Screening Assessment for The Challenge (2024)

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Phenol, 4,4' -(1-methylethylidene)bis-
(Bisphenol A)

Chemical Abstracts Service Registry Number
80-05-7

Environment Canada
Health Canada

October 2008

(PDF Version - 910 KB)

• Synopsis
• Introduction
• Substance Identity
• Physical and Chemical Properties
• Sources
• Uses
• Releases to the Environment
• Environmental Fate
• Persistence and Bioaccumulation Potential
• Potential to Cause Ecological Harm
• Potential to Cause Harm to Human Health
• Conclusion
• References
• Appendix A: Robust Study Summary- Inherent Toxicity
• Appendix B: Robust Study Summary- Inherent Toxicity
• Appendix C: Robust Study Summary - Inherent Toxicity
• Appendix D: Summary of Studies Investigating Neurobehavioural Effects in Rodents

Synopsis

Pursuant to section 74 of the Canadian Environmental Protection Act, 1999 (CEPA 1999), the Ministers of the Environment and of Health have conducted a screening assessment of phenol, 4,4' -(1-methylethylidene)bis- or bisphenol A, Chemical Abstracts Service Registry Number 80-05-7. This substance was identified as a result of the categorization of the Domestic Substances List as a high priority for action under the Ministerial Challenge, since it was considered to pose the greatest potential for exposure to individuals in Canada and had been classified by the European Commission on the basis of reproductive toxicity. Bisphenol A also met the ecological categorization criterion for inherent toxicity to aquatic organisms, but it did not meet the ecological categorization criteria for persistence or bioaccumulation potential.

Bisphenol A is a high volume chemical, with global production at 4 billion kg in 2006. In the United States, production quantities increased from 521 million kg in 1990 to 736 million kg in 1995. Estimated production in the U.S. in 2007 was 1 billion kg. Canadian market values may be lower than those for the U.S; however, approximately 12 million kg of bisphenol A were reported as manufactured, imported or in commerce in Canada during the calendar year 1986. In 2006, no bisphenol A was manufactured in Canada at quantities equal to or greater than a reporting threshold of 100 kg. However, bisphenol A was used in Canada in the range of 100 000 to 1 000 000 kg and approximately half a million kg was imported into Canada either alone, in a product, in a mixture or in a manufactured item.

The available data indicate that bisphenol A does not persist significantly under aerobic conditions. However, the substance has been found not to degrade or to degrade only slowly under conditions of low or no oxygen. Bisphenol A has been detected in Canadian and U.S. surface waters, sediment, groundwater and soil, as well as in municipal and industrial waste treatment products. Studies from North America, Europe and Japan document detectable levels in several species of aquatic biota. The data demonstrate that bisphenol A is present in a wide range of environmental media.

Most data indicate relatively low bioaccumulation potential and a capacity for metabolism in various species. Most measured bioaccumulation and bioconcentration factors range only up to about 150 L/kg, with one study reporting a bioaccumulation factor of 650 L/kg in lower trophic levels. These studies confirm that bisphenol A is bioavailable and can accumulate in tissues to some degree. Bisphenol A is acutely toxic to aquatic organisms and has been shown to adversely affect growth and development in both aquatic and terrestrial species. There is evidence that low-level exposure to bisphenol A, particularly at sensitive life cycle stages, may lead to permanent alterations in hormonal, developmental or reproductive capacity. In laboratory testing, these effects have occurred within the range of concentrations measured in Canada, indicating that there is potential for adverse effects in populations, particularly close to point sources.

On the basis of expected continued or increasing exposure of biota, and information indicating the potential for long-term adverse effects to organisms within the range of concentrations currently measured in the environment, it is considered appropriate to apply a precautionary approach when characterizing risk. As such, it is concluded that bisphenol A is entering the environment in a quantity or concentration or under conditions that have or may have an immediate or long-term harmful effect on the environment or its biological diversity.

Human exposure to bisphenol A in Canada can result from dietary intake (e.g., migration from food packaging, migration from repeat-use polycarbonate containers), from environmental media (i.e., ambient air, indoor air, drinking water, soil and dust), from use of consumer products and other sources. Dietary intake is the primary source of exposure. Exposure estimates for the general population of Canada range from 0.08 µg/kg body weight/day to 4.30 µg/kg-bw per day. Specific exposure estimates for the most highly exposed subpopulation (i.e., infants) range from an average of 0.50 µg/kg body weight/day (maximum 4.30 µg/kg-bw per day) for infants aged 0 to 1 month to an average of 0.27 µg/kg body weight/day (maximum 1.75 µg/kg body weight/day) for infants aged 12 to 18 months. A critical effect for characterization of risk to human health is reproductive and developmental toxicity. The neurodevelopmental and behavioural dataset for rodents, though highly uncertain, is suggestive of potential effects at doses the same order of magnitude to 1-2 orders of magnitude higher than exposures. Given that toxico*kinetics and metabolism data from experimental animal and limited human studies indicate potential sensitivity to the maternal-fetal unit and infant, and that animal studies suggest a trend towards heightened susceptibility during stages of development in rodents, it is considered appropriate to apply a precautionary approach when characterizing risk. As such, it is concluded that bisphenol A be considered as a substance that may be entering the environment in a quantity or concentration or under conditions that constitute or may constitute a danger in Canada to human life or health.

In addition and where relevant, research and monitoring will support verification of assumptions used during the screening assessment and, where appropriate, the performance of potential control measures identified during the risk management phase.

Based on the information available, bisphenol A meets one or more of the criteria set out in section 64 of the Canadian Environmental Protection Act, 1999.

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Introduction

The Canadian Environmental Protection Act, 1999 (CEPA 1999) (Canada 1999) requires the Minister of the Environment and the Minister of Health to conduct screening assessments of substances that have met the categorization criteria set out in the Act to determine whether these substances present or may present a risk to the environment or human health. Based on the results of a screening assessment, the Ministers can propose to take no further action with respect to the substance, to add the substance to the Priority Substances List (PSL) for further assessment, or to recommend that the substance be added to the List of Toxic Substances in Schedule 1 of the Act and, where applicable, the implementation of virtual elimination.

Based on the information obtained through the categorization process, the Ministers identified a number of substances as high priorities for action. These include substances that

• met all of the ecological categorization criteria, including persistence (P), bioaccumulation potential (B) and inherent toxicity to aquatic organisms (iT), and were believed to be in commerce; and/or
• met the categorization criteria for greatest potential for exposure (GPE) or presented an intermediate potential for exposure (IPE), and had been identified as posing a high hazard to human health based on classifications by other national or international agencies for carcinogenicity, genotoxicity, developmental toxicity or reproductive toxicity.

The Ministers therefore published a notice of intent in theCanada Gazette, Part I, on December 9, 2006 (Canada 2006), that challenged industry and other interested stakeholders to submit, within specified timelines, specific information that may be used to inform risk assessment, and to develop and benchmark best practices for the risk management and product stewardship of those substances identified as high priorities.

The substance, phenol, 4,4' -(1-methylethylidene)bis- or bisphenol A was identified as a high priority for assessment of human health risk because it was considered to present greatest potential for human exposure (GPE) and had been classified by other agencies on the basis of reproductive toxicity. The Challenge for bisphenol A was published in the Canada Gazette on May 12, 2007. A substance profile was released at the same time (Canada 2007). The substance profile presented the technical information available prior to December 2005 that formed the basis for categorization of this substance. As a result of the Challenge, submissions of information were received.

Although bisphenol A was determined to be a high priority for assessment with respect to human health, and it also met the ecological categorization criterion for inherent toxicity to aquatic organisms, it did not meet the categorization criteria for potential for persistence or bioaccumulation. Therefore this assessment focuses principally on information relevant to the evaluation of risk to human health.

Under CEPA 1999, screening assessments focus on information critical to determining whether a substance meets the criteria for defining a chemical as toxic as set out in section 64 of the Act, where

"64. [...] a substance is toxic if it is entering or may enter the environment in a quantity or concentration or under conditions that

(a) have or may have an immediate or long-term harmful effect on the environment or its biological diversity;
(b) constitute or may constitute a danger to the environment on which life depends; or
(c) constitute or may constitute a danger in Canada to human life or health."

Screening assessments examine scientific information and develop conclusions by incorporating a weight of evidence approach and precaution.

This screening assessment includes consideration of information on chemical properties, hazards, uses and exposure, including the additional information submitted under the Challenge. Data relevant to the screening assessment of this substance were identified in original literature, review and assessment documents, stakeholder research reports and from recent literature searches, up to September 2008 for human health effects and exposure sections of the document. Key studies were critically evaluated; modelling results may have been used to reach conclusions.

Evaluation of risk to human health involves consideration of data relevant to estimation of exposure (non-occupational) of the general population, as well as information on health hazards (based principally on the weight of evidence assessments of other agencies that were used for prioritization the substance). Decisions for human health are based on the nature of the critical effect and/or margins between conservative effect levels and estimates of exposure, taking into account confidence in the completeness of the identified databases on both exposure and effects, within a screening context. The screening assessment does not represent an exhaustive or critical review of all available data. Rather, it presents a summary of the critical information upon which the conclusion is based.

This screening assessment was prepared by staff in the Existing Substances Programs at Health Canada and Environment Canada and incorporates input from other programs within these departments. The draft of this screening assessment was subject to a 60-day public comment period. Additionally, this assessment has undergone external written peer review/consultation. Comments on the technical portions relevant to human health were received from Toxicology Excellence for Risk Assessment (TERA), Dr. Sergio Pellis of University of Lethbridge, Alberta, and Dr. Alfonso Abizaid of Carleton University, Ottawa. While external comments were taken into consideration, the final content and outcome of the screening risk assessment remain the responsibility of Health Canada and Environment Canada. The critical information and considerations upon which the assessment is based are summarized below.

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Substance Identity

For the purposes of this document, this substance will be referred to as bisphenolA.

Table 1. Substance identity

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Physical and Chemical Properties

Table 2 contains experimental and modelled physical and chemical properties of bisphenolA that are relevant to its environmental fate.

Table 2. Physical and chemical properties for bisphenolA

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Sources

There is no reference in the published literature to the natural occurrence of bisphenolA in the environment.

Global production of bisphenol A was 4 billion kg in 2006. In the United States (U.S.), production quantities increased from 521 million kg in 1990 to 736 million kg in 1995. Estimated production in the U.S. in 2007 was 1 billion kg (SRI Consulting, 2007).

Based on a survey conducted under section 71 of CEPA 1999, no bisphenol A was manufactured in Canada in 2006 at quantities greater than or equal to 100 kg. However, 26 companies reported using bisphenol A in Canada in the range of 100 000 1 000 000 kg or importing approximately half a million kg into Canada either alone, in a product, in a mixture or in a manufactured item (Environment Canada 2007a). The extent to which the reported values represent quantities of bisphenol A present in finished and semi-finished goods entering Canada from other parts of the world is unknown as these uses would be unlikely to meet the reporting criteria for the survey.

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Uses

BisphenolA is primarily used as a monomer in the production of polycarbonates and as a precursor or a starting material for monomers of certain epoxy resins (EFSA2006). In 2003, consumption patterns of bisphenolA in the U.S. indicated that approximately 72% of bisphenolA was used to manufacture polycarbonates, 21% was used in epoxy resins and 6% was used in other applications (NTP2007).

The type of mixture, product or manufactured item reported for 2006 in response to a survey of Canadian industry under section 71 of CEPA 1999 included resins, curing agents, epoxy curing agents, hardeners, plastic resin formulations, monomer, paperboard packaging, metal cans, phenolic resins, industrial coatings, plasticizers, adhesives, two part epoxy adhesives, chain oil, brake fluid, heat transfer fluid and lubricant formulations. Information voluntarily submitted in 2007 in response to the Challenge Questionnaire and other information submitted by industry additionally include use in epoxy polymer flooring, as a laminating adhesive, in custom colour powder coating and as a curing agent for resurfacing concrete.

The literature indicates that polycarbonate is used in the manufacture of compact discs, food and beverage contact containers (e.g., baby bottles, repeat use water bottles, pitchers, water carboys, tableware and storage containers), water pipes, medical devices, and in glazing applications and film. Polycarbonate blends find application in the electric and electronics industry (e.g., alarm devices, mobile phone housings, computer parts, household electric equipment, lamp fittings, power plugs) and the automotive industry (e.g., car headlight and rear light reflectors and coverings, bumpers, radiator and ventilation grills, safety glazing, inside lights, motorcycle windshields and protective helmets) (NTP2007; EFSA2006). Medical devices which contain polycarbonate (e.g., hemodialyzers, hemofilters, blood oxygenators and respiratory devices) have been approved or licensed for use in Canada under the Medical Devices Regulations of theFood and Drugs Act (Health Canada, Medical Devices Bureau, Health Products and Food Branch, pers. comm., 2008Feb5, unreferenced).

Epoxy resins made with bisphenol A are commonly used in protective coatings, structural composites, electrical laminates, electrical applications and adhesives (NTP 2007). Epoxy-phenolic resins are used as liners in metal cans for foods and beverages and as coatings on metal lids for glass jars and bottles, as well as surface coatings on residential drinking water storage tanks (EFSA 2006).Cans used for food and beverages, including ready-to-use, concentrate and powdered infant formula, that were evaluated by Health Canada are lined with epoxy resin (Health Canada, Chemical Health Hazard Assessment Division, Health Products and Food Branch, pers. comm., 2008 Feb 26, unreferenced).

Epoxy resins are used in automotive electro-coating and in automotive and industrial primer and topcoat applications. Epoxy resins are also used industrially in powder coatings used on reinforcing bars for concrete, oil pipelines, wire racks and shelving, as automotive primers and as general industry primers and topcoats. Aerospace coatings (e.g., decals or logos on the sides or tails of planes) also contain bisphenolA (Environment Canada2007b).

Dental products containing bisphenolA-based polymers (e.g., resin-based composite restorative materials and sealants) have been approved or licensed for use in Canada under the Medical Devices Regulations of the Food and Drugs Act (Health Canada, Medical Devices Bureau, Health Products and Food Branch, pers. comm., 2008 Feb5, unreferenced).

BisphenolA is also used in the production of phenolplast, phenolic and unsaturated polyester resins, thermal paper, polyvinylchloride (PVC), alkoxylated bisphenolA and polyols/polyurethane (NTP2007). Additionally, it may be used in the manufacture of brake fluids, tires, modified polyamide, and tetrabromobisphenolA (ECB2003). Other products of bisphenolA include protective window glazing, building materials, optical lenses and development of dyes (NTP 2007).

The U.S. National Library of Medicine Household Products Database reports that bisphenolA-based polymers are used in coatings, adhesives and putties available to the general public for use in automobiles, home maintenance and repair, and hobbies (NTP 2007). Response to a survey conducted under section71 of CEPA1999 indicated that these products were also reported to be available for use in Canada. BisphenolA-based polymers may also be used in the production of cosmetics, such as lipsticks, face and eye makeup and nail lacquers (Health Canada, Cosmetics Division, pers. comm., 2008March18, unreferenced).

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Releases to the Environment

There are no known natural sources of bisphenolA and potential releases to the environment are restricted to those associated with anthropogenic activities.

Releases of bisphenolA may occur during production, processing, use or disposal of the substance or products containing it. Results from a section 71 Notice with respect to certain Batch 2 Challenge substances conducted for 2006 (Environment Canada 2007a) indicated that bisphenolA was not manufactured in Canada in that year in an amount equal to or greater than 100 kg, although it was imported into the country for use in processing. As production of bisphenolA is not known to be occurring in Canada, potential releases from this source will not be considered further in this assessment.

Based on its moderate water solubility and low vapour pressure, wastewaters and washing residue generated during production and processing of application materials such as polycarbonates and epoxy resins are the most likely industrial sources of release of bisphenol A into the Canadian environment. Unintentional release of fugitive dust from closed systems during handling and transportation of the substance may also occur. Bisphenol A has low vapour pressure at typical environmental temperatures; however, elevated temperatures occurring during some processing operations may increase the vapour pressure, resulting in formation and possible emission of gaseous bisphenol A from manufacturing facilities.

BisphenolA released to wastewater would likely be transported to a treatment facility. Once at the plant, the moderate organic carbon-water partition coefficient (log K_oc 2.53) suggests that a fraction of the release will be sequestered by sludge. However, the moderate water solubility of 120 to 300mg/L indicates that a portion will likely remain in the water phase and will therefore be present in final effluents discharged to receiving waters. Lee et al. (2004) measured concentrations of 1.0 x 10^-5 to 1.73 x 10^-2µg/L in effluents and 0.07 to 10.6 mg/kg dry weight in raw sludge samples collected in 1999 and 2000 from four Toronto wastewater treatment plants. The results provide evidence that bisphenolA partitions into both the solid and liquid phases within treatment plants. A more detailed review of concentrations measured in waste treatment products is provided in Table 9a. BisphenolA present in sewage sludge could be released into the soil compartment through application of sludge biosolids to agricultural and pasture lands. Lee and Peart (2000a) examined the efficiency of bisphenolA removal by the sewage treatment plant process. Overall reduction rates of from<1% to 99% were determined from influent/effluent sample pairs collected from 36 Canadian sewage treatment plants, with a median reduction value of 68%. Plants exhibiting greater than 50% reduction rates were those employing secondary waste treatment.

