance-associated protein-dependent mechanism, and that venous excretion increases and biliary excretion decreases during pregnancy.

Miyakoda et al. (2000) examined the production of bisphenol A glucuronide in fetal and adult rats. Bisphenol A was orally administered at 10 mg/kg bw to pregnant Wistar rats on GD 19 and to 10-week-old adult male Wistar rats. [The number of animals exposed was not reported. In some legends for study figures, it was stated that the data were from 4 experiments, suggesting that 4 pregnant rats and adult males may have been exposed.] Fetuses were removed at 1 hr following dosing. Blood was drawn and testes were removed from adult males at 1, 3, and 8 hr following dosing. GC/MS was used to measure bisphenol A concentrations in 19 fetuses and in testis of adult rats before and following homogenization with β-glucuronidase. In fetal extracts, there were no differences in bisphenol A concentrations before or after treatment with β-glucuronidase, suggesting that bisphenol A glucuronide was not present at detectable concentrations. The study authors noted the possibility that bisphenol A glucuronide was not transferred from dams to fetuses and stated that glucuronidation by the rat fetus is unlikely. At 1 hr following dosing of adult male rats, 90% of bisphenol A was detected as glucuronide in plasma and testis. Bisphenol A glucuronide concentrations gradually decreased and bisphenol A concentrations increased slightly in testis over the 8-hr sampling period. In plasma, bisphenol A-glucuronide decreased to 55% of the maximum observed concentration at 3 hr following dosing and increased to 100% of maximum observed concentration at 8 hr following dosing. Based on concentrations of bisphenol A glucuronide in testis and blood (40 ppb [μg/kg] and 600 ppb [μg/L]) at 8 hr, the study authors concluded that bisphenol A glucuronide passage through the testicular barrier was unlikely. It was thought that bisphenol A passed through the testicular barrier, was converted to the glucuronide within the testis, and was then gradually released following digestion of the glucuronide by β-glucuronidase.

Matsumoto et al. (2002), studied developmental changes in expression and activity of the UDPGT isoform UGT2B toward bisphenol A in Wistar rats. Activity toward other compounds was also examined but this summary focuses on bisphenol A. Microsomes were prepared from livers of fetuses, neonates on PND 3, 7, 14, and 21, and pregnant rats on GD 10, 15, and 19. Activity toward the bisphenol A substrate was measured using an HPLC method. Expression of UGT2B1 protein was examined by Western blot and messenger ribonucleic acid (mRNA) expression was examined by Northern blot. Little to no UGT2B activity toward bisphenol A was detected in microsomes of fetuses. Activity increased linearly following birth and reached adult concentrations by PND 21. [No data on UGT2B activity for non-pregnant adult rats were shown and it was not clear if activity in adults was examined in this study.] The same developmental patterns were observed for expression of UGT2B1 protein and mRNA. Activity and protein expression of UGT2B1 were also found to be reduced in pregnant rats.

The European Union (2003) reviewed an unpublished study by Sipes that compared clearance of bisphenol A by hepatic microsome from fetal (n=8/sex), immature (n=4/sex), and adult (n=4) rats. The clearance rate in microsomes from male and female GD 19 rat fetuses (0.7–09 mL/min/mg) was lower than clearance rates in microsomes from 4-day-old males and females (1.2–2.6 mL/min/mg), 21-day-old males and females (2.4–2.7 mL/min/mg), and their dams (2.6 mL/min/mg). The European Union concluded that clearance rate was lower in fetuses but reached adult concentrations by 4 days of age.

In a qualitative study of bisphenol A metabolites in pregnant mice injected with 0.025 mg/kg bw bisphenol A, 10 radioactive peaks were observed in urine by Zalko et al. (2003). The major metabolites detected in urine were bisphenol A glucuronide and a hydroxylated bisphenol A glucuronide. Unchanged bisphenol A was the major compound detected in feces (>95%). Bisphenol A glucuronide represented >90% of the compounds detected in bile. Additional compounds detected in urine, feces, digestive tract, or liver included a double glucuronide of bisphenol A and sulfate conjugates. Unchanged bisphenol A, bisphenol A glucuronide, and “metabolite F” (disaccharide conjugate of BPA) were the major compounds detected in all tissues. [Authors state that formation of glucuronic acid conjugate of BPA, several double conjugates, and conjugated methoxylated compounds, demonstrate the formation of potentially reactive intermediates.] The most abundant compound in all tissues was bisphenol A glucuronide, except in placenta where bisphenol A and metabolite F were the major compounds detected. Concentrations of bisphenol A decreased rapidly in all tissues. It was determined that metabolite F was most likely bisphenol A glucuronide conjugated to acetylated galactosamine or glucosamine. Distribution of bisphenol A and its metabolites in maternal and fetal tissues in summarized in Table 31. Additional details of this study are included in Section 2.1.2.2.

Jaeg et al. (2004) reported metabolites observed following incubation of CD-1 mouse liver microsomes or S9 fractions with bisphenol A at 20–500 μM [4.6–114 mg/L]. The metabolites included isopropyl-hydroxyphenol, bisphenol A glutathione conjugate, glutythionyl-phenol, glutathionyl 4-isopropylphenol, 2,2-bis-(4-hydroxyphenyl)1-propanol, 5-hydroxy bisphenol A, and bisphenol A dimers. It was postulated that bisphenol A-ortho-quinone, produced from 5-hydroxy bisphenol A (catechol), may be the reactive intermediate leading to the formation of these metabolites.

Kurebayashi et al., (2002) examined metabolism of bisphenol A in monkeys. Three adult male and female Cynomolgus monkeys were dosed with 0.1 mg/kg bw 14C-bisphenol A/non-radiolabeled bisphenol A by i.v. injection on Study Day 1 and by gavage on Study Day 15 (Kurebayashi et al., 2002). Additional details of the study are included in Section 2.1.2.2. Up to five peaks were identified in urine. Analysis by radio-HPLC suggested that the major peaks in both sexes treated by either exposure route were mono- and diglucuronides. Five peaks were identified in plasma, and some differences were noted in comparisons of i.v. to oral exposure. In the 2 hr following dosing, most of the radioactivity in plasma was represented by bisphenol A glucuronide after i.v. dosing (57–82%) and oral dosing (89–100%). The percentage of radioactivity represented by unchanged bisphenol A was higher following i.v. (5–29%) than oral (0–1%) dosing.

Kang et al. (2006) reviewed studies that provided some information about metabolism of bisphenol A in fish and birds. One study reported bisphenol A sulfate and bisphenol A glucuronide as the major metabolites detected in zebra fish exposed to bisphenol A. A second study conducted in carp reported an increase in UDPGT activity for bisphenol A in microsomes and metabolism of bisphenol A to bisphenol A glucuronide in intestine. In quail embryos, metabolism and excretion of bisphenol A was reported, but specific metabolites were not indicated. Another study reported that 14C-bisphenol A administered orally or i.v. to laying quail was rapidly removed via bile and excreted through feces.

2.1.2.4 Elimination:

Elimination of bisphenol A and its metabolites was examined in Sprague–Dawley rats that were gavaged with bisphenol A and 14C-bisphenol A at 10 mg/kg bw (Domoradzki et al., 2003). One group of rats was not pregnant, and three additional groups were treated on either GD 6 (early gestation), 14 (mid gestation), or 17 (late gestation). More details of this study are available in Section 2.1.2.2. Most of the radioactivity (65–78%) was eliminated in feces. Elimination in urine accounted for 14–22% of the dose, and considerable variability for urinary elimination among animals was evident by the large standard deviations, which were 50% of means. The authors stated that bisphenol A glucuronide represented 62–70% of radioactivity in urine and bisphenol A represented 19–23% of radioactivity in urine [data were not shown by authors]. Nine peaks were identified in urine. In feces, 83–89% of radioactivity was represented by bisphenol A and 2–3% was represented by bisphenol A glucuronide; 7 peaks were identified in feces. The study authors concluded that urinary elimination and fecal elimination of radioactivity were similar in pregnant and non-pregnant rats.

Difference in excretion following oral or i.v. exposure of rats to a low bisphenol A dose was examined by Kurebayashi et al. (2003). Three male rats/group were exposed to 0.1 mg/kg bw 14C-bisphenol A (>99% radiochemical purity) in phosphate buffer vehicle by oral gavage or i.v. injection. Radioactivity levels were measured in urine and feces, which were collected over a 48-hr period. Additional details of the study are included in Section 2.1.2.2. Results of that study are summarized in Table 44. With both oral and i.v. dosing, fecal excretion was the main route of elimination.

Table 44. Excretion of Radioactivity Following Oral or Intravenous Dosing of Rats With 0.1 mg/kg bw 14C-Bisphenol Aa
 Percent radioactive dose excreted
Time post-dosing, hrUrineFecesTotal
  • Values presented as mean±SD.

  • a

    aKurebayashi et al. (2003).

Oral
 0–246.3±1.149.3±2.155.7±2.8
 24–483.8±1.032.3±2.136.1±3.0
 Total10.1±1.681.6±3.791.8±5.0
Intravenous
 0–248.4±1.846.2±1.854.6±3.4
 24–484.1±0.931.4±1.535.4±1.8
 Total12.5±0.977.6±1.890.1±2.7

Kurebayashi et al. (2005) examined elimination of radioactivity in 3 adult male and female F344 rats that were orally dosed with 0.1 mg/kg bw 14C-bisphenol A. Urine and feces were collected over a 168-hr period and analyzed by liquid scintillation counting. Total radioactivity excreted in urine and feces over the 168-hr period was ∼98% in males and females. In male rats, ∼10% was excreted in urine and ∼88% was excreted in feces. Female rats excreted ∼34% of the radioactivity in urine and ∼64% in feces. [The majority of radioactivity, ∼90%, was excreted over 48 hr by males and 72 hr by females.]

Snyder et al. (2000) compared toxicokinetics of bisphenol A in CD and F344 rats. Four CD and F344 rats were gavaged with 100 mg/kg bw 14C-bisphenol A in propylene glycol vehicle. Disposition of radioactivity in urine, feces, and carcass was examined over a 144-hr period. Samples were analyzed by scintillation counting, HPLC, or nuclear magnetic resonance. Data were analyzed by ArcSin transformation of the square root of the mean and using two-sample t-test. Recovery of radioactivity was 93% in both strains. The highest concentrations of radioactivity were detected in feces (70% of dose in CD rat and 50% of dose in F344 rats) followed by urine (21% of dose in CD rat and 42% of dose in F344 rats). The percentages of the dose excreted in urine and feces differed significantly by strain. Much lower percentages of radioactivity were detected in the carcass (∼1%). Bisphenol A glucuronide, representing 81–89% of the dose, was the major urinary metabolite detected in both strains. A much lower percentage (2.2–10%) of the dose was represented by urinary bisphenol A.

Kim et al. (2002b) reported urinary excretion of bisphenol A in 4-week-old male F344 rats given bisphenol A in drinking water at 0 (ethanol vehicle), 0.1, 1, 10, or 100 ppm (equivalent to 0.011, 0.116, 1.094, or 11.846 mg/kg bw/day) for 13 weeks. Urine samples were collected for 24 hr following administration of the last dose and analyzed by HPLC before and after digestion with β-glucuronidase. The focus of the study was male reproductive toxicity; the study is described in detail in Section 4.2.2.1. Bisphenol A was not detected in the urine of rats from the control and 2 lowest dose groups. [At the 2 highest doses, free bisphenol A represented 60 and 30% of the total urinary bisphenol A concentrations.]

In rats exposed to 10 or 100 mg/kg bw/day 14C-bisphenol A through the oral, i.p., or s.c. routes, fecal elimination represented the highest percentage of radioactivity in all exposure groups (52–83%) (Pottenger et al., 2000). Elimination of radioactivity through urine was ∼2-fold higher in females (21–34%) than males (13–16%) in all dose groups. Additional details of this study are included in Section 2.1.2.3.

Elimination of bisphenol A and metabolites was examined in 3 adult male and female Cynomolgus monkeys dosed with 0.1 mg/kg bw 14C-bisphenol A/non-radiolabeled bisphenol A by i.v. injection on Study Day 1 and by gavage on Study Day 15 (Kurebayashi et al., 2002). Additional details of the study are included in Section 2.1.2.2. Following oral or i.v. exposure, the percentage of radioactivity recovered in excreta and cage washes was 81–88% over a 1-week period. Most of the radioactivity was recovered in urine (combination of urine and cage washes), with most of the radioactivity excreted in urine within 12 hr and essentially all of the dose excreted within 24 hr following treatment. Percentages of radioactive doses recovered in urine within 1 week after dosing were ∼79–86% following i.v. dosing and 82–85% following oral dosing. Much smaller amounts were recovered in feces during the week following i.v. or oral exposure (∼2–3%). The study authors concluded that because fecal excretion was very low following oral exposure, absorption was considered to be complete. The authors also noted that there were no obvious route or sex differences in excretion of radioactivity. The study authors concluded that terminal elimination half-lives were longer following i.v. than oral exposure. A limited amount of information was presented for the fast phase, defined as the 2 hr following i.v. injection. Fast-phase elimination half-life of bisphenol A following i.v. exposure was significantly lower in females (0.39 hr) than males (0.57 hr). There were no sex-related differences in fast-phase half-life for bisphenol A glucuronide (0.79–0.82 hr) or total radioactivity (0.61–0.67 hr).

2.1.3 Comparison of humans and experimental animals.

Studies comparing toxicokinetics and metabolism of bisphenol A in humans and laboratory animals were reviewed and are summarized below. In most cases the data were from original sources, but information from secondary sources was included if the information was not new or critical to the evaluation of developmental or reproductive toxicity.

Elsby et al. (2001) compared bisphenol A metabolism by rat and human microsomes. Microsomes were obtained from 8 immature Wistar rats (21–25 days old) and histologically normal livers from 4 male (25–57 years old) and 4 female (35–65 years old) Caucasian donors who were killed in accidents. Human microsomes were pooled according to sex of the donor. Glucuronidation was examined following exposure of microsomes to bisphenol A concentrations of 0–1000 μM [0–228 mg/L] for 30 min with human microsomes and 10 min with rat microsomes. Metabolites were identified by HPLC or LC/MS. Data were obtained from 4 experiments conducted in duplicate. Data were analyzed by Mann–Whitney test. Maximum velocity (Vmax) and the rate constant (Km) values are summarized in Table 45. The study authors reported a significant difference between the Vmax for glucuronidation in immature rats and humans. No sex-related difference was reported for glucuronidation by human microsomes. As a result of less extensive glucuronidation by human than rat microsomes, the study authors noted that estrogen target tissues in humans may receive higher exposure to bisphenol A than tissues of immature female rats used in estrogenicity studies. Lastly, oxidation of bisphenol A by female rat or human microsomes was examined following incubation with 200 μM [46 mg/L] bisphenol A and NADPH. The only metabolite identified was 5-hydroxybisphenol A.

Table 45. Glucuronidation Kinetics in Microsomes From Immature Rats and Adult Humansa
Sex/speciesVmax, nmol/minute/mg proteinKm, μM
  • Data presented as mean±SEM.

  • a

    aElsby et al. (Elsby et al., 2001).

Male/human5.9±0.477.5±8.3
Female/human5.2±0.366.3±7.5
Female/immature rat31.6±8.127.0±1.2

The European Union (2003) reviewed a series of studies by Sipes that compared metabolism of bisphenol A in microsomes from male and female humans (15 pooled samples/sex and 3–5 individual samples/sex), rats (4/sex), and mice (4/sex). It was concluded that the studies generally agreed with the findings of Elsby et al. (2001). Clearance rates (Vmax/Km) in human microsomes (0.4–0.9 mL/min/mg for pooled samples and 0.3–0.5 mL/min/mg in individual samples) were lower than those observed in rats (1.0–1.7 mL/min/mg) and mice (1.3–3.0 mL/min/mg).

Pritchett et al. (2002) compared metabolism of bisphenol A in hepatocyte cultures from humans, rats, and mice. Cell cultures were prepared from adult male and female F344 rats, Sprague–Dawley rats, and CF1 mice. Human hepatocyte cultures were obtained from 3 females and 2 males. [No information was provided about the age of human donors.] Cells were exposed to 14C-bisphenol A (99.3% purity)/bisphenol A (>99% purity) in a DMSO vehicle. In a cytotoxicity assessment, lactate dehydrogenase activity was measured in rat cells following incubation for 18 hr in 5–100 μM [1.1–23 mg/L] bisphenol A, and cytotoxicity was observed at ≥75 μM bisphenol A. Bisphenol A concentrations tested and times of exposure were 5–20 μM [1.1–4.6 mg/L] for up to 6 hr in time-dependent metabolism studies and 2.5–30 μM [0.57–6.8 mg/L] for 10 min in concentration-dependent metabolism studies. Metabolites in cell media were analyzed by HPLC and LC-MS/MS.

Analysis of media from human hepatocytes incubated with bisphenol A indicated that the major metabolite was bisphenol A glucuronide, and compounds found at lower concentrations were bisphenol A glucuronide/sulfate diconjugate, and bisphenol A sulfate conjugate. Table 46 summarizes percentages of each type of metabolite detected in media following incubation with 20 μM [4.6 mg/L] bisphenol A for 3 hr in human cells and 6 hr in rodent cells. In cells from all sexes and species except male F344 rats, bisphenol A glucuronide was the major metabolite detected. The glucuronide/sulfate diconjugate was the major metabolite detected in cells from male F344 rats. In concentration-dependent studies conducted in F344 rat hepatocytes, a biphasic curve was obtained following a 10-min incubation, with a Vmax of 0.36 nmol/min at bisphenol A concentrations of 20–30 μM [4.6–6.8 mg/L] and a Vmax of ∼0.15 nmol/min at bisphenol A concentrations of 2.5–10 nM [0.57–2.3 mg/L]. Table 47 summarizes the higher Vmax values obtained with cells from human, rat, and mouse livers. Total hepatic capacity was determined by multiplying Vmax by total number of hepatocytes/liver in vivo. [The only graphical data presented were for male F344 rats]. The authors noted that Vmax values were highest in mice>rats>humans. However, when adjusted for total hepatocyte number in vivo, the values were predicted to be highest in humans > rats > mice.

Table 46. Metabolites Obtained From Incubation of Human, Rat, and Mouse Hepatocyte Cultures With 20 μM [4.6 mg/L] Bisphenol Aa
 Percentage of parent compound or metabolites
Sex and speciesGlucuronide/sulfateSulfateGlucuronideBisphenol A
  • Human cells were incubated for 3 hr, and animal cells were incubated for 6 hr.

  • a

    aPritchett et al. (2002).

Human samples
 Female–140930
 Female–220842
 Female–3432550
 Male–110850
 Male–207.5750
Rodent samples
 Male F344 rat700300
 Female F344 rat100860
 Male Sprague–Dawley rat302580
 Female Sprague–Dawley rat001000
 Male CF1 Mouse001000
 Female CF1 mouse00930
Table 47. Rates of Bisphenol A Glucuronide Formation Following Incubation of Human, Rat, and Mouse Hepatocytes With Bisphenol Aa
Species and sexVmax, nmol/min/0.5×106 hepatocytesHepatic capacity, μmol/hrb
  • a

    aPritchett et al. (2002).

  • b

    bHepatic capacity was estimated by multiplying Vmax by total numbers of hepatic cells in vivo.

Human female0.278000
F344 rat female0.4646.5
F344 rat male0.3661.8
Sprague Dawley female0.3954.5
Sprague Dawley male0.4579.9
CF1 mouse female0.5013.8
CF1 mouse male0.8223.6

Data from Pritchett et al. (2002) appeared to be included in a series of unpublished studies by Sipes that were reviewed by the European Union (2003). In their review, the European Union noted that metabolic patterns appear to be similar in humans, rats, and mice. It was stated that the biphasic kinetic profile indicated involvement of a high-affinity glucuronidase enzyme at low concentrations and a high-capacity enzyme at high concentrations. In the interpretation of kinetic profiles in humans and experimental animals, the authors of the European Union report noted that the study calculations did not consider in vivo conditions such as varying metabolic capacity of hepatic cells, relationship of hepatic size to body size, and possibly important physiological endpoints such as blood flow. In addition, it was noted that calculations were based on limited data that did not address inter-individual variability in enzyme expression.

Cho et al. (2002) examined toxicokinetics of bisphenol A in mouse, rat, rabbit, and dog and used that information to predict toxicokinetic values in humans. Bisphenol A was administered by i.v. injection at 2 mg/kg bw to 5 male ICR mice and at 1 mg/kg bw to 7 male Sprague–Dawley rats, 7 male New Zealand White rabbits, and 5 male beagle dogs. Blood samples were drawn before dosing and at multiple time points between 2 min and 6 hr following injection. Serum bisphenol A concentrations were measured by HPLC. Toxicokinetic endpoints in animals are summarized in Table 48. The study authors noted that clearance and volume of distribution increased with increasing animal weight but that terminal half-life remained relatively constant across the different species. Simple allometric scaling and species-invariant time methods were used to predict values for a 70-kg human, and those values are summarized in Table 49. Regression analyses of estimates using the species-invariant time methods demonstrated that data from the 4 animal species were superimposable (r=0.94–0.949).

Table 48. Toxicokinetic Endpoints for Bisphenol A in Mice, Rats, Rabbits, and Dogs Intravenously Dosed With 2 mg/kg bw Bisphenol Aa
EndpointMousebRatRabbitDog
  • Data are presented as mean±SD.

  • a

    aCho et al. (2002).

  • b

    bVariances not reported.

Systemic clearance, L/hr0.31.9±0.412.6±4.927.1±8.0
Volume of distribution, L0.11.3±0.47.1±2.320.0±5.4
Half-life, min39.937.6±12.840.8±17.143.7±21.9
Table 49. Predicted Bisphenol A Toxicokinetic Endpoints in Humans Based on Results From Experimental Animal Studiesa
 Prediction method
EndpointAllometric scalingKallynochronsApolysichronsDienetichrons
  • a

    aCho et al. (2002).

Systemic clearance, L/hr127.1123120.746.0
Volume of distribution, L125.3229.7138.0149.3
Half-life, min43.6110.467.8196.2

Teeguarden et al. (2005) developed a physiologically based pharmacokinetic (PBPK) model for bisphenol A. Rat toxicokinetic data for the model were obtained from the studies by Pottenger et al. (2000) and Upmeier et al. (2000). Human toxicokinetic data were obtained from the study by Völkel et al. (2002). The model was developed to simulate blood and uterine concentrations of bisphenol A following exposure of humans through relevant routes. Correlations were determined for simulated bisphenol A binding to uterine receptors and increases in uterine wet weight, as determined by an unpublished study by Twomey. Although intestinal metabolism of bisphenol A to the glucuronide metabolite had been demonstrated recently, the model attributed bisphenol A metabolism entirely to the liver. Plasma protein binding was considered in both the rat and human model. The model accurately simulated plasma bisphenol A glucuronide concentrations in humans orally administered 5 mg bisphenol A, with the exception of underpredicting bisphenol A glucuronide concentrations at the 24–48-hr period following exposures. Cumulative urinary elimination of bisphenol A glucuronide in human males and females was simulated accurately. Less accurate simulations were observed for toxicokinetics in orally exposed rats, and the study authors indicated that a likely cause was oversimplification of the rat gastrointestinal compartment. Comparisons in metabolic clearance rates for i.v. and oral exposure suggested significant intestinal glucuronidation of bisphenol A. Enterohepatic recirculation strongly affected terminal elimination in rats but not humans. Consideration of bound versus unbound bisphenol A was found to be important in simulating occupancy of the estrogen receptor (ER) and uterine weight response. No increase in uterine weight was reported with simulated receptor occupancy of ∼1–15%. An increase in uterine weight was reported with ∼25% receptor occupancy, and doubling of uterine weight was reported with 63% receptor occupancy.

Shin et al. (2004) developed a PBPK model to predict the tissue distribution (lung, liver, spleen, kidneys, heart, testes, muscle, brain, adipose tissue, stomach, and small intestine) and blood pharmacokinetics of bisphenol A in rats and humans. The model was based on experimentally determined steady state blood-to-serum and tissue-to-blood partition ratios and does not include parameters to account for elimination via glucuronidation or differences in metabolism between rats and humans (e.g., enterohepatic circulation). Predicted concentration-time profiles were then compared to actual rat toxicokinetic data and to a profile for a simulated 70-kg human. Rat toxicokinetic information was obtained by administering multiple i.v. injections of bisphenol A (0.5 mg/kg) to adult male rats to achieve steady state. Bisphenol A concentrations were determined by a modified HPLC method with fluorescence detection. The authors noted good agreement between predicted and observed concentration-time profiles for blood and all tissues but did not present any statistical analysis or evaluate the performance of alternative models in order to establish goodness of fit. Based on the figures presented in the article, the PBPK model appeared to more accurately predict concentrations of bisphenol A in some tissues (e.g., blood, lung, and liver) better than others such as the small intestine and adipose tissue. The model was then applied to predict blood and tissue levels of bisphenol A in a 70 kg human after single i.v. injection (5-mg dose) and multiple oral administrations to steady state (100-mg doses every 24 hr). Tissue volumes and blood flow rates for a 70 kg human were taken from the literature. The authors concluded that simulated steady-state human blood levels (0.9–1.6 ng/ml) were comparable to blood levels of bisphenol A reported in the literature (1.49 ng/ml). In addition, the authors noted the similarity of predicted toxicokinetic endpoints obtained from their PBPK model to those predicted by Cho et al. (2002) based on simple allometric scaling on rat data.

2.2 General Toxicity, Estrogenicity, and Androgenicity

This section includes information on general toxicity as well as information on estrogenicity and androgenicity; however, results of estrogenicity and androgenicity testing are not considered a priori evidence of toxicity.

2.2.1 General toxicity.

The European Union (2003) reported there were no adequate studies for assessing acute toxicity of bisphenol A in humans.

In an acute toxicity study in rats orally dosed with bisphenol A at ≥2000 mg/kg bw, clinical signs included lethargy, prostration, hunched posture, and piloerection [reviewed by (European-Union, 2003)]. Gross signs in animals that died included pale livers and hemorrhage in the gastrointestinal tract. In a study in which male and female rats were subjected to whole body inhalation exposure to 170 mg/m3 bisphenol A dust for 6 hr, there were no gross signs of toxicity [reviewed by (European-Union, 2003)]. Effects observed in the respiratory tract at 2 but not 14 days following exposure included slight inflammation of nasal epithelium and slight ulceration of the oronasal duct. LD50 reported in studies with oral, dermal, inhalation, or i.p. exposure are summarized in Table 50. The European Union (2003) concluded that bisphenol A is of low acute toxicity through all exposure routes relevant to humans.

Table 50. LD50 s for Bisphenol A
SpeciesExposure routeLD50 (mg/kg bw)
  • a

    aNational Toxicology Program (NTP, 1982).

  • b

    bReviewed by the European Union (2003).

  • c

    cReviewed in ChemIDplus (2006).

RatOral3300–4100a
  5000b
  3250c
 Inhalation>170 mg/m3b
MouseOral4100–5200a
  2400c
 Intraperitoneally150c
Guinea pigOral4000c
RabbitOral2230bc
 Dermal>2000b
  3 mL/kgc

The European Union (2003) noted limited anecdotal data reporting skin, eye, and respiratory tract irritation in workers exposed to bisphenol A, but concluded that the reports were of uncertain reliability. It was noted that a recent, well-conducted study in rabbits demonstrated that bisphenol A is not a skin irritant. Other studies conducted in rabbits demonstrated eye irritation and damage, and it was concluded the bisphenol A can potentially cause serious eye damage. Slight respiratory tract inflammation occurred in rats inhaling ≥50 mg/m3 bisphenol A, and it was concluded that bisphenol A had limited potential for respiratory irritation. Based on the results of the studies described above, the European Union concluded that bisphenol A is not corrosive.

The European Union (2003) reviewed studies examining possible sensitization reactions in humans exposed to products containing bisphenol A, and those studies reported mixed results. In studies reporting positive findings, it was unclear if bisphenol A or epoxy resins were the cause of hypersensitivity. Cross-sensitization responses in individuals exposed to compounds similar to bisphenol A were also reported. Animal studies were determined unreliable for assessing sensitization. Based on the results of human studies, it was concluded that bisphenol A may have potential for sensitization in individuals exposed to resins. Human studies suggested that bisphenol A can induce dermal photosensitization responses. Photosensitization studies in mice resulted in reproducible positive results. Mechanistic studies in mice suggested that sensitization occurs through an immune-mediated process. The overall conclusion of the European Union was that it was somewhat unclear if bisphenol A induces orthodox skin sensitization, photosensitization, or responses in individuals previously sensitized to another substance, such as epoxy resins. No information was available on potential respiratory sensitization by bisphenol A.

The European Union (2003) summarized systemic toxicity reported in subchronic, chronic, and reproductive toxicity studies of rats, mice, and dogs. CERHR also reviewed the studies that examined reproductive organs, and those studies are summarized in detail in the appropriate section of this report. A relevant study by Yamasaki et al. (2002a) was published subsequent to the European Union review and was reviewed in detail by CERHR.

In studies reviewed by the European Union (2003) and in a study by Yamasaki et al. (2002a), rats were orally exposed to bisphenol A for periods of 28 days to 2 years. Cecal enlargement occurring at doses ≥25 mg/kg bw/day was the effect observed most frequently in those studies but was not considered toxicologically significant by the European Union. Histological alteration in the cecum consisting of mucosal hyperplasia was only reported in one study at doses ≥200 mg/kg bw/day. Histopathological changes in liver and kidney were reported at doses ≥500 mg/kg bw/day. The changes in liver were characterized by prominent hepatocyte nuclei or inflammation. Histopathology in kidney was characterized by renal tubule degeneration or necrosis. Testicular toxicity (degeneration of seminiferous tubules and arrested spermatogenesis) was observed in 1 study at doses ≥235 mg/kg bw/day.

The European Union (2003) found subchronic and chronic studies conducted by the NTP (NTP, 1982) to be the only reliable studies for assessing systemic toxicity in mice orally exposed to bisphenol A. The liver was found to be the target organ of toxicity, with multinucleated giant hepatocytes observed in male mice exposed to ≥120 mg/kg bw/day and female mice exposed to 650 mg/kg bw/day.

In a 90-day dietary study in dogs reviewed by the European Union (2003), an increase in relative liver weight with no accompanying histopathological alterations was found to be the only effect at doses ≥270 mg/kg bw/day. This finding was considered by the European Union to be of doubtful toxicological significance.

In a subchronic inhalation exposure study in rats reviewed by the European Union (2003), cecal enlargement as a result of distention by food was observed at ≥50 mg/m3. Also observed at ≥50 mg/m3 were slight hyperplasia and inflammation of epithelium in the anterior nasal cavity.

A limited number of repeat-dose systemic toxicity studies were summarized in detail by CERHR because they included examination of reproductive organs. Those studies are summarized in detail below.