The National Pollutant Release Inventory (NPRI) has tracked the release of bisphenolA from industrial facilities in Canada since the 1990s. Table 3 provides a summary of NPRI on-site release and total disposal data over the period 2001 to 2005. Where information was provided, on-site releases were exclusively to the air compartment. In recent years, total annual releases have varied between about 44 and 8770kg, with no clear temporal trend apparent. Most recent reporting (2005) indicates total on-site releases of 120kg, a decrease from that seen in previous years (NPRI 2007). It should be noted that only facilities meeting established criteria are required to report to the NPRI, and therefore NPRI data are likely to underestimate total Canadian releases of bisphenolA. For example, all of the following reporting criteria must have been met before NPRI reporting of bisphenolA was required for the 2005 calendar year: (1) employees at the facility worked a total of 20000 hours or more or the facility was used for an activity to which the 20000-hour employee threshold does not apply; and (2) the facility manufactured, processed or otherwise used 10tonnes (10000kg) or more of the substance in one year; and (3) the substance was manufactured, processed or otherwise used at a concentration greater than or equal to 1% by weight, with the exception of NPRI substances considered to be by-products. Facilities not meeting all of these criteria were not required to submit information relating to the release and/or disposal of bisphenolA at their facility. While the number of such facilities is not known, any such releases would not be reported and would therefore be over and above those reported through the NPRI.

BisphenolA also appears on the United States Environmental Protection Agency Toxics Release Inventory (TRI), a database of chemicals managed through disposal and other releases, recycling, energy recovery and waste treatment. In 2005, a total of approximately 675000kg (1.5×10⁶pounds) of bisphenolA were reported to the TRI as disposed or released, with land disposal predominating (about 550000kg) and releases occurring to air (about 90000kg) and surface waters (about 4500 kg; USEPA2007).

Table 3. NPRI release and disposal data (kg) for bisphenolA from 2001 to 2005

Release over the service life of end products may occur through volatilization or leaching. The majority of bisphenolA appears to be effectively retained within the polymer matrix of materials such as polycarbonates and therefore losses through leaching from the product surface are expected to be limited (ECB2003). As well, the low vapour pressure suggests bisphenolA will have little tendency to volatilize from products at normal environmental temperatures. Some losses could occur at elevated temperatures, for example, during heating of end products. Losses may also take place through weathering and breakdown of end products, particularly those with outdoor applications.

BisphenolA may enter the environment through physical and chemical degradation of end products during disposal and recycling operations. Releases would be primarily to soil, and to a lesser extent, to water and air. Moderate water solubility suggests there is potential for leaching from disposal sites under some circ*mstances (e.g., during rainfall events) and concentrations up to 1.41 x 10^-3 mg/L have been reported in groundwater samples collected in the vicinity of municipal landfills (Rudelet al.1998; see section on Ecological Exposure Assessment). Higher concentrations would be expected should the leachates have an elevated pH.

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Environmental Fate

Based on its physical and chemical properties (Table 2) and compartments to which it is released, the results of Level III fugacity modelling (Table 4) suggest that bisphenol A is expected to partition predominantly to soil or water, depending on the compartment of release.

As a moderately hydrophobic substance with some water solubility, bisphenolA can be expected to partition to organic phases such as sediments and soils; however, an appreciable fraction will also likely be present in the dissolved phase (Cousins et al. 2002). This suggests that the medium of release may be particularly important in predicting the partitioning behaviour and fate of bisphenolA in the environment.

Level III fugacity modelling (Table 4) predicts that when bisphenolA is released exclusively into water, the majority (96.9%) will remain within the water column with only a small fraction (3.1%) moving into sediment. When released to air or soil, or when equal proportions are released to all three media, most bisphenolA (78.7% to 99.3%) will partition to the soil compartment, with smaller amounts expected to be present in water and sediment. BisphenolA is not predicted to remain within air, even when released exclusively into that compartment.

Table 4. Results of the Level III fugacity modelling (EQC 2003)^[1]

^[1] Input values used in the model were: Type 2 substance; molar mass 228 g/mol; data temperature 25°C; reaction half-life (h) 3.1 in air, 81.6 in water, 326 in soil and sediment; partition coefficients 4.5 × 10^-10 (dimensionless) for air-water, 9.7 L/kg for soil-water, 19.4 L/kg for sediment-water, 97 L/kg for suspended particles-water, 5.5 L/kg for fish-water, 100 (dimensionless) for aerosol-water

Eisenreich et al. (1981) predicted that bisphenolA released into the atmosphere would exist almost entirely in the particulate phase and would be subject to removal through dry deposition or photolysis. The small fraction present as vapour would react with photochemically generated hydroxyl radicals (half-life approximately 0.13day; AOPWIN2000) or undergo photolysis. Photodegradation products formed include phenol, 4-isopropylphenol, and a semiquinone derivative of bisphenolA (PhysProp2006).

Biodegradation is expected to be the dominant loss process for bisphenol A in most aquatic and terrestrial environments. Biodegradation half-lives of less than four days were measured in natural waters following an adaptation period (acclimation) from one to four days (Dorn et al. 1987). A slower rate of biodegradation is expected to occur in non-acclimated waters; other processes such as sorption to suspended solids and sediments and photolysis may also take place. Recent research (e.g., Chin et al. 2004; Peng et al. 2006; Zhan et al.2006) documents both direct and indirect photochemical transformation of bisphenol A in aquatic media. Chin et al. (2004) reported that the rate of direct photolysis was significantly slower than that of indirect photolysis, which includes the presence of dissolved organic matter (DOM), and hypothesized that indirect photodegradation may proceed via multiple pathways involving DOM-derived phototransient substances. Bisphenol A is not expected to undergo hydrolysis due to a lack of hydrolyzable functional groups (Lyman et al. 1982). The low Henry's Law constant suggests that volatilization is unlikely to be a significant removal process.

The moderate water solubility of bisphenol A implies that the substance will exhibit low to moderate mobility when released to soil (Dorn et al. 1987). However, Loffredo and Senesi (2006) examined the adsorption kinetics and adsorption/desorption isotherms of bisphenol A onto four soil samples collected from the surface (0-30 cm) and deep (30-90 cm) horizons of two acidic sandy soils. The researchers concluded that soil adsorption of bisphenol A is generally reversible, with desorption occurring quickly and completely. Thus, the substance is expected to move down the soil profile and could possibly contaminate groundwater in acidic sandy soils (Loffredo and Senesi 2006). The movement of bisphenol A through the soil matrix and/or into groundwater will be influenced by the physical and chemical properties of the receiving soil and of water travelling through it, and in particular the pH of these media, as well as the nature and properties of organic material present in the soil and water. Quantities of bisphenol A available for transport will be determined by factors such as the loading rate of the substance into the receiving medium and degradation processes. Potential contamination of groundwater from soil-based bisphenol A contamination is thus site-specific and difficult to predict across the Canadian landscape. Rudel et al. (1998) reported concentrations of up to 1.41 × 10^-3 mg/L in groundwater samples collected in the vicinity of municipal landfills and wastewater treatment plants. Bisphenol A was present in monitoring wells located downgradient of secondary wastewater treatment plant infiltration beds to which no wastewater had been discharged for six months, suggesting no or only slow degradation of the substance in subsurface waters.

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Persistence and Bioaccumulation Potential

Data relevant to the evaluation of potential environmental persistence of bisphenolA are summarized in Tables 5a and 5b.

Table 5a. Empirical data for persistence

Aerobic biodegradation is likely to be the dominant loss process for bisphenolA in most aquatic and terrestrial environments (Staples et al. 1998). BisphenolA has demonstrated rapid biodegradation in standard aerobic 28-day ready biodegradability testing using OECD Test Guideline No. 301F (West et al.2001), but is indicated to be non-biodegradable using another aerobic OECD method, Test Guideline 301C (NITE 1977). Klecka et al. (2001) reported biodegradation half-lives of less than 3.5 days in river surface water samples, including those to which 0.05% surface sediment mixtures (top 1-2 cm) were added. A lag phase of from two to eight days was required prior to the onset of degradation.

A number of studies report resistance to biodegradation under conditions of low oxygen, suggesting there may be circ*mstances under which bisphenolA remains stable in the environment. Kang and Kondo (2002) measured rapid biodegradation in spiked river water samples incubated under aerobic conditions, but a less than 10% decrease in samples incubated anaerobically for an equivalent time period of 10 days. The researchers postulated that anaerobic bacteria may have little or no capability of degrading bisphenolA, although chemical degradation of the substance (for example, by lipoperoxide and sodium chloride) could still occur. Ying and Kookana (2005) documented rapid degradation of bisphenol A in soil within 7 days, but no degradation of the substance under anaerobic conditions in the soil. BisphenolA was reported to persist in an anaerobic slurry of river bed sediment, with no degradation evident after 3 months of incubation (Ronen and Abeliovich 2000). The substance also remained stable in anoxic estuarine sediments, with no degradation observed after 120 days (Voordeckers et al. 2002). Concentrations of bisphenolA remained unchanged in 70-day anaerobic biodegradation studies conducted using spiked samples of seawater and marine sediment (Ying and Kookana 2003), four different soil types (Ying and Kookana 2005), and water and sediment collected from an aquifer (Ying et al. 2003). Little or no degradation was reported for periods of up to 4 months, suggesting that half-lives of bisphenolA in anoxic and anaerobic media could well exceed one year.

Sarmah and Northcott (2008) examined degradation of bisphenol A in river water-sediment and groundwater-aquifer test systems. A biphasic degradation pattern was observed under both aerobic and anaerobic testing conditions, with rapid initial degradation of greater than 90% occurring in the first four days and trace amounts still remaining at the end of the 70-day study period. The formation of degradation products was monitored for the duration of the study; however, no potential metabolites were detected. Measured data were fitted using a first-order double exponential decay model, with results being presented as dissipation times (i.e., losses from the test system) rather than degradation half-lives. Reported 50% dissipation times (DT50s) were 0.57 to 1.212 days for aerobic conditions and 1.38 to 2.315 days for anaerobic conditions. Times for 90% dissipation were 2.26 to 2.75 days and 4.901 to 306 days for aerobic and anaerobic conditions, respectively. The researchers noted the incongruency of their results with those of previous studies which found little or no biodegradation of bisphenol A under anaerobic conditions, and suggested that lower test concentrations used in the study and variations in experimental protocols may have contributed to the observed differences. In addition, losses of up to 40% of the initial amount applied occurred in the sterile (control) treatments; this was considered significantly high and attributed to the possible occurrence of abiotic degradation processes or irreversible sorption of bisphenol A to test vessels.

The modelled data generally support the empirical evidence, with BIOWIN (2000) predicting a half-life for ultimate aerobic biodegradation of 37.5 days but no or very slow anaerobic biodegradation (Table 5b). The results from Table 5b show that the majority of the probability models suggest this substance does not biodegrade fast. However, only the MITI non-linear model produced a result less than 0.3, the cut-off suggested by Aronson et al. (2006) in identifying substances as having a half-life >60 days (based on the MITI probability models). CPOPs (2008), which uses the CATABOL model, predicts that bisphenol A will fail ready biodegradation testing based on OECD method 301C, supporting the experimental results reported by NITE (1977).

Table 5b. Modelled data for persistence

^[1] In all domains of applicability (global parameter, structural and metabolic)
^[2] Values were derived from the modelled half-life in water using the extrapolation factors of Boethling et al.(1995): t_{1/2 soil} : t_{1/2 water} : t_{1/2 sediment} = 1: 1: 4

The predicted half-life for atmospheric degradation of bisphenolA due to reaction with the hydroxyl radical is 0.133 day (Table 5b), indicating the substance will likely be rapidly oxidized in air. AOPWIN (2000) predicts reaction with atmospheric nitrate radicals may also occur.

There is conflicting evidence on the aerobic biodegradation potential of bisphenol A. Laboratory tests show positive and negative results of ready biodegradation potential. However, there is insufficient detail for the only laboratory aerobic biodegradation test showing a negative ready biodegradation test (NITE 1977) to determine whether the lack of biodegradation was the result of a lag phase. This test received lower weighting overall. There is greater concordance with the remaining laboratory and field evidence showing a fairly rapid loss of bisphenol A under aerobic conditions. Field tests are expected to be more indicative of the biodegradation potential of bisphenol A under natural conditions than the laboratory studies or QSARs based on laboratory data. Therefore, field data received greater weight. Results for the majority of aerobic biodegradation models also suggest a fairly short half-life, but some models (e.g., CPOPS and MITI non-linear model) suggest greater persistence of bisphenol A. Both these models, however, are trained on the NITE ready biodegradation data which, as already suggested, may not adequately account for a potential lag phase of bisphenol A. Therefore, these predictions are subject to test error which will not be captured by bisphenol A's structural characteristics alone and thus receive a lower weight. When combined laboratory, field and model evidence are considered, there is greater reliable evidence to suggest that bisphenol A does not meet persistence criterion for air (half-life in air ≥ 2 days), water and soil (half-life in soil and water ≥ 182 days) as set out in the Persistence and Bioaccumulation Regulations (Canada 2000). There are many studies showing that bisphenol A does not degrade or degrades only slowly under conditions of low or no oxygen and the measured presence in sediment, a medium to which there is no direct release, is further evidence of this slow degradation. Therefore, bisphenol A meets criteria for persistence in sediments (half-life in sediments ≥ 365 days) under these conditions.

The empirical and modelled log K_ow values of 3.32 and 3.64, respectively, suggest that bisphenolA may have some potential to accumulate in organisms. However, the substance demonstrates low bioaccumulation potential based on laboratory-derived bioconcentration factor (BCF) values obtained in testing with fish and other aquatic species (Table6a). Measured BCF values range from 3.5 to 68 in fish, but are higher in freshwater clams (107 to 144; Heinonen et al. 2002) and frogs (131 to 147; Kopenen et al. 2007). Takahashi et al. (2003) calculated bioaccumulation factors (BAF; i.e., incorporating consideration of uptake through food) of from 18 to 650 for periphyton (presence of detritus not reported, uncertain whether steady-state concentrations achieved) and from 8 to170 for benthos (procedures for sampling/collection, species not specified), and postulated that food may be an important uptake route for organisms in the aquatic environment.

Table 6a. Empirical data for bioaccumulation

Heinonen et al. (2002) reported very slow depuration of bisphenolA in freshwater clams, Pisidium amnicum, exposed at temperatures ranging from 2°C to 12°C, with statistically insignificant depuration occurring at the lowest test temperature of 2°C. This suggests that in addition to environmental exposure concentrations, body clearance rates and metabolism are important in determining final tissue levels of the substance. Metabolism of bisphenolA has been reported in fish (Lindholst et al. 2001), with the formation of glucuronides identified as the primary metabolic pathway. Kang et al (2006c) conducted a review of existing literature and also found that various organisms, including plants, fish, birds and mammals have capacity to metabolize bisphenol A.

A range of QSAR modelled bioaccumulation and bioconcentration values are available, with BAF or BCF values of from 5.8 (Baseline BCF Model) to 955 (OASIS Forecast; Table 6b). Most modelled values support a conclusion of low bioaccumulation potential.

Table 6b. Modelled data for bioaccumulation

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Potential to Cause Ecological Harm

There is modelled and experimental evidence that bisphenolA causes harm to aquatic organisms at relatively low concentrations. Although modelled predictions for aquatic toxicity were performed for this substance, given the numerous experimental data available (Tables 7a, 7b and 7c).

Table 7a. Empirical data for aquatic toxicity

^[1] Median effective concentration
^[2] Endpoint value based on cell count (2.7 mg/L) or total cell volume (3.1 mg/L)
^[3] Endpoint value based on cell count (1.0 mg/L) or chlorophyll a content (1.8 mg/L)
^[4] Median lethal concentration
^[5] Inhibiting concentration for a specified percent effect. A point estimate of the concentration of a test substance that causes 25% reduction in a quantitative biological measurement such as growth rate.
^[6] No observed effect concentration; highest concentration in a toxicity test not causing a statistically significant effect in comparison to the controls. In this case, the study identified that the highest concentration tested did not result in statistically significant results. Since the NOEC could be higher, the NOEC is described as being greater than or equal to the highest test concentration.
^[7] Low observed effect concentration; lowest concentration in a toxicity test that caused a statistically significant effect in comparison to the controls.
^[8] Minimum concentration in a toxicity test that just causes an observable effect.

Alexander et al. (1988) conducted a series of standard freshwater and saltwater toxicity tests on bisphenolA. Lowest endpoint values were a 96-hour EC₅₀ of 1.0 mg/L for the marine diatom, Skeletonema costatum, based on cell counts, and a 96-hour LC₅₀ of 1.1 mg/L for the mysid shrimp,Mysidopsis bahia. The mysid 96-hour LC₅₀value agrees well with that of 1.03 mg/L reported by Hirano et al. (2004).

Brennan et al. (2006) examined the potential for acute and chronic effects in two generations of water flea, Daphnia magna. The measured acute value was in the range of that reported by Alexander et al. (1988; see Table 7a). However, in chronic testing, higher mortality was observed in second generation daphnids compared with first generation animals exposed to the same test concentrations. The researchers postulated that elevated mortalities in the second generation may indicate weakening of the offspring relative to the first (parental) generation, leading to greater susceptibility to lethal effects in the second generation. This trend could indicate action at the cellular level of organisms, possibly DNA damage.

The same study also examined potential reproductive effects. No significant impacts were observed in the fecundity of first and second generation daphnids exposed to concentrations up to 1.0 mg/L. Mu et al. (2005). However, reported reduced reproductive success for Daphnia magna at a threshold concentration of 1.30 mg/L and proposed that the substance could elicit chronic effects through interference with ecdysis (moulting).

Seven-day NOEC and LOEC values of 0.94 and 1.88 mg/L, respectively, were reported for reproduction in the daphnid species, Ceriodaphnia dubia (Tatarazako et al.2002); the IC₂₅ for the study was 3.92 mg/L. These values are similar to results obtained using Daphnia magnaand provide further evidence of possible chronic toxicity.

Johnson et al. (2005) investigated potential impacts to the freshwater amphipod, Gammarus pulex. No statistically significant effects were evident on female moulting and juvenile production following a 14-day exposure period, although adult survival was reduced at the highest test concentration of 1.0mg/L.

Potential effects following long-term exposure (164days) were studied in the fathead minnow, Pimephales promelas(Sohoniet al. 2001). The growth of adult male fish and hatchability of offspring were significantly inhibited at a lowest concentration of 0.64 mg/L, while more subtle reproductive effects (i.e., induction of vitellogenin synthesis, alterations to sex cells) occurred at 0.016 mg/L. The researchers concluded that bisphenolA acts as a weak estrogen mimic to fish exposed via the water.

Lahnsteiner et al. (2005) exposed male and female brown trout, Salmo trutta f. fario, during the late pre-spawning and spawning period, and investigated effects on maturation, quantity and quality of sem*n and eggs. BisphenolA was added into the flow-through test system by means of an injection pump, with final reported test concentrations of 0.00175,0.00240 and 0.005mg/L calculated based on water flow and test chemical injection rates. Similar numbers of males (6-8) provided sem*n in the control group and at the two lowest concentrations, while only one male from the high test group gave sem*n. sem*n quality was lower at estimated concentrations of 0.00175 and 0.00240mg/L compared with the controls (reduced sperm density, motility rate, and swimming velocity), both at the beginning and in the middle of spawning. Production of high quality sem*n was restricted to the end of the spawning season and delayed for approximately 4 weeks in comparison to the control. The percentage of ovulated females was similar in the control group and two lowest exposure groups; however, females at the highest exposure group did not ovulate over the 103-day investigation period. Ovulation was delayed by approximately 2 weeks in the 0.00175mg/L exposure group and by approximately 3 weeks in the 0.00240mg/L group. Therefore, the tested bisphenolA concentrations affected the percentage of ovulated females and the time point of ovulation. No effect were observed on the quality of eggs (egg mass, percentile mass increase during hardening, egg fertility).

Acute aquatic toxicity estimates derived from QSAR models are in the range of 3 to 9mg/L (Table7b), slightly above results obtained experimentally. However, the predicted values are sufficiently close to those of laboratory studies to support a decision of potential acute aquatic toxicity at low concentrations. The weight of experimental and modelled evidence indicate that bisphenolA can be considered highly hazardous to the aquatic environment (i.e., (L(E)C₅₀ values at or approaching 1 mg/L and/or chronic NOEC equal to or less than 0.1 mg/L)..

Table 7b. Modelled data for aquatic toxicity

^[1] Median lethal concentration
^[2] Median effective concentration
^[3] CPOPS has identified bisphenol A as having an unspecified reactive model of toxicity in the fish 96hr LC₅₀ QSAR. The ≤ denotes that the median lethal concentration is expected to be lower than the reported default value which is based on a baseline narcotic mode of action
^[4] ± 95% confidence interval
^[5] Maximum Allowable Toxicant Concentration, generally presented as the range between the No Observed Effect Concentration and Lowest Observed Effect Concentration or as the geometric mean of the two measures

Evidence of disruption to reproductive and developmental processes following exposure to bisphenolA at concentrations below those causing acute effects has been reported in fish, aquatic invertebrates, amphibians and reptiles. A selection of reported endpoint values relevant to potential hormonal disruption is presented in Table7c. While there is a widespread variation in reported values, many fall in the range of 0.001 to 1 mg/L. In addition, differing sensitivities are evident between groups of organisms, with endpoint values for fish generally higher than those for aquatic invertebrates. Considered together, the data provide strong evidence that bisphenolA is capable of eliciting adverse effects: (1) following prolonged exposure at levels below those usually seen to elicit effects in standard toxicity tests (i.e., tests based on recognized methods which evaluate endpoints such as survival, reproduction and growth); (2) following brief low-dose exposure, particularly at sensitive developmental stages, with effects apparent later in the life cycle; (3) on filial generations following parental exposure; and (4) using more than one mode of action.

Table 7c. Selected endpoint values relating to potential hormonal effects

^[1] Significant effects occurred at the lowest test concentration
^[2] One test concentration
^[3] Test concentration was reported as 4.7×10^-7 M

Only limited toxicity data are available for non-mammalian terrestrial species; these are summarized in Table 8.

Table 8. Empirical data for terrestrial toxicity

^[1] No observed effect concentration; highest concentration in a toxicity test not causing a statistically significant effect in comparison to the controls.
^[2] Low observed effect concentration; lowest concentration in a toxicity test that caused a statistically significant effect in comparison to the controls.
^[3] Study identified that the highest concentration tested did not result in statistically significant results. Since the NOEC could be higher, the NOEC is described as being greater than or equal to the highesttest concentration.
^[4] The concentration which causes an effect on 25% of the test organisms.
^[5] Median effective concentration.
^[6] No observed adverse effect level; highest dose in a toxicity test not causing a statistically significant effect in comparison to the controls.
^[7] Lowest observed adverse effect level; lowest dose in a toxicity test that caused a statistically significant effect in comparison to the controls.

No statistically significant effects on reproduction were evident in earthworm, Eisenia sp., exposed for 56 days to concentrations of up to 100 mg/kgdw of soil; however, statistically significant and concentration-dependent mortality occurred at concentrations of 100mg/kgdw of soil and above (up to 10000mg/kg dw; Johnson et al. 2005).

A "limit test" using concentrations of 150 and 1000mg/kg soil was performed using six species of terrestrial plants (cabbage, corn, oat, soybean, tomato and wheat) (Dow Chemical Canada Inc.2007). Effects were evident at both concentrations tested. No information on statistical endpoints was provided in the summary report submitted to Environment Canada.

Ferrara et al. (2006) reported phytotoxicity and morphological anomalies in four plant species (broad bean, tomato, lettuce and durum wheat) grown hydroponically for 21days at doses of 10 and 50mg/L bisphenolA.

ECB (2008) reported results from three unpublished studies conducted using springtail (Folsomia candida; ECT 2007a), enchytraeid worm (Enchytraeus crypticus; ECT 2007b), and six terrestrial plant species (Springborn Smithers 2007). In springtail testing, no significant effects on survival occurred up to a highest exposure level of 1000 mg/kg dw; however, reproduction was significantly reduced at this concentration and the NOEC for this endpoint was therefore 500 mg/kg dw (EBC 2008). No significant effects on survival and reproduction were reported in enchytraeid worms up to a maximum exposure concentration of 100 mg/kg dw. Emergence in six species of terrestrial plants was significantly inhibited at a lowest exposure level of 320 mg/kg dw (cabbage, tomato), while the lowest effect level for growth was 50 mg/kg dw (tomato; ECB 2008).

Furuya et al. (2006) reported on the development of male chicken phenotypes (comb, wattle and testes) in White Leghorn,Gallus domesticus, chicks administered oral doses of 0.002 to 200mg/kg-bw bisphenolA in corn oil/alcohol carrier every two days from 2 to 25weeks of age. Dose-dependent growth inhibition of all three endpoints occurred even at the minimum dosage, although recovery was evident at all but the highest dosage. This suggests development was delayed at lower doses and permanently altered at the highest dose of 200mg/kg-bw per 2 days (i.e., 100 mg/kg-bw per day).

While there is little information on potential effects in wildlife species, a number of studies examine toxicity in rodents (see Health Effects Assessment section). Chronic oral exposure to low levels of contaminants, for example through ingestion in food, is more likely to be the scenario of concern for wildlife and there are some studies that, though limited when subject to a weight of evidence analysis, are suggestive of potential effects at doses well below the NOAELs of 5 and 50 mg/kg-bw per day established for bisphenolA in rodent species. There is also evidence that the pregnant female/fetus and developing offspring may be a sensitive subpopulation (see Health Effects Assessment section). A more detailed consideration of potential impacts in mammalian species is provided in the Health Effect Assesment section of this document.

Table 9a provides a summary of Canadian and North American water and sediment concentrations, as well as levels reported in water and wastewater treatment products.

Table 9a. Concentrations of bisphenolA in the ambient environment and waste treatment products

^[1] Sediment or sludge concentrations are reported on a dry weight basis
^[2] Municipal wastewater treatment plant
^[3] Primary effluent
^[4] Influent to secondary treatment
^[5] Secondary effluent
^[6] Values are reported as a combined range of both influent and effluent concentrations
^[7] Values estimated from graph
^[8] Median and maximum values are for different sample sets collected from two rivers in the same region

The highest reported values for surface water and sediment were 0.012 mg/L and 0.0061 mg/kg dw, respectively. Most surface water concentrations were much lower than the maximum reported value, however, with levels more in the range of 0.0001 to 0.001 mg/L. In their study on sediment concentrations in Lake Erie, Chu et al. (2005a) reported that highest concentrations of bisphenol A (up to 0.0061 mg/kg dw) were measured in the western basin of the lake, likely reflecting the substantial loading and input from numerous wastewater treatment plants in the region into Lake Erie and the Detroit River, which empties into it.

Rudel et al. (1998) reported a maximum concentration of 0.00141 mg/L in groundwater samples collected in the vicinity of municipal landfills, while up to 2.9×10-5 mg/L was detected in monitoring wells located downgradient of secondary wastewater treatment plant infiltration beds to which no wastewater had been discharged for six months.

High levels have been measured in some industrial wastewaters, most notably those associated with paper and allied products (maximum 0.14923 mg/L; median 0.00872 mg/L), chemicals and chemical products (maximum 0.09127 mg/L; median 0.00150 mg/L) and commercial laundries (maximum 0.04345 mg/L; median 0.00656 mg/L) (Lee et al. 2002).

BisphenolA has also been detected in influent, effluent and sludge collected from municipal wastewater treatment plants across Canada. In general, effluent levels are lower than those found in influents from the same plant, indicating some removal from wastewater streams during processing. Lee and Peart (2000a) estimated a median reduction rate of 68% (range<1-99%) based on concentration differences measured in 36 influent/effluent sample pairs collected from Canadian municipal wastewater treatment plants over the period 1999-2000. The values incorporate consideration of losses through degradation and from partitioning to sludge. The presence of high concentrations in some sludge samples (up to around 40mg/kg dw) confirms that bisphenolA partitions into this medium, even while some portion remains within the water phase. This supports the prediction made by Cousinset al. (2002; see Environmental Fate section) that based solely on properties of moderate hydrophobicity and water solubility, bisphenolA will partition into organic phases such as sediments and soils while an appreciable fraction remains in the dissolved phase.

High levels of up to 4.6 mg/kg dw were measured in biosolid products produced by municipal wastewater treatment plants in seven different U.S. states (Kinney et al. 2006). The maximum concentration was found in a wet cake product intended for use in agriculture.

BisphenolA has been measured in surface waters, sediments, and wastewater treatment products in many European countries. These data are well summarized in ECB (2003).

Limited information on measured levels in soil was found in the published literature, and this was more relevant to the assessment of human health (see Environmental Media section of Human Health Exposure Assessment).

Only limited North American data were found for levels in biota; however, some European and Japanese data are available (see Table 9b).

Table 9b. Concentrations of bisphenolA in biota

^[1] Concentrations are reported on a dry weight basis
^[2] Concentrations are reported on a wet weight basis
^[3] Pooled samples muscle tissue from 50 fish
^[4] Pooled samples liver tissue sample from10 fish

The approach taken in this ecological screening assessment was to examine available scientific information and develop conclusions based on a weight-of-evidence approach and using a precautionary approach, as required under section 76.1 of CEPA 1999. Lines of evidence with respect to bisphenol A relate to persistence, bioaccumulation potential, toxicity, environmental occurrence, and trends in production and use.

The available data indicate that bisphenolA does not persist under aerobic conditions. However, the substance has been found not to degrade or to degrade only slowly under conditions of low or no oxygen. This stability, combined with significant anthropogenic production and use, could lead to the formation of areas of accumulation or "sinks" in the environment, with the potential to act as future sources of exposure for organisms

The available data generally point to low bioaccumulation potential and a capacity for metabolism in various species. Most measured bioaccumulation and bioconcentration factors range upwards to approximately 147 L/kg; however, a lower trophic level study suggests a potential bioaccumulation factor of up to 650. These studies confirm that bisphenol A is bioavailable and can accumulate in tissues to some degree.

BisphenolA has been detected in surface waters, sediment, and groundwater over wide areas in Canada and the United States. It has also been measured in municipal and industrial wastewaters, sludge and biosolids. The published literature contains no reference to North American air concentrations; however, releases reported to the NPRI (see Table 3) were exclusively to the air compartment and therefore at least some presence in air can be expected from steady emission to the atmosphere. In addition, some limited data are available for concentrations in North American soils (see Exposure Assessment for Human Health). While few data exist for levels in North American biota, concentrations of up to 0.075 mg/kg dw are documented in fish and invertebrates from Europe and Japan. The data indicate that bisphenolA is present in a wide range of environmental media.

Bisphenol A has shown high toxicity to aquatic organisms, with acute toxicity values falling below 13 mg/L and chronic values below 2 mg/L. The empirical and modelled data demonstrate that bisphenol A can be considered highly hazardous to the aquatic environment.

A quantitative evaluation of exposure and ecological effects was conducted for bisphenolA as part of the weight of evidence evaluation of its potential to cause harm. Predicted Environmental Concentrations (PECs) were determined based on an analysis of exposure pathways, and these exposure pathways were used to identify sensitive receptor organisms. For bisphenolA, PECs were derived using Canadian environmental monitoring data, or if monitoring data were not available, the PECs were based on simple calculation procedures that took into account some degree of local environmental conditions, but largely relied on generic environmental parameters. A Predicted No-Effects Concentration (PNEC) was determined for each exposure pathway, by dividing a Critical Toxicity Value (CTV) by an application factor. CTVs typically represented the lowest ecotoxicity value (e.g., an EC 10, LOEC, etc.) from an acceptable data set. A risk quotient analysis, integrating estimated potential exposures with potential adverse effects, was performed and the resulting risk quotient (PEC/PNEC) was used to estimate the potential for ecological risk.

An effluent concentration of 0.0173 mg/L was reported for a municipal wastewater treatment plant in the Toronto area (Leeet al. 2004). This value is selected for derivation of the Predicted Exposure Concentration (PEC) for pelagic organisms as it represents the highest reliable Canadian effluent concentration from a municipal treatment plant providing both primary and secondary wastewater treatment and is therefore deemed both realistic and conservative. Applying a dilution factor of 10 (to account for exposures in the immediate mixing zone) yields a PEC value of 0.00173 mg/L. Lahnsteiner et al. (2005) reported significantly reduced sem*n quality and delayed ovulation in the brown trout, Salmo trutta f. fario, at a Lowest Observed Effect Concentration (LOEC) concentration of 0.00175 mg/L and this value is selected as the CTV. To derive a Predicted No-Effects Concentration (PNEC), the CTV is divided by an application factor of 10 in order to account for variability in interspecies and intraspecies sensitivity and extrapolation from laboratory to field conditions. The resulting PNEC for the pelagic compartment is 0.000175 mg/L and the risk quotient (PEC/PNEC) is 0.0173/0.000175 = 9.9. The results indicate that bisphenol A concentrations in water have the potential to cause adverse effects on populations of pelagic organisms in Canada.

For the sediment compartment, Johnson et al. (2005) reported a LOEC of 1.0 mg/L for reduced adult survival in the freshwater amphipod, Gammarus pulex, and this value is selected as the CTV. While this testing used water only as the exposure medium, the results reflect adverse effects on an epibenthic organism and are therefore considered appropriate for use in estimating potential risk to sediment species. A higher assessment factor of 100 is selected due to the paucity of data for this environmental compartment, and the resulting PNEC is 0.01 mg/L.

While measured concentrations are available for sediment (e.g., Chu et al. 2005), the PEC value of 0.00173 mg/L derived for the pelagic compartment will also be used in the calculation of a risk quotient for the sediment compartment. This approach allows direct comparison with the PNEC value, which was determined in water-only testing, and is reasonable given that a significant exposure route for sediment species will be by means of interstitial water or water in the boundary layer directly above the sediment surface. The risk quotient for benthic species is then 0.00173 / 0.01 = 0.173. The risk quotient analysis results suggest there is currently a low risk of direct adverse effects to sediment organisms in Canada.