NTP (1982), conducted acute, subacute, and subchronic bisphenol A toxicity studies in F344 rats and B6C3F1 mice. Animals were randomly assigned to treatment groups. Purity of bisphenol A was <98.2%. Concentration and stability of bisphenol A in feed were verified. In acute studies, single doses of bisphenol A in a 1.5% acacia vehicle were administered by gavage to 5 rats/group/sex at doses of 2150, 3160, 4640, or 6810 mg/kg bw/day and 5 mice/group/sex at 1470, 2150, 3160, 4640, 6810, or 10,000 mg/kg bw. LD50 values for that study are summarized in Table 50.

In a 14-day repeat dose study, survival and body weight gain were evaluated in 5 rats and mice/sex/group that were fed diets containing bisphenol A at 0, 500, 1000, 2500, 5000, or 10,000 ppm. Survival was unaffected by treatment. Weight gain was reduced by 60% or more in male rats exposed to ≥2500 ppm and 40% or more in female rats exposed to ≥5000 ppm bisphenol A. Survival and weight gain in mice were not affected by bisphenol A exposure.

In subchronic studies, 10 rats and mice/sex/group were exposed to bisphenol A in diet for 13 weeks. Dietary doses were 0, 250, 500, 1000, 2000, or 4000 ppm for rats and 0, 5000, 10,000, 15,000, 20,000, or 25,000 ppm for mice. A review by the European Union (2003) estimated bisphenol A intake at 0, 25, 50, 100, 200, and 400 mg/kg bw/day for rats, 0, 600, 1200, 1800, 2400, and 3000 mg/kg bw in male mice, and 0, 650, 1300, 1950, 2600, and 3250 mg/kg bw/day in female mice. Animals were observed and weighed during the study and killed and necropsied on Day 91 of the study. [Histopathological evaluations were conducted but it was not clear if all dose groups and all animals/dose group were examined. There was no mention of statistical analyses.] In rats, the only deaths occurred in 2/10 males of the 1000 ppm group. Weight gain was reduced by 18% or more in male rats and 10% or more in female rats exposed to ≥1000 ppm. There were no effects on feed intake. Hyaline masses in the bladder lumen were not observed in control male rats but were observed in 5 of 10 males exposed to 250 ppm, 3 of 10 exposed to 500 ppm, 3 of 10 exposed to 1000 ppm, 6 of 10 exposed to 2000 ppm, and 4 of 10 exposed to 4000 ppm. Cecal enlargement, which was observed in rats at a rate of 60–100% in each dose group with the exception of females exposed to 250 ppm was considered to be treatment-related. No histological alterations were observed in the cecum. Death in mice was limited to 2 of 10 females in the 5000 ppm group. Weight gain was reduced by at least 14% in male mice exposed to ≥15,000 ppm. Non-dose-related decreases in weight gain of 17% or more occurred in female mice of all dose groups. A dose-related increase in multinucleated giant hepatocytes was observed in all dose groups of male mice; the only incidence data reported for multinucleated giant hepatocytes were 0 of 10 female controls and 9 of 10 male mice of the 25,000 ppm group. [A complete set of data for histopathological findings was not presented for rats or mice.]

Yamasaki et al. (2002a) examined the effects of bisphenol A exposure on male and female CD rats in a study conducted according to Good Laboratory Practices (GLP). [Because this study included a number of reproductive organ and hormone endpoints, it is also discussed inSections 4.2.1.1and4.2.2.1.] Rats were fed a commercial diet (MF Oriental Yeast Co.) and housed in stainless steel wire-mesh cages. Rats were groups according to body weight and then randomly assigned to treatment groups. Ten 7-week-old rats/sex/group were gavaged with bisphenol A at 0 (olive oil vehicle), 40, 200, or 1000 mg/kg bw/day for 28 days. Due to the death of 1 animal exhibiting clinical signs in the 1000 mg/kg bw/day group, the high-dose was reduced to 600 mg/kg bw/day on Study Day 8. In an additional study, rats were exposed to ethinyl estradiol at 0, 10, 50, or 200 μg/kg bw/day for 28 days. Endpoints examined during the study were clinical signs, body weight gain, and food intake. Estrous cyclicity was examined in females for 2 weeks beginning on Study Day 15. Males were killed on Study Day 29 and females were killed in diestrus on Study Day 30, 31, or 32. Hematology and clinical chemistry endpoints were assessed, and blood hormone concentrations were measured by immunoassay systems. Sperm motility and viability were evaluated. Organs, including those of the reproductive system, were weighed and subjected to histopathological evaluation. With the exception of the testis and epididymis, which were fixed in Bouin solution, the organs were fixed in 10% neutral buffered formalin. Statistical analyses included Bartlett test for homogeneity of variance, ANOVA, Dunnett test, and/or Kruskall–Wallis test.

One female and 3 males from the high-dose group died; clinical signs observed in those animals included soft stools, decreased mobility, reduced respiration rate, and decreased body temperature. Soft stools were also observed in surviving males and females of the mid- and high-dose groups. Results of the study are summarized in Table 51. Terminal body weights were lower in females of the mid- and high-dose groups and males of the high-dose group. During the first week of study, food intake was decreased in both sexes of the mid- and high-dose group. [Data were not shown by study authors.] As noted in Table 51, some alterations in hematological and clinical chemistry endpoints were observed, mainly at the high-dose. [Data were not shown by study authors.] There were no treatment-related abnormalities in sperm or alterations in blood concentrations of thyroid hormones, follicle stimulating hormone (FSH), luteinizing hormone (LH), 17β-estradiol, prolactin, or testosterone. Number of females with diestrus lasting 4 or more days was increased in the high-dose group. Changes in relative organ weights [assumed to be relative to body weight] included decreased heart weight in females from the mid- and high-dose groups. At the high-dose, there were decreases in relative weight of ventral prostate and increases in relative weights of testis and adrenals in males and thyroid and liver in females. Gross signs observed in animals that died included enlarged kidney, elevated mucosa in the forestomach, and atrophied spleen and thymus. In surviving animals, the cecum was enlarged in the mid- and high-dose group and forestomach mucosa was elevated in the high-dose group. As described in more detail in Table 51, histopathological alterations were observed in the intestine, cecum, and colon of males and intestine and cecum of females in the mid- and high-dose groups. Additional histopathological alterations were observed in the high-dose group in the kidney, forestomach, and adrenals of males and females and livers of females.

Table 51. Toxicological Effects in Rats Gavaged With Bisphenol A for 28 Daysa
EndpointBisphenol A dose (mg/kg bw/day)
 40200600–1000c
  • a

    aYamasaki et al. (2002a).

  • b

    bData were not shown by study authors.

  • c

    cThe dose was 1000 mg/kg bw/day at the beginning of the study, but was decreased to 600 mg/kg bw/day in the second week of the study due to excessive toxicity.

    ↑,↓ Statistically significant increase, decrease compared to controls; ↔ no statistically significant effects compared to controls.

Males
 Terminal body weight↓ 17%
 Relative testes weight↑ 21%
 Relative ventral prostate weight↓ 28%
 Relative adrenal weight↑ 19%
 Feed intakeb
 Prothrombin timeb
 Glutamic-oxaloacetic transaminaseb
 Triglycerideb
 Alkaline phosphataseb
 γ-Glutamyl transpeptidaseb
 Chlorideb
 Renal tubular degeneration and necrosis0/100/107/7
 Forestomach squamous epithelial cell hyperplasia0/100/106/7
 Lacteal dilatation in duodenum0/1010/102/7
 Lacteal dilation in jejunum0/100/102/7
 Mucosal hyperplasia in cecum0/103/106/7
 Mucosal hyperplasia in colon0/102/107/7
 Adrenal cortical vacuolization0/100/103/7
Females
 Terminal body weight↓ 7%↓ 5%
 Relative thyroid weight↑ 22%
 Relative liver weight↑ 10%
 Relative heart weight↓ 9%↓ 15%
 Feed intakeb
 Hemoglobin and hematocrit valuesb
 Cholinesteraseb
 Glutamic-oxaloacetic transaminaseb
 Albumin and albumin: globulin ratsb
 Diestrus ≥4 days0/100/103/9
 Prominent hepatocyte nuclei0/100/104/9
 Renal tubular degeneration and necrosis0/100/109/9
 Forestomach squamous epithelial cell hyperplasia0/100/105/9
 Lacteal dilatation in duodenum0/107/106/9
 Mucosal hyperplasia in cecum0/106/104/9
 Adrenal cortical vacuolization0/100/103/9

Male rats from the mid- and high-dose ethinyl estradiol groups experienced decreased prostate, seminal vesicle, and pituitary weights, increased testis weight, and histopathological alterations in prostate, seminal vesicle, mammary gland, and testis. Females from the mid- and high-dose ethinyl estradiol group experienced alterations in estrous cyclicity. Females from the high-dose group experienced decreased ovary weight, increased uterine weight, and histopathological changes in ovary, uterus, and vagina.

General Electric (1984) conducted a subchronic toxicity study in Beagle dogs orally dosed with bisphenol A [purity not reported]. Dogs weighing 6.5–13.4 kg were housed in metal metabolism cages and fed Purina Dog Chow. During a 90-day period, 4 dogs/sex/group were given feed containing bisphenol A at 0, 1000, 3000, or 9000 ppm. The European Union (2003) estimated bisphenol A intake at 0, 28, 74, or 261 mg/kg bw/day in males and 0, 31, 87, or 286 mg/kg bw/day in females. Dogs were observed for body weight gain, food, intake, and clinical signs. Ophthalmoscopic examination was conducted before and following the treatment period. Hematology, clinical chemistry, and urinalysis evaluations were conducted before treatment and at 1, 2, and 3 months into the study. Dogs were killed at the end of the treatment period. Organs were weighed and fixed in 10% neutral buffered formalin. Histopathological evaluations were conducted in organs from the control and high-dose groups; prostate, uterus, testis, and ovary were among organs evaluated. [Procedures for statistical analyses were not described.] No treatment-related clinical signs (conducted monthly), ophthalmological changes, or death were observed during the study. Bisphenol A treatment did not affect body weight gain or food intake. There were no treatment-related effects on hematology, biochemistry, or urinalysis. Relative liver weight was significantly increased [by 18% in males and 26% in females] in the high-dose group, and the study authors considered the effect to be treatment-related. No treatment-related gross or histopathological lesions were observed in the high-dose group.

Nitschke et al. (1988) conducted a subchronic inhalation toxicity test with bisphenol A in F344 rats. Rats were fed Purina Certified Rodent Chow 5002 and housed in stainless steel wire cages. At 7 weeks of age, rats were stratified according to body weight and randomly assigned to treatment groups. Thirty rats/sex/group received whole-body exposures to polycarbonate grade bisphenol A dust (99.7% purity) at 0, 10, 50, or 150 mg/m3 for 6 hr/day, 5 days/week, for 13 weeks. Mass median aerodynamic diameter of bisphenol A dust was measured at ≤5.2 microns. Stability and concentrations of bisphenol A were verified. Rats were observed for clinical signs, body weight gain, and food intake. Ten rats/sex/group in each time period were killed and necropsied on the day following and at 4 and 12 weeks following exposure. At each necropsy period, hematological and clinical chemistry endpoints were examined. The lungs, brain, kidneys, and testes were weighed. Numerous organs were preserved in 10% phosphate-buffered formalin. In most cases, histological examinations were conducted in organs from the control and high-dose groups. Respiratory organs and organs with lesions or signs of toxicity were histologically examined at all dose levels. Included among organs undergoing histopathological examination immediately after the exposure period were the epididymis, mammary gland, ovary, oviduct, prostate, seminal vesicles, testis, uterus, and vagina. No reproductive organs were examined following the recovery periods. Statistical analyses included Bartlett's test, ANOVA, Dunnett test, Wilcoxon Rank-Sum test, and Bonferroni correction for multiple comparisons. Gross pathology and histopathology data did not appear to have been statistically analyzed.

During the exposure period, a reddish material around the nose (most likely porphyrin) was observed in 2–10 of 10 animals/sex in the 50 and 150 mg/m3 groups. Perineal soiling was observed in 2 of 10 females in the 10 mg/m3 group and 9–10 of 10 animals/sex in the 50 and 150 mg/m3 groups. Decreased body weight gain during treatment was observed in males from all dose groups and females in the 50 and 150 mg/m3 groups. Immediately following the treatment period, terminal body weights were reduced by ∼5% in males and ∼11% in females from the 150 mg/m3 group. [Body weights were ∼4% lower in males from the 50 mg/m3 group.] No differences in feed intake were observed at this or any other time period in the study. The only hematological effect observed was slightly increased hemoglobin in males exposed to 10 mg/m3, but the study authors did not consider the effect to be biologically significant. Clinical chemistry observations in the 150 mg/m3 group included decreased serum glutamic pyruvic transaminase activity, serum glutamic oxaloacetic transaminase activity, and glucose in males and decreased total protein and albumin and increased alkaline phosphatase activity in females. Alkaline phosphatase activity was also increased in females exposed to 50 mg/m3. The study authors did not consider any of the clinical chemistry changes to be biologically [toxicologically] significant. Absolute liver weight was decreased in males exposed to ≥10 and 150 mg/m3, and relative brain weight was increased in females exposed to ≥50 mg/m3. Additional organ weight changes observed in females from the 150 mg/m3 group included decreased absolute liver and kidney weights and increased relative lung weights. Because the organ weight changes were not associated with microscopic changes in organs, the study authors concluded that the effects reflected decreases in body weight and were not toxicologically significant. Cecal size was increased as a result of distention by food in all (10/dose/sex) males and females exposed to ≥50 mg/m3, and the effect was considered to be treatment-related. No histopathological alterations were observed for cecal wall morphology. Hemolyzed blood was observed in the stomachs of three to seven of 10 males/group exposed to 50 and 150 mg/m3, but there were no signs of histopathological alterations in the gastrointestinal tract. Slight histopathological alterations, consisting of hyperplasia in stratified squamous and ciliated epithelium lining and inflammation of submucosal tissues was observed in the anterior nasal cavities of all (10/dose/sex) males and females exposed to ≥50 mg/m3. Slight-to-moderate hyperplasia of goblet cells was also observed in the lateral nasal wall. No other treatment-related histopathological alterations were observed, including in reproductive organs.

During the 4-week recovery period, body weights remained lower in males and females of the 50 and 150 mg/m3 groups At 4 weeks following exposure, terminal body weights of males and females in the 150 mg/m3 group were ∼6% lower than control values. A decrease in white blood cell count in females from the 10 and 150 mg/m3 groups was the only hematological effect observed. The clinical chemistry effects that were somewhat consistent with effects observed immediately following treatment were increased alkaline phosphatase activity in females exposed to 10 and 150 mg/m3 and decreased serum glutamic pyruvic activity transaminase activity in females exposed to 150 mg/m3; the study authors did not consider the clinical chemistry changes to be treatment-related. The study authors concluded that an increase in relative brain weight in males of the 150 mg/m3 group was related to decreased body weights in those animals. Enlarged cecal size was observed in 5 of 10 males of the 150 mg/m3 group, a decreased incidence compared to the period immediately following treatment. Nasal histopathology was observed in the 150 mg/m3 but was reduced in magnitude and severity compared to rats observed immediately following exposure.

In rats examined following 12 weeks of recovery, body weights of males in the 150 mg/m3 group remained lower than controls, and terminal body weight was decreased by ∼6%. An increase in white blood cell counts but not differential counts was observed in male rats of the 10 and 150 mg/m3 group. The only clinical chemistry finding consistent with earlier observations was decreased total protein and globulin in females from the 150 mg/m3 group, but the study authors did not consider the effect to be biologically significant. Organ weight changes in the 150 mg/m3 group included decreased absolute kidney and lung weights in males and decreased absolute and relative kidney weights in females. No histopathological alterations were observed in kidney or lung. No other gross or histopathological alterations were observed, including cecal enlargement and nasal histopathology, which were observed at earlier time periods.

2.2.2 Estrogenicity.

The first identification of bisphenol A as an estrogen has been attributed to Dodds and Lawson (1936), who reported that 100 mg injected by an unspecified route twice daily for 3 days resulted in maintenance of 5 of 5 rats in vaginal estrus for 40 days. The estrogenicity of bisphenol A has since been evaluated using several different kinds of assays. In vitro studies are summarized in Table 52, and in vivo studies are summarized in Table 53 using comparisons with 17β-estradiol, ethinyl estradiol, diethylstilbestrol, and, in one study, estrone. There is considerable variability in the results of these studies with the estrogenic potency of bisphenol A ranging over about 8 orders of magnitude, but similar means (Fig. 2).

Figure 2.

In vitro estrogenic potency (log10) in ER α and β binding and transcriptional assays and estrogen-dependent cell proliferation assays) distributions of bisphenol A and estrogen responses in vivo in rats, mice and fish. Each data point represents one bisphenol A study in which bisphenol A was compared to a reference estrogen in rats, mice, fish, or in vitro. Data summarized from Table 52 and Table 53, midrange values used when a range is given in the table.

Table 52. In Vitro Estrogenicity Testing of Bisphenol A
EndpointMolar potency relative to 17β-estradiolReference
  • a

    aProgesterone receptor was increased and androgen receptor was decreased by 17β-estradiol 10−10 M.

Binding assays
 Frog liver cytosol binding[1.4×10−3]Lutz and Kloas (1999)
 Carp liver cytosol binding[1.3×10−3]Segner et al. (2003)
 Rainbow trout ER binding5.8×10−5Olsen et al. (2005)
 Rainbow trout ER binding2.1×10−3Matthews et al. (2000)
 Anole ER binding1.3×10−3Matthews et al. (2000)
 Chicken ER binding4.4×10−4Matthews et al. (2000)
 Mouse ERα binding8.6×10−5Matthews et al. (2000)
 Mouse uterine cytosol binding[1.2×10−4]Matthews et al. (2001)
 Rabbit uterine ER binding[1.3×10−5]Andersen et al. (1999a)
 Rat uterine cytosol binding∼5×10−4Krishnan et al. (1993)
 Rat uterine cytosol binding8×10−5Blair et al. (2000)
 Rat uterine cytosol binding1–2×10−4Kim et al. (2001a)
 Rat ERα binding[2.5×10−4]Strunck et al. (2000)
 ER binding in rat lactotroph1–10×10−5Chun and Gorski (2000)
 Rat ERα binding5×10−4Kuiper et al. (1997)
 Rat ERβ binding3.3×10−4Kuiper et al. (1997)
 Rat uterine ERα and β binding6.2×10−5Washington et al. (2001)
 Rat uterine Type II estrogen-binding site4×10−3Washington et al. (2001)
 ER binding in MCF-7 lysates1×10−2Dodge et al. (1996)
 Human ERα binding4×10−4Bolger et al. (1998)
 Human ERα binding1×10−4Kuiper et al. (1998)
 Human ERβ binding1×10−4Kuiper et al. (1998)
 Human ER binding5.6×10−4Perez et al. (1998)
 Human ER binding[1.3×10−4]Andersen et al. (1999a)
 ER binding in ECC-1 cells3×10−3Bergeron et al. (1999)
 Human ERα binding8×10−5Matthews et al. (2000)
 Human ERα binding[2.5×10−3 diethylstilbestrol]Nakagawa and Suzuki (2001)
 Human ERα binding7.3×10−4Routledge et al. (2000)
 Human ERβ binding7.5×10−3Routledge et al. (2000)
 Human ER binding[7.1×10−5]Sheeler et al. (2000)
 Human ERα binding[8×10−5]Matthews et al. (2001)
 Human ERβ binding[3.8×10−3]Matthews et al. (2001)
 Human ERα binding5×10−2Paris et al. (2002)
 Human ERβ binding4×10−2Paris et al. (2002)
 Human ER binding[3×10−4]Stroheker et al. (2004)
 Human ERα binding[2.4×10−4]Seidlová-Wuttke et al. (2005)
 Human ERβ binding[2.8×10−2]Seidlová-Wuttke et al. (2004)
 Human ERα binding[1.1×10−4]Takemura et al. (2005)
 Human ERβ binding[4.4×10−4]Takemura et al. (2005)
 Human ER binding3.15×10−3Olsen et al. (2005)
 ERα binding[9.4×10−4]Takayanagi et al. (2006)
 ERβ binding[9.6×10−4]Takayanagi et al. (2006)
Recombinant yeast reporter systems
 Human ER activation5×10−5Coldham et al. (1997)
 Human ER activation6.7×10−5Gaido et al. (1997)
 Human ER activation[2.5×10−5]Harris et al. (1997)
 Human ER activation[4–8×10−5]Andersen et al. (1999a)
 Human ER activation[3.9×10−5]Sheeler et al. (2000)
 Human ER activation∼1×10−4Sohoni and Sumpter (1998)
 Human ER activation3.7×10−5Metcalfe et al. (2001)
 ERα activation6.2×10−5Silva et al. (2002)
 ERα activation[1×10−4]Nishihara et al. (2000)
 ERα activation[∼1×10−4]Beresford et al. (2000)
 Human ERα[3.3×10−5]Rajapakse et al. (2001)
 Human ERα, no microsomes[5.5×10−5]Elsby et al. (2001)
 Human ERα, human liver microsomes[6.6×10−6]Elsby et al. (2001)
 ER activation∼10−5Chen et al. (2002)
 Human ER activation[8.1×10−5]Segner et al. (2003)
 Human ER activation9×10−5Li et al. (2004)
 ERα activation[4×10−5]Singleton et al. (2006)
 Human ERα, with denatured rat S9[2.4×10−6]Yoshihara et al. (2004)
 Human ERα, with active rat S9[9.2×10−6]Yoshihara et al. (2004)
 Human ERα, with denatured mouse S9[3.0×10−6]Yoshihara et al. (2004)
 Human ERα, with active mouse S9[7.8×10−6]Yoshihara et al. (2004)
 Human ERα, with denatured monkey S9[2.4×10−6]Yoshihara et al. (2004)
 Human ERα, with active monkey S9[6.0×10−6]Yoshihara et al. (2004)
 Human ERα, with denatured human S9[2.2×10−6]Yoshihara et al. (2004)
 Human ERα, with active human S9[4.6×10−6]Yoshihara et al. (2004)
 Human ERα activity[2.3×10−5]Terasaki et al. (2005)
 Medaka ERα activity[3.3×10−4]Terasaki et al. (2005)
 “Estrogenic activity”3.4×10−5Kawagoshi et al. (2003)
 ERα activation[2.3×10−4]Singleton et al. (2006)
 Fish ERα activation4.1×10−4Fu et al. (2007)
 Fish ERβ2 activation3.2×10−5Fu et al. (2007)
Other cell-based recombinant reporter systems
 ER activation in trout gonad cell line5.4×10−3Ackerman et al. (2002)
 Mouse ERα in HeLa cells[<1×10−5]Ranhotra and Teng (2005)
 Mouse ERβ in HeLa cells[∼1×10−2]Ranhotra and Teng, 2005
 HepG2 cells, human ERα[3.0×10−3]Snyder et al. (2000)
 HepG2 cells, human ERβ[1.1×10−2]Snyder et al. (2000)
 Rat ERα in HeLa cells[1.6×10−7]Yamasaki et al. (2002b)
 ER activation in HeLa cells[8.8×10−4]Takahashi et al. (2004)
 ERα activation in HeLa cells[2.5×10−2]Hiroi et al. (1999)
 ERβ activation in HeLa cells[2.3×10−2]Hiroi et al. (1999)
 ERα activation in HeLa cells[6.1×10−1]Vivacqua et al. (2003)
 ERβ activation in HeLa cells[5.6×10−1]Vivacqua et al. (2003)
 ERα activation in HeLa cells[7.7×10−1]Recchia et al. (2004)
 ERβ activation in HeLa cells[1.2]Recchia et al. (2004)
 ERα activation in T47D cells[6.2–7.9×10−1]Recchia et al. (2004)
 Proliferation in T47D cells[6.6×10−1]Recchia et al. (2004)
 Human ER in hepatoma cells[3×10−2]Gould et al. (1998)
 Human ERα, human embryonal kidney[4.8×10−3]Kurosawa et al. (2002)
 Human ERβ, human embryonal kidney[4.6×10−3]Kurosawa et al. (2002)
 Human ERα, endometrial carcinoma[5.4×10−3]Kurosawa et al. (2002)
 Human ERβ, endometrial carcinoma[4.9×10−3]Kurosawa et al. (2002)
 Human ERα, osteosarcoma[7.3×10−3]Kurosawa et al. (2002)
 Human ERβ, osteosarcoma[7.7×10−3]Kurosawa et al. (2002)
 Human ERα, human hepatoma cells[2.7×10−1]Gaido et al. (2000)
 Human ERβ, human hepatoma cells[1.8×10−1]Gaido et al. (2000)
 Human ERα, 239HEK cells2×10−4 diethylstilbestrolLemmen et al. (2004)
 Human ERβ, 239HEK cells7×10−4 diethylstilbestrolLemmen et al. (2004)
 Human ERα, endometrial carcinoma[6.1×10−3]Singleton et al. (2006)
MCF-7 cells
 G6PD activity[1×10−1]Kim et al. (2003a)
 Expression of proteins[1×10−3]Perez et al. (1998)
 Progesterone receptor mRNANot increased at 10−6 MaDiel et al. (2002)
 Androgen receptor mRNANot decreased at 10−6 MaDiel et al. (2002)
 Progesterone receptor∼2×10−4Krishnan et al. (1993)
 ER binding, serum-free3.3×10−4Samuelsen et al. (2001)
 ER binding, 100% human serum1.7×10−4Samuelsen et al. (2001)
 ER binding3.2×10−3Olsen et al. (2003)
 ER activation[1.4×10−5]Kitamura et al. (2005)
 ERα expression[7.5×10−5]Matthews et al. (2001)
 ERβ expression[1.8×10−4]Matthews et al. (2001)
 ERα activation[4.7–6.9×10−1]Vivacqua et al. (2003)
 ERα activation[5.5–6.7×10−1]Recchia et al. (2004)
 pS2 induction[1.8×10−6]Leffers et al. (2001)
 ER production[7×10−8]Olsen et al. (2003)
 Progesterone receptor production[6.8×10−8]Olsen et al. (2003)
 pS2 production[10−7]Olsen et al. (2003)
 pS2 mRNA[1.1]Vivacqua et al. (2003)
 pS2 mRNA[8.9×10−1]Recchia et al. (2004)
 Cathepsin D mRNA[8.2×10−1]Recchia et al. (2004)
 Transcription of human telomerase reverse transcriptase[∼10−2]Takahashi et al. (2004)
 Proliferation[3.8×10−4]Krishnan et al. (1993)
 Proliferation1×10−3Brotons et al. (1995)
 Proliferation1×10−4Soto et al. (1997)
 Proliferation[∼1×10−3]Dodge et al. (1996)
 Proliferation[1×10−4]Perez et al. (1998)
 Proliferation[9.8×10−4]Schafer et al. (1999)
 Proliferation (3 different laboratories)5–100×10−7Andersen et al. (1999a)
 Proliferation6×10−5Körner et al. (2000)
 Proliferation3×10−5Kim et al. (2001a)
 Proliferation[2.5×10−6]Suzuki et al. (2001)
 Proliferation2×10−5Samuelsen et al. (2001)
 Proliferation[9.2×10−4]Nakagawa and Suzuki (2001)
 Proliferation[∼1×10−3]Shimizu et al. (2002)
 Proliferation[7×10−9]Diel et al. (2002)
 Proliferation1.6×10−5Olsen et al. (2003)
 Proliferation[4.5–5×10−1]Vivacqua et al. (2003)
 Proliferation[1.1×10−4]Stroheker et al. (2004)
 Proliferation[6×10−1]Recchia et al. (2004)
 Proliferation2×10−5Olsen et al. (2005)
 Proliferation, with denatured rat S9[6.5×10−5]Yoshihara et al (2001)
 Proliferation, with active rat S9[3.4×10−4]Yoshihara et al (2001)
Rat pituitary cells
 Proliferation1–10×10−6Chun and Gorski (2000)
 Proliferation[∼8.4×10−3]Steinmetz et al. (1997)
 Prolactin release1×10−5Chun and Gorski (2000)
 Prolactin release (GH3 cell)[6×10−3]Steinmetz et al. (1997)
 Prolactin release (F344 pituitary)2–10×10−4Steinmetz et al. (1997)
 Prolactin gene expression[∼1×10−3]Steinmetz et al. (1997)
Rat uterine adenocarcinoma cells
 Induction of complement C3 mRNA[8×10−3]Strunck et al. (2000)
Human uterine adenocarcinoma cells
 Progesterone receptor mRNA/protein[∼1×10−2]Bergeron et al. (1999)
 ProliferationNo effect at 10−5 MBergeron et al. (1999)
Vitellogenin production, fish hepatocytes
 Carp1×10−4Smeets et al. (1999)
 Carp[3.1×10−3]Segner et al. (2003)
 Carp[1×10−5]Letcher et al. (2005)
 Carp[3×10−4]Rankouhi et al. (2002)
 Trout2×10−5Shilling and Williams (2000)
 Trout[8×10−4]Segner et al. (2003)
 Trout2.9×10−5Olsen et al. (2005)
Frog hepatocytes
 Vitellogenin mRNA expression[∼1×10−3]Kloas et al. (1999)
 Vitellogenin productionNo effect at 100 μMRankouhi et al. (2004)
 ER mRNA expression∼10−2Lutz et al. (2005)
Table 53. In Vivo Estrogenicity Tests of Bisphenol A
Model and exposureHusbandryaEndpointMolar potency/comparatorbReference
  • a

    aHusbandry information for rodent studies includes caging and bedding materials and diet when indicated by the authors.

  • b

    bEstimates include comparison of administered dose, magnitude of effect, and molecular weight.

Rat uterus
 Adult ovariectomized Sprague–Dawley, gavage×4 daysTD89222 diet, metal cageUterine wet weight[3.9×103]/ethinyl estradiolDodge et al. (1996)
 Immature Sprague–Dawley, bisphenol A given “orally”×3 days; 17β-estradiol i.p.×3 daysNot indicatedUterine weightNot affected by bisphenol A at up to 150 mg/kg bw/day; 17β-estradiol was positive at 0.005 mg/day [0.089 mg/kg bw/day]Gould et al. (1998)
  Progesterone receptor[5.9×103]/17β-estradiol 
  Peroxidase activity[7.6×103]/17β-estradiol 
 Adult ovariectomized Crl:CD BR, gavage×4 daysPurina 5002 diet, steel cageUterine weight[3.5×105]/17β-estradiolCook et al. (1997)
  Stromal cell proliferation[4.1×105]/17β-estradiol 
 Adult ovariectomized F344, i.p.×1Not indicatedcfos expression[2.1×104]/17β-estradiolSteinmetz et al. (1998)
 Adult ovariectomized F344 or Sprague–Dawley, silastic implant×3 daysNot indicatedUterine wet weight: Steinmetz et al. (1998)
  F344[8.2×103]/17β-estradiol 
  Sprague–Dawley[6.0×103]/17β-estradiol 
  Uterine cell height:  
  F344[1.1×102]/17β-estradiol 
  Sprague–Dawley[9.2×103]/17β-estradiol 
 Juvenile ovariectomized DA/Han, Wistar, or Sprague–Dawley, gavage×3 daysNot indicatedUterine wet weight: Diel et al. (2004)
  DA/Han[1.8×105]/ethinyl estradiol 
  WistarNo response to 200 mg/kg/d 
  Sprague–Dawley[1.7×105]/ethinyl estradiol 
  Uterine epitheliumNo response to 200 mg/kg/day 
  Vaginal epitheliumNo response to 200 mg/kg/day 
  Clusterin mRNANo response to 200 mg/kg/day 
 Immature Alpk:AP, s.c.×3 daysRM3 diet, wire cageUterine wet weight[2.6–2.7×105]/diethylstilbestrolAshby and Tinwell (1998)
  Uterine dry weight[2.5–3.0×105]/diethylstilbestrol 
 Immature Alpk:AP, gavage×3 daysRM3 diet, wire cageUterine wet weight[2.3–3.1×105]/diethylstilbestrol 
  Uterine dry weight[2.7–3.6×105]/diethylstilbestrol 
 Immature Long–Evans, gavage×3 daysPurina 5001 dietUterine wet weight 6 hr after dosing[1.4×105]/17β-estradiolLaws et al. (2000)
  Uterine wet weight 24 hr after dosingNo effect at bisphenol A at ≤400 mg/kg bw/day 
 Adult ovariectomized Long–EvansPurina 5001 dietUterine wet weightNo effect of bisphenol A at ≤100 mg/kg bw/dayLaws et al. (2000)
 Juvenile ovariectomized DA/Han, gavage×3 daysSsniff R-10 dietUterine wet weight relative to bw[1.2×105]/ethinyl estradiolDiel et al. (2000)
  Expression of:  
  Androgen receptor[3.9×104]/ethinyl estradiol 
  ER[1.9×104]/ethinyl estradiol 
  Progesterone receptorbisphenol A and ethinyl estradiol produced opposite effects 
  Complement C3[2.2×105]/ethinyl estradiol 
  ClusterineNo bisphenol A effect at 200 mg/kg bw/day; ethinyl estradiol showed an effect at 0.1 mg/kg bw/day 
  Glyceraldehyde phosphate dehydrogenase  
 Adult ovariectomized Alpk:ApfSD, s.c.×3 daysNot indicatedUterine wet weight[1.7×104]/17β-estradiolAshby et al. (2000)
  Uterine dry weight[1.8×104]/17β-estradiol 
 Immature Crj:CD (SD), s.c.×3 daysMF diet, steel cageWet and blotted uterine weightEffect noted at ≥8 mg/kg bw/day bisphenol A/no comparatorYamasaki et al. (2000)
 Immature Crj:CD (SD), gavage×3 daysMF diet, steel cageWet and blotted uterine weightEffect noted at ≥160 mg/kg bw/day bisphenol A/no comparator 
 Adult ovariectomized Wistar, s.c.×7 daysNot indicatedBlotted uterine weightIncreased relative weight compared to placebo at ≥11 mg/kg bw/day; uterus reached 83% of weight of sham-ovariectomized control at bisphenol A dose of 250 mg/kg bw/day.Goloubkova et al. (2000)
 Adult ovariectomized Sprague–Dawley, exposed in drinking water×3 daysGlass water bottles, plastic cage (negative E-Screen of ethanol cage washes)Uterine wet weightNo effect of bisphenol A at up to 16.9 mg/kg bw/day; estrone positive at 0.12 mg/kg bw/dayRubin et al. (2001)
 Adult ovariectomized Sprague–Dawley, s.c.×3 daysPMI Certified Rodent Diet, polycarbonate cage, elm beddingUterine wet weight[1.7×106]/17β-estradiolKim et al. (2001a)
  Uterine dry weight[2.3×106]/17β-estradiol 
 Immature Alpk:ApfSD, s.c.×3 daysRM1 dietUterine wet weight[2.9×104]/17β-estradiolMatthews et al. (2001)
  Uterine dry weightNo effect of bisphenol A at 800 mg/kg bw/day; 17β-estradiol positive at 0.4 mg/kg bw/day 
 Immature Alpk:ApfSD, gavage×3 daysRM1 dietUterine wet weight[2.3–5.5×104]/17β-estradiol 
  Uterine dry weight[2.4–7.1×104]/17β-estradiol 
 Immature Sprague–Dawley, s.c.×3 daysSoy-free diet, polycarbonate cageUterine wet weightNo effect of bisphenol A at ≤1000 mg/kg bw/day; 17β-estradiol was positive at 0.04 mg/kg bw/dayAn et al. (2002)
  Calbindin D9k expression[8.4×106]/17β-estradiol 
  ERα expression[3.4×105]/17β-estradiol 
 Immature Crj:CD (SD), s.c.×3 daysMF diet, steel cageUterine wet weight[5.1×105]/ethinyl estradiolYamasaki et al. (2002b)
 Immature Sprague–Dawley, s.c.×3 daysSoy-free diet, polycarbonate cage, corncob beddingBlotted uterine weight[8×107]/ethinyl estradiolWade et al. (2003)
  Epithelial cell height[1.2×106]/ethinyl estradiol 
 Pubertal Sprague–Dawley, gavage PND 22–42/43Purina 5002 diet, polycarbonate cage, chip beddingBlotted uterine weightAbsolute organ weight decreased with increase dose (400 and 600 mg/kg bw/day); no effect on relative organ weightGeorge et al. (2003)
  Vaginal openingNo effect at 400 and 600 mg/kg bw/day 
 Pregnant Sprague–Dawley, s.c. bisphenol A on GD 17–19 (17β-estradiol s.c.×1)Soy-free diet, polycarbonate cageMaternal uterine weight[1.8×105]/17β-estradiolHong et al. (2003)
 Pregnant Sprague–Dawley, s.c. bisphenol A on GD 17–19 (17β-estradiol s.c.×1)Soy-free diet, polycarbonate cageMaternal uterine calbindin D9k protein[1.7×105]/17β-estradiolHong et al. (2003)
 Lactating Sprague–Dawley, s.c. bisphenol A×5 days (17β-estradiol s.c.×1)Soy-free dietMaternal uterine calbindin D9k mRNA calbindin D9k protein [2.2×105]/17β-estradiol [6.9×105]/17β-estradiolHong et al. (2004)
 Immature and adult ovariectomized Wistar, gavage×4 daysAO4C diet, wire cageUterine wet and dry weightNo effect in either model of bisphenol A at ≤200 mg/kg bw/day/17β-estradiol positive at 0.025–0.035 mg/kg bw/dayStroheker et al. (2003)
 Immature Sprague–Dawley, s.c.×3 daysSoy-free feed, polycarbonate cageCalbindin D9k protein[5.1×105]/17β-estradiolAn et al. (2003)
 Immature Sprague–Dawley, s.c.×3 daysShinchon dietUterine wet weight[1.5×106]/17β-estradiolKim et al. (2003a)
  Uterine wet weight relative to bw[1.3×106]/17β-estradiol 
  Glutathione peroxidase activity[4.2×103]/17β-estradiol 
 Immature Alpk:ApfSD, gavage×3 daysRM1 diet, polycarbonate cageBlotted uterine weight[2.5×104]/17β-estradiolAshby and Odum (2004)
  Expression of:  
  Progesterone receptor A[3.8×104]/17β-estradiol 
  Progesterone receptor B[4.2×104]/17β-estradiol 
  Complement C3[1.8×104]/17β-estradiol 
  Lipocalcin[2.3×104]/17β-estradiol 
 Immature AP, s.c.×3 daysRM1 diet, polypropylene cages, sawdust and shredded paper beddingUterine wet weight[1.0×106]/ethinyl estradiolTinwell and Ashby (2004)
  Uterine dry weight[1.2×106]/ethinyl estradiol 
 Adult ovariectomized Sprague–Dawley, diet×3 monthsPhytoestrogen-free dietUterine weight, endometrial thickness, ERα, ERβ expressionNo bisphenol A effect at 0.37 mg/kg bw/day; estradiol benzoate positive controlSeidlová-Wuttke et al. (2004)
  Complement C3 expressionBisphenol A and estradiol benzoate produced opposite effects 
 Immature Sprague–Dawley, s.c.×3 daysPMI Certified Rodent DietUterine wet weight[4.5×107]/ethinyl estradiolKim et al. (2005)
  Uterine dry weight[4.9×107]/ethinyl estradiol 
 Adult ovariectomized Crj:CD (SD), s.c.×3 daysEstrogen-free NIH-07PLD diet, aluminum cage, paper beddingUterine wet weight, relative to bw[2.1×105]/17β-estradiolKoda et al. (2005)
  Blotted uterine weight, relative to bw[1.7×106]/17β-estradiol 
 Adult Holzman, progesterone-treated to delay implantation, given test agent s.c. on GD 7Unspecified Purina rodent chow, plastic cage, pine shavingsImplantation[4–34×106]/estroneCummings and Laws (2000)
Rat vagina
 Adult ovariectomized F344, i.p.×1Not indicatedBrdU labelingIncreased at bisphenol A dose of 37.5 but not 18.5 mg/kg bw/no comparatorSteinmetz et al. (1998)
  cfos expression[1.3×104]/17β-estradiol 
 Adult ovariectomized Long Evans, bisphenol A by gavage×11 days; 17β-estradiol by s.c.Purina 5001 dietVaginal cytologyNo effect at bisphenol A dose of 100 mg/kg bw/day; 17β-estradiol 0.005 mg/kg bw/day resulted in persistent estrusLaws et al. (2000)
 Long–Evans treated PND 21–35 by gavagePurina 5001 dietVaginal openingNo effect at bisphenol A dose ≤400 mg/kg bw/day; ethinyl estradiol was active at 0.01 mg/kg bw/dayLaws et al. (2000)
 Adult ovariectomized F344 and Sprague–Dawley, i.p.×1Not indicatedBrdU labelingF344: [4.5×106]/17β-estradiol Sprague–Dawley: [1.4×106]/17β-estradiolLong et al. (2000)
 Immature Wistar, gavage×4 daysAO4C diet, wire cageVaginal cornification[3.8×104]/17β-estradiol 
 Adult ovariectomized Wistar, gavage×4 daysAO4C diet, wire cageVaginal cornificationNo effect at bisphenol A dose of 100 mg/kg bw/day; 17β-estradiol was positive at 0.1 mg/kg bw/dayStroheker et al. (2003)
 Immature Sprague–Dawley, s.c.×3 daysPMI Certified Rodent DietVaginal weight[5.3×107]/ethinyl estradiolKim et al. (2005)
Other rat organs
 Ovariectomized Sprague–Dawley, daily gavage for 5 weeksTD89222 diet, metal cagePrevention of bone mineral density declineNo effect at bisphenol A dose up to 10 mg/kg bw/day; no standard estrogen comparatorDodge et al. (1996)
 Adult ovariectomized Sprague–Dawley, treated in feedPhytoestrogen-free dietPrevention of bone mineral density declineNo effect at bisphenol A dose ≤ 370 μg/kg bw/day; estradiol benzoate was effective at 1.18 mg/kg bw/daySeidlová-Wuttke et al. (2004)
 Adult ovariectomized Sprague–Dawley and F344, by s.c. implant×3 daysNot indicatedSerum prolactinF344: [1.7×102]/17β-estradiol Sprague Dawley: no effect of bisphenol A at 40–45 μg/day or 17β-estradiol at 1.2–1.5 μg/daySteinmetz et al. (1997)
 Adult ovariectomized Wistar, s.c.×7 daysNot indicatedPituitary weightIncreased compared to vehicle control at 128 but not 78 mg/kg bw/dayGoloubkova et al. (2000)
  Serum prolactinIncreased compared to vehicle control at 128 mg/kg bw/day 
Mouse uterus
 Immature CFLP, s.c.×3 daysNot indicatedRelative uterine weightNo response at up to 0.5 mg [50 mg/kg bw/day]Coldham et al. (1997)
 Adult ovariectomized CD-1, s.c.×1Not indicatedIGF1 expression[8.4×104]/17β-estradiolKlotz et al. (2000)
 Juvenile-adult ovariectomized B6C3F1, s.c.×4 daysPurina 5001, polypropylene cage, chip beddingUterine wet weight[2.3×105]/17β-estradiolPapaconstantinou et al. (2000)
  Endothelial proliferation[6.9×106]/17β-estradiol 
 Juvenile-adult ovariectomized B6C3F1, s.c.×4 daysPurina 5001, polypropylene cage, cellulose fiber beddingInduction of grp94[2.4×105]/17β-estradiolPapaconstantinou et al. (2001)
  Induction of hsp72[3.5×106]/17β-estradiol 
  Induction of hsp90[5.3×106]/17β-estradiol 
 Juvenile-adult ovariectomized B6C3F1, s.c.×4 daysPurina 5001, polypropylene cage, cellulose fiber beddingUterine weight[5.3×106]/17β-estradiolPapaconstantinou et al. (2002)
  Induction of hsp90α[1.2×105]/17β-estradiol 
  Induction of grp24[8.4×106]/17β-estradiol 
 Juvenile-adult ovariectomized B6C3F1, s.c.×1Purina 5001, polypropylene cage, cellulose fiber beddingBlotted uterine weight, 6 hr after dose[8.4×106]/17β-estradiolPapaconstantinou et al. (2003)
  Blotted uterine weight, 12 hr after dose[4.2×106]/17β-estradiol 
 Adult ovariectomized transgenic ER-reporter, s.c.×1Purina 5001, polystyrene cageUterine wet weight[2.9×105]/diethylstilbestrolNagel et al. (2001)
  ER activation[1.0×104]/diethylstilbestrolNagel et al. (2001)
 Immature AP, s.c.×3 daysRM1 diet, plastic cage, sawdust and shredded paper beddingBlotted uterine weight[2.3×105]/diethylstilbestrol in 4 of 8 trials; other trials showed no effect at bisphenol doses up to 300 mg/kg bw/dayTinwell and Joiner (2000)
 Immature AP, gavage×3 daysRM1 diet, plastic cage, sawdust and shredded paper beddingBlotted uterine weightNo effect at bisphenol A doses up to 300 mg/kg bw/day; diethylstilbestrol produced response at 10 μg/kg bw/dayTinwell and Joiner (2000)
 Immature CD-1, s.c.×3 daysRM1 dietLactoferrin expressionNo effect at bisphenol A doses up to 1000 mg/kg bw/day; diethylstilbestrol showed effect at 0.1 μg/kg bw/dayMehmood et al. (2000)
  Uterine weight, BrdU incorporation, peroxidase productionNo effect at bisphenol A doses up to 100 mg/kg bw/day; diethylstilbestrol showed effect at 1–5 μg/kg bw/day 
 Immature CD-1, s.c. minipump×3 daysRMH 3000 diet, cage, and bedding estrogen-negative by E-ScreenUterine wet weight[1.6×105]/17β-estradiolMarkey et al. (2001b)
  Epithelial cell height[3.8×105]/17β-estradiol 
  Lactoferrin expression[3.9×105]/17β-estradiol 
 Ovariectomized adult B6C3F1, i.p.×3 daysNot indicatedRelative uterine to body weight[3.6–74×105]/17β-estradiolKitamura et al. (2005)
 Ovariectomized adult Swiss, s.c.×1Economy Rodent Maintenance dietIncreased uterine vascular permeability∼1×10−4/17β-estradiolMilligan et al. (1998)
Other mouse organs
 Juvenile-adult aromatase knock-out, diet×4 monthsNMF dietUterine and ovarian histology, bone mineral densityDietary bisphenol A (0.1%) exerted estrogenic effects. Mean±SD serum bisphenol A 84.3±8.7 μg/L. No comparator estrogen was used for these endpointsToda et al. (2002)
Fish
 Immature rainbow trout, injected Plasma vitellogenin[3×104]/17β-estradiolChristiansen et al. (1997)
 Juvenile rainbow trout, injected Plasma vitellogenin[5.6×103]/17β-estradiolAndersen et al. (1999a)
 Juvenile rainbow trout, exposed in water Plasma vitellogenin[8.4×105]/17β-estradiolLindholst et al. (2000)
 Male medaka, exposed in feed Plasma vitellogenin[1.4×104]/ethinyl estradiolChikae et al. (2003)
 Male medaka, exposed in water Hepatic vitellogenin and ERα mRNA[8.4×106]/17β-estradiolYamaguchi et al. (2005)
 Male killfish, injected Plasma vitellogenin[2.7×104]/17β-estradiolPait and Nelson (2003)
 Male zebrafish, juvenile rainbow trout, exposed in water Plasma vitellogenin[0.2]/ethinyl estradiolVan den Belt et al.(2003)
Invertebrates
 Mud snail, exposed in water New embryo production[1.5×104]/ethinyl estradiolJobling et al. (2004)
 Ramshorn snail, exposed in water Egg productionIncreased (EC10 13.9 ng/L); blocked by Faslodex and tamoxifen. No comparison to reference estrogenOehlmann et al. (2006)

The most common method of comparing potency is to test responses over a range of concentrations and to compare the concentrations producing the half-maximal (or other fractional) response of the comparator estrogen. An alternative is to compare the magnitude of the response at an equimolar concentration of the 2 estrogens. The difference in these two methods is illustrated in Figure 3. An example of the difference in potency estimations according to comparison method is the study of Vivacqua et al. (2003), in which the fold-increase in reporter activity for an estrogen-responsive gene was compared over a range of concentrations for bisphenol A and for 17β-estradiol. This study's Figure 3 presents curves analogous to Figure 3, but also presents a bar graph comparing response of the reporter at a 10−7 M concentration of each estrogen. Based on the half-maximal response to 17β-estradiol, bisphenol-A appeared 1000 times less potent than 17β-estradiol, but based on the fold-difference in reporter activity at 10−7 M, bisphenol A was about half as potent. Data for other estrogenicity comparisons in this study and in many other studies are presented only using bar graphs comparing responses at the same molar concentrations of the 2 estrogens, thereby overestimating the estrogenic potency of bisphenol A compared to studies in which comparisons are based on the half-maximal response.

Figure 3.

Alternative approaches to comparing estrogenic potency. In this example, the half-maximal response to the comparator estrogen occurs at 10−8 M. A similar response occurs with the test estrogen at 10−5 M, suggesting a 1000-fold difference in potency. If the magnitudes of response at equimolar concentrations are compared, the apparent potency may be much different. The response to the test estrogen at 10−7 M (a) is about half the response to the comparator estrogen at 10−7 M (a+b).

Competitive binding assays, which evaluate the concentration at which bisphenol A displaces labeled 17β-estradiol from ER, are summarized in the top part of Table 52. The receptor binding of bisphenol A in these assays varies over 3 orders of magnitude. Bisphenol A competes for human ER binding at molar concentrations 20–10,000 times that of the native ligand. When bisphenol A binding to ERα and ERβ was compared in the same study, 3 reports found little difference by receptor subtype (Kuiper et al., 1998; Paris et al., 2002; Takayanagi et al., 2006), and 3 studies found binding to ERβ to be 4, 10, 47, and 254 times greater than binding to ERα (Routledge et al., 2000; Matthews et al., 2001; Seidlová-Wuttke et al., 2004, 2005; Takemura et al., 2005). Yeast reporter systems, which reflect activation of post-receptor pathways, show less variability; these studies show bisphenol A activity to be 10,000–26,000 times less than that of 17β-estradiol.

Some variability in estimating bisphenol A potency appears to be due to differences between laboratories. Andersen et al. (1999a) reported results from 3 laboratories that evaluated the proliferative response of MCF-7 breast cancer cells to bisphenol A. The laboratories, which were in the U.S., Spain, and Denmark, were sent samples of the same stock of bisphenol A, 17β-estradiol, and MCF-7 cells. Procedures were similar in the labs, although two different counting methods were used. The bisphenol A potencies relative to 17β-estradiol were 5 × 10−7, 3 × 10−6, and 1 × 10−5. Laboratory variability may underlie some of the large differences in cell-based assays for ER activation; in those studies bisphenol A molar potency compared to 17β-estradiol were reported to vary by over 7 orders of magnitude (Table 52). Another explanation for this wide range of reported values is the difference in defining relative potency in some assays, as discussed above. [According to a study author, the wide variability in relative bisphenol A potency was due to a wide fluctuation in the 17β-estradiol dose at which half-maximal proliferation was achieved (0.1–70 pM) (A. Soto, personal communication, March 2, 2007).]

A study using ERα- and ERβ-reporting systems in 3 human cell lines found that bisphenol A had a small antagonistic effect on ERα activation in the presence of 17β-estradiol in human embryonal kidney and endometrial carcinoma cells (Kurosawa et al., 2002). There were no significant interactions between bisphenol A and 17β-estradiol on ERα activation in human osteosarcoma cells or on ERβ activation in any tested cell type. By contrast, a study using a recombinant yeast assay for ERα activation found 17β-estradiol and bisphenol A to have additive effects (Rajapakse et al., 2001), and a study using MCF-7 cell proliferation found 17β-estradiol and bisphenol A to have synergistic effects (Suzuki et al., 2001).

The data in Table 52 are applicable only to unconjugated bisphenol A. Estrogenic activity has not been identified for bisphenol A glucuronide (Matthews et al., 2001) or sulfate (Shimizu et al., 2002).

In vivo tests (Table 53) have been conducted principally in rats and mice. Most endpoints in these studies involved the uterus, and effects on uterine weight in immature or ovariectomized animals are the uterine endpoints reported most commonly. The potency of bisphenol A in increasing uterine weight varies over ∼4 orders of magnitude. Some of this variation may be related to the short half-life of bisphenol A. Uterotrophic evaluations are typically performed 24 hr after the last dose of the test agent is administered. Laws et al. (2000) showed no significant effect of bisphenol A at doses ≤400 mg/kg bw/day given orally on uterine wet weight assessed 24 hr after administering the last dose. When assessed 6 hr after the last oral dose, bisphenol A 200 mg/kg bw/day increased uterine wet weight to ∼2.5 times the control [estimated from a graph], which was about the same as the increase produced by administering 17β-estradiol 0.005 mg/kg bw/day sc. Increase in uterine weight in the first 6 hr after treatment represents fluid inbibition and not true tissue growth. A dose-related decrease in blotted uterine weight and body weight, with no effect on weight-adjusted uterine weight, was shown in pubertal rats treated on PND 22–42/43 with bisphenol A by gavage at 400 or 600 mg/kg bw/day (George et al., 2003).

For studies showing an increase in uterine weight after bisphenol A treatment, dose route affects response; bisphenol A given by gavage increased uterine weight by approximately 25% while the same dose given s.c. increased uterine weight by approximately 170% (Laws et al., 2000). A greater response by the s.c. than oral route was also shown by Yamasaki et al. (2000) and Kanno et al. (2003b) in the OECD multilaboratory study who showed a lowest effective bisphenol A dose of 8 mg/kg bw/day by the s.c. route and 160 mg/kg bw/day by the oral route. The greater activity per unit dose of s.c. than oral bisphenol A is due presumably to glucuronidation of the orally administered compound with consequent loss of estrogenicity (Matthews et al., 2001). A few studies could not confirm the greater effect of s.c. compared to oral bisphenol A on uterine weight. Ashby and Tinwell (1998) concluded that the magnitude of uterine weight response was similar for s.c. and oral routes. [The Expert Panel notes a greater numerical magnitude of response after s.c. than oral exposure in most of the experiments reviewed in this report, and that statistical comparison of the dose routes was not reported.] Matthews et al. (2001) found a similar increase in uterine weight in rats given s.c. or oral bisphenol A at 800 mg/kg bw/day.

Nagel et al. (1997, 1999) noted that 17β-estradiol is extensively protein-bound in vivo and bisphenol A is minimally protein-bound. A recent study indicated more extensive binding of bisphenol A to plasma binding proteins (Teeguarden et al., 2005). Nagel suggested that estrogenicity of BPA (as well as other steroid hormones) can be predicted more accurately in rats by considering the free fraction of a chemical in human serum. [The Expert Panel notes that Figure 2does not suggest that bisphenol A is more potent than 17β-estradiol in vivo than in vitro. The developmental effects of bisphenol A in the prostate are discussed inSection 3.2.]

Inter-strain variability in rats has been evaluated as a source of variability in estrogenicity assays. Inspection of Table 53 does not suggest large sensitivity differences between Sprague–Dawley, Wistar, and Long-Evans rats. Greater sensitivity of F344 than Sprague–Dawley rats has been shown with respect to uterine weight and epithelial cell height (Steinmetz et al., 1998), where 17β-estradiol-adjusted potencies differed by 20–37% between the strains. BrDu labeling of vaginal epithelium was 3 times greater in F344 than Sprague–Dawley rats in another study (Long et al., 2000), and a third study (Steinmetz et al., 1997) showed that both bisphenol A and 17β-estradiol increase serum prolactin in ovariectomized F344 but not ovariectomized Sprague–Dawley rats. Diel et al. (2004) evaluated estrogenic response to bisphenol A in juvenile ovariectomized DA/Han, Sprague–Dawley, and Wistar rats. After 3 days of treatment with bisphenol A 200 mg/kg bw/day, there were small statistically significant increases in uterine weight in DA/Han and Sprague–Dawley rats but not in Wistar rats. There were no alterations in uterine or vaginal epithelium or in uterine clusterin mRNA expression in any of the strains after bisphenol A treatment.

Inter-laboratory variation in the uterotrophic assay was evaluated by the Organization for Economic Cooperation and Development (OECD) (Kanno et al., 2003b). Coded chemicals, including bisphenol A, were sent to up to 212 different laboratories. Four assay protocols were evaluated including oral treatment of intact immature rats for 3 days, s.c. treatment of intact immature rats for 3 days, s.c. treatment of ovariectomized 6–8-week-old rats for 3 days, and s.c. treatment of ovariectomized 6–8-week-old rats for 7 days. Not all laboratories used all protocols or tested all compounds. Rat strains and suppliers were not standardized across laboratories. Comparisons were made between labs based on the lowest dose level at which body weight-adjusted blotted uterine weight was significantly different from the control. Results are summarized in Table 54. The lowest effective dose of bisphenol A was uniformly identified for the assays performed in ovariectomized adults. Assays performed in immature animals varied in identification of the lowest effective bisphenol A dose level. There was no apparent effect of strain on sensitivity of the uterotrophic response in immature (intact or castrate) or adult female rats.

Table 54. Differences Between Laboratories in Rat Uterotrophic Assay With Bisphenol Aa
LaboratoryRat strainLowest effective dose level (mg/kg bw/day)
Immature, gavage×3 days 2003756001000 
 2CD(SD)IGS ×   
 7CD(SD)IGS  ×  
 12CD(SD)IGS BR ×   
 13Wistar   × 
Immature, s.c.×3 days 101003006001000
 2CD(SD)IGS ×   
 6CD(SD)IGS BR  ×  
 7CD(SD)IGS ×   
 8Alpk:ApfSD ×   
 12CD(SD)IGS BR   × 
 13Wistar  ×  
 15Wistar  ×  
 18Sprague–Dawley×    
 20Sprague–Dawley×    
 21CD(SD) BR×    
Adult, s.c.×3 days 101003006001000
 2CD(SD)IGS ×   
 6CD(SD)IGS BR ×   
 7CD(SD)IGS ×   
 8Alpk:ApfSD ×   
 12CD(SD)IGS BR ×   
Adult, s.c.×7 days 101003006001000
 2CD(SD)IGS ×   
 7CD(SD)IGS ×   

Intra-laboratory variability has been noted for the bisphenol A uterotrophic assay in immature mice (Tinwell and Joiner, 2000). Of 8 studies performed over a 2-year period at s.c. bisphenol A dose levels up to 200 or 300 mg/kg bw/day, 4 showed a significant increase in uterine weight at 200 mg/kg bw/day. The other 4 studies, including the 2studies that went to 300 mg/kg bw/day, showed no effect of bisphenol A treatment on uterine weight despite the expected response to diethylstilbestrol. Study authors noted that reducing the permissible body weight of the mice selected for study resulted in lower and less variable control uterine weights and greater likelihood of bisphenol A effect (Tinwell and Joiner, 2000; Ashby et al., 2004). [The Expert Panel notes that these studies all used high s.c. doses of bisphenol A.]

Markey et al. (2001b) proposed that the rodent uterotrophic assay is relatively insensitive to the estrogenic effects of bisphenol A. These authors treated immature CD-1 mice with bisphenol A in s.c. minipumps and evaluated uterine weight, relative area of uterine compartments, epithelial height, expression of lactoferrin and proliferating cell nuclear antigen (PCNA), and induction of vaginal opening. Dose–response curves for the endpoints that showed significant changes from control are illustrated in Figure 4. The study authors also noted that significant alterations in some endpoints were observed at much lower doses (0.1 mg/kg bw/day for vaginal opening and 5 mg/kg bw/day for epithelial cell height), giving rise to a U-shaped dose–response curve. [The assertions of some investigators notwithstanding, the Expert Panel notes that oral bisphenol A does not consistently produce robust estrogenic responses and, when seen, estrogenic effects after oral treatment occur at high-dose levels.]

Figure 4.

Dose–response curves for endpoints of estrogenic activity in s.c.-dosed mice. On pair-wise testing, body weight was increased at 0.5 mg/kg bw/day and decreased at 100 mg/kg bw/day; vaginal opening was advanced at 0.1 and 100 mg/kg bw/day; epithelial cell height was increased at 5, 75, and 100 mg/kg bw/day; PCNA labeling was increased at 75 and 100 mg/kg bw/day; and uterine wet weight was increased at 100 mg/kg bw/day. Data from Markey et al. (2001b).

Transgenic reporter mice have permitted in vivo identification of activation of the estrogen response element. Eight hr after i.p. injection on GD 13.5 of wild-type dams carrying transgenic fetuses, luciferase reporter activity was increased for bisphenol A 1 and 10 mg/kg bw (Lemmen et al., 2004). The luciferase response after bisphenol A was about 50% of that after a similar dose of estradiol dipropionate and ∼25% of that after a 10-fold higher dose of diethylstilbestrol [estimated from a graph]. Use of an in vitro reporter system showed bisphenol A potency to be 3–4 orders of magnitude less than that of diethylstilbestrol (Table 52). The authors concluded that the in vivo estrogenic potency of bisphenol A may be greater than predicted by in vitro assays.

Nagel et al. (2001) developed a transgenic mouse with a thymidine kinase-lacZ reporter linked to 3 copies of the vitellogenin estrogen response element. This model showed an increase in ER activity after a single s.c. bisphenol A dose of 25 μg/kg bw (P=0.052), with further increases in activity after 0.8 and 25 mg/kg bw. Uterine weight was only increased at the 25 mg/kg bw dose level. Normalized to the diethylstilbestrol response, uterine weight response to bisphenol A 25 mg/kg bw was less than one-third the response in ER activity [estimated from a graph].