The limited soil concentration data were deemed more relevant for the assessment of exposure to humans (see Exposure Assessment, Human Health); however, several studies provide Canadian data on measured levels of bisphenol A in sewage sludge. Application of sewage sludge to agricultural lands represents a direct pathway for bisphenol A into soil. A maximum concentration of 36.7 mg/kg dw was reported for digested sewage sludge collected from a municipal wastewater treatment plant in southern Ontario (Lee and Peart 2000a,b), and this value is selected for derivation of the PEC. The PEC is estimated for tilled agricultural soil and pastureland based on an approach described in Bonnell Environmental Consulting (2001). This approach uses the same basic approach as outlined in the EU's Technical Guidance Document (European Communities 1994) but modified for Canadian conditions. The following assumptions were used: dry sludge is applied annually to soil at the rate of 0.5 kg/m²; soil has a bulk density of 1700 kg/m³; sludge is mixed in soil to a depth of 0.2 m (depth of tillage) in agricultural soil and 0.1 m in pastureland (Bonnell Environmental Consulting 2001 and European Communities 1994); no loss of the target substance due to erosion; no input from atmospheric deposition; and no background bisphenol A accumulations in the soil. It is also assumed that bisphenol A persists in the soil for a period of at least 56 days, the exposure period of the study used for the derivation of the PNEC value (see below).

The PEC is then:

PEC_{agricultural soil} = (36.7×0.5) / (0.2×1700) = 0.054 mg/kg dw

PEC_pastureland = (36.7×0.5) / (0.1×1700) = 0.11 mg/kg dw

A 56-d LOEC of 100 mg/kg dw was reported for significant mortality in the earthworm, Eisenia sp. (Johnson et al. 2005), and this value is selected as the CTV. A lower toxicity value of 10 mg/L was reported for significantly reduced growth in several terrestrial plant species (Ferrara et al. 2006), however, this value is not suitable for comparison with the PEC as the plant study was reported in terms of water concentrations only. An application factor of 1 00 is applied to the CTV in order to account for extrapolation from laboratory to field conditions and interspecies and intraspecies variations in sensitivity. The application factor is larger than that used for pelagic organisms as fewer data are available for soil species and therefore the uncertainty in extrapolating between species is greater. The resulting PNEC value is 1.0 mg/kg dw. No information is available on the organic carbon (OC) content of the sludge samples or the soil used in the earthworm study, therefore both will be assumed to have a standard OC content of 2%. The risk quotient, PEC/PNEC, for soil organisms is therefore 0.054/1.0 = 0.054 for tilled agricultural soil and 0.11/1.0 = 0.11 for pastureland. The risk quotient analysis results suggest there is currently a low risk of direct adverse effects to soil species in the Canadian environment.

A PEC for mammalian wildlife was estimated based on a calculation of the total daily intake of the target substance by mink and otter. An energetics model based on the general exposure model for wildlife from the U.S. Environmental Protection Agency's (EPA) Exposure Factors Handbook (USEPA 1993) was used:

TDI = [FMR ((Ci × Pi) / (GEi × ATi))] × Pt

TDI=total daily intake (mg/kg-bw per day)
FMR=normalized free metabolic rate of wildlife receptor of interest (236 kcal/kg-bw per day for mink and 183 kcal/kg-bw per day for river otter)
C_i=concentration of contaminant in the ith prey species (mg/kg-bw) (see below)
P_i=proportion of the ith prey species in the diet (unitless) (default=35% for mink; 100% for otter)
GE=gross energy of the ith prey species (default=1240 kcal/kg-bw prey)
AEi=assimilation efficiency of the ith prey species by the wildlife receptor (default=0.91)
P_t=proportion of the time the receptor spends in the contaminated area (50% for mink and 50% for otter)

The model incorporated the metabolic rate of the wildlife receptors of interest (mink and otter), the proportion of food uptake by the receptors and the amount of time the animals spend in the contaminated area, which is based on the typical habitat range of the wildlife receptors.

As there are no Canadian data for bisphenolA concentrations in fish, concentrations of the target substance in fish (Ci) were estimated based on the highest PEC_water and a BAF. The BAF was estimated using the GOBAS BAF three trophic level model (Arnot and Gobas 2003; see Bioaccumulation section). The BAF represents a benthic/pelagic food chain and estimates the accumulation from all sources in a mid-trophic-level fish that would typically be eaten by a mammalian piscivore.

C_i = PEC_water · BAF

C_i= concentration in a prey fish (mg/kg-bw)
PEC_water = PEC of 0.00173 mg/L calculated above for the pelagic compartment
BAF= bioaccumulation factor of 148 L/kg estimated for mid-trophic level fish based Arnot and Gobas (2003)
C_i= 0.00173 · 148 = 0.256 mg/kg-bw

The model estimated PECs of 0.009 mg/kg-bw per day and 0.021 mg/kg-bw per day for mink and otter, respectively.

As identified in Characterization of Risk to Human Health, Health Effects Section, the dataset of neurodevelopmental and behavioural studies, though limited when subject to a weight of evidence analysis, is suggestive of potential effects at doses in the range of 0.01 to 0.1 mg/kg-bw per day and higher (study details are provided in Appendix D). The range of 0.01 - 0.1 mg/kg-bw per day is selected as an estimate for CTVs in mammalian wildlife. The CTVs for predictive sentinel wildlife species (in this case, mink and otter) can then be calculated from the range of CTVs of the mammalian test species, using the formula (Sample et al.1996):

CTV_wildlife = CTV_ts· (bw_ts/bw_w)^1/4

CTV_wildlife = critical toxicity value for wildlife
CTV_ts = critical toxicity value for test species
bw_ts = mean body weight of test species (0.35 kg)
bw_w = mean body weight of predictive sentinel species (0.807 kg for mink, 6.01 kg for otter; Canadian Wildlife Service, Ontario Region, pers. comm., 2004, unreferenced).

Therefore, the CTV_wildlife ranges are then 11 to 112 mg/kg-bw per day for mink and 42 to 420 mg/kg-bw per day for otter. Applying an application factor of 10 to account for extrapolation from laboratory to field conditions yields PNEC values of 0.0008 to 0.008 mg/kg-bw per day and 0.0005 to 0.005 mg/kg-bw per day for mink and otter, respectively. The resulting risk quotients (PEC / PNEC) are 1.25 to 12.50 for mink and 4.2 to 42.0 for otter. It is concluded that bisphenol concentrations in biota have the potential to cause adverse effects in populations of wildlife in Canada.

A similar methodology was used to derive a risk quotient relevant to avian wildlife species. The total daily intake (TDI) was calculated for a sentinel avian wildlife species, the lesser scaup (Aythya affinis). The lesser scaup was selected as a suitable avian sentinel species as it is abundant through much of North America and consumes clams as part of its normal diet of aquatic invertebrates and vegetation (USEPA 1993). As clams are present in the diet of lesser scaup, the tissue concentration of 0.0283 mg/kg dw reported in clam collected off the Massachusetts coast (Stuart et al. 2005) is selected for use in the development of a PEC for avian wildlife. This concentration is converted to 0.00535 mg/kg wet weight (ww) using a dry weight to wet weight ratio of 5.29 for clams (Vinogradov 1953). The total daily intake (TDI) for lesser scaup is then:

TDI = [FMR ((Ci × Pi) / (GEi × ATi))] × Pt

TDI = total daily intake (mg/kg-bw per day)
FMR = normalized free metabolic rate of wildlife receptor of interest (216 kcal/kg-bw per day for adult female scaup; USEPA 1993)
C_i = concentration of contaminant in the ith prey species (mg/kg-bw) (0.00535 mg/kg ww)
P_i = proportion of the ith prey species in the diet (unitless) (19.2% for adults; Dirschl 1969)
GE_i = gross energy of the ith prey species (default = 800 kcal/kg-bw prey)
AE_i = assimilation efficiency of the ith prey species by the wildlife receptor (default = 0.77)
P_t = proportion of the time the receptor spends in the contaminated area (assume 50%)

The model estimated a PEC value of 0.00018 mg/kg-bw per day for adult lesser scaup.

A LOAEL of 100 mg/kg-bw per day, based on permanently altered development in male White Leghorn chicks, Gallus domesticus, was reported by Furuya et al. (2006), and this value will be used as the CTV for the derivation of a PNEC for avian wildlife. The CTV for lesser scaup can then be calculated using the same formula as for mammalian wildlife (Sample et al. 1996). Sample et al. (1996) recommend using a scaling factor of 1 for interspecies extrapolation among birds; however, as the reported mean body weight for the lesser scaup considers the adult bird, while the toxicity value for Leghorn was derived using juveniles, an extrapolation which considers body weight was deemed appropriate in order to incorporate consideration of weight differences relating to life stage.

The CTV for avian wildlife is then:

CTV_wildlife = CTV_ts · (bw_ts/bw_w)

CTV_wildlife = critical toxicity value for wildlife
CTV_ts = critical toxicity value for test species (100 mg/kg-bw per day)
bw_ts = mean body weight of test species (1.3 kg; estimated from Leeson and Caston 1991)
bw_w = mean body weight of predictive sentinel species (0.77 kg; Nelson and Martin 1953)

The CTV for avian wildlife is therefore 169 mg/kg-bw per day. Applying an application factor of 100 to account for extrapolation from laboratory to field conditions and interspecies and intraspecies variations in sensitivity yields a PNEC value of 1.69 mg/kg-bw per day and the risk quotient is then 0.00018 / 1.69 = 0.0001. The results suggest there is currently a low risk of direct adverse effects to avian wildlife species in Canada.

Considered together, the various lines of evidence indicate that the potential impacts of bisphenolA in the Canadian environment are of sufficient magnitude to warrant use of a precautionary approach in response to uncertainties in the evaluation of risk.

There is evidence that bisphenolA is capable of altering hormonal, development or reproductive function: following prolonged exposure at levels below those usually seen to elicit effects in standard toxicity tests (i.e., tests based on recognized methods which evaluate endpoints such as survival, reproduction and growth); following brief low-level exposure, particularly when this occurs at sensitive developmental stages, with effects apparent later in the life cycle; in filial generations following parental exposure; and using more than one mode of action. Studies conducted using fish, aquatic invertebrates and amphibians demonstrate impacts in the range of 0.001 to 1 mg/L, levels that have been measured in Canadian and U.S. wastewaters, receiving waters and sediments. For this reason, potential exists for adverse effects in populations of aquatic organisms, particularly those residing in close proximity to treatment plant outfalls or other point sources. While the ability of bisphenolA to cause adverse low-dose effects appears clear, there is uncertainty respecting the extent to which organisms in the Canadian environment are exposed to levels sufficient to elicit the types of impacts observed in laboratory studies. Further information is needed to more clearly delineate the relationship between exposure levels measured in the Canadian environment and the potential for adverse effects in organisms.

BisphenolA is one of the highest volume chemicals produced worldwide and global demand for the substance is predicted to continue increasing at a rate of 6 % to 10% annually. The total quantity of bisphenolA reported as imported or in use in Canada during the calendar year 2006 was approximately 0.5 million kg. Actual quantities available for release to the Canadian environment could be much higher, however, as these values likely do not fully capture amounts entering Canada in finished and semi-finished products. This is particularly so, given the wide application of the substance and number of bisphenolA derivative products.

In summary, the lines of evidence relating to persistence, bioaccumulation, toxicity, environmental occurrence and trends in production and use raise concerns regarding the presence of bisphenol A in the environment. The substance is predicted to degrade fairly rapidly under aerobic conditions (half-life in the order of days), suggesting that bisphenol A released into receiving waters should remain only a short time in this medium. However, it has been widely detected in surface waters. Possible explanations for this may be that because inputs to the environment are sufficiently large and continuous, measurable concentrations are always present and/or that the substance is not degrading as rapidly as laboratory testing would indicate. Measurable levels in sewage sludge and effluent are also suggestive of some level of stability of this substance. Further, bisphenolA is present in media to which there is no direct release, such as sediment and groundwater. This implies that the substance remains sufficiently long enough in the environment to move from its point of release into other environmental media. There is evidence that bisphenolA is stable in some environmental compartments, most notably under conditions of low oxygen, such as those found beneath the surface layers of sediment and soil and in marshes and bogs. Levels in the environment are expected to increase given that bisphenolA is a high production chemical and global demand for the substance is increasing. BisphenolA has been detected in aquatic species, confirming that the substance can be taken up by organisms and stored in tissues. The potential for food web impacts is still undetermined, although bioaccumulation factors of 650 have been measured in lower trophic level species, suggesting that possible exposure of predator species could occur through consumption of prey. BisphenolA is acutely toxic to aquatic organisms and adversely affects reproduction in earthworms, growth in terrestrial plants, and development in birds. Results from an analysis of risk quotients, which compare the potential for exposure with possible effects in organisms, indicate that bisphenolA concentrations in the Canadian environment have the potential to cause adverse effects in populations of pelagic organisms and mammalian wildlife but are unlikely to cause adverse effects in benthic and soil organisms and bird species in Canada. In addition to acute toxicity, bisphenolA can impact the normal development of individual organisms and influence the development of their offspring. In laboratory testing, these effects have occurred at concentrations below those shown to cause acute effects, and these effect concentrations have been measured in the Canadian environment. Considered together, the various lines of evidence indicate that the potential impacts of bisphenolA in the Canadian environment are of sufficient magnitude to warrant use of a precautionary approach in response to uncertainties in the evaluation of risk.

There is uncertainty regarding sources of release and release quantities of bisphenol A into the Canadian environment. An industry survey was used to collect information on Canadian manufacture, import and use of bisphenol A for the calendar year 2006. However, it is not clear whether potential releases from partially-finished and finished goods being imported into Canada and from products currently in use in Canada have been adequately captured. This information is needed to more accurately estimate potential exposure concentrations for organisms in the Canadian environment. The presence of a recent database of measured environmental concentrations for some locations in Canada and the United States assisted with characterizing exposure potential in this assessment. However, quantitative exposure estimates were possible only for pelagic organisms. Increasing the size of this database by including a wider range of Canadian locations, more information on possible time trends, and more data on sediment, soil and surface water concentrations, as well as possible presence in leachates and groundwater, would increase the level of certainty in exposure estimation and allow estimation of potential exposure for benthic and terrestrial species.

There is uncertainty about the measured presence and accumulation potential of bisphenolA in biota. The available information indicates that bisphenolA does not meet bioaccumulation criteria as specified in the Persistence and Bioaccumulation Regulations of CEPA 1999. However, bioaccumulation factors of up to 650 have been determined for lower trophic level aquatic species, and this suggests there may be circ*mstances or conditions under which bisphenolA may accumulate within organisms. Lower trophic aquatic organisms form an important bridge between primary producers (i.e., phytoplankton) and higher trophic species such as fish. Therefore, while high tissue levels could indicate the presence of a site-specific condition, such as very high local concentrations, it is also possible that bioaccumulation is occurring with the subsequent potential for food chain transfer and secondary poisoning of predator species. An increased database of measured concentrations in Canadian biota, including trophic magnification studies, would provide greater clarity on the potential for accumulation within individual organisms and along food webs.

There is uncertainty regarding the potential for adverse effects in Canadian biota resulting from single-dose or prolonged (i.e., lifetime and multigenerational) low-level exposure to bisphenolA, particularly during sensitive life cycle stages. This uncertainty reflects a need for greater clarification of minimum effects levels in pelagic, sediment, soil and wildlife species, and of exposure concentrations in the Canadian environment. Based on the weight of available data, the assessment indicates that there is a potential for adverse impacts in aquatic organisms, particularly those residing in the vicinity of treatment plant outfalls or other point sources. However, information is needed in order to more clearly delineate the relationship between exposure levels in the Canadian environment and possible effects in organisms.

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Potential to Cause Harm to Human Health

Dietary Intake

With respect to dietary intake, two sources of potential exposure are (i) migration into food and beverages from packaging materials that directly contact food, specifically food and beverage containers with internal epoxy resin coatings and (ii) migration from polycarbonate repeat use containers such as baby bottles and drinking bottles.

(i) Migration from Food Packaging

Epoxy resins are used as an interior protective lining for food and beverage cans and for metal closures for some jars and bottles. As a result of these food contact uses, very small quantities of the monomer bisphenolA can migrate into the food and beverage contents of the containers.

Exposure to bisphenolA from food packaging uses has been characterized based on bisphenolA concentrations reported in a survey of canned foods on the U.S. market in 2007 (EWG 2007). EWG (2007) tested foods and beverages from 97 cans purchased in grocery stores in three states. Twenty-eight different types of foods were analyzed for bisphenolA, including canned fruits, vegetables, pasta, beans, infant formula, meal replacement and canned milk. Results are presented in Table 10. Reported concentrations were considered to be representative of additional data available to Health Canada and of data reported in the literature (EFSA 2006).