Gene expression profiles have been performed to compare the presumably ER-mediated response to bisphenol A with the response to reference ER agonists. Naciff et al. (2002) evaluated expression in the uteri and ovaries of Sprague–Dawley fetuses after s.c. dosing of dams on GD 11–20 with ethinyl estradiol 0, 0.5, 1, or 10 μg/kg bw/day or bisphenol A 0, 5, 50, or 400 mg/kg bw/day. The high-dose of both compounds induced nipples and areolae in male and female fetuses. There were 366 genes in which expression was altered by ethinyl estradiol and 397 genes in which expression was altered by bisphenol A. Expression of 66 genes was changed in the same direction with high-doses of ethinyl estradiol, bisphenol A, and genistein (which was also tested in this model). Of the 40 genes with at least a 1.8-fold change in expression, 17 responded similarly to ethinyl estradiol and bisphenol A. The authors identified 50 mg/kg bw/day as the lowest dose level at which estrogen-like gene expression activity could be identified, which is lower than the 400–800 mg/kg bw/day dose range at which uterotrophic activity is typically reported in rats (Ashby and Tinwell, 1998).

Terasaka et al. (2006) used expression of 120 estrogen-responsive genes (based on previous work) in MCF-7 cells to compare the profiles of bisphenol A and 17β-estradiol. Response was highly correlated (R=0.92) between the 2 compounds. Another gene array study (Singleton et al., 2004) used MCF-7 cells that had lost ER and were re-engineered to express ERα. Among 40 estrogen-responsive genes, 12 responded to both bisphenol A and 17β-estradiol, 9 responded only to bisphenol A, and 19 responded only to 17β-estradiol. In the ER-deficient MCF-7 cell line from which these cells had been engineered, 1 gene responded to both bisphenol A and 17β-estradiol and 14 responded to bisphenol A alone, suggesting ER-independent activity. The same group reported the response of an additional 31 genes, associated with growth and development, from the same chip (Singleton et al., 2006). In the ERα-containing cells, 5 of these genes showed regulation with both 17β-estradiol and bisphenol A, 13 were regulated only by bisphenol A, and 13 were regulated by only 17β-estradiol.

Differences in the estrogenic activity of bisphenol A and reference estrogens may be due to differences in recruiting by the liganded receptor of co-regulatory proteins. Singleton et al. (2006) used a co-regulator-independent yeast reporter system to evaluate the estrogenicity of bisphenol A and 17β-estradiol. Bisphenol A activity was more than 3 orders of magnitude less than 17β-estradiol in the yeast system, compared to about a 2-order-of-magnitude difference in an MCF-7 cell assay, leading the authors to postulate that mammalian co-activators may be involved in enhancing bisphenol A activity. In a comparison of ER binding and co-activator recruitment, Routledge et al. (2000) showed bisphenol A to bind the receptor more avidly than the liganded receptor recruited 2 co-activator proteins, normalized to 17β-estradiol (Table 55).

Table 55. Bisphenol A Receptor Binding and Recruitment of Co-Activator Proteinsa
 Activity relative to 17β-estradiol
AssayERαERβ
  • a

    aRoutledge et al. (2000).

Receptor binding7.3×10−47.5×10−3
TIF2 recruitment< 1×10−65×10−4
SRC-1a recruitment3×10−42×10−4

The classical ERs are receptors that, when bound, produce their activity through alterations in genomic transcription. In contrast, a membrane-bound ER has been described in murine pancreatic islet cells (Nadal et al., 2000, 2004; Quesada et al., 2002; Alonso-Magdalena et al., 2005). This membrane-bound receptor regulates calcium channels and modulates insulin and glucagon release. Bisphenol A has been shown to activate this receptor in vitro at a concentration of 1 nM, which is similar to the active concentration of diethylstilbestrol (Nadal et al., 2000; Alonso-Magdalena et al., 2005). Treatment of mice with bisphenol A or 17β-estradiol s.c. at 10 μg/kg bw acutely or daily for 4 days resulted in decreased plasma glucose and increased insulin (Alonso-Magdalena et al., 2006). By contrast, Adachi et al. (2005) reported that exposure of rat pancreatic islets to 0.1–1 μg/L [0.4–4.4nM] bisphenol A did not alter insulin secretion over a 1-hr period. Exposure of islets to bisphenol A 10 μg/L [44nM] for 24 hr increased insulin release. This response was prevented by actinomycin D and by ICI 182,780, supporting the conclusion that bisphenol A insulin release occurs through interaction with the cytoplasmic ER rather than the membrane-bound receptor.

A membrane-bound ERα in the pituitary could be related to regulation of the release of stored prolactin in response to estrogens, a non-genomic response mediated by calcium influx. Using a rat prolactinoma cell line, bisphenol A was shown to promote calcium influx and release prolactin over a concentration range similar to that for 17β-estradiol (Wozniak et al., 2005; Watson et al., 2007). The response to bisphenol was bimodal, with maximal responses at concentrations of 10−12 and 10−8 M and little-to-no response at intermediate concentrations. Calcium influx in MCF-7 cells has been shown to occur rapidly after exposure to bisphenol and 17β-estradiol concentrations of 10−10 M through a non-ER-mediated mechanism (Walsh et al., 2005).

Recently, bisphenol A was identified as competitor to 17β-estradiol for binding to the GPR30 receptor; a novel seven-transmembrane receptor that mediates nongenomic estrogen actions to upregulate adenylyl cyclase and MAPK activities (Thomas and Dong, 2006). Similar to findings reported previously with nuclear estrogen receptors and membrane estrogen receptors, bisphenol A was identified as a relatively effective competitor of 17β-estradiol binding, with relative binding affinities of 2.8% that of the natural estradiol ligand and an IC50 of 630 × 10−9 M. Bisphenol A, at a concentration of 200 nM significantly increased cAMP levels in transfected cells 30 min after compound addition.

Bisphenol A has been found to bind estrogen-related receptor γ, a nuclear receptor with no known natural ligand that shows little affinity for 17β-estradiol (Takayanagi et al., 2006). Estrogen-receptor γ demonstrates high constitutive activity that is maintained by bisphenol A in the presence of 4-hydroxytamoxifen, which otherwise blocks nuclear ER activity. This observation led to the suggestion that bisphenol A may maintain estrogen-related receptor γ activity in the presence of a yet-to-be-identified natural antagonist and that cross talk between the estrogen-related receptor and ER systems could be responsible for the estrogenic activity of bisphenol A in spite of low binding affinity for ERα and β (Takayanagi et al., 2006).

In addition to the studies reviewed for this section, there are studies in which the putative estrogenicity of environmental samples or synthetic products were evaluated using one or another assay. For example, Olea et al. (1996) evaluated resin-based dental composites in an MCF-7 culture system. The response of the system was attributed to the bisphenol and its methacrylate detected in the composites, but bisphenol A was not specifically tested. These articles were not reviewed for this section.

2.2.3 Androgen activity.

Transfected cell-based assays have not identified bisphenol A as having androgenic activity (Sohoni and Sumpter, 1998; Gaido et al., 2000; Kitamura et al., 2005; Xu et al., 2005). However, bisphenol A is mitogenic in cultured human prostate carcinoma cells at a concentration of 1 nM (Wetherill, 2002). Based on stimulated cell growth in this system, the potency of bisphenol A is about 5% that of dihydrotestosterone [estimated from a graph]. This bisphenol A activity was shown to be mediated by interaction with a mutant tumor-derived androgen receptor called AR-T877A. Anti-androgenic activity has been demonstrated using cells transfected with androgen receptor reporting systems (Table 56). The anti-androgenic activity of bisphenol A is expressed as the concentration needed to halve the androgen reporter response to a reference androgen. Studies in transfected cells have shown that bisphenol A interferes with the binding of dihydrotestosterone to the androgen receptor, interferes with translocation of the liganded receptor to the nucleus, and prevents transactivation at the androgen-response element (Lee et al., 2003a).

Table 56. Anti-Androgenicity Studies of Bisphenol A in Cells Transfected With Androgen Receptor Reporter
Cell typeReference androgen concentration (nM)Bisphenol A median inhibitory concentration (IC50) μM [mg/L]Reference
  1. a

    aEstimated from a graph.

Human prostate adenocarcinomaR1881 0.17 [1.6]Paris et al. (2002)
Chinese hamster ovaryR1881 0.119.6 [4.5]Roy et al. (2005)
YeastTestosterone 101.8 [0.4]Lee et al. (2003a)
YeastDihydrotestosterone 1.252a[0.5]Sohoni and Sumpter (1998)
Monkey kidneyDihydrotestosterone 10.746 [0.2]Xu et al. (2005)
Monkey kidneyDihydrotestosterone 12.14 [0.5]Sun et al. (2006)
Mouse fibroblastDihydrotestosterone 0.014.3 [1.0]Kitamura et al. (2005)
Human hepatomaDihydrotestosterone 100No anti-androgenic activityGaido et al. (2000)

Kim et al. (2002a) conducted a Hershberger assay to determine the effects of bisphenol A exposure on reproductive organs of rats. Sprague–Dawley rats were fed PMI Certified Rodent LabDiet and housed in polycarbonate cages. No information was provided about bedding materials. One experiment was conducted to determine the optimum dose and age for observing testosterone exposure effects. In a second experiment, 10 rats/group rats were castrated at 5 weeks of age and 7 days later gavaged with bisphenol A (99% purity) at doses of 0 (ethanol/corn oil vehicle) 10, 100, or 1000 mg/kg bw/day for 7 days. A second group of castrated 6-week-old males rats was gavaged with bisphenol A at 0, 50, 100, 250, or 500 mg/kg bw/day for 7 days. In a third experiment, 10 castrated 6-week-old rats/group were treated with 0.4 mg/kg bw/day testosterone by s.c. injection in addition to gavaged bisphenol A at 50, 100, 250, or 500 mg/kg bw/day or flutamide at 1, 5, 10, or 25 mg/kg bw/day for 7 days. A positive control group was given 0.4 mg/kg bw/day testosterone for 7 days. [There is some confusion in the article regarding ages at castration and start of treatment. For the first group of bisphenol A-treated rats, it is reported that rats were castrated at 5 weeks of age and treated at 6 weeks of age. For the other groups of bisphenol A-treated rats, the Methods section reported that treatment began at 6 weeks of age, but tables in the Results section indicated that rats were castrated at 6 weeks of age.] During the study, clinical signs were observed and body weights were measured. Blood was collected and rats were killed ∼24 hr after administration of the last dose. Accessory reproductive organs were removed and weighed. Serum luteinizing hormone (LH) and testosterone concentrations were measured by radioimmunoassay (RIA). Statistical analyses included Bartlett test, analysis of covariance (ANCOVA), Dunnett test, and Bonferroni test. Exposure to bisphenol A did not affect weights of the ventral prostate, seminal vesicles, glans penis, or levator ani plus bulbocavernosus muscle; or serum concentrations of LH or testosterone. Testosterone increased the weights of accessory reproductive organs. Flutamide increased serum LH concentrations and inhibited testosterone-induced increases in accessory reproductive organ weights. Study authors concluded that bisphenol A did not exhibit androgenic or antiandrogenic effects in rats.

Yamasaki et al. (2003) conducted a Hershberger assay in rats exposed to bisphenol A or 1 of 29 other chemicals. In this study, which was conducted according to GLP, animals were housed in stainless steel wire-mesh cages. Assuming these males were fed the same diets as rats used in an uterotrophic assay also described in this study, they received MF Oriental Yeast feed. Rats were randomly assigned to treatment groups. Beginning at 56 days of age and continuing for 10 days, 6 castrated male Brl Han: WIST Jcl (GALAS) rats/group were administered bisphenol A by stomach tube at doses of 0 (olive oil vehicle), 50, 200, or 600 mg/kg bw/day. An additional group of rats was administered the same vehicle and doses of bisphenol A in addition to 0.2 mg/kg bw/day testosterone propionate by s.c. injection. Dose selection was based on results of preliminary studies. A positive control group was given 10 mg/kg bw/day flutamide in addition to 0.2 mg/kg bw/day testosterone propionate. Rats were killed 24 hr after receiving the final dose. Ventral prostate with fluid, seminal vesicles with fluid, bulbocavernosus/levator ani muscle, glans penis, and Cowper gland were collected and weighed. Data were analyzed by Student t-test. Bisphenol A did not affect body weight and there were no clinical signs of toxicity. The only statistically significant effect on relative organ weight was a [24%] increase in glans penis weight in rats given 600 mg/kg bw/day bisphenol A without co-administration of testosterone. In contrast, rats treated with flutamide plus testosterone propionate experienced increases in weights of ventral prostate, seminal vesicle, bulbocavernosus/levator ani muscle, glans penis, and Cowper gland. [Absolute organ weights were not reported. It is assumed but was not stated that relative weights were based on body weight.] Study authors noted that because glans penis weights were variable in control rats and weights of other accessory reproductive organs were not affected, bisphenol A could not be clearly determined to have androgen agonistic properties.

Nishino et al. (2006) performed a Hershberger assay in Wistar rats. At 2 weeks of age, rats were given ssniffR 10 diet and housed in Makrolon cages with ssniff bedding. Seven days after orchiectomy, rats were placed in groups of 13 [randomization not discussed] and treated orally [gavage assumed] with bisphenol A [purity not indicated] in propylene glycol at 0, 3, 50, 200, or 500 mg/kg bw/day for 7 days or s.c. with testosterone propionate 1 mg/kg bw. Another group was given oral bisphenol A 500 mg/kg bw/day and flutamide 3 mg/kg bw/day. Rats were killed by decapitation after treatment. Seminal vesicles and prostates were weighed and fixed in 4% neutral buffered paraformaldehyde. Immunohistochemical evaluation of androgen receptor, PCNA, and MIB-5 was performed. Epithelial cell height and duct luminal area were determined morphometrically. Review by the Expert Panel indicated that this study was inadequate due to methodological issues.

2.3 Genetic Toxicity

Assessment of mutagenicity associated with bisphenol A was based primarily on reviews by the European Union (2003) and Haighton et al. (2002). CERHR summarized a limited number of studies that were not included in reviews. Results of in vitro genetic toxicity testing are summarized in Table 57, and results of in vivo genetic toxicity tests are summarized in Table 58.

Table 57. In Vitro Genetic Toxicity Studies of Bisphenol A
ConcentrationCellEndpointResultsReference
  • a

    aReviewed by Haighton et al. (2002).

  • b

    bReviewed by European-Union (2003).

  • c

    cAccording to the Haighton et al. (2002) review, positive results occurred at cytotoxic concentrations.

  • d

    dDiscrepancies noted between information presented by Haighton et al. (2002) and European-Union (2003).

  • e

    eConclusion by Haighton et al. (2002).

  • f

    fConclusion by European-Union (2003).

  • ↑,↓ increase, decrease.

3.3–333.3 μg/plate, with and without metabolic activationSalmonella typhimurium strains TA98, TA100, TA1535, TA1537MutagenicityNegativeHaworth et al. (1983)ab
50–500 μg/plate, with and without metabolic activationSalmonella typhimurium strains TA97a, TA98, TA100, TA102MutagenicityNegativeSchweikl et al. (1998)ab
≤5000 μg/plate with and without metabolic activationSalmonella typhimurium strains TA97, TA98, TA100, TA102MutagenicityNegativeTakahata et al. (1990)ab
≤1000 μg/mL, with and without metabolic activationSalmonella typhimurium strain TA1538 and Escherichia coli strains WP2 and WP2uvrAMutagenicityNegativeDean and Brooks (1978)ab
5–1250 μg/plate, with and without metabolic activationSalmonella typhimurium strains TA98, TA100, TA1535, TA1537, and Escherichia coli strain WP2uvrAMutagenicityNegativeJETOC (1996)ab
1 mM [228 mg/L], with and without metabolic activationSalmonella typhimurium strains TA98 and TA100MutagenicityNegativeMasuda et al. (2005)
0.1–0.2 mM [23–46 mg/L], without metabolic activationChinese hamster V79 cells, hprt locusMutagenicityNegativeSchweikl et al. (1998)ab
5–60 mg/L without metabolic activation, 25–200 mg/L with metabolic activation, or 5–60 mg/L with and without metabolic activationdMouse lymphoma L5178Y cells, tk+/− locusMutagenicityNegative (results questioned due to possible inability to count small colonies)Myhr and Caspary (1991)ab
Concentrations not specified, with and without metabolic activationMouse lymphoma L5178Y cells, tk+/− locusMutagenicityInconclusive without and negative with metabolic activationHonma et al. (1999)ab; Moore et al. (1999)ab
25–200 μM [5.7–46 mg/L], without metabolic activationSyrian hamster embryo cells (Na+/K+ ATPase and hprt loci)MutagenicityNegativeTsutsui et al. (1998)ab
10−8–10−5 M [0.002–2 mg/L], without metabolic activationHuman RSa cellsMutagenicity↑ at all dosesTakahashi et al. (2001)
≤500 mg/L, with and without metabolic activationSaccharomyces cerevisiae strain JDIMutagenicityNegativeDean and Brooks (1978)ab
10–8–10−4 M [0.002–23 mg/L], without metabolic activationMCF-7 cellsDNA damage (assessed by comet assay)↑ at ≥10−6 M [0.2 mg/L]Iso et al. (2006)
10−4 M [23 mg/L], without metabolic activationMDA-MB-231 cellsDNA damage (assessed by comet assay) 
20–40 mg/L, without metabolic activation and 30–50 mg/L with metabolic activationChinese hamster ovary (CHO) cellsChromosomal aberrationNegative (inconsistent ↑ at high dose with metabolic activation)Ivett et al. (1989)ab; Tennant et al. (1986, 1987)b
350–450 μM [80–103 mg/L], without metabolic activation and ≤250 μM [57 mg/L] with metabolic activationCHO cells, clone WBLChromosomal aberrationPositive at ≥400 μM [91.3 mg/L] without metabolic activationc; negative with metabolic activationHilliard et al. (1998)a
400 and 450 μM [91 and 103 mg/L], without metabolic activationCHO cells, clone WBLChromosomal aberrationPositivecGalloway et al. (1998)a
25–200 μM [5.7–46 mg/L], without metabolic activationSyrian hamster embryo cellsChromosomal aberrationNegativeTsutsui et al. (1998)ab
10–30 mg/LEpithelial-type rat liver cell line (RL1)Chromosomal aberrationNegativeDean and Brooks (1978)b
25–200 μM [5.7–46 mg/L], without metabolic activationSyrian hamster embryo cellsAneuploidy/polyploidyInconclusive (non-dose-related ↑ in aneuploidy at ≥50 μM [11 mg/L])e; apparently positivefTsutsui et al. (1998)ab
0.8–25 mg/L, without metabolic activation and 30–50 μg/mL, with metabolic activationCHO cellsSister chromatid exchangeNegative (one small ↑ was not reproducible)Ivett et al. (1989)ab Tennant et al. (1986)b
0.2–0.5 mM or nMd[46–114 mg/L or μg/L]Rat hepatocytesDNA strand breaksNegative (↑ noted but scored as negative by study authors due to excessive cytotoxicity)Storer et al. (1996)ab
10−9–10−5 M [0.0002–2 mg/L], without metabolic activationHuman RSa cellsUnscheduled DNA synthesis↑ at 10−6 M [0.2 mg/L] and ↓ at 10−7[0.02 mg/L] and 10−5 M [2 mg/L]Takahashi et al. (2001)
Not specified, but stated to cover range of cytotoxicityA31-1-13 clone of BALB/c-3T3 cellsTransformationNegativeMatthews et al. (1993)a
25–200 μM [5.7–46 mg/L], without metabolic activationSyrian hamster embryo cellsTransformationPositive at ≥50 μM [11.4 mg/L] (non-dose-related ↑)e; equivocalfTsutsui et al. (1998, 2000)ab
≤50 mg/L for 24 hr; ≤30 mg/L for 7 days, without metabolic activationSyrian hamster embryo cellsTransformationNegativeLeBoeuf et al. (1996)a
2–60 mg/LSyrian hamster embryo cellsTransformationNegativeJones et al. (1988)b
50–200 μM [11.5–46 mg/L], without metabolic activationSyrian hamster embryo cellsDNA adduct formationPositive at ≥50 μM [11 mg/L] (dose-related ↑)Tsutsui et al. (1998)ab
1000 μg presence of peroxidase and hydrogen peroxidePurified rat DNADNA adduct formationPositiveAtkinson and Roy (1995a)
10–100 μM [2.3–23 mg/L], metabolic activation unknownBovine brain microtubule proteinInhibited microtubule polymerizationPositiveMetzler and Pfeiffer (1995)a
50–200 μM [11.5–46 mg/L], no metabolic activationBovine brain microtubule proteinInhibited microtubule polymerizationPositive (dose-related)Pfeiffer et al. (1996)b
20–200 μM [4.6–46 mg/L], metabolic activation unknownBovine brain microtubule proteinInhibited microtubule polymerizationPositive (EC50=150 μM [34 mg/L])Pfeiffer et al. (1997)ab
200 μM [46 mg/L], without metabolic activation; 100 μM [23 mg/L] for metaphase arrest assayChinese hamster V79 cellsAneuploidogenic potential as assessed by micronuclei formation, microtubule assay, and metaphase arrestPositivePfeiffer et al. (1997)ab
100–200 μM [2346 mg/L], without metabolic activationChinese hamster V79 cellsAneuploidogenic potential as assessed by micronuclei formationPositiveOchi (1999)ab
10 or 30 μM [2.3 or 6.9 mg/L]Oocytes from Balb/c miceMeiotic spindle formationCentrosomes and spindles disorganizedCan et al. (2005)
0.05–0.4 mg/LOocytes from MF1 miceAneuploidyNo hyper haploidy but ↑ diploid metaphase II oocytes at 0.2 mg/LPacchierotti et al. (2007)
Table 58. In Vivo Genetic Toxicity Studies of Bisphenol A
Species and sexDose (route)CellsEndpointResultsReference
  • a

    aReviewed by Haighton et al. (2002).

  • b

    bReviewed by European-Union (2003).

Male rat85 mg/kg bw/day for 5 days (i.p.)GermDominant lethalityNegativeBond et al. (1980)ab (abstract only)
Male rat200 mg/kg bw (i.p.) and 200 mg/kg bw for 4, 8, 12, or 16 days (oral)DNAAdduct formationPositiveAtkinson and Roy (1995b)
Male and female mouse500–2000 mg/kg bw (oral)Bone marrowMicronucleiNegativeGudi and Krsmanovic (1999)a; Shell Oil Co. (1999)b
Male mouse1 mmol/kg bw [228 mg/kg bw] (oral)Peripheral blood reticulocyteMicronucleiNegativeMasuda et al. (2005)
20–22-day-old female mouse0.02–0.100 mg/kg bw/day (oral) for 6–8 days or 0.02 mg/kg bw for 3, 5, or 7 daysOocyteCongression failurePositive at all doses; statistically significant with 7-day exposureHunt et al. (2003)
Pregnant mouse GD 11.5–18.50.4 μg/day s.c. pellet [20 μg/kg bw/day]OocyteEvaluation of pachytene fetal oocyte and of ploidy in oocytes and 2-cell embryos from adults that were exposed in uteroIncomplete synapsis, end-to-end association of sister chromatids, ↑hyperploidySusiarjo et al. (2007)
Female mouse0.2 or 20 mg/kg bw acutely or daily for 7 days or 0.4 mg/L in drinking water for 7 weeksOocyteAneuploidyNegativePacchierotti et al. (2007)
Male (102/ElxC3H/El) F1 mouse0.002–0.2 mg/kg bw for 6 days (oral)SpermatocyteMeiotic delay and aneuploidyNegativePacchierotti et al. (2007)
Drosophila melanogaster10,000 ppm (oral)OffspringSex-linked recessive lethal testNegativeFoureman et al. (1994)ab
Turbot50 ppb in aquarium water for 2 weeksErythrocyteMicronucleiPositiveBolognesi et al. (2006)

The European Union (2003) noted that bisphenol A demonstrated aneugenic potential and micronuclei formation in in vitro tests without metabolic activation. However, there was no evidence of micronuclei formation in an in vivo mouse study. Other studies demonstrated disruption of microtubule formation and the presence of DNA adducts. In the studies reviewed by the European Union, there was no evidence of gene mutations or structural chromosomal aberrations in in vitro tests and negative results were obtained in a dominant lethal test in rats; however, the European Union noted several limitations for those studies. Based on their review of genotoxicity data and the lack of significant tumors reported in animal studies, the European Union (2003) concluded that bisphenol A does not appear to have significant mutagenicity potential in vivo. Because aneugenic potential was apparently observed only in in vitro tests, it was judged to be of no concern. The relevance of DNA adduct formation was unclear, but based on weight of evidence, i.e., negative findings for gene mutation and clastogenicity in cultured mammalian cells, DNA adduct formation was thought unlikely to be of concern for humans.

Haighton et al. (2002) concluded that results of standardized and validated genetic toxicity tests demonstrated the lack of mutagenic and genotoxic activity of bisphenol A in vivo. Studies demonstrating disrupted microtubule formation or DNA adduct formation were noted, but because the studies used high-doses, they were judged to be of limited relevance. The lack of activity in an in vivo micronucleus assay in mice was said to confirm negative results observed in in vivo tests. Lastly, it was concluded that bisphenol A (parent) had no structural features that suggested mutagenic activity.

Subsequent to the release of the European Union (2003) and Haighton et al. (2002) reviews, Hunt et al. (2003), published a study examining meiotic aneuploidy potential of bisphenol A in female mice. In 1998, a large increase in background rate of congression failure (from 1–2 to 40%) and in aneuploidy (from 0.7 to 5.8%) was observed in the study authors' laboratory. The increase was found to coincide with damage to polycarbonate caging material. Removal of the most damaged cages and change to polysulfone cages resulted in decreased background rates of congression failure. Intentionally damaging polycarbonate cages and water bottles resulted in increased rates of congression failure. As noted in Table 58, congression failure rates were increased in juvenile female mice orally exposed to ≥20 μg/kg bw/day bisphenol A for 6–8 days or 20 μg/kg bw/day for 7 days. The study authors concluded that bisphenol A was a potential meiotic aneugen.

In a follow-up study (Susiarjo et al., 2007), pregnant C57Bl/6 mice on GD 11.5 were implanted with s.c. pellets designed to release bisphenol A 0 or 0.4 μg/day. [The authors assume a 20 g bw, giving an estimated dose level of 20μg/kg bw/day.] Oocytes from GD 18.5 female fetuses showed an increase in pachytene synaptic abnormalities including incomplete synapsis and end-to-end associations of sister chromatids. There was also paradoxically an increase in recombinant foci in pachytene oocytes of bisphenol A-exposed females. Some female offspring of bisphenol A-treated dams were fostered to untreated dams. Eggs or 2-cell embryos from these female offspring at 4–5 weeks of age showed an increase in hyperploidy. Pachytene oocyte abnormalities similar to those identified in fetuses exposed to bisphenol A were seen in oocytes obtained from ERβ knock-out mice, suggesting to the authors that bisphenol A may exert adverse effects on meiosis by blocking ERβ.

In response to the study of Hunt et al. (2003), Pacchierotti et al. (2007) investigated the aneugenic effects of bisphenol A in mouse somatic and germ cells. C57Bl/6 female mice were superovulated using pregnant mare serum and hCG after which they were gavaged with bisphenol A 0.2 or 20 mg/kg bw. Metaphase II oocytes were collected after 17 hr and evaluated using C-banding. Additional female mice were gavaged with bisphenol A 0.04 mg/kg bw/day for 7 days or were given bisphenol A in drinking water at a concentration of 0.4 mg/L for 7 weeks. These mice were superovulated at the end of the 7-day or 7-week treatment period and housed overnight with untreated males. Females without vaginal plugs were killed for evaluation of oocytes by C-banding. Females with vaginal plugs were treated with colchicine to prevent the first embryonic cleavage, and zygotes were collected the next morning for evaluation by C-banding. There were no bisphenol A effects on induction of aneuploidy. There was a statistically significant increase in premature centromere separation in the group treated for 7 weeks, but there was no effect of bisphenol A treatment on the proportion of zygotes with structural or numeric chromosome changes. Male mice were treated with bromodeoxyuridine 8 days before being treated with bisphenol A 0.2 mg/kg bw/day for 6 days. Evaluation of sperm after 21–25 days did not show a significant mitotic delay in spermatocytes. Additional male mice were given bisphenol A orally at doses of 0, 0.002, 0.02, and 2 mg/kg bw/day for 6 days. Epididymal sperm were collected 22 days after the end of bisphenol A treatment and multicolor fluorescent in situ hybridization was used to evaluate decondensed sperm for aneuploidy. Sperm count was decreased by bisphenol A 0.002 mg/kg bw/day, but there was no increase in the frequency of hyperhaploidy or diploidy. Bisphenol A was negative in a bone marrow micronucleus test at dose levels up to 2 mg/kg/day for 2 days.

2.4. Carcinogenicity

No human data examining the carcinogenicity of bisphenol A were identified.

NTP (1982) and Huff (2001) examined carcinogenicity of bisphenol A in F344 rats and B6C3F1 mice. Animals were randomly assigned to treatment groups. Bisphenol A (<98.2% purity) was administered through feed for 103 weeks to 50 rats/sex/dose at 0, 1000, or 2000 ppm, 50 male mice/group at 0, 1000, or 5000 ppm, and 50 female mice/group at 0, 5000, or 10,000 ppm. NTP estimated mean bisphenol A intakes of 74 and 148 mg/kg bw/day for male rats and 74 and 135 mg/kg bw/day for female rats. [Data on bisphenol A intake, food intake, and body weights were not provided for mice.] Using default values, the European Union (2003) estimated bisphenol A intakes of 120 and 600 mg/kg bw/day in male mice and 650 and 1300 mg/kg bw/day in female mice. Concentration and stability of bisphenol A in feed were verified. Body weights and clinical signs were observed during the study. Following the exposure period, animals were killed and necropsied. Organs, including seminal vesicle, prostate, testis, ovary, and uterus, were preserved in 10% neutral buffered formalin and examined histologically. Statistical analyses included Cox and Tarone methods, 1-tailed Fisher exact test, Bonferroni inequality criterion, Cochran-Armitage test, and life table methods for linear trend.

In rats, body weights of males and females from both dose groups were lower than controls throughout the study. Feed intake was decreased in females of both dose groups beginning at Week 12. No adverse effects on survival were observed. There were no non-neoplastic lesions [including in male and female reproductive organs] that appeared to be treatment-related. The incidence of leukemia was increased in males (13 of 50, 12 of 50, and 23 of 50 in control and each respective dose group) and females (7 of 50, 13 of 50, and 12 of 50). In males the trend for leukemia was significant by Cochran-Armitage test, but statistical significance was not shown by life table analysis for trend or incidence in the high-dose group, according to the unpublished version of the study. The published version of the study indicated statistical significance at the high-dose. Statistical significance was not attained for leukemia incidence in female rats. An increased incidence of testicular interstitial cell tumors (35 of 49, 48 of 50, 46 of 49) was statistically significant in both dose groups. An increased incidence of mammary fibroadenomas in males of the high-dose group (0 of 50, 0 of 50, and 4 of 50) achieved statistical significance for trend by Cochran-Armitage test but not by Fisher exact test. In bisphenol A groups, there were decreased incidences of adrenal pheochromocytomas in males, adrenal cortical adenomas in females, and uterine endometrial stromal polyps. The NTP concluded that none of the increases in tumor incidence in rats was clearly associated with bisphenol A exposure.

In mice, body weights were lower in high-dose males and in females of both dose groups. Feed intake could not be accurately determined because of spillage. Bisphenol A did not affect the survival of mice. Incidence of multinucleated hepatocellular giant cells was increased in treated males (1 of 49, 41 of 49, and 41 of 50). [A review of the data indicated no increases in incidence of non-neoplastic lesions in the reproductive organs of male or female mice.] The incidence of leukemia or lymphoma in male mice by dose group (2 of 49, 9 of 50, and 5 of 50) was not statistically significant. The published version of the report indicated an increasing trend for lymphoma. The linear trend for increased pituitary chromophobe carcinomas in male mice (0 of 37, 0 of 36, 3 of 42) was reported to be statistically significant by Cochran-Armitage test but statistical significance was not shown by Fisher exact test. The study authors concluded that none of the increases in tumor incidence in mice could be unequivocally associated with bisphenol A exposure.

NTP concluded that under the conditions of this study, there was no convincing evidence the bisphenol A was carcinogenic in F344 rats or B6C3F1 mice. However, study authors stated that there was suggestive evidence of increased cancer in the hematopoietic system based on marginally significant increases in leukemia in male rats, non-statistically significant increases in leukemia in female rats, and a marginally significant increase in combined incidence of lymphoma and leukemia in male mice. A statistically significant increase in testicular interstitial cell tumors in aging F344 rats was also considered suggestive evidence of carcinogenesis. The effect was not considered conclusive evidence because of the high incidence of the testicular neoplasm in aging F344 rats (88% incidence in historical controls).

The NTP study was reviewed by the European Union (2003) and Haighton et al. (2002). For increases in leukemia, mammary gland fibroadenoma, and Leydig cell tumors in male rats, both groups noted the lack of statistical significance using the appropriate analyses and the common occurrence of these tumor types in F344 rats. The European Union (2003) concluded, “Overall, all of these [tumor] findings in rats and mice are not considered toxicologically significant. Consequently, it is concluded that bisphenol A was not carcinogenic in this study in both species.” Haighton et al. (2002) concluded, “Overall, the results of this bioassay did not provide any compelling evidence to indicate that [bisphenol A] was carcinogenic in F344 rats or in B6C3F1 mice.” Based on the experimental animal data, the European Union concluded that “…the evidence suggests that bisphenol A does not have carcinogenic potential.” Using a weight of evidence approach, Haighton et al. (2002) concluded that bisphenol A was not likely to be carcinogenic to humans. This conclusion was based on NTP study results; lack of activity at noncytotoxic concentrations in both in vitro genetic toxicity tests and in an in vivo mouse micronucleus test; and data from metabolism studies that show rapid glucuronidation and no formation of possibly reactive intermediates, with the possible exception of reactive intermediates potentially generated as a result of saturated detoxification pathways at high-doses.

2.5 Potentially Susceptible Subpopulations

As noted in Section 2.1.1.3, one pathway of bisphenol A metabolism in humans and experimental animals is glucuronidation. Studies in experimental animals demonstrated that both the intestine and liver can glucuronidate bisphenol A. UGT2B1 was identified as the isoform involved in bisphenol A glucuronidation in rat liver (Yokota et al., 1999). The UDPGT isoform involved in human intestinal glucuronidation of bisphenol A is not known to have been identified. Despite uncertain isoform identification, studies in humans and experimental animals demonstrate developmental changes in expression of activities of several UDPGT isoforms that potentially affect bisphenol A metabolism.

Coughtrie et al. (1988) examined the ontogeny of UDPGT activity in human liver microsome samples obtained postmortem from adults and premature or full-term infants. Results of this analysis are listed in Table 59. Activities for isoenzymes catalyzing glucuronidation of bilirubin, testosterone, and 1-napthol were very low at birth in premature and full-term infants. Activities increased with age for the isoenzymes catalyzing glucuronidation of bilirubin (∼80% of adult levels by 8–15 weeks of age) and 1-naphthol (∼30% of adult levels at 8–15 weeks of age). During the first 55 weeks of life, no consistent increase in activity was noted for the isoenzyme catalyzing glucuronidation of testosterone. Using an immunoblot technique with antibodies developed toward liver testosterone/4-nitrophenol and kidney naphthol/bilirubin, 1 immunoreactive protein was observed in microsomes of 18- and 27-week-old fetuses and 3 immunoreactive proteins were observed in microsomes of full-term infants. Most isoenzymes present in adults were observed in infants within 3 months of age at levels ∼25% those of adults.

Table 59. Development of UDPGT Activity in Humansa
 UDPGT activity, nmol/min/mg protein
AgeBilirubinTestosterone1-Napthol
  • a

    aCoughtrie et al. (1988).

  • Data presented as individual values or mean±SD.

30 weeks gestation0.0500.56
30 weeks gestation with 10 weeks survival0.4; 10.14; 0.853.0; 1.8
Full-term infants surviving 1–10 days (n=7)0.07±0.040.10±0.060.75±0.68
Full-term infants surviving 8–15 weeks (n=6)0.64±0.320.12±0.052.4±1.1
Full-term infants surviving 22–55 weeks (n=5)0.99±1.10.09±0.063.6±2.1
Adult males (n=3)0.76±0.430.46±0.617.2±2.2

Strassburg et al. (2002) used a reverse transcript (RT)-polymerized chain reaction (PCR) technique to examine developmental changes in expression for 13 UDPGT genes in liver samples obtained from 16 pediatric patients undergoing liver transplant for extrahepatic biliary atresia (6–24 months old) and 12 adults undergoing liver transplant for carcinoma (25–75 years). Changes in gene expression were also assessed in hepatic RNA samples for two 20-week-old fetuses. No transcripts for UDPGT were detected in samples from 20-week-old fetuses. In infant and adult livers, transcripts were detected for UGT1A1, UGT1A3, UGT1A4, UGT1A6, UGT1A9, UGT2B4, UGT2B7, UGT2B10, and UGT2B15; there were no age-related differences in expression. Expression of UGT1A9 and UGT2B4 mRNA was lower in the pediatric samples. Western blot analyses of protein expression for UGT1A1, UGT1A6, and UGT2B7 were consistent with findings for mRNA expression. Activities toward 18 specific substrates were assessed in microsomes. In 13–24-month-old children compared to adults, glucuronidation activity was lower for ibuprofen (24-fold), amitriptyline (16-fold), 4-tert-butylphenol (40-fold), estrone (15-fold), and buprenorphine (12-fold).

Cappiello et al. (2000) compared uridine 5′-diphosphoglucuronic acid concentrations in livers and kidneys of human fetuses and adults and in placenta. In adults undergoing surgery, liver samples were obtained from 1 man and 4 women (23–72 years of age) and kidney samples were obtained from 1 woman and 4 men (55–63 years of age). Fetal livers and kidneys were obtained from 5 fetuses legally aborted between 16 and 25 weeks gestation. Five placenta samples were obtained on delivery at 17–25 weeks gestation. Compared to adults, fetal uridine 5′-diphosphoglucuronic acid concentrations were 5-fold lower in liver and 1.5-fold lower in kidney. Concentrations of uridine 5′-diphosphoglucuronic acid in placenta were 3–4-fold lower than in fetal liver. Based on these findings, study authors concluded that glucuronidation is potentially limited in the human fetus.

As noted in Sections 2.1.2.2 and 2.1.2.3, rat fetuses appear to have no or low ability to glucuronidate bisphenol A (Miyakoda et al., 2000; Matsumoto et al., 2002; Domoradzki et al., 2003). Although rats glucuronidate bisphenol A at birth, glucuronidation capacity appears to increase with age (Matsumoto et al., 2002; European-Union, 2003; Domoradzki et al., 2004).

Some possible interindividual or sex-related differences in the ability to produce the bisphenol A sulfate conjugate were identified in a limited number of human studies. As discussed in more detail in Section 2.1.1.3 and shown in Table 8, higher amounts of urinary bisphenol A sulfate were detected in 15 adult women than in 15 adult males (Kim et al., 2003b). In a study examining bisphenol A metabolism by human hepatocytes, an ∼10-fold higher concentration of a bisphenol A glucuronide/sulfate conjugate was observed in the sample from 1 female than in samples from 2 other females and 2 males (Pritchett et al., 2002).

Yang et al. (2003) examined the effects of polymorphisms in sulfotransferase enzymes on urinary excretion of total bisphenol A (conjugated and free) in Korean volunteers. Urinary bisphenol A concentrations were measured by HPLC and a PCR method was used to determine sulfotransferase genotype. The SULT1A1*1 allele was reported to have greater enzyme activity than the SULT1A1*2 enzyme and it was expected that individuals with the SULT1A1*1 allele would be able to rapidly eliminate bisphenol A. However, no significant differences in urinary bisphenol A concentrations were observed between 57 individuals with the SULT1A1*1 allele (geometric mean±SD=10.10±8.71 μg/L) and 15 individuals with the SULT1A1*2 enzyme (6.31±8.91 μg/L). Adjustment for possible bisphenol A exposure through vinyl wrap use also did not result in significant differences between the 2 groups. The study authors concluded that additional enzymes involved in bisphenol A metabolism should be studied to determine possible sensitivity differences.

One animal study demonstrated sex-related differences in sulfation. Male versus female Sprague–Dawley and F344 rats were found to produce higher amounts of a bisphenol A glucuronide/sulfate conjugate (Pritchett et al., 2002).

As noted in Table 7, one human study reported ∼2-fold higher blood bisphenol A concentrations in Japanese men than women (Takeuchi and Tsutsumi, 2002). Based on positive correlation between serum bisphenol A and testosterone concentrations, authors speculated that sex-related differences in bisphenol A concentrations might be due to androgen-related metabolism (Takeuchi and Tsutsumi, 2002). There are no known human studies showing inter-individual or sex-related variations in metabolism that could lead to higher bisphenol A concentrations in blood. Experimental animal studies have not consistently demonstrated higher concentrations of bisphenol A or radioactive dose in one sex (Pottenger et al., 2000; Kurebayashi et al., 2005). In Wistar rats orally administered 1 mg bisphenol A every 2 days for 2 or 4 weeks, liver microsomal UDPGT activity toward 17β-estradiol and testosterone and expression of UGT2B1 protein and mRNA were reduced in males but not females (Shibata et al., 2002). One study reported an ∼3-fold higher concentration of blood bisphenol A in male than in female Wistar–Imamichi rats that were apparently not treated, but there were was no sex-related difference in percent glucuronidated bisphenol A in serum (Takeuchi et al., 2004b). However, in an in vitro study conducted with hepatic microsomes, glucuronidation of bisphenol A and expression of UGT2B1 mRNA were higher in microsomes from female than male rats. As described in more detail in Section 2.1.2.3, one study showed reduced capacity to glucuronidate bisphenol A in livers from pregnant than in non-pregnant rats (Inoue et al., 2004).

2.6 Summary of General Toxicology and Biologic Effects

Analytical considerations.

Free concentrations of BPA measured in various matrices can be affected by analytic techniques and methodology. Free bisphenol A contamination from reagents and plastic ware may contribute to the measured free concentration of bisphenol A (Tsukioka et al., 2004; Völkel et al., 2005). Analytical techniques employed may incorrectly overestimate the free concentration of measured bisphenol A. HPLC with ultraviolet, fluorescence, or electrochemical detection is unable to make definitive identification of bisphenol A or bisphenol A glucuronides, because similar retention times may occur for the metabolites of other endogenous and exogenous compounds (Völkel et al., 2005). Bisphenol A glucuronide can also be hydrolyzed and in some cases degraded to unknown components either in acidic or basic pH solutions of diluted urine, adding another potential source of error in the measurement of sample levels of bisphenol A and its conjugates (Waechter et al., 2007). These considerations taken together, suggest that it is possible that free bisphenol A concentrations reported in biological samples may be overestimated.

2.6.1 Toxicokinetics and metabolism.

Human toxicokinetic data for bisphenol A are summarized in Table 60. In humans ingested bisphenol A is rapidly glucuronidated and circulated as bisphenol A glucuronide (Völkel et al., 2002). There is no evidence of enterohepatic circulation (Völkel et al., 2002). Most of the bisphenol A dose is excreted by humans through urine; bisphenol A recoveries in urine were reported at ≥84% within 5 hr of dosing (Völkel et al., 2005) and 100% within 42 hr of dosing (Völkel et al., 2002). Human urinary profiles were reported at ∼33–70% bisphenol A glucuronide, ∼10–33% parent compound, and ∼5–34% bisphenol A sulfate conjugate (Kim et al., 2003b; Ye et al., 2005). The presence of bisphenol A in human fetal tissues or fluids demonstrates that bisphenol A is distributed to the human conceptus (Ikezuki et al., 2002; Schönfelder et al., 2002b; Yamada et al., 2002; Kuroda et al., 2003; Tan and Mohd, 2003; Engel et al., 2006) (Table 61). Results from a limited number of studies indicated that fetal bisphenol A concentrations are within the same order of magnitude as maternal blood concentrations (Schönfelder et al., 2002b; Kuroda et al., 2003) and amniotic fluid bisphenol A concentrations are ∼1 order of magnitude lower than maternal blood concentrations (Yamada et al., 2002). Significantly higher mean bisphenol A concentrations were reported in the blood of male than female fetuses (3.5±2.7 vs. 1.7±1.5 ng/mL, P=0.016). Bisphenol A concentrations were measured in placenta samples at 1.0–104.9, median 12.7 μg/kg (Schönfelder et al., 2002b). There were no differences between pregnant and non-pregnant blood levels (median in μg/L 0.44, range=0.22–0.87; mean+SD=0.46+0.20) (Kuroda et al., 2003). Median bisphenol A concentrations in human milk were reported to be ≤1.4 μg/L (Calafat et al., 2006; Ye et al., 2006). One of the studies reported a milk/serum ratio of 1.3 (Sun et al., 2004).

Table 60. Human Toxicokinetic Values for Total Bisphenol A Dose
EndpointValueReference
Oral absorption, %≥84%Völkel et al.(2002, 2005)
Dermal absorption, in vitro, %∼10%European-Union (2003)
Tmax, min80Völkel et al. (2002)
Elimination half-life, hr4–5.4Völkel et al. (2002, 2005)
Table 61. Concentrations of Bisphenol A in Maternal and Fetal Samplesa
 Bisphenol A concentrations, μg/L, median (range) or mean±SD 
 Serum or plasma  
Study description; analytical methodMaternalFetalOther fetal compartmentsReference
  • a

    aAs discussed in Section 1.1.5, ELISA may overestimate bisphenol A concentrations so only results from studies based on HPLC, GC/MS, and LC/MS are presented.

21 samples collected in women in the U.S. before 20 weeks gestation; LC with electrochemical detection  0.5 (Non-detectable <0.5–1.96) 10% of amniotic fluid samples detectableEngel et al. (2006)
37 German women, 32–41 weeks gestation; GC/MS3.1 (0.3–18.9); 4.4±3.92.3 (0.2–9.2); 2.9±2.512.7 (ng/g) (1.0–104.9) 11.2±9.1) Placental tissueSchönfelder et al. (2002b)
9 sets of maternal and umbilical cord blood samples obtained at birth in Japanese patients; HPLC0.43 (0.21–0.79) 0.46±0.20.64 (0.45–0.76) 0.62±0.13 Kuroda et al. (2003)
180 Malaysian newborns; GC/MS Non-detectable (<0.05) to 4.05 88% of samples detectable Tan and Mohd (2003)

Animal toxicokinetic data for bisphenol A are summarized in Table 62. Following oral intake of bisphenol A by rats, most of the dose (≥77%) is glucuronidated and circulated as bisphenol A glucuronide (Miyakoda et al., 2000; Domoradzki et al., 2003; Kurebayashi et al., 2005). Most bisphenol A (90–95%) circulates bound to plasma proteins (Kurebayashi et al., 2003) [reviewed in (Teeguarden et al., 2005)]. In a single study in mice injected with a low dose (0.025 mg/kg bw), the most abundant compound in most tissues was bisphenol A glucuronide (Zalko et al., 2003). Most of a bisphenol A dose is circulated as the glucuronide in monkeys (Kurebayashi et al., 2002). As noted in Table 63, free bisphenol A in blood represents ≤8% of the dose in rats and ≤1% of the dose in monkeys following oral dosing; higher concentrations of free bisphenol A in blood were observed following parenteral dosing. The presence of 2 or more Cmax values for radioactivity or bisphenol A, an indication of enterohepatic circulation, was noted in some rat studies (Upmeier et al., 2000; Domoradzki et al., 2003; Kurebayashi et al., 2005). In rats, glucuronidation of bisphenol A was shown to occur in intestine (Sakamoto et al., 2002; Inoue et al., 2003a) and liver (Inoue, 2004). UGT2B1 was identified as a liver enzyme capable of glucuronidating bisphenol A, and possible involvement of other liver isoforms was noted (Yokota et al., 1999). There are some data indicating that glucuronidation capacity is very limited in fetuses and lower in immature than adult animals. Little-to-no UGT2B activity toward bisphenol A was detected in microsomes of rat fetuses; activity of the enzyme increased linearly following birth (Matsumoto et al., 2002). In an in vitro study comparing clearance of bisphenol A by hepatic microsomes from rats of different ages, activity was lower in microsomes from fetuses than in those from immature animals and adults [reviewed in (European-Union, 2003)]. As noted in Table 63, immature rats are capable of glucuronidating bisphenol A, but activity appears to increase with age. One study demonstrated that neonatal rats were able to glucuronidate a larger fraction of a lower (1 mg/kg bw) than higher (10 mg/kg bw) bisphenol A dose (Domoradzki et al., 2004).

Table 62. Toxicokinetic Values for Bisphenol A in Non-Pregnant Animals
ModelEndpointValueReference
  1. a

    aResults presented for low and high dose.

Rats orally exposed to ≤100 mg/kg bwTmax, hr0.083–0.75Domoradzk et al. (2004); Pottenger et al. (2000); Negishi et al. (2004b); Takahashi and Oishi (2000); Yoo et al. (2001)
Ovariectomized, adult rats gavaged with bisphenol A at 10 and 100 mg/kg bwTmax1 /Tmax2, hr0.5–1.5 / 3–6Upmeier et al. (2000)
Immature rats orally dosed with ≤10 mg/kg bwTmax hr0.25–3Domoradzk et al. (2004)
Monkeys orally dosed with ≤100 mg/kg bwTmax, hr0.7Negeshi et al. (2004b)
Chimpanzees orally dosed with 10 mg/kg bwTmax, hr0.5Negeshi et al. (2004b)
Rats s.c. dosed with ≤100 mg/kg bwTmax, hr1Negeshi et al. (2004b)
Monkeys s.c. dosed with ≤100 mg/kg bwTmax, hr2Negeshi et al. (2004b)
Chimpanzees s.c. dosed with 10 mg/kg bwTmax, hr2Negeshi et al. (2004b)
Ovariectomized, adult rats orally dosed with bisphenol A at 10 and 100 mg/kg bwBioavailability, %16.4 and 5.6aUpmeier et al. (2000)
Rats orally dosed with 10 mg/kg bwBioavailability, %5.3Yoo et al. (2001)
RatPlasma protein binding, %90–95%.Kurebayashi et al. (2003); Teeguarden et al. (2005)
Rats orally dosed with 10 mg/kg bwCmax, μg/L14.7–63Domoradzk et al., (2004); Yoo et al. (2001)
Rats orally dosed with 100 mg/kg bwCmax, μg/L580Negeshi et al. (2004b).
Ovariectomized, adult rats orally dosed with (mg/kg bw):Cmax1/Cmax2, μg/L Upmeier et al. (2000)
 10  30/40
 100  150/134
Oral doses (mg/kg bw) in immature rats at each age:Cmax (μg/L)Range of values for males and females:Domoradzki et al. (2004)
 1 (PND 4)  30–60 
 10 (PND 4)  10,200–48,300 
 1 (PND 7)  40–80 
 10 (PND 7)  1100–1400 
 1 (PND 21)  5–6 
 10 (PND 21)  200 
Monkeys orally dosed with 10 and 100 mg/kg bwCmax, μg/L2793 and 5732aNegeshi et al. (2004b)
Monkeys orally dosed with 10 mg/kg bwCmax, μg/L96–325Negeshi et al. (2004b)
Rats s.c. dosed with 10 and 100 mg/kg bwCmax, μg/L872 and 3439aNegeshi et al. (2004b)
Monkeys s.c. dosed with 10 and 100 mg/kg bwCmax, μg/L57,934 and 10,851aNegeshi et al. (2004b)
Chimpanzees s.c. dosed with 10 mg/kg bwCmax, μg/L1026–2058Negeshi et al. (2004b)
Oral doses (mg/kg bw) in immature rats at each age:AUC, μg-hr/LRange of values for males and females:Domoradzki et al. (2004)
 1 (PND 4)  100–200
 10 (PND 4)  7200–23,100
 1 (PND 7)  100
 10 (PND 7)  1700–1900
 10 (PND 21)  1000–1100
Rats orally dosed with 10 mg/kg bwAUC, μg-hr/L85.6Yoo et al. (2001)
Rats orally dosed with 100 mg/kg bwAUC0–24 h, μg-hr/L1353Negeshi et al. (2004b).
Monkeys orally dosed with 10 and 100 mg/kg bwAUC0–24 h, μg-hr/L3247 and 52,595aNegeshi et al. (2004b).
Chimpanzees orally dosed with 10 mg/kg bwAUC0–24 h, μg-hr/L813–1167Negeshi et al. (2004b).
Rats s.c. dosed with 10 and 100 mg/kg bwAUC0–24 h, μg-hr/L3377 and 23,001aNegeshi et al. (2004b).
Monkeys s.c. dosed with 10 and 100 mg/kg bwAUC0–24 h, μg-hr/L3247 and 39,040aNegeshi et al. (2004b).
Chimpanzees s.c. dosed with 10 mg/kg bwAUC0–24 h, μg-hr/L12,492–21,141Negeshi et al. (2004b).
Rats orally dosed with 10 mg/kg bwApparent volume of distribution, L/kg4273Yoo et al. (2001)
Immature rats orally dosed with ≤10 mg/kg bwHalf-life, hr4.3–21.8Domoradzki et al. (2004)
Rats orally dosed with 10 mg/kg bwTerminal elimination half-life, hr21.3Yoo et al. (2001)
Rats orally dosed with 10 mg/kg bwOral clearance, mL/min/kg2352.1Yoo et al. (2001)
Table 63. Age and Route Factors Affecting Free Bisphenol A Concentrations in Blood
Model and regimenFindings for free bisphenol A in bloodReference
Age effects of rat oral dosing at 1 or 10 mg/kg: Domoradzki et al. (2004)
 4 days of age1.5–56.8 mg/L 
 7 days of age1.1–12.2 mg/L 
 21 days of age0.8–8 mg/L 
 Adulthood0.07–0.6 mg/L 
S.C. or gavage dosing of 18–21-day-old rats with 160 mg/kg bw/day[93% lower] with oral than s.c. dosing 2.9 mg/L s.c. (plasma) 0.2 mg/L oral (plasma)Yamasaki et al. (2000)
Route effects in rats administered 10 or 100 mg/kg bw: Pottenger et al. (2000)
 Oral[<2–8%] BLQ (males); 0.04 mg/L (females) (at 10 mg/kg) 
 s.c.[65–76%] 0.69 (males); 0.87 mg/L (females) (at 10 mg/kg) 
 i.p.[27–51%] 0.39 (males); 0.34 mg/L (females) (at 10 mg/kg) 
Route effects in monkeys:Percent of dose:Kurebayashi et al. (2002)
 i.v.5–29 
 Oral0–1 

Kurbayashi et al. (2005) evaluated fetal and maternal rat bisphenol A during different stages of pregnancy. Bisphenol A labeled with carbon-14 was administered p.o. to male and female Fischer (F344) rats at relatively low doses (20, 100, and 500 μg/kg), and i.v. injected at 100 and 500 μg/kg). 14C-bisphenol A (500 μg/kg) was also administered orally to pregnant and lactating rats to examine the transfer of radioactivity to fetuses, neonatal rats, and milk (Table 64). Radioluminographic determination using phosphor imaging plates was employed to achieve highly sensitive determination of radioactivity. Absorption ratios of radioactivity after three oral doses were high (35–82%); parent 14C- bisphenol A in the circulating blood was quite low, however, suggesting considerable first-pass effect. After an oral dose of 100 μg/kg 14C- bisphenol A, the radioactivity was distributed and eliminated rapidly, but remained in the intestinal contents, liver, and kidney for 72 hr. The major metabolite in the plasma and urine was bisphenol A glucuronide, whereas most of the bisphenol A was excreted with the feces as free bisphenol A. A second peak in the time-course of plasma radioactivity suggested enterohepatic recirculation of bisphenol A glucuronide. There was limited distribution of 14C- bisphenol A to the fetus and neonate after oral administration to the dam. Significant radioactivity was not detected in fetuses on GD 12 and 15. On GD 18, however, radioactivity was detected in the fetal intestine and urinary bladder 24 hr after oral dosing of 14C- bisphenol A to the dam. The distribution pattern of radioactivity in pregnant rats was essentially the same as that in non-pregnant female rats. The distribution levels were dose-dependent in most of the tissues. There was limited distribution of 14C-bisphenol A to the fetus. Radioactivity in fetal tissues was undetectable except on GD 18 in the fetal urinary bladder and intestine. On GD 18, the amount of radioactivity in fetal tissues at 24 hr was about 30% that in maternal blood, and the yolk sac contained a much higher level of radioactivity than the maternal blood. The Expert Panel thought these differences were a consequence of the routes of administration, i.v. or p.o., because only trace amounts of parent bisphenol A dosed orally appeared in the plasma.

Table 64. Tissue Radioactivity in Pregnant and Fetal Rats After Oral Administration of 500 μg/kg 14C-Bisphenol A to Dams
 Radioactivity concentration (ng bisphenol A eq. g−1 or mL−1)
 2 days of gestation15 days of gestation18 days of gestation
Dam and fetal tissues30 mina124 hr30 mina24 hr30 mina24 hr
  • a

    aEach time shows the sacrifice time after oral administration of 14C-bisphenol A to each pregnant rat.

  • ND, not determined (indistinguishable); NQ, nonquantifiable (below LOQ).

Dams
Amniotic fluidNDNDNQNQNQNQ
Blood43.324.3337.513.8330.9910.79
Ovary21.943.9613.91NQ15.673.49
Placenta15.43NQ18.12NQ9.913.86
Uterus22.68NDNQNQ15.31NQ
FetusNQNQNQNQNQ3.28
Fetal membraneNQNQNQNQNQ10.87
Yolk sacNQNDNDNDNQ54.14

The major metabolite of bisphenol A is the glucuronide conjugate. Another metabolite that has been commonly detected in urine is bisphenol A sulfate. Excretion patterns for bisphenol A are summarized in Table 65. As noted in Table 65, the major elimination routes for bisphenol A in rodents are feces and bile; a lower percentage of the dose is eliminated through urine. The major compound detected in feces is bisphenol A and the major compound detected in bile and urine is bisphenol A glucuronide. Excretion patterns and metabolic profiles observed in rodents dosed orally or parenterally with low (<1 mg/kg bw/day) or high-doses (10–100 mg/kg bw/day) were similar. In contrast to rodents and similar to humans, most of the dose in orally- or i.v.-exposed monkeys was eliminated through urine.

Table 65. Summary of Elimination Information for Bisphenol A
ModelElimination routeDose eliminated (%)Compound and metabolite profileReference
Pregnant or non-pregnant rats orally, i.p., or s.c. dosed with <100 mg/kg bwFeces50–83Bisphenol A (83–93%); Bisphenol A glucuronide (2–3%)Domoradzki et al. (2003); Snyder et al. (2000); Pottenger et al. (2000)
 Urine13–42Bisphenol A (3–23%); Bisphenol A glucuronide (57–87%); Bisphenol A sulfate (2–7%) 
Rats orally or i.v. dosed with 0.1 mg/kg bwFeces64–88Not reportedKurebayashi et al. (2003)
 Urine10–34  
Rats orally or i.v. dosed with 0.1 mg/kg bwBile45–66 within 5 hrBisphenol A glucuronide (84–88%)Kurebayashi et al. (2003)
Rats orally dosed with 100 mg/kg bw/dayFecesNot reportedBisphenol A (61% of dose)Kurebayashi et al. (2003)
 Urine Bisphenol A and bisphenol A sulfate (≤1.1% of dose); Bisphenol A glucuronide (6.5% of the dose) 
 Bile Bisphenol A glucuronide (41% of dose) 
Pregnant mice injected with 0.025 mg/kg bw bisphenol AFecesNot reportedBisphenol A (>95%)Zalko et al. (2003)
 Urine Major metabolites: bisphenol A glucuronide and hydroxylated bisphenol A glucuronide 
 Bile Bisphenol A glucuronide (>90%) 
Monkeys orally or i.v. dosed with 0.1 mg/kg bwFeces2–3Not reportedKurebayashi et al. (2002)
 Urine79–86  

Toxicokinetics of bisphenol A were examined in pregnant rats and are summarized in Table 66 for free bisphenol A and Table 67 for total dose. One study demonstrated similar disposition, metabolism, and elimination of bisphenol A in pregnant and non-pregnant rats (Domoradzki et al., 2003). A number of rodent studies demonstrated distribution of bisphenol A or radioactive dose to fetuses following oral dosing of the dam (Miyakoda et al., 1999; Takahashi and Oishi, 2000; Domoradzki et al., 2003; Kim and Hwang, 2003; Kabuto et al., 2004; Kurebayashi et al., 2005). Bisphenol A distribution to fetus was also demonstrated with i.v. dosing of rats (Shin et al., 2002) and s.c. dosing of mice or monkeys (Uchida et al., 2002; Zalko et al., 2003). In a study in which bisphenol A was orally administered to rats on GD 19, bisphenol A glucuronide was not detected in fetuses (Miyakoda et al., 2000); study authors noted the possibilities that bisphenol A glucuronide was not likely transferred from dams to fetuses and that fetuses do not likely possess glucuronidation ability. Some of the studies demonstrated slower elimination of bisphenol A from fetuses than maternal blood following oral dosing (Miyakoda et al., 1999; Takahashi and Oishi, 2000).