Table 10. Average bisphenolA concentrations for various canned food types

^[1] Average is geometric mean
^[2] Not detected at the limit of detection 1 ppb

Potential daily intakes from the use of epoxy resins as interior protective lining for food and beverage cans were estimated using average and maximum concentrations of bisphenol A, and daily intake values for the food commodities (Health Canada,Food Directorate, Health Products and Food Branch [HPFB], pers. comm., 2007 March14 unreferenced) and are presented in Table 11. Leaching from metal closures was considered to contribute minimally to these exposures.

Table 11. Potential daily intake of bisphenolA (µg/kg-bw per day) from use of epoxy resins as interior protective lining for food and beverage cans

^[1] Based on the average concentration of bisphenolA found in foods as reported in Table 10.
^[2] Based on the maximum concentration of bisphenolA found in foods as reported in Table 10.
^[3] Based on body weight of 14.4 kg for 1-4 year olds; 26.4 kg for 5-11 year olds; 53.8 kg for 12-19 year olds; and 60 kg for adults (Nutrition Canada 1980)

Analyses of bishenol A in other food items are limited. Miyamoto and Kotake (2006) identified bisphenolA in noncanned foods, but attributed this to additives in ink in paper used in the food packaging. Table 9b presents concentrations of bisphenolA found in aquatic biota. Seafood purchased in a Singapore market was analysed by Basheer et al. (2004) and mean concentrations (n=5 for each product) were reported as 13.3 ng/g in prawns, 213.1 ng/g in crabs, 56.5 ng/g in blood co*ckle, 27.4 ng/g in white clam, 118.9 ng/g in squid, and 65.6 ng/g in carangid fish (limit of detection 1.357 ng/g). Vivacqua et al (2003) detected bisphenolA in the range of 0.25 to 1.0 l g/g in fresh market vegetables in Italy (8/14 detects).

As there are no measured concentrations of bisphenol A in fish caught in Canada or North America, or of fish available on the Canadian market, the maximum concentration reported in fish from European countries (see Table 9b) was used to derive estimates of intake from fish consumption. Using a maximum concentration of 61.97 µg/kg as measured by Fjeld et al. (2004) and based on rates of consumption of fish for Canadians (Health Canada 1998), the estimated intake of bisphenol A from this source would range from 0.11 µg/kg-bw per day (for 12-19 year olds) to 0.24 µg/kg-bw per day (for 1-4 year olds). Studies have also indicated that concentrations of bisphenol A tend to be higher in fish liver and other organs (Belfroid et al. 2002) so consumption of these organs may slightly increase exposure. Confidence in these estimates is low as they are based on a number of assumptions; however it is likely that they are an overestimate of actual exposure from this source.

As infant formula (ready-to-use, concentrate, powder) available on the Canadian market is packaged in cans, a separate estimate of the dietary intake of bisphenol A by infants was conducted. Due to the nature of the extracting medium, migration into liquid is considered to be greater than migration into powder and analysis therefore focussed on liquid formulations. Preliminary research conducted at Health Canada supports the conclusion that migration of BPA is significant only for liquid infant formulas (Health Canada, Food Directorate, HPFB, pers. comm., 2008 Sept 04, unreferenced). In 2007, Health Canada analyzed canned liquid infant formula to characterize residual levels of bisphenol A. Samples packaged in cans lined with epoxy-based coatings were collected in October, 2007 in Ottawa, Canada from 21 cans representing eight brands of ready-to-drink or concentrated liquid infant formulas. Bisphenol A was detected and quantified in all samples, with concentrations ranging from 2.27 to 10.2 ppb (Cao et al.2008). These data are consistent with other data reported in the literature. For instance, EWG (2007) reported average bisphenol A concentrations of 2.4 ppb (range from limit of detection of 1 ppb to 17 ppb) in 6 cans of infant formula available on the US market. The Canadian data were considered to be the most relevant for use in the assessment. As a worst-case scenario it was assumed that liquid infant formula is the main source of food for infants in the first 18 months of their life, and that their formula intake is equal to their daily total fluid intake (i.e., formula, milk, water, juice). A dilution factor of 0.5 was applied to the concentrated liquid formulas as they would be diluted with water prior to the feeding (1 part water to 1 part concentrated formula). No dilution factor was applied to the ready-to-use infant formulas. Assumptions regarding formula intakes are based on recommendations of the Institute National de Santé Publique du Québec (INSPQ 2001). Potential daily intake of bisphenol A from liquid infant formula packaged in containers with epoxy resin interior protective lining is provided in Table 12.

Table 12. Potential daily intake of bisphenolA (µg/kg-bw per day) from liquid infant formula packaged in containers with epoxy resin interior protective lining.

^[1] Based on the average and maximum concentration of bisphenolA measured in twenty-one cans representing 8 brands of ready-to-drink or concentrated liquid infant formula purchased in Ottawa, Ontario (see paragraph above)) with selective grouping according to the age of use intended by the manufacturer.
^[2] Based on average female infant weight of 3.9 kg (NCHS 2000) and an average formula intake of 644 g/day and maximum formula intake of 1080 g/day (INSPQ 2001).The average and maximum concentrations of bisphenolA found in products intended for use by this age group were 2.72 and 4.89 ppb respectively after accounting for dilution of concentrated formulas.
^[3] Based on average female infant weight of 5.5 kg (NCHS 2000) and an average formula intake of 1080 g/day and maximum formula intake of 1470 g/day (INSPQ 2001).The average and maximum concentrations of bisphenolA found in products intended for use by this age group were 2.72 and 4.89 ppb respectively after accounting for dilution of concentrated formulas.
^[4] Based on average female infant weight of 7.2 kg (NCHS 2000) and an average formula intake of 1050 g/day and maximum formula intake of 1470 g/day (INSPQ 2001).). The average and maximum concentrations of bisphenolA found in products designed for this age group were 2.70 and 5.12 ppb respectively after accounting for dilution of concentrated formulas.
^[5] Based on average female infant weight of 9.0 kg (NCHS 2000) and an average formula intake of 735 g/day and maximum formula intake of 960 g/day (INSPQ 2001).The average and maximum concentrations of bisphenolA found in products intended for use by this age group were 2.70 and 5.12 ppb respectively after accounting for dilution of concentrated formulas.
^[6] Based on average female infant weight of 10.6 kg (NCHS 2000) and an average formula intake of 750 g/day and maximum formula intake of 900 g/day (INSPQ 2001).). The average and maximum concentrations of bisphenolA found in products intended for use by this age group were 3.10 and 5.44 ppb respectively after accounting for dilution of concentrated formulas.

Infants can also be exposed to bisphenol A through the consumption of human breast milk, however no Canadian data on concentrations of bisphenol A in breastmilk were identified. In a survey of 20 U.S. lactating women with no known occupational exposure, Ye et al. (2006) identified concentrations of total bisphenol A (free + conjugate) ranging from the limit of detection (1 ng/ml) to 7.3 ng/mL (7.1 ppb - assuming breastmilk has a density of 1.03 g/mL (Health Canada 1998)), with a mean concentration of 1.9 ng/mL (1.8 ppb). Sun et al. (2004) reported a mean bisphenol concentration in breast milk of 0.61 ng/ml (range 0.28 to 0.97ng/ml) in a survey of 23 Japanese women (limit of detection 0.11 ng/ml). In another Japanese study, Otakaet al. (2003) quantified levels ranging from the limit of detection (0.09 ng/g) to 0.70 ng/g (ppb) of bisphenol A in the breast milk of three women. Based on the U.S. data (Ye et al. 2006), potential daily intakes for infants exposed to bisphenol A through breastmilk are provided in Table 13.

Table 13. Potential daily intake of bisphenolA (µg/kg-bw per day) from consumption of human breastmilk containing either average or maximum bisphenol A concentrations.

^[1] Average and maximum concentrations of bisphenolA as measured in breastmilk from 20 lactating women in the United States (Ye et al.2006).
^[2] Based on average female infant weight of 3.9 kg (NCHS 2000) and an estimated breast milk intake of 596 g/day (Arcus-Arthet al. 2005).
^[3] Based on average infant weight of 5.5 kg (NCHS 2000) and an estimated breast milk intake of 649 g/day (Arcus-Arth et al. 2005).
^[4] Based on average infant weight of 7.2 kg (NCHS 2000) and an estimated breast milk intake of 744 g/day (Arcus-Arth et al. 2005).
^[5] Health Canada recommends exclusive breastfeeding of infants from 0-6 months, at which time they would be introduced to solid foods. Although exposure through breastmilk may continue past the age of six months, the intake of breast milk past this age is expected to decrease. Therefore intake from this source was not estimated past seven months of age.

Migration of small amounts of bisphenolA from polycarbonate repeat-use containers such as baby bottles, drinking bottles, pitchers and carboys can contribute to oral exposure. Numerous studies have reported on migration of bisphenolA from polycarbonate bottles into water or milk simulant under different conditions (Brede et al. 2003; Biedermann-Brem et al. 2007; Biles et al. 1997; Cao and Coriveau (2008); Ehlert et al. 2008; Kawamura et al. 1998; Leet al. 2008; Maragou et al. 2007; Mountfortet al. 1997; Environmental Defence, 2008; Sun et al. 2000; Takao et al. 1999; Wong et al.2005, D'Antuono et al. 2001, Miyamoto and Kotake 2006) and data have been summarized by others (EFSA 2006; ECB 2003; NTP 2007). These studies were conducted under diverse conditions using food simulants such as water, 10% ethanol, 50% ethanol and 3% acetic acid. In considering these results, it is important to differentiate between study designs intended to simulate realistic exposure scenarios versus those designed to characterize maximum concentration that may migrate from the product under extreme conditions or over the life of a container.

Five studies of bottles obtained from the Canadian market are available. In 2000, Health Canada conducted a series of limited migration studies(Page et al. 2006). No bisphenolA was detected in milk simulant (50% ethanol/water) following storage in polycarbonate baby bottles for up to 7 days at 4 or 22°C (detection limit of <0.1 µg/kg). In additional trials, bottles were subjected to the following treatments: (i) storage for 7 days at 4°C then storage for an additional 7 days at 22°C, followed by 6 days at 70°C; (ii) storage for 7 days at 22°C then storage for 7 days at 4°C, followed by 6 days at 70°C C; (iii) storage for 6 days at 70°C. In these 3 trials, aliquots of stimulant were collected after 1, 2, 3 and 6 days at 70°C and analyzed for bisphenolA. For all three trials, bisphenolA concentrations in the simulant on day 1 were 0.4 to 1 ppb, and concentrations increased on subsequent days of incubation at 70°C. Sustained incubation at 70°C is not considered a meaningful study design for characterization of migration under realistic use conditions.

A second study also measured the migration of bisphenol A from polycarbonate baby and drinking bottles into water. Cao and Corriveau (2008) tested five polycarbonate bottles, three baby bottles and two refillable drinking bottles, bought in Ottawa, Canada in 2007. The bottles were filled to capacity with boiling water and then left at room temperature for 24 hours. Bisphenol A concentrations in the water ranged from 1.7 to 4.1 µg/L (ppb).

Preliminary raw data are available from a study conducted by Health Candada in 2008 (Health Canada, Safe Environments Programme, Healthy Environments and Consumer Safety Branch (HECSB), pers. comm., 2008 March 6, unreferenced ). Migration was characterized by filling 14 brands of baby bottles available in Canada with water to simulate migration to aqueous and acidic foods, and a 50% ethanol/water solution to simulate migration to fatty foods such as infant formula. The bottles were incubated for 8, 24 or 240 hours at 40°C to simulate use at room temperature, as per recommendations from the US FDA (2007). The 240-hour scenario (10 days) was included to estimate migration from both repetitive use and worst-case scenarios. Average migration for these tests ranged from 0.095 µg/L (ppb) in water incubated for 8 hrs, to 2.05 µg/L in the 50:50 ethanol/water blend incubated for 240 hours. The scenario considered to represent the most realistic use consisted of filling bottles with the 50:50 ethanol/water blend and incubating them for 8 hours. This scenario resulted in average concentrations of 0.15 µg/L.

In another study of polycarbonate baby bottles on the Canadian market (Environmental Defence 2008), new baby bottles (n=9) were filled with water, sealed and allowed to sit for 24 hours; initially at room temperature and then in a subsequent trial at 80°C. The study investigators specify that the 80°C treatment was considered to simulate repeat washing of bottles (i.e., approximately 60-100 washes). As such, the 80°C trials are not considered to represent a realistic migration from a single use. Results from the room temperature trials showed that bisphenol A levels ranged from below the limit of detection, i.e. 0.05 to 0.063 ng/mL (ppb). Results from the 80°C trial ranged from 4.294 to 8.323 ppb.

Most recently, the migration of bisphenol A from various types of new and used baby bottles into 3.25% milk and apple juice was analyzed under normal use conditions (INSPQ pers. comm., 2008 September 8, unreferenced). Thirty bottles were washed in a dishwasher, filled with milk and refrigerated at 4°C for 16 hours. Bottles were then heated in a microwave oven for 30 seconds to 37^oC before the milk was tested for bisphenol A. The same bottles were then rewashed and placed under similar experimental conditions to test migration into apple juice; the only difference in protocol was that the juice was brought to room temperature. The concentration range of bisphenol A measured in milk migrating from new, hard polycarbonate bottles ranged from 0.44 to 0.53 µg/L (n=10) and was reported to be 0.29 µg/L in used bottles (n=5). The migration concentrations measured in apple juice ranged from 0.29 to 0.46 µg/L in new bottles (n=10) and was 0.43 µg/L when measured in the used bottles (n=5). For both liquids, bisphenol A concentrations from the disposable plastic bags and glass bottles were below the detection limit (0.2 µg/L).

Maragou et al. (2007) assessed migration potential from 31 new polycarbonate baby bottles (from 6 different brands) available on the Greek market under a variety of conditions, including those which study investigators considered to be representative of actual use. To determine the impact of the repeated cleaning of bottles, either with the dishwasher or with brushing, the bottles were cleaned using either method, rinsed and sterilized prior to being filled with water and incubated at 70°C for 2 hours. This series of treatments did not lead to detectable bisphenolA migration with a reported detection limit of 2.4 ppb. Migration was only detected when bottles were filled with boiling water (100°C). Specifically, 5 repeat cycles of cleaning the bottles by brushing with detergent, sterilizing in boiling water for 10 minutes, then filling with boiling water (100°C) and leaving at ambient temperature for 45 minutes resulted in an average bisphenolA concentration of 10 ppb (ranged from limit of detection, 2.4ppb, to 14.3 ppb).

Additionally, EFSA (2006) summarizes two studies conducted on commercially available baby bottles, conducted under experimental conditions designed to represent realistic conditions of use, that measured concentrations of up to approximately 50 µg/L (ppb) of bisphenol A in the food simulant contents. In a study of bottles filled with boiling water and kept overnight at room temperature Hanai (1997) measured bisphenol A migration varying from 3 to 55 µg/L (detection limit of 2 µg/L). Earls et al. (2000) detected bisphenol A in 5 out of 12 bottles filled with either boiling water or 3% acetic acid, placed in a refrigerator for 24 hours, and then heated to approximately 40°C by immersion in boiling water for approximately 2 minutes. Bisphenol A concentrations in detected samples ranged from 20 to 50 µg/L (detection limit of 10 µg/L).

However, the Norwegian Food Safety Authority (Biedermann-Bremet al. 2007) investigated the effect of migration of bisphenol A from polycarbonate under extreme washing conditions such as washing using very strong alkali detergents (80^oC for 1 hour, followed by drying of unrinsed bottles at 90^oC for 30 minutes). They concluded that even under aggressive conditions, concentrations of bisphenol A in contents are unlikely to exceed 10 µg/L (ppb).

A recent study investigated migration behaviour of bisphenolA into water during microwaving of polycarbonate baby bottles obtained from the European market (Ehlert et al. 2008). Polycarbonate bottles filled with water were heated to 100°C during 3 microwave cycles. Migration of bisphenolA into water was shown to be in the range of <0.1 to 0.7 µg/L (ppb).

Miyamoto and Kotake (2006) reported bisphenolA concentrations of 0.05 to 3.9 ppb (limit of detection 0.05 ppb) in new unwashed polycarbonate bottles, which were exposed to water at 95°C for 30 minutes.

Polycarbonate carboys can be used to store drinking water. Bileset al. (1997) measured bisphenol A in a 5-gallon carboy by filtering a 1 L aliquot of water through an SPE cartridge (flow rate: 3 ml/min) followed by a DDH20 wash (2x20ml), with analysis by GC-MSD. The range of bisphenol A was 0.1 to 4.7 ng/L (0.0001 to 0.0047 ppb) (detection limit < 0.05 ng/L).

Le et al. (2008) measured bisphenol A migration from polycarbonate drinking bottles used for consumption of water and other beverages. New and used polycarbonate bottles were filled with water and left for up to 7 days at room temperature (22^oC). Samples were collected on days 1, 3, 5 and 7. On day 1, the calculated mean concentration of bisphenol A detected in the water was 0.24 ng/mL (ppb). Mean concentrations across all sampling days, for both new and used polycarbonate containers, ranged from 0.08 to1.33 ng/mL (ppb). In general, the concentration of bisphenol A released from polycarbonate bottles into water at room temperature increased with time, but no significant difference was observed between the concentrations resulting from the use of new versus used bottles. Repeating the study using boiling water (100^oC) and leaving the samples for 24 hr, during which time the water samples slowly cooled, increased the amount of bisphenol A migration from 15-fold to 55-fold.