Table 66. Toxicokinetic Values for Free Bisphenol A in Pregnant Rats and Fetuses
DoseEndpointMaternalFetalReference
1000 mg/kg bw orally on GD 18Cmax, μg/L14,7009220Takahashi and Oishi (2000)
10 mg/kg bw orally on GD 19Concentration 1-hr post-dosing, μg/L3411Miyakoda et al. (1999)
2 mg/kg bw i.v. on 1 day between GD 17–19Cmax, μg/L927.3794Shin et al. (2002)
1000 mg/kg bw orally on GD 18Tmax, min2020Takahashi and Oishi (2000)
2 mg/kg bw i.v. on 1 day between GD 17–19Tmax, hrNo data0.6±0.3Shin et al. (2002)
1000 mg/kg bw orally on GD 18AUC, μg · hr/L13,10022,600Takahashi and Oishi (2000)
2 mg/kg bw i.v. on 1 day between GD 17–19AUC, μg · hr/L905.51964.7Shin et al. (2002)
1000 mg/kg bw orally on GD 18Mean retention time, hr10.620.0Takahashi and Oishi (2000)
1000 mg/kg bw orally on GD 18Variance in retention time, hr2203419Takahashi and Oishi (2000)
2 mg/kg bw i.v. on 1 day between GD 17–19Mean residence time, hr3.03.0Shin et al. (2002)
1000 mg/kg bw orally on GD 18Half-life, hr:  Takahashi and Oishi (2000)
 From 20–40 min0.09520.55 
 From 40 min–6 hr2.581.60 
 From 6–48 hr4.65173 
2 mg/kg bw i.v. on 1 day between GD 17–19Elimination half-life, hr2.52.2Shin et al. (2002)
Table 67. Toxicokinetic Values for Radioactive Dose in Pregnant Rats (Total Bisphenol A)a
EndpointValue
  • a

    aDormoradzki et al. (2003).

  • Dams were gavaged with 10 mg/kg bw/day on GD 6–10, 14–18, or 17–21.

  • GD, gestation day.

Cmax1/ Cmax2, μg eq/L370–1006/211–336
Tmax1/ Tmax2, hr0.25/12–24
Time to non-quantifiable concentration, hr72–96
AUC 14C, μg –eq · hr/L7100–12,400
AUC Bisphenol A glucuronide,  μg –eq · hr/L6800–12,300

Toxicokinetics data in lactating rats are summarized in Table 68 for free bisphenol A and Table 69 for total dose. Distribution of bisphenol A to milk and/or nursing pups was demonstrated in rodent studies with oral or i.v. exposures (Snyder et al., 2000; Yoo et al., 2001; Kurebayashi et al., 2005). One study reported that most of the bisphenol A dose is present as bisphenol A glucuronide in milk of lactating rats (Snyder et al., 2000). In a study that compared bisphenol A concentrations in maternal serum, milk, and offspring after rat dams were administered low oral doses (0.006 or 6 mg/kg bw/day), a significant increase in bisphenol A concentration was only observed in the serum of dams from the high-dose group on PND 21; no increase was observed in milk or pups (Yoshida et al., 2004). Another study demonstrated higher concentrations of bisphenol A in milk compared to maternal serum after i.v. dosing of rat dams (Yoo et al., 2001).

Table 68. Toxicokinetic Values for Free Bisphenol A in Lactating Ratsa
EndpointBlood valueMilk value
  • Rats were i.v. injected 0.47, 0.94, or 1.88 mg/kg bw and then infused over a 4-hr time period with 0.13, 0.27, 0.54 mg/hr.

  • a

    aYoo et al. (2001).

  • b

    bEffect at each dose, from low to high dose.

Systemic clearance, mL/min/kg119.2/142.4/154.1b 
Steady state bisphenol A concentration, ng/mL66.1/120.0/217.1b173.1/317.4/493.9b
Milk/serum ratio2.7/2.6/2.4b 
Table 69. Toxicokinetic Values for Radioactive Dose in Lactating Rats (Total Bisphenol A)a
EndpointBlood valueMilk value
  • a

    aKurebayashi et al. (2005).

  • Rats were orally dosed with 0.5 mg/kg bw on PND 11.

Cmax, μg – eq/L27.24.46
Tmax, hr48
Elimination half-life, hr3126
AUC (0–48 hr), μg – eq · hr/L)689156

A number of in vitro studies compared bisphenol A metabolic velocity rates in microsomes or hepatocytes from rodents and humans. Generally, faster rates were demonstrated by rodent than human hepatocytes and microsomes (Elsby et al., 2001; Pritchett et al., 2002) [reviewed in (European-Union, 2003)]. One of the studies noted that adjustment for total hepatocyte number in vivo resulted in higher predicted rates for humans than rodents (Pritchett et al., 2002). The European Union (2003) noted that the interpretation of such studies should included knowledge about in vivo conditions such as varying metabolic capacity of hepatic cells, relationship of hepatic size to body size, and possibly important physiological endpoints such as blood flow.

2.6.2 General toxicity.

Gross signs of toxicity observed in rats acutely exposed to bisphenol A included pale livers and gastrointestinal hemorrhage [reviewed by the (European-Union, 2003)]. Acute effects of inhalation exposure in rats included transient and slight inflammation of nasal epithelium and ulceration of the oronasal duct. Based on LD50 observed in animals, the European Union (2003) concluded that bisphenol A is of low acute toxicity through all exposure routes relevant to humans. According to the European Union (2003), there is evidence that bisphenol A is irritating and damaging to the eye and is irritating to the respiratory tract and possibly the skin. Findings regarding sensitization potential were not clear.

Possible target organs or systems of toxicity identified in repeat-dose animal studies with oral dosing included intestine, liver, kidney, and male, and female reproductive systems [reviewed in (NTP, 1982; Yamasaki et al., 2002a; European-Union, 2003)]. Intestinal findings (effect levels) in rats included cecal enlargement (≥25 mg/kg bw/day) and cecal mucosal hyperplasia (≥200 mg/kg bw/day). Hepatic effects included prominent hepatocyte nuclei or inflammation in rats (≥500 mg/kg bw/day), multinucleated giant hepatocytes in mice (≥120 mg/kg bw/day), and increased weight with no evidence of histopathology in dogs (≥270 mg/kg bw/day). Renal tubule degeneration or necrosis was observed in rats dosed with ≥500 mg/kg bw/day. Reproductive findings are discussed in Section 4.0. Effects in subchronic inhalation studies in rats included cecal enlargement resulting from distention by food and transient, slight hyperplasia and inflammation of epithelium in the anterior nasal cavity; both effects occurred at (≥50 mg/m3).

2.6.3 Estrogenicity.

Estrogenicity of bisphenol A has been evaluated using in vitro (Table 52) and in vivo (Table 53) assays. In those studies estrogenic potency was compared to 17β-estradiol, ethinyl estradiol, diethylstilbestrol, and, in one study, estrone. There is considerable variability in the results of these studies with the estrogenic potency of bisphenol A ranging over about 8 orders of magnitude (Fig. 2). On the other hand, the average potency only differs by 1 order of magnitude and there is very little difference between rat and mouse means.

Most in vivo estrogenicity studies examined effects on uterine weights of intact weanling or ovariectomized adult rats or mice. The potency of bisphenol A in increasing uterine weight varied over ∼4 orders of magnitude. Uterine weight findings can be affected by the time period between dosing and measurement. Most, but not all studies, showed a greater effect on uterine weight with s.c. than with oral dosing. The greater activity of s.c. than oral bisphenol A is presumably due to glucuronidation of the orally administered compound with consequent loss of estrogenicity (Matthews et al., 2001). Inter-strain variability in rats has been evaluated as a source of variability in estrogenicity assays. (see Section 4.0 for additional discussion) Inter-laboratory variability has been noted for uterotropic effects in intact weanling mice exposed to bisphenol A (Tinwell and Joiner, 2000); one factor that can result in variability is body weight of the animal. Use of mice with lower body weights results in lower and less variable control uterine weights and greater likelihood of bisphenol A effect (Tinwell and Joiner, 2000; Ashby et al., 2004). In in vivo studies examining gene expression profiles, some but not all gene expression changes were consistent between bisphenol A and reference estrogens (Tinwell and Joiner, 2000; Naciff et al., 2002; Singleton et al., 2004; Terasaka et al., 2006); ER-independent activity was suggested by 1investigator (Singleton et al., 2004). [Based on one comprehensive study of the effects of bisphenol A orally delivered from 601000 mg/kg for 37 days, the Expert Panel concludes that the uterotrophic responses were only found at higher does (Ashby,2002; Kanno et al.,2003a) whereas s.c. dosing produced consistent uterine weight increases at lower doses.]

2.6.4 Androgenic activity.

In the majority of in vitro tests conducted, bisphenol A was not demonstrated to have androgenic activity (Sohoni and Sumpter, 1998; Gaido et al., 2000; Kitamura et al., 2005; Xu et al., 2005). Anti-androgenic activity was demonstrated in systems using cells transfected with three different androgen receptor reporting systems (ARE-luc, MMTV-lacZ, and C3-luc) (Table 56). No consistent effects were observed on male accessory reproductive organ weights in 3 in vivo studies in which rats were dosed with bisphenol A at ≤600 mg/kg bw/day; the study authors concluded that bisphenol A does not have anti-androgenic or androgenic activity (Kim et al., 2002a; Yamasaki et al., 2003; Nishino et al., 2006).

2.6.5 Genetic toxicity.

In in vitro genetic toxicity studies reviewed by the European Union (2003) and/or Haighton et al. (2002), evidence of aneugenic potential, chromosomal aberration, micronuclei formation, and DNA adducts was observed (Table 57). Because of the lack of chromosomal effects in in vivo studies (Table 58) and unknown relevance of DNA adduct formation, which only occurred at high-doses, both groups concluded that bisphenol A is not likely to have genotoxic activity in vivo.

2.6.6 Carcinogenicity.

Carcinogenic potential of bisphenol A was evaluated in rats and mice by the NTP (1982) and Huff (2001). NTP concluded that under the conditions of the study, there was no convincing evidence that bisphenol A was carcinogenic in F344 rats or B6C3F1 mice. However, NTP stated that there was suggestive evidence of increased cancer in the hematopoietic system based on marginally significant increases in leukemia in male rats, non-statistically significant increases in leukemia in female rats, and a marginally significant increase in combined incidence of lymphoma and leukemia in male mice. A statistically significant increase in testicular interstitial cell tumors in aging F344 rats was also considered suggestive evidence of carcinogenesis. The effect was not considered conclusive evidence because of the high incidence of the testicular neoplasm in aging F344 rats (88% incidence in historical controls). Both the European Union (2003) and Haighton et al. (2002) stated that the evidence does not suggest carcinogenic activity of bisphenol A in rats or mice. Conclusions by the European Union (2003) and Haighton et al. (2002) were based on factors such as lack of statistical significance for leukemia, mammary gland fibroadenoma, and Leydig cell tumors, lack of activity at noncytotoxic concentrations in both in vitro genetic toxicity tests and an in vivo mouse micronucleus test, and unlikely formation of reactive intermediates at doses that do not saturate detoxification pathways.

2.6.7 Potentially sensitive subpopulations.

Studies in humans and laboratory animals demonstrated developmental changes in UDPGT gene expression or enzyme activity that could potentially affect the concentration of free bisphenol A reaching target organs because of a differential capacity for bisphenol A glucuronidation. In humans, activities for some UDPGT isozymes were reported to be very low at birth but increased with age (Coughtrie et al., 1988). No transcripts for UDPGT were detected in samples from 20-week-old human fetuses and activity for some UDPGT enzymes was lower in children than adults (Strassburg et al., 2002). Compared to adults, human fetal uridine 5′-diphosphoglucuronic acid concentrations were 5-fold lower in liver and 1.5-fold lower in kidney (Cappiello et al., 2000). It is not clear if any of the isozymes examined are involved in bisphenol A glucuronidation by humans. Human findings were consistent with rodent studies that demonstrated no or limited glucuronidation capacity by fetuses (Miyakoda et al., 2000; Matsumoto et al., 2002; Domoradzki et al., 2003) and lower glucuronidation capacity in immature than adult rats (Matsumoto et al., 2002; European-Union, 2003; Matsumoto et al., 2004).

Some studies suggested possible gender-related differences in sulfation capacity in humans (Pritchett et al., 2002; Kim et al., 2003b) and laboratory animals (Pritchett et al., 2002). One study in humans demonstrated no differences in urinary bisphenol A concentrations in individuals carrying a sulfotransferase genotype associated with greater activity (Yang et al., 2003).

3.0 DEVELOPMENTAL TOXICITY DATA

The Panel attended to multiple design and analysis characteristics in judging the acceptability of individual studies. It was our consensus that for a study to be acceptable for this review process, several conditions had to be met. First, effects related to litter of origin needed to be accounted for in design and statistical procedures. Second, animals needed to be dosed via the dam or directly under individual housing conditions. Concern that multiple exposures within a cage to different animals could cause cross-animal contamination across cage-mates led to the determination that this design was not acceptable. Third, a minimum of 6 animals per treatment condition needed to be used to provide minimal confidence in results. Fourth, if similar tests were conducted at multiple ages, the statistical analyses needed to account for repeated measurement in order not to inflate degrees of freedom. The Panel carefullyconsidered the merits of each study according to these primary criteria, and the related design characteristics represent the most common reasons for judging a study to be unacceptable for our review process. Our intent was to have our review depend most heavily on studies that would have reduced risks for false negative or false positive findings.

In addition, the Panel carefully considered the value of studies where bisphenol A was administered anywhere other than to the mouth or stomach of the experimental animal. Human exposure is overwhelmingly oral, and oral exposure produces an internal metabolite profile which is overwhelmingly dominated by the (inactive) glucuronide in both rats and humans. Subcutaneous or parenteral injections result in blood levels of active parent compound which are much higher than those seen after oral exposure. In light of these pharmacokinetic differences, the Panel concluded that injection studies, unless they proved otherwise, would produce irrelevantly high internal doses of the active parent compound, and would tend to produce “false positive” effects from the point of view of the human oral situation. Thus, the Panel viewed those otherwise adequate studies that injected bisphenol A as providing “supplemental” information (i.e., of limited utility), unless they also analyzed the levels of parent compound and metabolites after the injection. The intent of this approach is limit the impact of those studies which produced an unrealistic and irrelevant internal metabolite profile (i.e., one which is significantly different from that experienced by humans). Thus, the closer any given study came to replicating the human situation, the more weight it had in the final analysis.

The report below mentions “dosing procedures” as reasons for limiting the adequacy or utility of various studies. This has been used to mean non-gastric administration (s.c. injection, intramuscular [i.m.] injection, i.p. injection, or intracerebroventricular injection).

The Panel also had extensive discussion about dosing vehicles. Dimethyl sulfoxide (DMSO) has significant biological activities of its own (Santos et al., 2003), and the experience of the Panel is that DMSO can help move solutes into cells. Increasing the DMSO concentration can produce a greater solute effect, even when holding that solute concentration stable. The real impact of this for in vivo injections is uncertain, and this effect is likely to be small at the dosing volumes administered in the studies considered here. The use of 100% DMSO as a vehicle for ALZET mini-pump studies is a clear contravention of the directions for mini-pump use3, as it accelerates the breakdown of the mini-pumps and produces blood levels that are not predictable and therefore not useful for the Evaluative Process. Various oils each can bring their own potential issues, such as oxidative damage, but these were considered and discussed by a sub-team of the Panel and not considered to be consequential for this analysis

The Panel also examined the issue of data that would be expected to result when positive controls were employed. While we did not feel that positive controls were required for studies, when they were used, expected effects needed to be demonstrated to validate that the experimental model was capable of responding to a certain stimulus. This is of even more value when there is no response to the main exposure under study. When looking for estrogenic responses, investigators often use 17β estradiol or diethylstilbestrol. These must be used at adequate doses to produce the desired response. Inadequate challenge by the positive control, resulting in no response, leaves the reader uncertain whether the lack of response is due to the selection of too low a dose, or whether the experimental model is incapable of responding to a sufficient challenge. Even though the Panel, based on its own scientific experience, might conclude that inappropriately low doses had been selected and thus a lack of response is not surprising, the Panel was left with little choice in such situations but to give much less weight to studies where non-effective doses of a positive control compound were used.

The Panel is confident in our assessment of those studies judged adequate and useful, and are focusing our limited time on the consistency and utilization of these data.

3.1 Human

No studies were located on possible human developmental effects of bisphenol A.

3.2 Experimental Animal

Studies are presented by species (rat, mouse, other), route (oral, parenteral), and by whether exposure was during pregnancy or the postnatal period. Studies in which exposures were started during pregnancy and continued after pregnancy are discussed with studies in which exposures occurred postnatally.

3.2.1 Rat—oral exposure only during pregnancy.

3.2.1.1 Evaluation of pre- or perinatal growth and development:

Morrissey et al. (1987), supported by NTP/NCTR, examined the effects of prenatal bisphenol A exposure in rats and mice in a study conducted according to GLP. Studies are also available as NTP publications for rats (NTP, 1985c) and mice (NTP, 1985b). The study was conducted in two sets of rats and mice, and data were pooled for each species. [The data for mice are discussed inSection 3.2.5.1.] Pregnant CD rats were randomly assigned to groups of ≥10 animals in each set of the study, for a total of ≥20 animals/dose. On GD 6–15 (GD 0=sperm or plug), rats were gavaged with bisphenol A at 0 (corn oil vehicle), 160, 320, 640, or 1280 mg/kg bw/day. Doses were based on results of preliminary studies and were expected to result in 10% maternal mortality at the high-dose and no toxicity at the low dose. Purity of bisphenol A was >95% and 2,4′-bisphenol A was reported as an impurity. Dosing solution concentrations were verified. Pregnant animals were weighed during the study. Rats were killed on GD 20. Liver and uterus were weighed, and corpora lutea and implantation sites were examined. Fetuses were sexed, weighed, and examined for viability and external, visceral, and skeletal malformations. Data were analyzed by Bartlett test for homogeneity of variance, ANOVA and/or William multiple comparison, Dunnett, or Fisher exact probability tests. [Data were presented and analyzed on a per litter basis.]

An unexpectedly high number of dams (7 of 27) died in the 1280 mg/kg bw/day group, with most deaths occurring in the second set of animals. Because of the high death rate, the study authors decided not to evaluate data in the 1280 mg/kg bw/day group. Clinical signs that occurred most frequently in dams from the 640 mg/kg bw/day group included lethargy, piloerection, pica, rough coat, wet urogenital area, weight loss, and alopecia. Significant and dose-related decreases in maternal body weights were observed during the entire gestation period and thus were not confined to the GD 6–15 treatment period in rats from the 160, 320, and 640 mg/kg bw/day groups. Body weight corrected for gravid uterine weight was also decreased in all three dose groups. Effects on maternal body weight were most pronounced during the treatment period. [During the treatment period, dam body weights were 35, 53, and 54% lower in the 160, 320, and 640 mg/kg bw/day groups than in control groups; estimated benchmark doses4in mg/kg bw/day were BMD10 113, BMDL10 94, BMD1SD 416, BMDL1SD 321.] Despite this large effect on maternal body weight, there were no effects on numbers of implantation sites or resorptions, gravid uterine weight, or liver weight. The numbers of litters available for evaluation in the control and 160, 320, and 640 mg/kg bw/day dose group were 23, 26, 24, or 29. There were no significant effects on fetal body weight or viability, percentage males/litter, or malformed fetuses/litter. Study authors concluded that bisphenol A was not teratogenic in rats at doses that cause maternal toxicity.

Strengths/Weaknesses: This study used adequate sample sizes to evaluate the effects of GD 6–15 exposure on maternal body weight during gestation and on implantation and resorption sites/dam, fetal body weight, and fetal viability to GD 20. Strengths are the verification of dosing solutions, use of GLP, adequate n, sensitive evaluation of soft and hard-tissue structures. Weaknesses include no postnatal examination, as well as the absence of data from the 1280 mg/kg bw/day group, the absence of a no-effect dose. The absence of effects on fetal endpoints despite marked reductions in maternal body weight corrected for gravid uterine weight warrants the appropriate conclusion that bisphenol was not teratogenic when based on GD 20 data. Further, a gross visceral exam is likely insensitive to certain abnormalities of the reproductive tract and brain, as noted above.

Utility (Adequacy) for CERHR Evaluation Process: This study is adequate and of high utility for the evaluation process.

Kim et al. (2001b), support not indicated, examined the effects of prenatal bisphenol A exposure on developmental toxicity in rats. Sprague–Dawley rats were fed commercial rodent chow (Jeil Feed Co., Daejon, Korea) and housed in polycarbonate cages; no information was provided about bedding. Twenty dams/group were gavaged with 0 (corn oil vehicle), 100, 300, or 1000 mg/kg bw/day bisphenol A [purity not provided] on GD 1–20 (GD 0=first 24 hr after detection of vaginal sperm or plug). Dose selection was based on the results of a preliminary study that demonstrated maternal and developmental toxicity at doses ≥400 mg/kg bw/day and a lack of effect at doses ≤200 mg/kg bw/day. Endpoints examined in dams during the study were clinical signs, body weight gain, and food intake. Dams were killed on GD 21 and examined for corpora lutea and implantation sites. Fetuses were sexed, weighed, and examined for viability and external abnormalities. Anogenital distance was measured and alternate fetuses were examined for visceral and skeletal malformations. The dam or litter was considered the statistical unit. Data were analyzed by ANOVA, Scheffé multiple comparison test, Kruskal–Wallis nonparametric ANOVA, Mann–Whitney U-test, and Fisher exact probability test.

Statistically significant effects are summarized in Table 70. Dose-dependent clinical signs observed in dams at the 2 highest doses included piloerection, dull fur, reduced locomotor activity, emaciation, sedation, red-colored tears, soft stool, diarrhea, urination, and perineal soiling. Pregnancy failure, as observed by lack of implantation sites, was increased in females from the high-dose group. Maternal body weight, body weight gain, and body weight corrected for gravid uterus weight were reduced at the mid- and high-dose. GD 4 was the only time period when food intake was significantly reduced at the mid- and high-dose. Expansion and congestion of stomach and/or intestines were observed in dams from the high-dose group. Body weights of male fetuses were decreased at the mid- and high-dose, and body weights of female fetuses were reduced at the high-dose. Increases in fetal death, early resorption, and post-implantation loss, accompanied by reduced number of live fetuses, were observed at the high-dose. Anogenital distance was significantly reduced in males from the mid- and high-dose groups, but there were no differences in anogenital distance of males or females when the values were normalized by the cube root of body weight. Significantly reduced ossification was observed in the high-dose group. There were no treatment-related differences in fetal sex ratio or external, visceral, or skeletal malformations. Study authors concluded that exposure of rats to a maternally toxic dose of bisphenol A during the entire gestation period resulted in pregnancy failure, post-implantation loss, reduced fetal body weight, and retarded fetal ossification but not dysmorphogenesis.

Table 70. Maternal and Developmental Effects in Rats Exposed to Bisphenol Aa
 Dose, mg/kg bw/day
Endpoint1003001000BMD10BMDL10BMD1SDBMDL1SD
  • a

    aKim et al. (2001b).

  • ↑,↓ Statistically significant increase, decrease compared to controls; ↔ No statistically significant effect compared to controls.

Dams
 No. pregnant↓30%    
 Body weight gain↓35%↓52%178152379304
 Corrected body weight↓14%↓15%631490566424
 Food intake on GD 4↓24%↓57%168147313257
 No. fetal deaths↑6.5-fold82713978585
 No. early resorptions↑6-fold82114980584
 Post-implantation losses↑11-fold1278394  
Fetuses
 No. live /litter↓36%929348982713
 Male body weight↓14%↓20%456339694497
 Female body weight↓21%439328682490
 Ossification    

Strengths/Weaknesses: This report presents a fairly standard embryo–fetal developmental toxicity study. One strength is that the doses utilized incorporated both a no-effect dose and a high maternally toxic dose, revealing fetal effects only at the high-dose that showed marked maternal toxicity. Measurement of anogenital distance is another strength. Weaknesses include the absence in all groups of information about postnatal viability, and postnatal function. Further, a gross visceral exam is likely insensitive to certain abnormalities of the reproductive tract and brain. However, this type of study does report on the ability of the exposure to cause structural malformations, which are notably absent.

Utility (Adequacy) for CERHR Evaluation Process: This study is adequate and of high utility for the evaluation process.

Kim et al. (2003), support not indicated, examined the effects of prenatal bisphenol A exposure on postnatal body and organ weights of Sprague–Dawley rats. Rats were housed in polycarbonate cages. [No information was provided on feed or bedding material.] Rats were grouped according to body weight and randomly assigned to dose groups. On GD 7–17 (GD 0=day of vaginal sperm or plug), at least 10 rats/dose group were gavaged with bisphenol A (>99.7% purity) at doses of 0 (corn oil vehicle), 0.002, 0.020, 0.200, 2, or 20 mg/kg bw/day. Dosing solution concentrations were verified. Dams were weighed and observed for clinical signs of toxicity during the study. Dams were killed on Day 21 of the postpartum period. Corpora lutea, implantation sites, resorptions, and fetal viability were assessed. Maternal liver, kidney, spleen, ovary, and gravid uterus were weighed. Live fetuses were weighed and examined for external and visceral abnormalities. Fetal liver, kidneys, spleen, and reproductive organs were weighed in half the fetuses. [These methods are produced here as written in the original; although dams were clearly stated to have been killed on PND 21, the “fetal” examinations described appear more consistent with killing of the dams on GD 21.] Data were analyzed by ANOVA and Student t-test. [It was not clear if the litter or fetus was considered the statistical unit in the evaluation of developmental toxicity data.]

A significant but non-dose-related increase in dam body weight occurred in the 0.2 mg/kg bw/day group on GD 0–15. Dam body weight was significantly increased on GD 21 in the 2 (by 53%) and 20 (by 43%) mg/kg bw/day groups. No significant differences in dam body weight were noted during the lactation period. Significant changes in dam relative organ weights (dose at which effects were observed) were: increased liver (0.002, 0.020, and 20 mg/kg bw/day); decreased right kidney (0.2 mg/kg bw/day); increased right kidney (2 mg/kg bw/day), and increased uterine (0.2 mg/kg bw/day). There was no effect on ovary weight of dams. The majority of dams were in diestrus when killed. One of 7 dams in the 0.2 mg/kg bw/day group was in proestrus. One of 7 dams in the 0.2 mg/kg bw/day, 1 of 6 dams in the 2 mg/kg bw/day group, and 2 of 8 dams in the 20 mg/kg bw/day group were in diestrus. Body weight effects in male and female offspring were reported in most treatment groups when evaluated at various time points between birth and PND 22. In general, when body weights effects were detected it was an increase in weight of ∼12–65%. [Changes occurred at most dose levels but were not consistent over time and there was little evidence of dose–response relationships. In general, effects appeared to be most pronounced in the lowest dose group.] Relative weights for several tissues attained statistical significance at 1 or more doses in offspring of both sexes: liver, spleen and right kidney. In addition, relative organ weights for were altered in males for the left kidney, both testes, right epididymis, left seminal vesicle, and prostate gland. There were no effects on ovary or uterus weights. [In most cases, there was little evidence of a dose–response relationship for organ weights, including male reproductive organs, in offspring.] Study authors concluded that bisphenol A had estrogenic effects on rat dams and offspring exposed during the gestation period.

Strengths/Weaknesses: While the verification of the dosing solution is a strength, this study is of unclear quality, to the point that there is real confusion about what was actually done. It is indicated that 10 dams were assigned to each dose group but numbers at sacrifice were 7, 7, 6, and 8 across the 4 doses. It is unclear whether fetal data were appropriately analyzed with litter as the unit. It is unclear when the dams were killed and analyzed. The absence of understandable dose-related effects complicates interpretation at these low doses; although the possibility of unusual low dose effects cannot be discounted.

Utility (Adequacy) for CERHR Evaluation Process: This study is inadequate for inclusion into the evaluation process, due to small sample size and poor documentation and communication about what was done.

3.2.1.2 Evaluation of reproductive organ development:

Talsness et al. (2000), supported by the German Federal Ministry for Environmental Protection and Radiation Security, examined the effect of prenatal bisphenol A exposure on the reproductive systems of male and female rats. [No information was provided about feed, caging, and bedding materials used.] On GD 6–21, Sprague–Dawley rats (n=18–20/group) were gavaged with 2% corn starch vehicle or bisphenol A [purity not indicated] at 0.1 or 50 mg/kg bw/day. A group of 11 dams was gavaged with 0.2 mg/kg bw/day ethinyl estradiol. Litters were weighed during the lactation period. Pups were weaned on PND 22 (according to Table 1 of the study, PND 1 was apparently the day of birth) and males and females were separated around PND 30. Vaginal opening was examined in 42–91 female offspring/group, and estrous cyclicity was monitored over a 3-week period in 42–53 females/group. At 4 months of age, 5–10 females/group were killed during diestrus and 20 females/group were killed while in estrus. A histopathological evaluation of vaginal tissue was conducted in 5 animals [assumed 5/group]. In 44–112 male offspring/group, anogenital distance was measured on PND 3, 15, and 21 and days of testicular descent and preputial separation were recorded. Males were killed on PND 70 (n=20/group) or 170 (n=17–20/group). Blood LH and testosterone concentrations were measured in 14–20 animals/group/time period. Sperm and spermatid numbers and sperm production and transit rates were determined in all offspring. Histopathological evaluation of the testis was conducted in 2 animals [assumed/group]. Body, reproductive organ, and liver weights were measured in all male and female offspring killed. Data from female rats were analyzed by ANOVA with post-hoc Dunnett test or Fisher test. Data from male rats were analyzed by ANOVA and Dunnett test. [It appears that offspring were considered the statistical unit.]

Pup body weights at birth were unaffected in the bisphenol A group, but on PND 22, pup body weights were lower [by 28%] in the low-dose group than in the control group. Study authors noted that the mean litter size in the low-dose group was larger by 2.6 pups than in the control group. Vaginal opening was delayed in the low-dose group and accelerated in the high-dose group. When estrous cyclicity data were evaluated according to total number of cycles, there was an increase in estrous phases lasting more than 1 day and prolongation of the cycle length in the high-dose group. Evaluation of estrous cycles by individual rat indicated a decrease in the percentage of low-dose females with 3 consecutive 1-day estrus phases. The only terminal body and organ weight effects occurred in the low-dose group and included decreased absolute liver weight in females killed in estrus and decreased body and uterus weights in females killed in diestrus or in estrus. There were no effects on relative organ weights. Histological observations in vaginal tissue of bisphenol A-exposed rats included less pronounced cornification during estrus and more pronounced mucification during diestrus, with magnitude of effect greater in the low- than the high-dose group. Observations in the animals exposed to ethinyl estradiol included decreased pup birth weight, delayed vaginal opening, near-persistent estrus, decreased absolute and relative uterus weights, and changes in vaginal histology similar to those described for the low-dose bisphenol A group.

Decreased anogenital distances was observed in the bisphenol A groups during all three time periods for male offspring, but the effect remained statistically significant only in the high-dose group when normalized for body weight. Testicular descent and preputial separation were delayed in the low-dose group. Organ weight effects that remained significant following adjustment for body weight included increased prostate weight in the high-dose group on PND 70 and increased testicular and epididymal weights in the low-dose group on PND 170. There was no effect on sperm morphology. Blood testosterone concentration was decreased in the high-dose group on PND 70, and blood LH concentration was increased in the high-dose group on PND 170. Testicular histopathology observations in the low-dose group on PND 70 included cellular debris in lumens, pyknotic nuclei in spermatids, and apoptotic debris in the region of the spermatogonia and primary spermatocyte. In testes of 70-day-old animals of the high-dose group, there were central necrotic masses, low numbers of meiotic figures in spermatocytes, and low spermatozoa numbers. On PND 170, observations in testes from the low-dose group included low spermatozoa numbers, a thin layer of spermatocyte meiotic figures, and apoptotic debris in region of spermatids. Low spermatocyte meiotic figures were the only testicular observation in the high-group on PND 170. Effects observed in the ethinyl estradiol group included increased anogenital distance, delayed testicular descent, accelerated preputial separation, decreased testis and prostate weights, decreased sperm counts and production, increased LH concentrations, increased testosterone concentrations on PND 170, apoptotic debris, and/or low sperm numbers in testes.

Study authors concluded that prenatal exposure to bisphenol A disrupts the reproductive systems of both male and female rats and that the effects do not occur according to a classic dose–response curve, which is generally observed in toxicology studies.

Strengths/Weaknesses: Strengths are the postnatal evaluation of various endpoints to “pup” adulthood and that the concentration of the dosing solutions was verified. Based on the description of numbers of pups contributing to various endpoints, however, the authors do not appear to have used the litter as the unit of analysis. These inflated numbers subjected to analysis complicate the interpretation of findings, especially for PND 1–21 measures. A weakness also is that only 2 dose levels were examined. The vaginal opening data for the controls were outside the normal range for Sprague–Dawley rats. It is unclear how the estrous cycle data were analyzed. The F1 data were not analyzed correctly. Data may be suggestive of developmental disruptions at both doses, but the magnitudes are likely unreliable, and the authors' statements about dose–response peculiarities must be viewed with caution until more complete dose–response assessments are published.

Utility (Adequacy) for CERHR Evaluation Process: This study is inadequate for the evaluation process.

Tinwell et al. (2002), support not indicated, examined the effects of in utero exposure to bisphenol A on sexual development of male rats. The study attempted to duplicate findings that were reported in several abstracts and as a full report (Talsness et al., 2000). Sprague–Dawley and Wistar-derived Alderley Park rats were housed in plastic-bottomed cages containing sawdust and shredded paper bedding. Rats were assigned to groups based on body weights and 6–7/group/strain were gavaged on GD 6–21 with bisphenol A (99% purity) at 0 (arachis oil vehicle), 0.020, 0.100, or 50 mg/kg bw/day. A positive control group initially received 200 μg/kg bw/day ethinyl estradiol, but the dose was reduced to 100 μg/kg bw/day between GD 11 and 14 due to maternal toxicity. Dosing solution concentrations and stability were verified. Dams were fed RM3 breeding diet (18.5% soybean protein; Special Diet Services, Ltd.) during gestation and lactation. At birth, pups were counted, sexed, and weighed. Anogenital distance was measured 24 hr following birth (PND 1). On PND 5, pups were culled to 8/litter, with equal numbers of males and females when possible. On PND 23, rats were weighed and housed according to sex. Following weaning, pups were fed RM1 feed (6.5% soybean protein). Pups were weighed throughout the post-lactation period. Ages at preputial separation, vaginal opening, and first estrus were assessed. Males were killed on PND 90–91 and females on PND 98. Liver and reproductive organs were weighed. Daily sperm production was determined. Data were analyzed using the litter and grouped individuals as the statistical unit. [Litter values are discussed below.] Data were analyzed by ANOVA, ANCOVA, and Dunnett test.

The only significant effect observed in female rats exposed to bisphenol A was a 1.6-day delay in vaginal opening in Alderley-Park rats of the high-dose group. The study authors stated that effect on vaginal opening was correlated with body weight. [Data were not shown by study authors.] In Alderley Park males of the high-dose group, significant reductions were observed for total sperm count/testis [12% lower than controls], sperm count/g testis [10% reduction], daily sperm count/testis [12% reduction], and daily sperm count/g testis [10% reduction]. Benchmark doses for the endpoints with statistically significant changes are shown in Table 71. In both strains, bisphenol A treatment had no effect on litter size, sex ratio, birth weight, anogenital distance, first day of estrus, or age of preputial separation. There were no significant effects on weights of liver, ovary, cervix, uterus, vagina, testis, epididymis, seminal vesicle, or prostate. Rats treated with ethinyl estradiol also experienced decreased sperm counts, in addition to decreased weights of male reproductive organs and advanced age of vaginal opening. Several findings (Talsness et al., 2000) were not duplicated in this study including: reduced anogenital distance; altered age of sexual maturation in males and females; variable changes in male reproductive organ weight, including prostate weight; and reduced sperm production at low doses. Study authors concluded that this study failed to confirm low-dose endocrine effects.

Table 71. Benchmark Doses for Rat Reproductive Organ Endpoints Affected by Prenatal Bisphenol Aa
 Benchmark dose, mg/kg bw/day
EndpointBMD10BMDL10BMD1SDBMDL1SD
  • a

    aCalculated from data in Tinwell et al. (2002).

Delayed vaginal opening68513516
Sperm count/testis55305731
Sperm count/g testis81416834
Daily sperm count/testis56315931
Daily sperm count/g testis83427034

Strengths/Weaknesses: Strengths of this study are the range and appropriateness of selected measures, the use of 2 strains of rat, the verification of dosing solutions, and the use of ethinyl estradiol, which produced expected responses. An unfortunate weakness is the small sample size of 6–7 dams/strain/group. Nevertheless, data were analyzed appropriately with the litter as the experimental unit, and significance judgments were apparently based on 7/group. Modest effects were noted in male and female offspring in the 50 mg/kg exposure group. While effects on the lowest doses in this study were not seen, it is important to recognize the effects seen at 50 mg/kg bw/day (the high-dose in this study) dosing on GD 6–21.

Utility (Adequacy) for CERHR Evaluation Process: This study is adequate and of high utility for the evaluation process.

Schönfelder et al. (2002a), supported by the German Federal Ministry for Education and Research, examined the effects of prenatal bisphenol A exposure on the rat vagina. Sprague–Dawley rats were gavaged on GD 6–21 with bisphenol A at 0 [2% corn starch vehicle (Mondamin)], 0.1, or 50 mg/kg bw/day. A positive control group was treated with 0.2 mg/kg bw/day 17α-ethinyl estradiol in a peanut oil vehicle. [No information was provided on the number of dams treated, the day of vaginal plug, purity of bisphenol A, or the type of chow, bedding, and caging materials used.] [According to the author the number of litters treated were: Mondamin=20, 0.1 mg/kg bw/day bisphenol A=20, 50 mg/kg bw/day bisphenol A=18, and 0.2 mg/kg bw/day 17α ethinyl estradiol=11; day of sperm positive smear was considered to be GD 0 and was used instead of day of vaginal plug; purity of bisphenol A was ≥98%; Altromin 1324 rodent chow was used (obtained from Altromin GmbH); bedding was wood shavings obtained from Altromin GmbH; caging was Type III macrolon cages (G. Schönfelder, personal communication, July 20, 2007).] At 3 months of age, estrous cyclicity was evaluated for 3 weeks in 42 female offspring of the control group, 21 offspring of the 0.1 mg/kg bw/day group, 18 offspring of the 50 mg/kg bw/day group, and 24 offspring of the 17β-estradiol group. [The number of litters represented was not stated.] At 4 months of age, female offspring were killed in either estrus or diestrus. [Authored clarified that each estrus group contained 22 offspring from 20 dams in the cornstarch group, 13 offspring from 13 dams in the 0.1 mg/kg/d and 12 offspring from 12 dams in the 50 mg/kg/d bisphenol A group, as well as 19 offspring from 11 dams in the 0.2 mg/kg/d 17α-ethinyl estradiol group (G. Schönfelder, personal communication, July 20, 2007)]. [Exact litter representation for animals collected during diestrus was not provided.] Vaginas were fixed in Bouin solution and a histopathological evaluation was conducted. Western blot analyses were conducted to measure expression of ERα and ERβ. [It does not appear that statistical evaluations were conducted.]

Qualitative descriptions of vaginal histopathology changes and ER expression were provided by the study authors. Low-dose animals killed during the estrous stage lacked keratinization of the surface epithelium and demonstrated reduced thickness of the total epithelium. Similar but less pronounced effects were observed in rats of the high-dose bisphenol A group. Vaginal findings were similar in the positive control group, and slight desquamation of the superficial layers was also observed. There were no differences in vaginal histopathology findings in rats killed during the diestrous stage. No ERβ was observed in vaginas of rats from any treatment group. Full-length ERα expression was not observed in either bisphenol A group during estrus, but ERα in the bisphenol A-exposed groups did not differ from the control group during the diestrous stage. ERα in vaginas obtained from the positive control group was either reduced or was not detected. The study authors concluded that altered vaginal morphology following bisphenol A treatment appears to be due to downregulation of ERα.

Strengths/Weaknesses: Vaginal histopathology of female offspring is of interest but the quality of the study cannot be judged due to unclear methodology. Uncertainty about the numbers of animals, the number of offspring examined and the lack of statistical accounting for litter effects are significant weaknesses.

Utility (Adequacy) of the CERHR Evaluation Process: This study is inadequate for the evaluation process for the reasons detailed above.

Schönfelder et al. (2004), supported by the German Federal Ministry for Environmental Protection and Radiation Security, examined the effects of prenatal bisphenol A exposure on the rat uterus. [No information was provided about composition of feed, caging, or bedding.] Sprague–Dawley rats [number treated not specified] were gavaged with bisphenol A [purity not reported] at 0 (2% corn starch vehicle), 0.1, or 50 mg/kg bw/day on GD 6–21. [Author clarified that the purity of bisphenol A was ≥98%; Altromin 1324 rodent chow was used (obtained from Altromin GmbH); bedding was wood shavings obtained from Altromin GmbH; caging was Type III macrolon cages (G. Schönfelder, personal communication, July 20, 2007).] The high bisphenol A dose was selected because it was reported to be the no observed effect level (NOEL) recommended by the Society of the Plastics Industry. A positive control group was gavaged with 0.2 mg/kg bw/day ethinyl estradiol on GD 6–21. Estrous cyclicity was examined for 3 weeks in 6 female offspring/group beginning at 3 months of age. Six female offspring/group were killed at 4 months of age on the day of estrus. Body and reproductive organ weights were measured. Uteri were fixed in methacarn solution and sectioned. Examinations of uterine morphology were conducted. Immunohistochemistry techniques were used to detect ERα and ERβ in the uterus, and results were verified by Western blot. Data were analyzed by Mann–Whitney test. [It was not clear if data were analyzed on a per litter or per offspring basis.] [Author states that each female came from a different litter so the data were analyzed on a per litter basis (G. Schönfelder, personal communication, July 20, 2007).] Statistically significant findings are summarized in Table 72. Effects observed at both dose levels were increased epithelial cell nuclei, epithelial nuclei with condensed chromatin, and epithelial cells with cavities and reduced ERβ-positive cells in uterine tissue. Additional effects observed only at the high-dose included decreased thickness of luminal epithelium and increased ERα-positive cells in the epithelium. Similar findings were observed following treatment with ethinyl estradiol. The study authors concluded that prenatal bisphenol A exposure causes uterine effects in rat offspring.

Table 72. Uterine Effects in Rats Exposed to Bisphenol A During Prenatal Development
 Dose, mg/kg bw/day
Endpoint0.150
  • a

    aIt is unclear if authors were referring to numbers of nuclei.

Thickness of luminal epithelium↓38%
Epithelial nucleia↑69%↑89%
Epithelial nuclei with condensed chromatin↑2.7-fold↑3.1-fold
Epithelial cells with cavities↑2.1-fold↑ 1.9-fold
ERα positive cells in epithelium↑67%
ERβ-positive cells in uterine tissue↓88%↓88%

Strengths/Weaknesses: A strength is the examination of effects on uterine indices in female offspring. A slight weakness is the use of only 6 females per group; however, the panel noted that the results appeared to be consistent across animals and across endpoints, especially in the 50 mg/kg bw/day treatment group.

Utility (Adequacy) for CERHR Evaluation Process: This study is adequate and of high utility for the evaluation process.

Wistuba et al. (2003), supported by the German Federal Ministry of Education and Science, examined the effects of prenatal exposure on testicular histology and sperm endpoints in rats. [No information was provided about chow, bedding, or caging.] Sprague–Dawley rats were gavaged with 0 (2% corn starch suspension vehicle), 0.1, or 50 mg/kg bw/day bisphenol A [purity not reported] on GD 6–21 (GD 0=day of sperm detection). A third group was treated with 0.02 mg/kg bw/day ethinyl estradiol. The high-dose was said to correspond to the current accepted no observed adverse effect level (NOAEL) and the lower dose was selected to determine if effects occurred at lower doses. It appears that the number of dams treated was 2 in the control group, 4 in the low-dose group, 1 in the high-dose group, and 4 in the ethinyl estradiol group. Litters were weighed during the lactation period. Pups were weaned on PND 22 [day of birth not defined]. Male offspring were killed between the ages of ∼9–12 months. The number of males killed was 5 from 2 litters in the control group, 15 from 4 litters in the low-dose group, 5 from 1 litter in the high-dose group, and 10 from 4litters in the ethinyl estradiol group. Testes were fixed in Bouin solution, and Sertoli cells were counted. Spermatogenesis was evaluated by examining germinal epithelia for cell death and distribution of various cell populations. Data were analyzed by ANOVA. [It appears that at least some data were analyzed on a per litter basis. In addition, analyses were done to determine intralitter variability and thus the numbers of animals per group that needed to be analyzed.]

Examination of tubule cross sections revealed qualitatively normal spermatogenesis in the bisphenol A groups. A comparison of Sertoli cell numbers in littermates revealed high variability (20–27%) in the 0.1 mg/kg bw/day group. A comparison of Sertoli cell numbers in the 4 litters from the 0.1 mg/kg bw/day group revealed almost identical results between litters. Sertoli cell numbers/organ were significantly increased by 19.4% in the low-dose group and 19% in the high-dose group. Bisphenol A had no significant effect on Sertoli cell numbers/g testis weight. The opposite situation occurred in the ethinyl estradiol group, with no significant effects on Sertoli cell numbers/organ but a significant increase in Sertoli cell numbers/g testis weight. Testis weight was not affected by bisphenol A treatment but was decreased in the ethinyl estradiol group. The study authors concluded that the study does not support the hypothesis of disruption of the male reproductive system by bisphenol A exposure.

Strengths/Weaknesses: The conceptual strength is the focus on the male reproductive tract/function. However, a weakness is that there were too few animals to provide reliable data.

Utility (adequacy) for the CERHR Evaluation Process: This study is inadequate based on insufficient sample size (n=2–4).

Thuillier et al. (2003), supported by National Institute of Environmental Health Sciences (NIEHS), examined a possible role for the platelet-derived growth factor system in estrogenic effects induced by bisphenol A in rats exposed during gestation. The effects of other compounds such as genistein and coumestrol were also examined but will not be discussed here. Pregnant Sprague–Dawley rats were gavaged with bisphenol A at 0 (corn oil vehicle) or 0.1, 1, 10, or 200 mg/kg bw/day from GD 14 through birth (PND 0). Additional rats were s.c. injected with diethylstilbestrol at 0.01–2 μg/kg bw/day during the same period. [No information was provided about number of rats treated, purity of bisphenol A, feed, or materials used in bedding and caging.] Male offspring were killed on GD 21 or PND 3 and testes were collected. Expression of mRNA or protein for platelet-derived growth factor receptor-α and platelet-derived growth factor receptor-β were determined in testes using RT-PCR, in situ hybridization, or immunohistochemistry. Statistical analyses included unpaired t-test with Welch correction. [It was not clear if the litter or offspring were considered the statistical unit.]

Expression of mRNA for platelet-derived growth factor receptor-α and -β was significantly increased at bisphenol A doses ≥1 mg/kg bw/day in testes from 3-day-old rats. All other experiments with bisphenol A were conducted with a single dose of 200 mg/kg bw/day. In situ hybridization examination of testes from 3-day-old rats from the bisphenol A group revealed an increase in expression of platelet-derived growth factor receptor-α mRNA in testicular interstitium and platelet-derived growth factor receptor-β mRNA in interstitium and seminiferous cords. Exposure to bisphenol A resulted in slightly increased platelet-derived growth factor receptor-α protein expression and strong expression of platelet-derived growth factor receptor-β in gonocytes from 3-day old rat testes. Immunolocalization studies in testes from 21-day-old fetuses revealed that exposure to 200 mg/kg bw/day bisphenol A did not affect expression of platelet-derived growth factor receptor-α protein in gonocytes, but platelet-derived growth factor receptor-β protein appeared to be increased in gonocytes and Sertoli cells. Diethylstilbestrol tended to have a biphasic effect with increased expression of platelet-derived growth factor receptor-α and -β mRNA in 3-day-old rat testis at low doses and decreased expression at the high-dose. Treatment with 1 μg/kg bw/day diethylstilbestrol decreased mRNA expression of platelet-derived growth factor receptor-α in interstitium and increased platelet-derived growth factor receptor-β mRNA expression in seminiferous cords. Immunoreactivity for platelet-derived growth factor receptor-α protein was maintained but there was a minimal level of platelet-derived growth factor receptor-β protein expression in 3-day-old rat testes following exposure to 1 μg/kg bw/day diethylstilbestrol. In testes obtained from 21-day-old fetuses, expression of platelet-derived growth factor receptor-α protein was decreased in Sertoli and interstitial cells and expression of platelet-derived growth factor receptor-β protein was apparently increased following exposure to diethylstilbestrol. The study authors concluded that the platelet-derived growth factor receptor pathway may be a target for estrogens in the testis, but the findings do not exclude the possibility that effects may have occurred through an ER-independent mechanism.

Strengths/Weaknesses: Endpoints are a strength, but inadequate methodological detail (i.e., sample size or adequate control for litter effects) precludes any informed judgment of study quality.

Utility (Adequacy) for CERHR Evaluation Process: This study is inadequate for the evaluation process based on insufficient methodological details.

Wang et al. (2004), supported by NIEHS, examined the effects of prenatal bisphenol A exposure on expression of ER-associated proteins in rat testis. The effects of genistein and coumestrol were also examined but will not be discussed here. Pregnant Sprague–Dawley rats [apparently 3/group] were gavaged with corn oil vehicle or bisphenol A at 0.1–200 mg/kg bw/day from GD 14 (14 days post-coitum) through birth. Additional rats were s.c. injected with 0.01–2 μg/kg bw/day diethylstilbestrol during the same time period. [No information was provided about feed, caging and bedding material, or compound purity.] Male offspring from three independent litters were killed on GD 21, PND 3, or PND 21. Western blot, RT-PCR, and immunohistochemistry techniques were used to measure expression of protein or mRNA for Hsp90, Hsp90α, p23, CYP40, Hsp70, and/or ERβ. Spermatogonia were quantified in PND 21 rat testis. Data were analyzed by unpaired t-test. The dam was considered the statistical unit.

In testes from 3-day-old rats, RT-PCR revealed significant increases in mRNA for hsp90 at bisphenol A dose levels of 10 and 200 mg/kg bw/day, and significant decreases in expression of CYP40 at 200 mg/kg bw/day and p23 at 1 mg/kg bw/day. In situ hybridization analyses in 3-day-old rat testes revealed that bisphenol A tended to increase expression of hsp90 throughout the testis, with patterns indicating increased expression in gonocytes and interstitial Leydig cells. Examination of protein in testes from 3-day old rats exposed to 200 mg/kg bw/day bisphenol A revealed significantly increased levels of hsp90 and hsp70, but no effect on levels of CYP40, p23, or ERβ. Immunohistochemistry revealed that hsp90 protein in testes from 3-day-old rats was most increased in gonocytes and less so in interstitium following exposure to 200 mg/kg bw/day bisphenol A. Use of a probe specific for hsp90α protein revealed that increased protein expression of hsp90 was due in a large part to the hsp90α isoform. Examination of testes from GD 21 fetuses and PND 21 pups revealed that the amount of hsp90 protein in the bisphenol A treatment group was similar to that observed on PND 3 but that the amount of protein did not differ from controls on PND 21. In 21-day-old rats from the bisphenol A group, the number of spermatogonia/tubule was significantly higher by ∼2-fold compared to the control group. [It is not clear which bisphenol A dose induced an increase in spermatogonia, but it was most likely 200 mg/kg bw/day, because that dose appeared to be used in all studies not examining dose–response relationships.] Effects following diethylstilbestrol exposure included increased expression of hsp90 mRNA at 1.0 μg/kg bw/day and decreased CYP40 mRNA expression at 0.01 and 1 μg/kg bw/day, but no effect on protein levels of those compounds was reported in testes from 3-day-old rats. The number of spermatogonia/tubule was also increased after prenatal exposure to diethylstilbestrol. The study authors concluded that prenatal exposure to bisphenol A affects hsp90 expression in gonocytes of rats, and because hsp90 interacts with several signaling molecules, changes in its expression could affect gonocyte development.

Strengths/Weaknesses: This study was generally well conceived, but the small sample size suggests it presents pilot data only. A full study is needed to provide reliable data.

Utility (Adequacy) for CERHR Evaluation Process: This study is inadequate based on insufficient sample size (n=3).

3.2.1.3 Neurodevelopmental endpoints

Funabashi et al. (2004a), supported in part by Yokohama City University, examined the effects of bisphenol A on the numbers of corticotropin-releasing hormone neurons in the preoptic area and bed nucleus of the stria terminalis of rats exposed during development. [No information was provided about chow or composition of bedding and caging.] Pregnant Wistar rats (n=8–11/treatment group) were given drinking water containing the 0.1% ethanol vehicle or 10 mg/L bisphenol A [purity not reported] until their offspring were weaned at 3 weeks of age. [It is implied but not stated that exposure occurred during the entire gestation period.] Bisphenol A intake was estimated by study authors at 2.5 mg/kg bw/day. Male and female offspring (n=8–11/group) were killed at 4–7 months of age, and immunocytochemistry techniques were used to determine the number of corticotropin-releasing hormone neurons in brain. Female rats were killed during proestrus. [Although the number of litters represented in each group was not specified, the number of rats examined suggests that 1 rat/sex/litter was examined.] Histological slides of brain were evaluated by an investigator blinded to treatment conditions. Two series of experiments were conducted, and data from both experiments were combined. Data were analyzed by ANOVA followed by Fisher protected least significant difference post-hoc test. [It was not stated if data were analyzed on a per litter or per offspring basis, but as stated earlier, it appears that 1 rat/sex/litter was examined.] In the control group, females had more corticotropin-releasing hormone neurons in the preoptic area and anterior and posterior bed nucleus of the stria terminalis than males. Bisphenol A treatment did not change the number of corticotropin-releasing hormone neurons in the preoptic areas of males. A loss in sex difference occurred in the anterior and posterior bed nuclei of the stria terminalis following bisphenol A treatment because differences in numbers of corticotropin-releasing hormone neurons between males and females were no longer evident. It appears that bisphenol A treatment increased the number of corticotropin-releasing hormone neurons in males and decreased the number in females. The study authors concluded that exposure to bisphenol A during gestation and lactation results in a loss of sex difference in corticotropin-releasing hormone neurons in the bed nucleus of the stria terminalis but not in the preoptic area.

Strengths/Weaknesses: This study was appropriately designed to examine effects on the development of brain areas known to be influenced by hormonal levels. Strengths include the relevance and subtleties of the endpoints measured; weaknesses include uncertainties about the numbers of animals examined and the duration of the dosing period. The results suggest a disruption of the normal pattern of sexually dimorphic neurons, a result of critical importance to concerns about disruptions relevant to reproductive function and sexually dimorphic behaviors. While the sample size was 8–11/group, the design and statistics appear to be appropriate. It is a weakness that the control for litter effects was not clear.

Utility (Adequacy) for CERHR Evaluation Process: This study is adequate for inclusion in the evaluation process, although of limited utility due to uncertainties about the sample size, duration of dosing, and control for litter effects.

Fujimoto et al. (2006), supported by the Japanese Ministry of Education, Culture, Sports, Science, and Technology, examined the effect of prenatal bisphenol A exposure on sexual differentiation of neurobehavioral development in rats. Wistar rats were fed CE-2 feed (CLEA, Japan). [Caging and bedding materials were not described.] From GD 13 (day of vaginal sperm not defined) to the day of birth (PND 0), 6 rats/group were given tap water containing bisphenol A [purity not reported] at 0 or 0.1 ppm. The study authors estimated the bisphenol A dose at 0.015 mg/kg bw/day. On PND 1, pups were weighed and litters were culled to 4 pups/sex. Pups were weaned on PND 21. Neurobehavioral evaluations conducted in 20–24 offspring/sex/group at 6–9 weeks of age included open-field, elevated plus maze, passive avoidance, and forced swimming tests. Statistical analyses included ANOVA, Fisher protected least significant difference test, and Mann–Whitney U-test. [It appears that offspring were considered the statistical unit.]

In the control group, rearing frequency and duration were significantly higher in females than males, but there were no sex-related differences in rearing frequency or duration in the bisphenol A group. Bisphenol A exposure caused an increase in rearing duration in males when compared to males from the control group. In the forced swim test, females in the control group struggled more than males but no sex-related differences in struggling were observed in the bisphenol A group. The duration of immobility in the swimming test was longer in males from the bisphenol A compared to males from the control group. Immobility was described as non-significantly increased in females exposed to bisphenol A compared to control females. Bisphenol A exposure had no effect on performance in passive avoidance and elevated plus maze test. The study authors concluded that exposure of male offspring to bisphenol A during the final week of gestation resulted in impaired sexual differentiation in rearing and struggling behaviors and facilitated depression-like behavior.

Strengths/Weaknesses: This study used a good choice of methods to examine functional disruptions in sexually dimorphic behaviors. Weaknesses include a lack of clarity about the nature of disruption of sexually dimorphic behavior patterns that was indicated in the authors' conclusions, the somewhat small sample size, the use of a single dose level, which was not confirmed, and the lack of clarity of the statistical methods regarding litter.

Utility (Adequacy) for CERHR Process: This study is inadequate for the evaluation process due to statistical methodology.

3.2.2 Rat—parenteral exposure only during pregnancy.

Ramos et al. (2001), supported by the Argentine National Council for Science and Technology, the Argentine National Agency for the Promotion of Science and Technology, and the Ministry of Health, examined the effects of bisphenol A exposure on the rat prostate. Wistar rats were housed in stainless steel cages. [No information was provided about chow or bedding material.] Four dams/group were exposed to bisphenol A [purity not reported] at 0 (DMSO vehicle), 0.025, or 0.250 mg/kg bw/day by s.c. pump on GD 8–23 (GD 1=day of vaginal sperm). Pups were weighed and sexed at birth. Litters were culled to 8 pups, with 4/sex when possible. Pups were weaned on PND 22 [day of birth not defined]. On PND 30, pups were injected with bromodeoxyuridine and killed 2 hr later. Ventral prostates were dissected and fixed in 10% neutral buffered formalin. Immunohistochemical techniques were used to measure proteins associated with cell proliferation and cell phenotypes. Morphometric measurements were taken. [It was not clear how many rats/treatment group were examined for each endpoint. Although a statement was made that males from a single dam were evaluated, it was later stated that siblings were excluded from the same experimental group. Therefore it appears that different litters were represented.] Data were analyzed by Kruskal–Wallis ANOVA and Mann–Whitney U-test. [It was not clear if the dam or offspring were considered the statistical unit.]

In the periductal stroma, the fibroblastic layer was increased, the smooth muscle layer was reduced, and androgen receptor-positive cells were decreased. Prostatic acid phosphatase-positive cells were reduced in epithelial cells. There were no effects on cell proliferation and ERα was not detected. No changes were observed in interductal stromal cells.

Strengths/Weaknesses: This study has an interesting design with respect to choice of endpoints. Certain design aspects are unclear and statistical approaches are inadequate. The sample size was small (4 dams/group) and there was considerable uncertainty about numbers of offspring examined and accounting for litter effects. The use of DMSO (% not specified) is of concern, as this can modify the effects of the solute. Of additional concern is the route of administration (s.c. pump).

Utility (Adequacy) for CERHR Evaluation Process: This study is considered inadequate.

Ramos et al. (2003), supported by the Argentine Ministry of Health, Argentine National Agency for the Promotion of Science and Technology, and the National University of Litoral, examined the effects of bisphenol A exposure on the prostate and the hypothalamic-pituitary-gonadal axis in Wistar rats. Rats were housed in stainless steel cages and 7–9/group were administered DMSO vehicle or bisphenol A at 0.025 or 0.250 mg/kg bw/day by s.c. pump on GD 8–23 (GD 1=day of vaginal sperm). [No information was provided on purity of bisphenol A, the type of feed used, or composition of bedding.] After birth, pups were weighed and sexed. Litters were culled to eight pups with equal numbers of male and female pups when possible. Pups were weaned on PND 22 [day of birth not defined]. During prepuberty (PND 15), peripuberty (PND 30), and adulthood (PND 120), 6–8 males/group were injected with bromodeoxyuridine and killed 2 hr later. Serum was collected for measurement of LH and prolactin by RIA. Immunohistochemistry techniques were used to evaluate markers of cell proliferation, estrogen/androgen receptors, and prostatic cells. Expression of mRNA for ERα and ERβ in the preoptic area and medial basal hypothalamus was determined by RT-PCR. Data were analyzed by Kruskal–Wallis 1-way ANOVA using Dunn post-test.

No significant effects were observed for ventral prostate weight. Numerous transient effects were observed in both bisphenol A dose groups. On PND 15, cellular proliferation was increased in the periductal stroma of the prostate, and serum testosterone levels were elevated. On PND 30, the fibroblasts (vimentin-positive cells) in the prostatic periductal stroma was increased and the area of smooth muscle cells α-smooth muscle actin) was decreased. Also observed on PND 30 was a reduction in androgen-receptor positive stromal cells, a decrease in epithelial cells positive for prostatic acid phosphatase, and an increase in serum prolactin levels. Expression of ERβ mRNA was increased in the preoptic areas on PND 30 and 120, and the study authors considered the effect to be permanent because it occurred on both days. The study authors concluded that prenatal exposure to environmental concentrations of bisphenol A during gestation results in transient and permanent changes in the male reproductive axis.