Directions for preparation of concentrate and powder formula typically specify that water be boiled for sterilization purposes and then be cooled before adding to the bottle. For infants 0-18 months it was considered appropriate to estimate bisphenol A intake from the use of polycarbonate baby bottles for both the recommended room temperature use and plausible high temperature use (boiling water) scenarios. Results calculated from Le et al (2008) indicate that filling polycarbonate bottles with room temperature water, could result in migration of approximately 0.24 ppb of bisphenol A into the contents of the bottle. This is supported by preliminary data from a migration study conducted by Health Canada indicating an average concentration of 0.15 ppb bisphenol A (Health Canada, Safe Environments Programme, HECSB, pers. comm., 2008 March 6, unreferenced).To estimate migration of bisphenol A from addition of boiling water [100°C] directly to the baby bottle, the results of Maragou et al. (2007), in which boiling water was added to polycarbonate bottles and cooled for 45 minutes, support a concentration of 10 ppb bisphenol A. As a worst-case scenario it was assumed that liquid infant formula is the main source of food for infants in the first 18 months of life, and so formula intake is considered to be equal to their daily total fluid intake (i.e., formula, milk, water, juice). Exposure estimates are provided in Table 14.

Table 14. Contribution of use of polycarbonate baby bottles to potential daily intake (PDI) of bisphenol A (µg/kg-bw per day) for infants.

^[1] Based on an average concentration of bisphenol A detected in new and used polycarbonate bottles that were filled with room temperature liquid and left for 24 hours (Le et al. 2008). This value is also supported by an average concentration of 0.15 ppb reported from filling bottles with liquid at room temperature and leaving for 8 hours. (Health Canada, Safe Environments Programme, HECSB, pers. comm., 2008 March 06, unreferenced).
^[2] Based on an average concentration of bisphenol A resulting from filling polycarbonate baby bottles with boiling water and leaving at room temperature for 45 minutes (Maragouet al. 2007)
^[3] Based on average female infant weight of 3.9 kg (NCHS 2000) and an average formula intake of 644 g/day and maximum formula intake of 1080 g/day (INSPQ 2001).
^[4] Based on average female infant weight of 5.5 kg (NCHS 2000) and an average formula intake of 1080 g/day and maximum formula intake of 1470 g/day (INSPQ 2001).
^[5] Based on average female infant weight of 7.2 kg (NCHS 2000) and an average formula intake of 1050 g/day and maximum formula intake of 1470 g/day (INSPQ 2001).
^[6] Based on average female infant weight of 9.0 kg (NCHS 2000) and an average formula intake of 735 g/day and maximum formula intake of 960 g/day (INSPQ 2001).
^[7] Based on average female infant weight of 10.6 kg (NCHS 2000) and an average formula intake of 750 g/day and maximum formula intake of 900 g/day (INSPQ 2001).

No studies of activity-use patterns for repeat use containers are available to help determine frequency and method of use of polycarbonate repeat-use bottles by other segments of the population (i.e., children aged 1-4 years, 5-11 years, 12-19 years and adults). Therefore exposure estimates were based on a migration of 0.24 ppb of bisphenol A as calculated from Le et al.(2008), resulting from the room temperature use of polycarbonate bottles, as it was assumed that contents would typically not be heated prior to serving for these age groups. This value is supported by preliminary results of a study conducted by Health Canada (Health Canada, Safe Environments Programme, HECSB, pers. comm., 2008 March 6, unreferenced), and represents an overestimate of concentrations reported in water carboys by Biles et al. (1997). It was considered, conservatively, that all drinking water consumed by these age groups was from polycarbonate containers (e.g., repeat use drinking bottles, pitchers and carboys). Exposure estimates for this use are presented in Table15.

Table 15. Contribution of use of polycarbonate drinking containers to potential daily intake of bisphenol A (µg/kg-bw per day) for children and adults (aged 1+ years)

^[1] Based on a calculated average bisphenol A concentration of 0.24 ppb (Le et al.2008), assuming contents would not be heated prior to serving
^[2] assumed to weigh 14.4 kg (Nutrition Canada 1980) and to drink 0.7 L of water per day (Health Canada 1998)
^[3] assumed to weigh 26.4 kg (Nutrition Canada 1980) and to drink 1.l L of water per day (Health Canada 1998)
^[4] assumed to weigh 53.8 kg (Nutrition Canada 1980) and to drink 1.2 L of water per day (Health Canada 1998)
^[5] assumed to weigh 60.0 kg (Nutrition Canada 1980) and to drink 1.5 L of water per day (Health Canada 1998)

Another potential source of exposure to bisphenolA is through the use of polycarbonate tableware and storage containers. Kawamuraet al. (1998) measured the migration of bisphenolA from tableware items commercially available in Japan (i.e., measuring cups, mugs and bowls) using either 20% ethanol maintained at 60°C degree C for 30 minutes (n=9) or water at 95°C degree C for 30 minutes (n=9). Under both sets of conditions, migration was below the limit of quantification (0.6 ppb) in 7 of 9 samples. In the 4 samples with detectable concentrations of bisphenolA, concentrations ranged from 1.7 to 4.5 ppb. Due to limited data on migration from polycarbonate tableware and storage containers, the lack of information on availability of these products on the Canadian market, and use patterns by the general population, exposure estimates cannot be derived. However, based on available information, under normal use conditions exposure from this source is expected to be limited.

Many recent reviews have summarized bisphenolA concentrations in environmental media (Wilhite et al. 2008; Vandenburget al. 2007; NTP 2007; Kang et al.2006b).

Data on concentrations of bisphenolA in Canadian waters are summarized in Table 9a. The only published Canadian drinking water values were measured by Chen et al. (2006) in Alberta at the Bearspaw (on the Bow River) and the Glenmore (on the Elbow River) water treatment plants with bisphenolA concentrations of 0.00045 and 0.00076 µg/L respectively (no detection limit presented). These values fall within the range reported by other countries, with concentrations between 300 pg/L to 2 ng/L (0.0003 to 0.002µg/L) (Kuch and Ballschmiter 2001) in Germany, and from below detection (0.01µg/L) to 0.06µg/L in Japan (Nakanishi et al. 2007). Boyd et al. (2003) was the only other Canadian study identified and although bisphenolA was detected (LOD 0.1 ng/L), it was not quantifiable, making this study unsuitable for deriving estimates of bisphenolA intake for the Canadian population.

No Canadian data were identified for ambient or indoor air. Although bisphenolA is released to air, due to its low vapour pressure it is more likely to partition to soil or water rather than remain in the air, where it has a half life of a few hours (see Environmental Fate section of Ecological Exposure Assessment). Concentrations in ambient air have been measured in two pilot studies conducted in 1997 (Wilson et al. 2001, 2003) and in a 2000-2001 survey (Wilson et al. 2007) of outside residences and daycare centres in North Carolina and Ohio. In the 2000-2001 survey, bisphenolA was detected in ambient air outside residences in 31% of samples in North Carolina, ranging from below the limit of detection (0.9 ng/m³) to 44.6 ng/m³ (50th percentile 0.9 ng/m³), and 34% of samples in Ohio, ranging from the limit of detection to 19.0 ng/m³ (50th percentile 0.920 ng/m³) (Wilsonet al. 2007). Concentrations of bisphenolA measured in ambient air outside daycare centres, ranged from below the limit of detection (0.01 ng/m³) to 51.5 ng/m³ (50th percentile below detection) in North Carolina (detected in 38% of samples), and below detection to 6.94 ng/m³ (50th percentile below detection) in Ohio (detected in 44% of samples). In deriving intake estimates for the Canadian population, it was considered appropriate to use the maximum and fiftieth percentile ambient air concentration in North Carolina (44.6 ng/m 3, 0.9 ng/m³), taking into consideration the residential location and larger sample size (n=127).

Concentrations of bisphenolA in indoor air were measured by Wilson et al. (2001, 2003, 2007) in 42 daycare centres and 153 residences across North Carolina and Ohio in 2000-2001. In Ohio residences, bisphenolA was detected in 63% of the samples within a range of concentration from below the limit of detection (0.01 ng/m³) to 30.3 ng/m³ (50th percentile 0.980ng/m³). In North Carolina, this range was extended to include a maximum concentration of 193 ng/m³ (50th percentile 1.82 ng/m³) with bisphenolA detected in 68% of samples. For daycare centres, indoor air concentrations ranged from below the detection limit (0.01 ng/m³) to 7.42 ng/m³ in Ohio, and 8.99 ng/m³ in North Carolina, with bisphenolA detected in 73% and 45% of the samples, respectively. In another survey of 120 homes in Cape Cod, Massachusetts, conducted by Rudel et al. (2003), concentrations in indoor air were below the reporting limit (18 ng/m³). In an earlier study, Rudel et al.(2001) measured bisphenolA in air in the range of 0.002 to 0.003 µg/m³ (2-3 ng/m³) in offices and residential locations (n=5). The maximum and fiftieth percentile indoor air concentrations in North Carolina (193 ng/m 3, 1.82 ng/m³) were considered the most appropriate for use in deriving estimates of intake based on the residential location and sample size (n=127).

No Canadian data were identified for soil. One study in the United States analyzed soil in four daycare facility play areas, and reported bisphenolA concentrations that ranged from 0.005 to 0.007 µg/g (mean of 0.006 µg/g, n=2), and in the vicinity of nine homes with levels ranging from 0.004 to 0.014 µg/g (mean of 0.007 µg/g, n=9) in North Carolina (detection limit 0.001 µg/g)(Wilsonet al. 2003). The maximum and mean concentrations in soil (0.007 µg/g; 0.006 µg/g) from residential locations were considered most appropriate for use in deriving estimates of intake.

Preliminary raw data are available for Phase 1 of the Canadian House Dust Study. Based on air-dried vacuum samples collected from 260 Canadian homes, median concentrations of bisphenolA in dust were reported at 1.60 µg/g dust (ranged from less than the method reporting limit [0.17 µg/g] to 23.84 µg/g). BisphenolA was detected in 99% of homes (Health Canada, Environmental Health Science and Research Bureau, pers. comm., 2008 Feb 7, unreferenced). These results are consistent with concentrations reported in dust from homes in other countries. In a survey of 118 homes in Cape Cod, Massachusetts, Rudel et al. (2003) identified a maximum value of 17.6 µg/g (mean of 1.65 µ/g) with 86% of samples above the reporting limit. Wilson et al.(2007) also reported dust levels in residences in North Carolina and Ohio. In North Carolina, 25% of homes had detectable levels, with a maximum value of 707 ng/g (50th percentile was reported as the limit of detection). In Ohio, 72% of homes had detectable levels, with a maximum value of 589 ng/g (50 th percentile is below detection limit). The maximum and median concentrations (23.84 µg/g; 1.60 µg/g) identified in dust measured in homes in Canada were considered most appropriate for use in deriving estimates of intake.

Tables 16 and 17 present mean and upper bounding estimates of intake for each age group in the general population of Canada based upon concentrations of bisphenolA identified in environmental media. Dust and soil concentrations have been summed for the purposes of the intake table.

Table 16. Average estimates of potential daily intake of bisphenolA by the general population in Canada from environmental media (µg/kg-bw per day by various age groups)

^[1] Assumed to weigh 5.9 kg (NCHS 2000), to breathe 2.1 m³ of air per day and ingest 30 mg of soil per day (Health Canada 1998).
^[2] Assumed to weigh 9.0 kg (NCHS 2000), to breathe 9.3 m³ of air per day and to ingest 100 mg of soil per day (Health Canada 1998).
^[3] Assumed to weigh 14.4 kg (Nutrition Canada 1980), to breathe 9.3 m³ of air per day, to drink 0.7 L of water per day and to ingest 100 mg of soil per day (Health Canada 1998).
^[4] Assumed to weigh 26.4 kg (Nutrition Canada 1980), to breathe 14.5 m³ of air per day, to drink 1.1 L of water per day and to ingest 65 mg of soil per day (Health Canada 1998).
^[5] Assumed to weigh 53.8 kg (Nutrition Canada 1980), to breathe 15.8 m³ of air per day, to drink 1.2 L of water per day and to ingest 30 mg of soil per day (Health Canada 1998).
^[6] Assumed to weigh 60.0 kg (Nutrition Canada 1980), to breathe 16.2 m³ of air per day, to drink 1.5 L of water per day and to ingest 30 mg of soil per day (Health Canada 1998).
^[7] Wilson et al. (2007) reported that t he fiftieth percentile ambient air concentration as measured in samples taken outside residences in North Carolina, U.S.A, was below the detection limit of 0.0009 µg/m³(detected in 31% of samples, n=127). Canadians are assumed to spend 3 hours outdoors each day (Health Canada 1998).
^[8] Wilson et al. (2007) reported that t he fiftieth percentile indoor air concentration as measured in samples inside residences in North Carolina, U.S.A, was 0.00182 µg/m³ (detected in 68% of samples, n=127). Canadians are assumed to spend 21 hours indoors each day (Health Canada 1998).
^[9] The only published Canadian values for bisphenolA in drinking water were 0.00045 and 0.00076 µg/L measured at two water treatment plants in March 2003 in Alberta, Canada (Chen et al. 2006) (sample size unknown, no detection limit presented). The maximum concentration value 0.00076 µg/L was used in generating the estimated average intake due to the lack of an average value presented in this reference.
^[10] NA - not applicable. Reported concentrations in drinking water (0.00076 µg/L) are significantly lower than those reported in liquid infant formula.
^[11] The estimated intake for soil and dust is based on the sum total of the average concentration identified in each medium.
^[12] The average concentration in soil of 7 µg/kg is based on a survey of 9 residences in North Carolina, U.S.A as part of a study by Wilson et al. (2003) which also included day care play facilities (detection limit of 1 µg/kg).
^[13] The median concentration in dust of 16 00 µg/kg was identified in the preliminary data of the Phase 1 Canadian House Dust Study of 260 homes (Health Canada, Environmental Health, Science and Research Bureau, pers. comm., 2008 Feb, unreferenced).

Table 17. Upper-bounding estimates of potential daily intake of bisphenolA by the general population in Canada from environmental media (µg/kg-bw per day by various age groups)

^[1] Assumed to weigh 5.9 kg (NCHS 2000), to breathe 2.1 m³ of air per day and to ingest 30 mg of soil per day (Health Canada 1998).
^[2] Assumed to weigh 9.0 kg (NCHS 2000), to breathe 9.3 m³ of air per day and to ingest 100 mg of soil per day (Health Canada 1998).
^[3] Assumed to weigh 14.4 kg (Nutrition Canada 1980), to breathe 9.3 m³of air per day, to drink 0.7 L of water per day and to ingest 100 mg of soil per day (Health Canada 1998).
^[4] Assumed to weigh 26.4 kg (Nutrition Canada 1980), to breathe 14.5 m³ of air per day, to drink 1.1 L of water per day and to ingest 65 mg of soil per day (Health Canada 1998).
^[5] Assumed to weigh 53.8 kg (Nutrition Canada 1980), to breathe 15.8 m³ of air per day, to drink 1.2 L of water per day and to ingest 30 mg of soil per day (Health Canada 1998).
^[6] Assumed to weigh 60.0 kg (Nutrition Canada 1980), to breathe 16.2 m³ of air per day, to drink 1.5 L of water per day and to ingest 30 mg of soil per day (Health Canada 1998).
^[7] Wilson et al. (2007) reported t he maximum ambient air concentration as measured in samples taken outside residences in North Carolina, U.S.A., as 0.0446 µg/m³(detected in 31% of samples, n=127). Canadians are assumed to spend 3 hours outdoors each day (Health Canada 1998).
^[8] Wilson et al. (2007) reported t he maximum indoor air concentration as measured in samples inside residences in North Carolina, U.S.A., as 0.193 µg/m³ (detected in 68% of samples, n=127). Canadians are assumed to spend 21 hours indoors each day (Health Canada 1998).
^[9] The only published Canadian values for bisphenolA in drinking water were 0.00045 and 0.00076 µg/L measured at two water treatment plants in March 2003 in Alberta, Canada (Chen et al. 2006) (sample size unknown, no detection limit presented). The maximum concentration value 0.00076 µg/L was used in generating the estimated upper-bounding intake.
^[10] NA - not applicable. Reported concentrations in drinking water (0.00076 µg/L) are significantly lower than those reported in liquid infant formula (10.2 µg/L) and so as a worst-case scenario it was assumed that all liquid consumed by these age groups (0-18 months) was in the form of liquid formula.
^[11] The estimated intake for soil and dust is based on the sum total of the maximum concentration identified in each medium.
^[12] The maximum concentration in soil of 14 µg/kg is based on a survey of 9 residences in North Carolina, U.S.A., in a study by Wilson et al. (2003) which also included day care play facilities (detection limit of 1 µg/kg).
^[13] The maximum concentration in dust of 23 840 µg/kg was identified in the preliminary data of the Phase 1 Canadian House Dust Study of 260 homes (Health Canada, Environmental Health, Science and Research Bureau, pers.comm., 2008 February, unreferenced).