Strengths/Weaknesses: The design appears reasonable as a means to address the study questions. Like many of these studies, altered values are given without addressing the normal range of variation or the likely functional significance of the changes. Weaknesses include use of the s.c. pump as a route of administration and use of DMSO as a vehicle.

Utility (Adequacy) for CERHR Evaluation Process: This study is inadequate for inclusion due to the use of 99.9% DMSO as a vehicle to administer BPA via s.c. pump. As discussed in earlier, the use of >50% DMSO as a vehicle for ALZET mini-pump studies is a clear contravention of the directions for mini-pump use, as it accelerates the breakdown of the mini-pumps

Naciff et al. (2002), from the Procter and Gamble Company, examined the effects of prenatal bisphenol A exposure on gene expression and, to a limited extent, development in female rat reproductive organs. Pregnant Sprague–Dawley rats were fed Purina 5K96, a casein-based soy- and alfalfa-free diet. [Composition of caging and bedding materials was not reported.] The rats were assigned to groups (≥7 rats/group) s.c. injected with bisphenol A (∼99% purity) in DMSO vehicle at 0, 5, 50, or 400 mg/kg bw/day on GD 11–20 (day of sperm detection=GD 0). Dams were killed on GD 20, and ovaries and uteri were removed from fetuses. In 4 litters/group, 1 female fetus/litter was examined for ovarian and uterine histopathology. In 5 litters/group, ovaries and uteri from at least 5 littermates were pooled for a microarray analysis of gene expression. Changes in gene expression were further quantified using RT-PCR. Data were analyzed by t-test, ANOVA, and Jonkheere-Terpstra test. Comparisons of gene expression among estrogenic compounds were made by Wilcoxon-Mann–Whitney and Jonkheere-Terpstra tests. Results of gene expression assays are discussed in Section 2. Vaginal bleeding and early parturition occurred in 1 of 8 dams in the high-dose group. Bisphenol A treatment had no effect on maternal body weight or number of live fetuses/litter, and there were no gross or histopathological effects on ovary or uterus. Prominent nipples and areolas were observed in males and females in the high-dose bisphenol A group [number of fetuses and litters affected were not reported].

Strengths/Weaknesses: Strengths are that these endpoints appear appropriate; weaknesses are the limited nature of the endpoints and the use of neat DMSO as vehicle. The sample sizes are 4–5/endpoint/group and judged to be inadequate. Of additional concern is the route of administration.

Utility (Adequacy) for CERHR Evaluation Process: This study is inadequate for the evaluation process.

Naciff et al. (2005), from The Procter and Gamble Company, examined the effect of prenatal bisphenol A exposure on male rat reproductive organ histology and gene expression. Pregnant Sprague–Dawley rats were fed Purina 5K96, a casein-based soy- and alfalfa-free diet. Rats were housed in stainless steel cages before mating. Rats were randomly assigned to groups (≥8 rats/group) and s.c. injected with bisphenol A (∼99% purity) in DMSO at 0, 0.002, 0.02, 0.5, 50, or 400 mg/kg bw/day on GD 11–20 (day of sperm detection=GD 0). Dams were killed on GD 20, and testes and epididymides were removed from fetuses. In 4 litters/dose group, 1 male fetus/litter was examined for testicular histopathology. In 5 litters/group, testes and epididymides from 5 littermates were pooled for a microarray analysis of gene expression. Changes in gene expression were further quantified using RT-PCR. Data were analyzed by t-test, ANOVA, and Jonkheere-Terpstra test. Comparisons of gene expression among estrogenic compounds were analyzed by Wilcoxon-Mann–Whitney and Jonkheere-Terpstra tests.

Bisphenol A treatment had no effect on maternal body weight or number of live fetuses/litter, and there were no gross or histopathological effects on the testis or epididymis. Prominent nipples/areolas were observed in male and female fetuses from the high-dose group [numbers of fetuses and litters affected were not reported]. In pooled testis and epididymis samples from the high-dose bisphenol A group, expression of 15 genes was significantly altered in a dose-related manner. When bisphenol A data were pooled with data obtained from ethinyl estradiol and genistein and globally analyzed, there were 50 genes that were significantly altered in the same direction by all three compounds. The study authors concluded that transplacental exposure to high-doses of bisphenol A alters the expression of certain genes in the testis and epididymis of fetal rats without causing malformations in those organs. The study authors noted that the dose response to bisphenol A was monotonic with no evidence of robust quantifiable responses at low doses.

Strengths/Weaknesses: Strengths are that these endpoints appear appropriate; weaknesses are the limited nature of the endpoints and the use of neat DMSO as vehicle. The sample sizes are 4–5/endpoint/group and judged to be inadequate. Of additional concern is the route of administration.

Utility (Adequacy) for CERHR Evaluation Process: This study is inadequate for the evaluation process.

Saito et al. (2003b), support not indicated, examined the effect of prenatal bisphenol A exposure on testosterone production during adulthood in rats. On GD 12–19 (day of vaginal plug not reported), 2 Wistar rats were s.c. injected with the corn oil vehicle, 4 rats were s.c. injected with 0.005 mg/day bisphenol A [purity not indicated], and 2 rats were injected with 5 μg/day 17β-estradiol. [Assuming a pregnant Wistar rat weights0.33 kg, 0.005 mg/day would be equivalent to 0.015 mg/kg bw/day bisphenol A.] Other materials found in dental composites were also evaluated but will not be discussed. During the lactation period, rats were housed in polypropylene cages with synthetic bedding. [No information was provided on feed.] Offspring were housed separately at 3 weeks of age and killed at 13 weeks of age. Body and testis weights were measured in all male offspring (22 in the bisphenol A group, 11 in the vehicle control group, and 5 in the 17β-estradiol group). Plasma testosterone level was measured by RIA, and plasma cholesterol level was measured using a kit. Activities of testicular enzymes involved in the production of testosterone from progesterone were also examined in an in vitro assay in which testicular microsomes were incubated with 14C-progesterone and 14C-δ4-androstenedione for 20 min. Data were analyzed using unspecified post-hoc tests. [Although not clear, it appears that offspring were considered the statistical unit for some analyses.]

Bisphenol A exposure had no effect on pup sex ratio. No effects on body weight or absolute testicular weight were observed in the bisphenol A group at 13 weeks of age. However, relative (to body weight) testicular weight was lower [by 6%] in rats of the bisphenol A compared to the control group. Also observed in the bisphenol A group was a reduction in plasma testosterone level [by28%]. No effect was observed on cholesterol level. In the ex vivo study, prenatal bisphenol A exposure increased activities of 17α -hydroxysteroid dehydrogenase [by140%] and 17β-hydroxysteroid dehydrogenase [by70%]. Observations in the 17β-estradiol compared to the control group included decreased numbers of offspring delivered, higher body weight of male offspring at 13 weeks of age, reduced plasma testosterone level, and increased testicular 17α-hydroxysteroid dehydrogenase activity. The study authors concluded that bisphenol A had an estrogenic effect on the testis but did not decrease activities of enzymes involved in testosterone production.

Strengths/Weaknesses: A strength of this study is the examination of testosterone levels at 13 weeks of age. This strength is negated by the sample size (n=2–4), which is too small to draw any firm conclusions.

Utility (Adequacy) for CERHR Evaluation Process: This study is inadequate based on insufficient sample size.

Murray et al. (2007), supported by NIH, examined the effect of prenatal bisphenol A exposure on in situ induction of mammary tumors. Wistar-Furth rats were fed Harlan Teklad 2018, which was reported to contain 20 fmol/g estrogen equivalents. Water was supplied in glass bottles. Caging and bedding materials were not reported, but they were stated that to test negative in the E-SCREEN. From GD 9 (GD 1=day of vaginal sperm) through PND 1 [The day of birth was PND 0 (A. Soto, personal communication, March 2, 2007)], rats received the 50% DMSO vehicle or bisphenol A [purity not reported] at 0.0025, 0.025, 0.250, or 1 mg/kg bw/day. Dosing solutions were delivered by implanted [assumed s.c.] osmotic pumps. [Number of dams treated was not reported. Based on a limited amount of information provided on the number of offspring examined, it appears that ≤6 dams/group were treated.] Pup viability was assessed on PND 1. On PND 2 pups were sexed and litters were culled to 8 pups. Anogenital distance was measured on PND 4. Litters were weighed during the lactation period. Female offspring were monitored for body weight and vaginal opening in the post weaning period. Female offspring were killed on PND 50 or 95. Mammary glands were collected and whole-mounted or sectioned for histopathological examination. Morphometric analyses were conducted to examine possible presence of preneoplastic lesions. Mammary glands were examined for ERα and Ki-67 protein by an immunohistochemistry technique. Maximal numbers of “maternal units” were represented in each dose group. One female/litter was included in histological examinations. [Apparently ≤6 offspring/group were examined in histopathological examinations. Number of offspring examined for other endpoints was not reported in the manuscript. According to an author, n=7–21 for the other endpoints (A. Soto, personal communication, March 2, 2007).] Statistical analyses included ANOVA followed by post-hoc tests (Bonferroni or t-test) when significant effects were observed by ANOVA. [It was not clear if dams or offspring were considered the statistical unit.]

Bisphenol A exposure did not affect offspring viability, sex ratio, age at vaginal opening, or female anogenital distance. Anogenital distance was reduced on PND 4 in males from the 0.250 mg/kg bw/day group. Percent hyperplastic ducts was increased in all dose groups on PND 50 and in the 0.0025 mg/kg bw/day group on PND 95; the study authors noted that the effect on PND 50 was quantitatively similar in all dose groups (i.e. 3–4-fold increase). Cribriform structures were observed in the 0.25 and 1 mg/kg bw/day groups. [Incidence was not reported for the control and lower dose groups.] The structures were classified as carcinomas-in-situ and were characterized by increased ductal size resulting from luminal epithelium proliferation, enlarged luminal epithelial cells, presence of a nucleolus, variable chromatin pattern, and rounded luminal spaces consisting of trabecular rods of cells perpendicularly aligned to the longer duct axis. Numbers of Ki-67- and ER-α positive cells were increased in aberrant compared to normal tissues, regardless of dose. [Results in treated compared to control groups were not reported.] The study authors concluded that fetal bisphenol A exposure is rats is sufficient to induce development of preneoplastic and neoplastic mammary lesions.

Strengths/Weaknesses: Relevance of endpoints is a strength, as is the use of multiple dose levels. Weaknesses include an unstated number of dams (and by inference, a small number of these, and thus, because of dam-related effects, a small overall n), the uncertainty of the response rate of histopathology in the controls, and the use of 50% DMSO as vehicle.

Utility/Adequacy for CERHR Evaluation: This study was inadequate due to small sample size, route of administration, and lack of clarity on statistical analysis.

Durando et al. (2007), supported by Universidad National del Litoral, Argentine National Agency for the Promotion of Science and technology, and NIH, examined the effects of prenatal bisphenol A exposure on susceptibility to mammary tumors in rats. Wistar rats were fed Cooperación (Buenos Aires, Argentina) and housed in stainless steel cages. [It was not clear if bedding was used.] On GD 8–23 (GD 1=day of vaginal sperm), 11–14 dams/group were s.c. dosed by osmotic pump with the DMSO vehicle or 0.025 mg/kg bw/day bisphenol A [purity not indicated]. Pups were delivered on GD 23 and weaned on PND 21. It was not indicated if day of birth was considered PND 0 or 1. During the study, body weights and day of vaginal opening were monitored. Offspring were killed before puberty (PND 30), after puberty (PND 50), or in adulthood (PND 110 and 180). In mammary gland stroma and epithelium, proliferation was assessed by BrdU incorporation and apoptotic cells were identified by TUNEL method. Morphometric analyses were conducted in sectioned mammary glands. Mast cells were identified by immunostaining for proteinase. At least 6 offspring/group/time point were evaluated. [No littermates were used in the evaluation at any given time point (A. Soto, personal communication, March 2, 2007).] Additional offspring were examined for responsiveness to chemically-induced mammary preneoplastic or neoplastic lesions. On PND 50, N-nitroso-N-methylurea was administered to 10–16 offspring from the vehicle control group at 25 or 50 mg/kg bw and 21 offspring from the bisphenol A group at 25 mg/kg bw. Based on findings from a pilot study, 25 mg/kg bw was considered a subcarcinogenic dose and 50 mg/kg bw was considered a positive control. During the study, rats were palpated for tumors. Rats that received 50 mg/kg bw N-nitroso-N-methylurea were killed on PND 180 and rats that received 25 mg/kg bw N-nitroso-N-methylurea were killed on PND 110 or 180. Whole-mounted mammary glands were examined for tumors. Immunostaining was conducted to identify cytokeratin 8 (an epithelial marker) and p63 (a myoepithelial marker). Data were statistically analyzed using the Mann–Whitney U-test.

Bisphenol A exposure did not affect successful pregnancies, dam weight gain, pregnancy duration, number of pups/litter, or percent females/litter. Anogenital distance on PND 1 or 5 and postnatal body weights were unaffected in pups exposed to bisphenol A. Vaginal opening was accelerated in pups from the bisphenol A group (mean 34 days of age compared to 39 days of age in controls). On PND 50, the BrdU/apoptosis ratio was significantly increased and apoptosis was significantly decreased in mammary parenchyma and stroma of bisphenol A-exposed animals; the effects were not observed on PND 30 or 110. Significantly increased percentages of hyperplastic ducts, density of stromal nuclei, and numbers of mast cells were observed in the bisphenol A group on PND 110 and 180. Exposure to bisphenol A resulted in formation of a dense stroma layer around mammary epithelial structures and replacement of normal adipose tissue with a fibroblastic stroma. In rats exposed to 25 mg/kg bw N-nitroso-N-methylurea on PND 50, incidence of hyperplastic lesions on PND 180 was significantly higher in the group with prenatal bisphenol A compared to DMSO exposure (mean incidence of 35.5% compared to 15.7% in controls). Although statistical significance was not achieved, exposure to 25 mg/kg bw N-nitroso-N-methylurea resulted in tumors in 2 of 15 rats in the prenatal bisphenol A group and 0 of 10 rats in the prenatal vehicle control group on PND 180. Cytokeratin 8 immunostaining revealed no invasion by stromal epithelial cells. The study authors concluded that rats prenatally exposed to environmentally relevant doses of bisphenol A may have an increased risk of developing mammary tumors.

Strengths/Weaknesses: Weaknesses include route of administration and the high single dose is a weakness as is the use of pure DMSO.

Utility (Adequacy) for CERHR Evaluation Process: This study is inadequate for inclusion due to the use of 99.9% DMSO as a vehicle to administer bisphenol A via s.c. osmotic pump.

Hong et al. (2005), sponsored by the Korea Research Foundation, investigated the effects of acute exposures to bisphenol A during late pregnancy on expression and protein level of calbindin-D9k, a putative biomarker of estrogen activity, in the uteri of offspring and lactating rats on PND 5. Pregnant Sprague–Dawley rats were given free access to water and a diet of soy-free pellets in polycarbonate caging. [Housing conditions (individual or group) and bedding material were not indicated.] On GD 17–19, pregnant rats were s.c. injected daily with 200, 400, or 600 mg/kg bw/day bisphenol A [purity not provided] in corn oil (n=5/group). Negative and positive control groups (n=10/group) were administered corn oil or 17β-estradiol 40 μg/kg bw/day. On PND 5, lactating dams and female pups were killed and their uteri harvested. Dose response changes in calbindin-D9k, expression levels in uteri of lactating dams and female offspring (3/group) were analyzed by Northern blot and RT-PCR, with appropriate housekeeping gene controls. Protein levels and localization of calbindin-D9k were performed by Western blot and immunohistochemistry for lactating dams only. Statistical analyses were performed using the Kruskall–Wallis and Dunnett tests. [It was not clear if dams or offspring were considered the statistical unit.]

Northern blot analysis revealed a significant increase [∼6.4-fold] in the level of calbindin-D9k expression in the uteri of lactating dams exposed to 600 mg/kg bw/day bisphenol A compared to oil controls. 17β-Estradiol treatment produced a significant [∼3.9-fold] increase in calbindin-D9k mRNA expression in the dam uterus that was not statistically distinct from the effect of the high bisphenol A dose. Uteri of offspring exposed to the highest dose level of bisphenol A also showed a significant upregulation [∼4.4-fold] in calbindin-D9k expression. Expression levels of ERα were unaffected in maternal uteri exposed to bisphenol A. However, ERα expression was increased significantly in uteri of pups exposed to 400 and 600 mg/kg bw bisphenol A [↑33% and 66%, estimated from a graph]. Protein levels of calbindin-D9k in lactating dam uteri were elevated significantly at all dose points [50, 40, and 50%, for 200, 400, and 600 mg/kg bw/day, respectively]. 17β-Estradiol-treatment was not associated with a significant increase in calbindin-D9k protein. The density of calbindin-D9k-immunopositive cells was increased in uterine sections from lactating dams exposed to all doses of bisphenol A relative to oil controls, correlating with Western blot results. Authors note insufficient material or low detectability of calbindin-D9k protein in offspring tissue, and protein analyses were not performed.

The authors suggest that calbindin-D9k can serve as a reliable biomarker of acute estrogenic exposure, particularly for insight into maternal-fetal metabolic exchange, given that calbindin-D9k is tightly regulated and rapidly induced by 17β-estradiol, diethylstilbestrol, alkylphenols, and now, bisphenol A. They further point out that calbindin-D9k expression is absent in immature rat and ovariectomized rat uteri.

Strengths/Weaknesses: This study supports the use of calbindin-D9k as a uterine biomarker of estrogenic effect in the perinatal period in the rat, and provides some dose–response information for bisphenol A induction of an estrogenic response. Limitations are the subcutaneous route of exposure, small sample size, high-doses and uncertain statistical analyses of the F1 data.

Utility (Adequacy) for CERHR Evaluation Process: While providing some dose–response information regarding bisphenol A-induced estrogenic effects following exposure of rats in the perinatal period, the lack of clarity regarding whether the dam or offspring was considered the statistical unit, route of exposure, and use of high doses render this study inadequate for consideration in the evaluation process.

3.2.3 Rat—oral exposure postnatally with or without prenatal exposure.

3.2.3.1 Reproductive studies:

The International Research and Development Corporation (General Electric,1976), sponsored by General Electric, examined the effects of bisphenol A exposure on CD rats and their offspring. Male and female F0 rats were housed in wire mesh cages and fed Purina Laboratory Chow. Ten rats/sex/group (body weights of 110–170 g for males and 100–151 g for females) were given feed containing bisphenol A [purity not specified] at 0, 1000, 3000, or 9000 ppm for 17 weeks. [It was not clear how long before mating that the dosing was started or if dosing was continued through the gestation and lactation periods.] The European Union (2003) estimated bisphenol A intake at 0, 70, 200, or 650 mg/kg bw/day in males and 0, 100, 300, or 950 mg/kg bw/day in females. F0 rats were mated at ∼100 days of age and assessed for fertility. F1 pups were counted and weighed at birth and on PND 21 (day of birth not defined). Fifteen male and female F1 rats/group/sex that were exposed in utero were selected for a 13-week feeding study and were fed diets containing the same concentration of bisphenol A as their parents. F1 rats were weighed and observed for clinical signs. Hematological, clinical chemistry, and urinalysis parameters were examined in 5 rats/sex/group in the control and 2 highest dose groups at 1, 2, and 3 months of F1 exposure. Ophthalmoscopic examinations were conducted at 3 months of F1 exposure. After 13 weeks of dosing, the F1 rats were killed and necropsied. Organs were weighed and fixed in 10% neutral buffered formalin. Included among organs weighed were testis and ovary. Histopathological examinations were conducted in tissues from 10 rats/sex/group in the control and high-dose group. Included among organs histologically examined were prostate, uterus, testis, and ovary. Statistical analyses included χ2 test with Yates correction, Fisher exact probability test, Mann–Whitney U-test, ANOVA, t-test, and Dunnett multiple comparison test.

Fertility was unaffected in F0 rats. Body weight gain was lower in F0 rats from the 3000 and 9000 ppm groups. Body weight at Week 17 followed the same patterns as body weight gain [6–7% decrease in the 3000 ppm group and 12–18% decrease in the 9000 ppm group compared to controls]. There were no differences in food intake. [Statistical significance for body weight effects was not reported. It was not clear if statistical analyses were not conducted or if the effects did not attain statistical significance.]

There were no effects on number of F1 pups/litter or survival of pups. Pup birth weights in the 9000 ppm group were slighter decreased but were said to be within normal range. Body weight gains on PND 21 were slightly decreased in pups from the 3000 and 9000 ppm dose groups. Body weights on PND 21 were significantly lower in pups from the 3000 and 9000 ppm groups [7 and 12% lower compared to controls; benchmark dose analysis not conducted because variances not reported]. One male F1 rat in the control group and 2 female F1 rats in each of the 3000 and 9000 ppm group died during the study. Post-weaning body weight gain was lower in females from all dose group and in males from the 3000 and 9000 ppm dose groups. Body weight at week 13 followed the same patterns as body weight gain [13% decrease in the 1000 ppm group, 11–17% in the 3000 ppm group, and 22–24% decrease in the 9000 ppm group compared to controls]. Food intake was decreased in females from all dose groups and in males from the 9000 ppm group. Examination by ophthalmoscopy revealed no treatment-related effects. No treatment-related effects were observed for hematology, biochemistry, or urinalysis. No changes in organ weights or gross or histopathological lesions were considered treatment-related. The study authors noted increases in mean weights of spleen, brain, thyroid, and adrenals in the treated groups but concluded that the effects resulted from decreased body weight. [With the exception of PND 21 pup weights, there was no discussion of statistical significance for effects observed in F1 rats. It was not clear if statistical analyses were not conducted or if statistical significance was not attained.]

Strengths/Weaknesses: This study is a conventional, state-of-the-art-at-the-time two-generation toxicity study. The inclusion of a breeding period and a second generation are strengths. Weaknesses are magnified in hindsight: these include the limited number of animals examined, the lack of close examination of the reproductive processes in the F1 animals, and uncertainty about the statistical significances. The study has not been peer-reviewed.

Utility (Adequacy) for CERHR Evaluation Process: Although this study was not designed to find non-linear dose–responses, it represents a conventional-for-the-time 2-generation toxicity study, and is adequate for the evaluation process but of limited utility because the high doses preclude evaluation of low dose effects and limit its utility in showing a lack of marked organ toxicity or gross reproductive toxicity in a limited number of animals at very high-doses.

The International Research and Development Corporation (General Electric,1978), sponsored by General Electric, examined the effects of bisphenol A exposure on male and female CD rats and their offspring. In the first part of the experiment, male and female rats were housed in wire mesh cages and were fed Purina Laboratory Chow containing bisphenol A [purity not specified] for 18 weeks. Ten rats/group (body weights of 135–179 g for males and 114–158 g for females) were assigned to each treatment group based on even distribution of body weight and litter mates. [Based on information provided in study tables, it appears that the rats were ∼30 days old at the start of dosing.] Bisphenol A was added to feed at concentrations of 0, 100, 250, 500, 750, or 1000 ppm. The European Union (2003) estimated bisphenol A intake at 0, 5, 15, 30, 50, and 60 mg/kg bw/day in males and 0, 10, 25, 50, 75, and 100 mg/kg bw/day in females. Rats were examined for clinical signs, body weight gain, and food intake throughout the study. Estrous cyclicity was examined in females for 3 weeks before breeding and during breeding. At 100 days of age (Week 10 of the study), rats were moved to plastic cages with corncob bedding and mated for 3 weeks. GD 0 was defined as the day that vaginal sperm or plug was observed. Rats were assessed for fertility and gestation length. Day of delivery was designated lactation day 0 (PND 0). Pups were counted, sexed, and weighed, assessed for viability at birth and through the lactation period. After weaning, 15 male and female F1 rats/group that were exposed in utero were selected for a 90-day feeding study. Parental rats and unselected F1 rats were killed and discarded.

During a 90-day period, F1 rats were fed diets containing the same concentration of bisphenol A as their parents. [Ages at the start of dosing were not reported, but based on body weight ranges reported (64–138 g for males and 57–118 g for females) it appears that rats were different ages at the start of dosing.] F1 rats were weighed and observed for clinical signs. Hematological, clinical chemistry, and urinalysis parameters were examined at Day 30, 60, and 90 of the study. Ophthalmoscopic examinations were conducted before initiation of and following 90 days of dosing. The rats were killed and organs weighed. Adrenals, pituitary, ovaries, and thyroid were weighed following fixation in 10% neutral buffered formalin. Histopathological examinations were conducted in tissues from 10 rats/sex/group in the control and high-dose groups. Organs histologically examined included prostate, uterus, testis, and ovary. Statistical analyses included χ2 test with Yates correction, Fisher exact probability test, Mann–Whitney U-test, ANOVA, t-test, and Dunnett multiple comparison test.

In parental rats, bisphenol A exposure did not affect general behavior, appearance, or survival. Mean body weight of males in the 1000 ppm group was 6% lower than control males. Food intake was increased [by ∼7–11%, no dose–response] in females of all dose groups. Bisphenol A exposure had no effect on estrous cyclicity or gestation length [data were not shown], male or female fertility, number of pups/litter, or pup survival Body weights of pups in the 750 ppm group were significantly higher [by ∼10%] compared to controls on PND 21, but the study authors did not consider the effect to be treatment-related.

In the F1 offspring, a slight decrease in body weight gain was observed for males in the 750 ppm group. [At the end of the study, body weights of males in the 750ppm group were ∼7% less than controls]. Food intake was similar in treated and control groups. Ophthalmoscope examinations did not reveal any treatment-related effects. Although mean blood urea nitrogen levels were slightly lower and mean serum glutamic-oxaloacetic transaminase values were sporadically increased in treated rats, the study authors noted that the values were within physiological ranges. There were no effects on hematological or urinalysis parameters. Some significant organ weight changes were noted by the study authors, but they stated that the biological significance of the effects was not known. [There did not appear to be dose–response relationships for any organ weight change.] The study authors stated that no compound-related lesions were observed in organs, including reproductive organs.

Strengths/Weaknesses: The use of multiple dose levels (going down to fairly low exposure levels) is a plus, as is a breeding phase. Weaknesses include the limited number of animals per group, discarding of the parental animals without examination, the fact that not all F1 animals were examined at least for structural effects, the lack of close examination of F1 animals for reproductive effects (cyclicity and sperm measures), and the use of the conventional “top-down” pathology evaluation, wherein the lower dose groups were examined only if effects were noted in the high-dose. The study has not been peer-reviewed.

Utility (Adequacy) for CERHR Evaluation Process: For what it is, this study is adequate and of limited utility for the evaluative process, showing no gross changes in the structure of a limited number of tissues in a limited number of F1 animals, exposed from pre-conception. This study was not designed to find unusual effects or non-linear dose–response relationships or to address the issue of low-dose functional responses or non-linear responses.

Ema et al. (2001), supported by the Japanese Ministry of Health and Welfare, examined developmental toxicity endpoints, in a 2 generation rats study described in detail in Section 4.2.3.1. Two generations of rats were gavaged with 0, 0.0002, 0.002, 0.020, or 0.200 mg/kg bw/day bisphenol A (99.9% purity) before and during mating and throughout the gestation and lactation period. These doses were based on previous studies that found effects at 0.002 and 0.020 mg/kg bw/day. There were some non-dose-related and sporadic effects, but the study authors concluded that none of the effects were related to bisphenol A treatment. Bisphenol A exposure did not adversely affect prenatal or postnatal growth or survival, developmental landmarks, anogenital distance, or age of puberty. In adult animals exposed to bisphenol A during development, there was no evidence of adverse effects on reproductive endpoints such as fertility, estrous cyclicity, or sperm counts. Prostate and other male reproductive organ weights were unaffected.

Strengths/Weaknesses: Strengths of this study were the thoroughness of the evaluation, the size of the dose range, the large number of animals, the litter-based analysis, and the verification of the dosing solution. A minor weakness is the lack of a positive control group, which leaves a question about the ability of this group of rats to respond.

Utility (Adequacy) for CERHR Evaluation Process: This study is adequate and of high utility for the evaluation process.

Tyl et al. (2002b), supported by The Society of the Plastics Industry, Inc., reported some developmental toxicity effects in a multigeneration bisphenol A study in Sprague–Dawley rats that is reported in detail in Section 4.2.3.1. In that study, F1, F2, and F3 rats were exposed to bisphenol A [99.70%-99.76% pure] indirectly during gestation and lactation and directly through feed after weaning. Dietary doses were 0, 0.015, 0.3, 4.5, 75, 750, or 7500 ppm, and target intakes were ∼0.001, 0.02, 0.30, 5, 50, and 500 mg/kg bw/day. At the 7500 ppm dose there were fewer pups and live pups/litter and body weight gain of pups was lower during the lactation period. Delayed puberty in both males and females of the 7500 ppm group was most likely related to reduced body weights according to the study authors. Bisphenol A exposure during development did not increase the weight of the prostate in adult rats. Although some decreases in epididymal sperm concentration and daily sperm endpoints were each observed in 1 generation of males from the high-dose group, the study authors concluded there were no treatment-related effects on sperm endpoints or reproductive function. The study authors identified an offspring and reproductive NOAEL of 750 ppm (∼50 mg/kg bw/day). A systemic NOAEL for adult rats was identified at 75 ppm (∼5 mg/kg bw/day) by the study authors; therefore, bisphenol A was not considered a selective developmental toxicant.

Strengths/Weaknesses: This study has numerous strengths, including the quality and number of the endpoints evaluated, the number of dose groups and generations examined, and the confirmation of dosing solutions. This study incorporated screening-level endpoints within the context of a multigeneration study. As such, it addresses gross issues but does provide helpful data regarding the NOAEL.

Utility (Adequacy) for CERHR Evaluation Process: This study is adequate and of high utility for the evaluation process.

3.2.3.2 Development of the reproductive or endocrine systems:

Cagen et al. (1999b), support not indicated (but all authors affiliated with industry), conducted a study to examine the effects of prenatal and lactational bisphenol A exposure on reproductive development of rats. The study attempted to replicate findings by Sharpe et al. that appeared in an unpublished meeting abstract. The protocol used by Cagen et al. (1999b) was the same as that used by Sharpe et al. with the exception that more dose levels were included, group sizes were larger, and a greater number of reproductive endpoints were examined. Animals were fed Certified Rodent Chow 5002. Music was played at a low volume to provide background noise. Female Han-Wistar rats were randomly assigned to groups. For 2 weeks before mating, during a 2-week mating period, and during the gestation and lactation periods, 28 rats/group were given drinking water containing bisphenol A (>99% purity) at 0.01, 0.1, 1.0, or 10