Dental materials manufactured from bisphenolA derived monomers include composite resins used to restore cavities in the teeth of children and adults and dental sealants. Sealants may be used, infrequently, to seal the biting surfaces of baby (primary) teeth but are mainly used to seal the surfaces of permanent first and second molars that erupt into the mouth at ages six and twelve respectively (Health Canada, Medical Devices Bureau, pers. comm., 2008 February 27, unreferenced)

While current Canadian data on the prevalence of dental decay in children are not available, it is known through studies of individual populations that it is a common childhood disorder, particularly in certain at-risk populations. Leake and Main (1996) found that children aged five in Ontario had 1.22 decayed, missing or filled teeth. Peressini et al. (2004) found that children aged seven in an Ontario First Nations Community had 6.20 decayed, missing or filled teeth. Leake et al. (2001) in a sample survey of the oral health of 3657 Toronto children aged five, seven or thirteen, found that 30.0% of five year olds, 41.3% of seven year olds and 39.3% of thirteen year olds children had one or more cavities.

The restoration of decayed teeth, both primary and permanent, may be accomplished using a variety of materials including dental amalgam and composite resin. While current data on the prevalence of usage of composite resins versus other restorative materials are not available, anecdotal information suggests that the use of composite resin is increasing and is currently used more frequently than other materials to restore primary and permanent teeth in children.

Santerre (2007) has stated the three sources of bisphenolA derived from dental materials are unpolymerized monomer on the surface of the restoration or sealant; hydrolyzed polymers of bisphenol-derived monomers in the material and fine particulate polymer produced by abrasion of the restoration. Santerre (2007) further states that BisGMA, one of the most commonly used bisphenolA-derived monomers used in the manufacture of dental restorative materials and sealants, shows no capacity to be hydrolyzed back into bisphenolA under normal oral conditions.

A number of investigators have shown that placement of composite resin restorations and sealants can result in acute, transient exposures to bisphenolA (Fung et al. 2000; Joskow et al. 2006; Sasaki et al. 2005)

Fung et al. (2000) collected saliva following the application of 8 mg (applied to one surface) and 32 mg (8 mg applied to four surfaces) of a dental sealant to a group of 18 and 22 adults, respectively. Concentrations of bisphenolA in saliva ranged from 5.8 to 105.6 ppb (detection limit of 5 ppb) at 1 hr and 3 hrs following treatment but was not detected immediately after treatment or at other time intervals (24 hr, 72 hrs and 120 hrs) post-treatment.

More recently, Joskow et al. (2006) measured concentrations of bisphenolA in saliva of 14 adults before and after the application of between 39.0 to 42.5 mg of two dental sealants. Mean concentrations of bisphenolA before treatment, immediately following treatment, and an hour post-treatment were reported as 0.22 ng/ml, 0.54 ng/ml, and 0.24ng/ml in five patients treated with one sealant and 0.30 ng/ml, 26.5 ng/ml and 5.12 ng/ml in nine patients treated with the other sealant.

Sasaki et al. (2005) measured bisphenolA concentration in saliva collected from 21 patients before and after dental treatments using 9 different commercial composite resins. The concentration of bisphenolA in saliva for patients treated with bis-GMA monomer-based composite resins (and bonding agents) labelled A (n=4) measured in the range of 0.3 to 2.0 ng/ml (mean 0.87ng/ml) prior to treatment, 21.0 to 60.1 ng/ml (mean 32.1 ng/ml) following treatment and 1.6 to 4.7 ng/ml (mean 3.1 ng/ml) after gargling for 30 seconds with water at 37°C. Patients treated with the composite labelled F (n=3) had the highest concentration of bisphenolA in saliva post-treatment of (100 ng/ml; range 60 to 100 ng/ml) and after gargling (19 ng/ml; range 5 to 19 ng/ml). The concentration of bisphenol in saliva following the use of the other composites labelled C (n=2), G (n=2), H (n=2), and I (n=2) ranged from 10 to 45 ng/ml.

Azarpazhooh and Main (2008) observed that no dental sealants that carried the Seal of the American Dental Association [product recognition program, now discontinued] in 2007 released detectable amounts of BPA and continued with recommendations for simple clinical procedures [commonly followed] to reduce the possibility of unpolymerized BPA-derived monomer remaining on the surface of the tooth following sealant placement.

A method to extrapolate from concentrations in saliva to daily estimates of exposure was not identified. However, consistent with the data presented above, other international assessments have characterized exposure to bisphenolA as a result of dental treatments as an acute event (ECB 2003) or as nominal relative to other oral sources of exposure (Nakanishi 2007).

Consumer products are also a potential source of exposure. Miyamoto and Kotake (2006) characterized exposure to bisphenolA for infants from contact with consumer products such as toys. The contact time was established using a survey of mothers who observed mouthing behaviour in a total of 50 infants (Tanimura 1999) and videorecordings of 25 infants in which time spent mouthing toys was monitored (Tanimura 2000). A minimum migration rate (of 0 µg/cm²/min (for products that do not contain bisphenolA) and a maximum migration rate 0.0162 µg/cm²/min identified in the Japanese literature [Kanagawa 2000]were used. The estimated intake of bisphenolA was 0.026 µg/kg per day (0-5 month old boys) and 0.069 µg/kg per day (6-11 month old boys). The authors noted that only a few toys are made of bisphenolA-based polymers and, as such, these are likely to be overestimates of exposure. Similarly, few toys on the Canadian market are expected to be made of bisphenolA-based polymers and, additionally, these hard plastics tend to have a lower migration rate than would be expected from soft plastic toys (Health Canada, Consumer Product Safety Bureau, pers. comm., 2008 March17, unreferenced) and therefore contribution to exposure amoung infants and children is expected to be limited.

The Household Products Database (National Institutes of Health 2007) indicates that bisphenolA may be found as a residual monomer in epoxy resins used in the manufacture of epoxy adhesives. The resins may be present in the adhesive products at concentrations ranging from 0.1%- 100%. However, only trace quantities of bisphenolA are likely to be present in the resins with a predicted residual level of 1 ppm (identified in ECB 2003). An estimate of exposure to bisphenolA from epoxy adhesives predicts air concentrations during use to range from 3.82×10^-5 µg/m³ to 6.08×10^-5 µg/m³ (Table 18). The chosen scenarios correspond to gluing a handle on a coffee mug or gluing a large vase. However it is expected that the majority of exposures will occur at the lower end of this range. Dermal exposure may also result from the use of this product and was estimated at 0.0167 µg/kg-bw per use (Table 18). Confidence in these estimates is low as they are based on a number of assumptions; however it is likely that they overestimate actual exposures from this source.

Table 18. Upper-bounding estimates of potential exposure to bisphenolA from consumer products

^[1] Possible exposure of teenagers (12-19 years old) and adults (20+). Scenarios were completed for adults only.

It has been reported to Health Canada that polymers manufactured with bisphenolA may be used in the production of cosmetic products, such as lipsticks, eye makeup, face makeup and nail lacquers (Health Canada, Cosmetics Division, pers. comm., 2008 March18, unreferenced). Residual bisphenolA in these products are not known but are expected to be low and are not expected to contribute significantly to the overall exposure to the general population.

For exposure among formula fed infants aged 0 to 18 months, exposures from multiple sources (food packaging of infant formula, migration from polycarbonate bottles, and through environmental media) were aggregated and are presented in Tables 19a and 19b.

Table 19a. Aggregated average estimates of potential exposure for formula fed infants aged 0 to 18 months (µg/kg-bwperday).

^[1] Based on the average concentration of bisphenol A as measured in liquid infant formula (Health Canada, Food Directorate, HPFB, pers. comm., 2007 Dec 14, unreferenced) and an average formula intake for each age group (see Table 12).
^[2] If drinking water is used to dilute the infant formula during preparation, it may affect the overall bisphenol A concentrations. However as concentrations of bisphenol A measured in drinking water were significantly lower than those measured in liquid infant formula, that source of exposure was not included in these estimates.
^[3] Intakes from room temperature water are minimal compared to boiling water and are not presented here (see Table 14).
^[4] Based on the average concentration of bisphenol A measured in water, 10 ppb, that resulted from filling polycarbonate bottles with boiling water (Maragou et al. 2007) (see Table 14)
^[5] Heath Canada recommends that solid food be introduced to infants after 6 months of age. For this assessment it was assumed that an infants main source of food is infant formula, however if infants consume other canned foods this may increase their total dietary exposure to bisphenol A.
^[6] Based on average concentrations measured in various environmental media (see Table 16).

Table 19b. Aggregated maximum estimates of potential exposure for formula fed infants aged 0 to 18 months (µg/kg-bwperday).

^[1] Based on the maximum concentration of bisphenol A as measured in liquid infant formula (Health Canada, Food Directorate, HPFB, pers. comm., 2007 Dec 14, unreferenced) and the maximum formula intake for each age group (see Table 12).
^[2] If drinking water is used to dilute the infant formula during preparation, it may affect the overall bisphenol A concentrations. However as concentrations of bisphenol A measured in drinking water were significantly lower than those measured in liquid infant formula, that source of exposure was not included in these estimates.
^[3] Intakes from room temperature water are minimal compared to boiling water and are not presented here (see Table 14).
^[4] Based on the average concentration of bisphenol A measured in water, 10 ppb, that resulted from filling polycarbonate bottles with boiling water (Maragou et al. 2007) (see Table 14).
^[5] Heath Canada recommends that solid food be introduced to infants after 6 months of age. For this assessment it was assumed that an infants main source of food is infant formula, however if infants consume other canned foods this may increase their total dietary exposure to bisphenol A.
^[6] Based on maximum concentrations measured in various environmental media (see Table 17).

For exposure among breastfed infants aged 0 to7 months, exposures from multiple sources (breast milk and through environmental media) were aggregated and are presented in Table20.

Table 20. Aggregate estimate of potential exposure for breastfed infants aged 0 to 7 months (µg/kg-bw per day)^[1]

^[1] If infants are fed pumped breastmilk via polycarbonate bottles, this may add to their overall exposure to bisphenol A.
^[2] Based on the average and maximum concentrations of bisphenol A identified in breastmilk by Ye et al. (2006) (see Table 13).
^[3] Based on the average and maximum concentrations of bisphenol A identified in various environmental media (see Tables 16 and 17)

For exposure among children and adults (aged 1+ years), exposure from multiple sources (food packaging of canned food and beverages, migration from polycarbonate drinking bottles and through environmental media) were aggregated and are presented in Table 21.

Table 21. Aggregate estimates of potential exposure for children (aged 1+ years) and adults (µg/kg-bw per day)

^[1] Based on the average and maximum concentrations of bisphenol A identified in canned food (EWG 2007) (see Tables 10 and 11)
^[2] Based on the calculated average concentration of bisphenol A detected in water, 0.24 ppb, that resulted from the room temperature use of polycarbonate bottles (Le et al.2008) (see Table 15).
^[3] Based on the average and maximum concentrations identified in various environmental media (see Tables 16 and 17)

In a limited study, Wilson et al. (2003) reported aggregate exposure estimates for nine children aged 2-5 year olds from concentrations of bisphenolA reported in indoor and ambient air, floor dust, play area soil and solid and liquid food collected at two daycare centres and nine households in North Carolina. Mean exposure was 42.981 ng/kg-bw per day and maximum exposure was 71.124 ng/kg-bw per day. The aggregate exposure was based on the following assumptions: 100% absorption, that a child spends 12 hours a day at the daycare centre, ingests soil and dust at a rate of 100 mg/day (Stanek and Calabrese 1995a,b), and inhales air at a rate of 8.3 m³/day (EFH 2000). In an expanded study by Wilson et al. (2007), 257 children aged 1.5 to 5 years old were surveyed in daycare centres and homes located within North Carolina and Ohio using the approach described above. A 50^th percentile aggregate exposure of 71.4 ng/kg-bw per day and a maximum of 1570 ng/kg-bw per day was reported in North Carolina. A 50th percentile aggregate exposure of 60.8 ng/kg-bw per day and maximum of 775 ng/kg-bw per day was reported in Ohio (an assumption of 50% absorption was made).

Concentrations in biological tissues and fluid have been reported by a number of investigators and are summarized in Vandenburg et al. (2007), Dash et al. (2006) NTP (2007) and Dekant and Volkel (2008). Tissues and fluid analyzed include sem*n, amniotic fluid, placental tissue and serum. Urine is considered the most appropriate body fluid for quantifying exposure to bisphenolA. Several studies have analyzed bisphenolA in urine. Most recently, bisphenolA (free and conjugated) was an analyte of the 2003-2004 National Health and Nutrition Examination Survey (NHANES) .The NHANES results are reported by Calafat et al. (2008) and present results of bisphenolA analysis (free and conjugated) in 2517 individuals aged 6 years and older in the United States. Using these data, Lakind and Naiman (2008) have derived estimates of daily intake of bisphenolA, assuming steady state excretion. Fiftieth and ninety-fifth percentiles for the following subpopulations of the U.S. (6-11 years, 12-19 years, 20-39 years, 40-59 years and 60+ years) have been derived based on standard values for urinary output and adjusted for individual bodyweights. For all individuals, the 50th percentile exposure estimate is 50.5 ng/kg-bw per day (the 95th percentile exposure estimate is 274.2 ng/kg-bw per day). For children aged 6 to 11, the 50th percentile exposure estimate is 67.4 ng/kg-bw per day (the 95th percentile exposure estimate is 310.5 ng/kg-bw per day).

Confidence in the exposure assessment is considered to be moderate. Canadian multimedia data are very limited. Biomonitoring data indicates that exposures may be less than those derived from aggregating exposure estimates based on concentrations in different media, but the absence of biomonitoring data for a key segment of the population (infants) precludes use of biomonitoring data to characterize risk. Additionally, applying precaution, derivation of exposure estimates based on concentrations in media captures exposures in the most exposed individuals. Furthermore, there are uncertainties associated with the use of spot urine samples from one point in time to predict daily exposure estimates.

With respect to metabolism of bisphenol A, it is considered that free bisphenol A, not the glucuronide conjugate, is the biologically active moiety. The route of exposure is significant as it affects the circulating levels of free bisphenol A. Following oral exposure, the glucuronidation of bisphenol A by hepatic uridine diphosphate glucuronyltransferase (UDPGT), appears to be a major in vivo metabolism pathway in rats (specifically the isoform, UGT2B1) (Pottenger et al. 2000; Yokota et al.1999), non-human primates (Kurebayashi et al.2002), and humans (Volkel et al. 2002). The UGT isoform(s) which may be involved in metabolism of bisphenol A in humans is/are not known (NTP 2008). In a recent draft, the European Food Safety Authority (EFSA) acknowledged glucuronidation as the major pathway of bisphenol A metabolism in adult humans and animals while sulfation was suggested as a minor pathway in humans and rats (EFSA 2008). The United States Food and Drug Administration (USFDA) also recognized glucuronidation as a major metabolic pathway in rodents, monkeys and humans, but the possibility of bisphenol A sulfation during metabolism was not sufficiently reviewed (USFDA 2008).

In humans and other non-human primates, bisphenolA administered orally enters first-pass metabolism in the gut wall and the liver and is quickly metabolized to bisphenolA-monoglucuronide, which has no endocrine activity and is rapidly excreted in urine with a half-life of less than 6 hours (Tominaga et al 2006; Volkel et al. 2002). Due to rapid elimination and plasma protein binding of bisphenolA, it is expected that very low concentrations of free bisphenolA would be available for receptor binding in humans (EFSA 2006).

In rats, bisphenol A administered orally also undergoes glucuronidation, but the resulting bisphenol A-glucuronide is excreted from liver into bile. In the gut, the bisphenol A-glucuronide is then cleaved into bisphenol A and aleuronic acid and bisphenol A is reabsorbed into the blood. This enterohepatic recirculation may cause slow elimination of bisphenol A in rodents (Pottenger et al. 2000; Snyder et al. 2000). An overview of toxico*kinetics or pharmaco*kinetics of bisphenol A by the EFSA and USFDA elaborated the similar stages of bisphenol A metabolism in rodents, non-human primates and humans (EFSA 2008; USFDA 2008).

Additionally, the toxico*kinetics and metabolism of bisphenol A in pregnant animals, fetuses or neonates may be different to that of non-pregnant adults. The absorption and distribution of bisphenol A was rapid in maternal F344 rat organs following a single oral exposure to a high dose of 1 g/kg, reaching the fetus by crossing the placenta within 20 minutes of exposure; the fetal bisphenol A concentration exceeded that of maternal blood 40 minutes after exposure (Takahashi and Oishi 2000). Also, in humans, bisphenol A was present in serum and follicular fluid of pregnant women, as well as fetal serum and amniotic fluid, demonstrating placental transfer. Bisphenol A (presumably bound and free) concentrations in amniotic fluid, mid-term, were 5-fold higher as compared to serum and follicular fluid (Ikezuki et al.2002). These data indicate that repeated maternal exposures could lead to an accumulation of fetal circulating levels of free bisphenol A and correspondingly elevated in uteroexposures (Ikezuki et al. 2002; Welshons et al.2006).

Free bisphenol A in human fetal plasma has also been reported to be 0.2 - 9.2 ng/mL (mean 2.9 ng/mL) and the mean bisphenol A plasma levels in the male human fetus was reported as 3.5 ng/mL; significantly lower levels of bisphenol A were reported in female fetal plasma with a mean of 1.7 ng/mL (Schönfelder et al.2002). In women, bisphenol A has been detected in biological fluids/tissue including breastmilk, colostrum and adipose tissue in parts per billion concentrations (Fernandez et al 2007; Kuruto-Niwa et al 2007; Sun et al 2004). A careful evaluation of biologically active or free bisphenol A concentration in various tissues of a sensitive population was previously identified as a research need and has also recently been identified as being of critical importance by the USFDA (2008).

A reduced UGT2B1 activity was seen in pregnant rats as compared to non-pregnant females. The UGT2B1 enzyme was not present in the fetus, its activity was slow in neonates, and reached the adult level of activity within three weeks after birth (Matsumoto et al. 2002). It has been suggested that an increase in glucuronidation capacity during pregnancy may result in negligible fetal exposure to xenobiotics (EFSA 2008). Human data show that glucuronide enzyme activity may be induced during pregnancy as it was evidenced by an increased clearance of orally administered lorazepam, paracetamol and lamotrigine in pregnant women (Minerset al. 1986; Papini et al. 2006; Pennell et al. 2008). It is important to note that based on the limited evidence of selective glucuronidation and considering other factors including individual variation, tissue localization, and substrate-specificity of these enzymes in animals, it cannot be assumed that glucuronidation activity would be increased in pregnant women following exposure to bisphenol A.

Limited activity of glucuronidation enzymes in the human fetal liver has also been demonstrated (Ring et al. 1999). In human liver microsomes, most of the UGT isozymes were not expressed prenatally. A full complement was present at three months after birth; however, levels were substantially reduced compared to the adult (Coughtrie et al. 1988). Studies indicate that, in humans, sulfotransferase (SULT) enzymes are fully functional at birth and it has been suggested that sulfoconjugation of bisphenol A may overcome the absence of or very low glucuronidation activity in newborns (Benedetti and Baltes 2003; EFSA 2008; Richard et al. 2001).

Some in vitro data showed that SULT enzymes may have the ability to metabolize some phenolic or estrogen-like compounds including bisphenol A in cell cultures (Nishiyama et al 2002; Shimizu et al 2002; Suiko et al 2000). This limited data needs to be interpreted carefully as different liver cells were studied under controlled conditions using different protocols. Whether the sulfotransferase activity reported in these studies would be sufficient to efficiently metabolize the bisphenol A in in vivo is not known at this time. A study conducted using post-mortem human (fetal, term and adults) liver microsomes, in general, showed reduced UDPGT activity towards some phenolic compounds, which were structurally similar to bisphenol A, in fetal and term samples as compared to adults. Also, the human fetal liver sulphotransferase activity was found to be less than 3% of that present in non-human primate fetal liver (unpublished work - Leakey). The authors suggested that glucuronic acid or sulphate conjugation pathway may not be efficient in human perinate against phenolic compounds (Leakey et al 1987).

It is therefore not clear, based on the data of Leaky et al. (1987) whether sulfotransferases would act as a major metabolic enzyme system in human neonates following exposure to bisphenol A. Together, the toxico*kinetics and metabolism differences suggest that the developing fetus and neonate may be more sensitive to bisphenol A due to reduced clearance, increased half-life of free bisphenol A (Welshons et al. 2006) and their inability to efficiently glucuronidate bisphenol A (Ringet al. 1999; Coughtrie et al. 1988).

Conversely, some evidence suggests that there were no significant differences between pregnant (at different stages of gestation) and non-pregnant Sprague-Dawley rats in the toxico*kinetics and metabolism of orally administered bisphenol A (10 mg/kg single dose), and fetal bisphenol A elimination half-life appeared to be dependent on maternal bisphenol A elimination (Domoradzki et al. 2003). It was reported that, in developing Sprague-Dawley rats, dose-dependent metabolism of 1 or 10 mg/kg (gavage) bisphenol A to its glucuronide conjugate occurred as early as postnatal day 4 through postnatal day 21 and no significant sex-related differences were noted in the metabolism and pharmaco*kinetics in neonatal to weanling rats (Domoradzkiet al. 2004).

In the absence of fully developed physiologically-based pharmaco*kinetic models to quantitatively characterize absorption, distribution, metabolism and excretion of bisphenol A and its metabolites in test animals and humans via different routes and at different life stages, it is considered appropriate to consider the pregnant woman/fetus and infant as potentially sensitive subpopulations.

Carcinogenicity

A lifetime (103-week) bisphenol A chronic toxicity/carcinogenicity study in male (0, 74 or 148 mg/kg/d) and female (0, 74 or 135 mg/kg per day) F344 rats reported reduced body weight gain in all exposed animals which was correlated with reduced feed consumption. A marginal increase in the incidence of leukemia and testicular interstitial tumors in male rats was considered statistically non-significant due to failure to meet the Bonferoni inequality criterion; the incidence of both tumor types was more common in the high-dose groups. Male and female B6C3F1 mice were administered doses of 0, 120 or 600 mg/kg-bw per day or 0, 650 or 1300 mg/kg-bw per day, respectively. Reduced body weight gain was observed in male and female B6C3F1 mice that received bisphenol A doses of 600 mg/kg per day or 650 or 1300 mg/kg per day respectively for 103 weeks. There was an increase in the combined incidence of leukemia and lymphoma in male mice, which was not considered dose-related. Overall, it was concluded that there was no convincing evidence that dietary exposure to bisphenol A causes carcinogenicity in rats or mice of either sex (NTP 1982). Some debate exists regarding the biological significance of the NTP (1982) findings. Specifically, it has been argued that the marginal increases in leukemias of rats and the combined incidence of lymphomas and leukemias in male mice should be considered to be bisphenol A related. It has been proposed as well that the observed testicular effects should also be considered as some evidence for carcinogenic activity (Huff 2001; 2002). However, others maintain that the 2-year study correctly concluded that the results demonstrate equivocal evidence of carcinogenicity in male rats and no evidence of carcinogenicity in female rats and mice of both genders (Haighton et al. 2002; Munro et al.2002a; 2002b). It was noted that the incidence of malignancies observed in the bisphenol A exposed groups of rats and mice were within the historical control ranges for the strains and species tested (Haighton et al. 2002). In support of the NTP (1982) conclusions, Tyl et al. (2002) observed no increase in cancer incidence in a three-generation assessment of bisphenol A toxicity in male and female Sprague-Dawley rats exposed to doses ranging from 0.001 to 500 mg/kg-bw per day (Tyl et al.2002). The available evidence suggests that bisphenol A is not a carcinogen following adult exposure (Keri et al.2007).

Recent reports have suggested that prenatal and/or neonatal exposure to bisphenol A in rats at doses of 25 or 250 ng/kg-bw (via osmotic pumps implanted into pregnant dams) and ranging from 2.5 µg/kg-bw per day to 1.0 mg/kg-bw per day (administered via maternal subcutaneous injection) during development, may increase susceptibility to neoplastic transformation in the prostate and mammary gland in adult rats (Ho et al. 2006; Murray et al. 2007; Durando et al. 2007; Markey et al.2003, 2005; Munoz-de-Toro et al. 2005). However, the limited evidence is insufficient to demonstrate that early bisphenol A exposure, acting independently, could lead to neoplastic events (Newbold et al. 2007; Prins et al. 2007; Ho et al. 2006; Keri et al. 2007; Ichihara et al. 2003). Further research is needed on the role of early life exposures to bisphenol A, in particular via routes most relevant to human exposure, in the process of carcinogenesis (Keri et al. 2007).

In vitro genotoxicity studies demonstrated a lack of mutagenic activity; however, it was noted that bisphenolA has aneugenic potential and induced morphological transformation and micronuclei formation in the absence of metabolic activation in cultured Syrian hamster embryo cells and Chinese hamster V79 cells, respectively (NTP 2007; ECB 2003; Haighton et al. 2002). In contrast, a large number of standard genotoxicity assays, including the in vivo mammalian dominant lethal assay and the mouse micronucleus assay, provided no evidence of mutagenic or clastogenic potential of bisphenolA (ECB 2003; Haighton et al. 2002). Using the 32 P-postlabeling technique, bisphenolA has been shown to form DNA adducts in vivo in the rat liver; the major adducts were not further characterized and the relevance to human health is unknown. Based on a review of the literature and the lack of significant tumour induction in animal studies, the European Union concluded that "it does not appear that bisphenolA has significant mutagenic potential in vivo" (NTP 2007; ECB 2003).

The aneugenic potential of bisphenolA was further investigated following the observation of a spontaneous increase in oocyte meiotic abnormalities in mice attributed to inadvertent exposure to bisphenolA from damage to caging materials (Hunt et al.2003). The authors subsequently investigated the impact of short-term bisphenolA exposure on meiotic processes during oogenesis. Oral bisphenolA administration to juvenile (20-22days old) female C57BL/6 mice (0, 20, 40, or 100 µg/kg-bw per day) for 6 to 8 days preceding oocyte collection induced a significant increase in chromosome congression failure (1.7%, 5.8%, 7.5% and 10.9% for 0, 20, 40, and 100 µg/kg-bw per day, respectively) and a dose-related increase in meiotic abnormalities (Hunt et al. 2003). Subsequent examination of the effects of bisphenolA on female and male germ cells and somatic cells did not confirm aneugenic potential. There was no evidence of increased hyperploidy or polyploidy, alterations on the timing of meiotic progression or micronuclei induction following bisphenolA exposure via drinking water (7 daily doses of 0.04 mg/kg or for 7 weeks at a concentration of 0.5 mg/L) to C57Bl/6 female mice or via gavage (0, 0.002, 0.02, 0.2 mg/kg) to (102/ElxC3H/E1) F1 male mice (Pacchierotti et al. 2008). Contrary to the induction of aneuploidy, it has been suggested that oocytes from bisphenolA exposed mice respond to disturbances in spindle formation by initiating meiotic arrest (Eichenlaub-Ritter et al. 2008). Thus, the lack of direct evidence of aneugenic induction in vivo and very limited data in vitro suggests that bisphenolA has no significant aneugenic potential (ECB 2003).

Multigeneration reproductive toxicity studies in Sprague-Dawley rats and CD-1 mice were evaluated by various jurisdictions and were considered to be valid (ECB 2008; Willhite et al. 2008; NTP 2007; Nakanishi et al. 2007; EFSA 2006; ECB 2003). Reproductive and developmental endpoints were examined following dietary exposure to bisphenolA in rats (0.001, 0.02, 0.30, 5, 50 and 500 mg/kg-bw per day) and mice (0, 0.003, 0.03, 0.3, 5, 50, 600 mg/kg-bw per day) (Tyl et al. 2002; 2007). NOAELs of 5 mg/kg-bw per day and 50 mg/kg-bw per day were identified for adult systemic effects and reproductive and developmental toxicity in rats and mice, respectively. The NOAEL of 5 mg/kg-bw per day for adult systemic effects was based on significantly reduced body weight and reduced body weight gain, reduced absolute and increased relative organ weights in rats, and an increased incidence of minimal to mild hepatocyte hypertrophy in adult male and female mice. The NOAEL of 50 mg/kg-bw per day for reproductive and developmental toxicity was based on effects on the offspring at the highest doses tested. These effects included reduced weanling body weight, exposure-related changes in absolute and relative organ weights in male and female rats, and hypoplasia of seminiferous tubules and delayed acquisition of preputial separation, as well as a slight increase in the incidence of undescended testes in weanling male mice. No changes of toxicologic significance were seen within the dose range of 0.001-5.0 or 0.003-5.0 mg/kg-bw per day in adult or developing rats and mice respectively (Tyl et al. 2002; 2007). The three-generation reproductive toxicity study by Tyl et al. (2002) in rats did not include an estrogenic reference control against which the negative results could be evaluated. However, the most recent multigeneration report in mice incorporated a dietary 17ß-estradiol reference group (Tylet al. 2007). Developmental toxicity in CD-1 mice and CD rats has been demonstrated to occur following prenatal exposure during major periods of organogenesis (gestational days 6-15), through gavage, at high doses of bisphenolA. Measures of developmental toxicity included percentage resorptions, number of live fetuses per ratio, sex ratio, average fetal body weight per litter and percentage of malformed fetuses.No developmental effects were observed following administration of 640 mg/kg-bw per day (highest dose reported due to severe maternal toxicity at 1280 mg/kg-bw per day) or 1000 mg/kg-bw per day in rats and mice, respectively. BisphenolA was not teratogenic at any of the administered doses (Morrissey et al. 1987).

Recent studies have shown that chemicals that alter endocrine function, including bisphenolA, can alter the cell function at levels below NOAELs identified in these standard toxicology studies (reviewed in Welshons et al. 2003; vom Saal and Welshons 2006; Welshons et al. 2006). These studies have been described as 'low dose studies'. Effects observed are not necessarily manifested following a linear dose-response relationship and, in several instances, have been found to follow a non-monotonic response curve. BisphenolA has been considered as a weak environmental estrogen, based on traditional bioassays, as it binds to the estrogen receptor alpha and beta (ERa and ERß) (Gouldet al. 1998; Kuiper et al. 1998; Pennie et al. 1998) with affinity, which is about 10000 to 100000-fold weaker than that of 17ß-estradiol (Vandenberg et al. 2007). A mode-of-action analysis is beyond the scope of this screening assessment; however, exposure to low doses of bisphenolA has been demonstrated to produce disruptive effects in endocrine organs including the androgen or estrogen responsive tissues, immune system, thyroid hormone function, and developing nervous system (reviewed in Richter et al. 2007; Vandenberg et al. 2007; Wetherill et al. 2007). Evidence suggests that bisphenolA can stimulate the aforementioned cellular responses at very low concentrations through genomic (nuclear estrogen receptor) or non-genomic (membrane-associated or intracellular transduction) mechanisms, with effects on cellular function at doses as low as 1 pM or 0.23 pg/ml (Anway et al. 2005, 2006; Wozniak et al.2005, Zoeller et al. 2005) (reviewed in Wetherill et al. 2007). Specifically, it has been demonstrated that some of the effects of bisphenolA may be mediated through the cell surface estrogen receptor (GPR30) being equipotent with 17ß-estradiol and diethylstilbestrol (Alonso-Magdelena et al. 2005).

There are many studies in the peer-reviewed literature that describe effects in experimental animals resulting from exposure to bisphenolA administered at doses below the NOAELS established in standard toxicological studies for systemic and reproductive/developmental endpoints. Effects observed include changes in prostate growth and development, mammary gland organization, protein induction in the uterus, organization of sexually dimorphic circuits in the hypothalamus, onset of estrus cyclicity, early puberty, body weight, genital malformations and others (reviewed in Richter et al. 2007; Wetherill et al. 2007). Thorough overviews of these studies have been summarized in several recent reviews (NTP 2007; Richter et al. 2007; Willhite et al. 2008). For purposes of the screening assessment, the primary focus will be placed on identification of alterations considered potentially 'adverse' for the purposes of human health risk assessment.

It should be noted that, due to the finding by Taylor et al. (2008) that the differences between oral and subcutaneous injection toxico*kinetics in adults are not likely conserved in neonates, those studies delivering bisphenol A during fetal development and/or early postnatal periods via the subcutaneous (sc) injection route were evaluated for the current assessment (Kato et al. 2003; 2006; Ho et al. 2006; Honmaet al. 2002; Nakamura et al. 2006; Suzuki et al. 2002). The neonatal data, using sc injection permits the evaluation of possible effects of ecposure during the early stages of development. While informative, these studies were not the basis for hazard characterization of bisphenol A for human health.

A number of reproductive and developmental studies provide evidence that in utero and/or neonatal exposure to doses below the established reproductive/developmental NOAEL of 50 mg/kg-bw per day may result in altered reproductive function and developmental endpoints. Prenatal exposure from gestational day 11 to 17, of CF-1 mice to 0, 2 or 20 µg/kg-bw perday bisphenolA administered orally, resulted in an increase in prostate size in male offspring beginning at 2 µg/kg-bw per day (Nagel et al. 1997), reduced epididymal weight and decreased daily sperm production at doses of 2 and 20 µg/kg -bw per day, respectively (vom Saal et al. 1998). It also resulted in increased postnatal growth in both males and females and early onset of sexual maturation in female offspring, observed following exposure to 2.4 µg/kg-bw per day (Howdeshell et al. 1999). Additionally, CD-1 mice exposed to 10 µg/kg-bw per day, from gestational day 14 to 18, displayed disrupted development of the prostate as shown by an increase in the number and size of the prostate ducts and an increase in the overall volume of the prostate; a malformation of the urethra was also observed (Timmset al. 2005). Honma et al. (2002), using subcutaneous injections of 0, 2 or 20 µg/kg-bw per day bisphenolA to ICR/Jcl mice from gestational day 11 to 17, demonstrated that these low doses administered via a non-oral route produced common alterations. Specifically, they produced altered postnatal growth and early vagin*l opening at 20 µg/kg-bw per day. However, female reproductive function was not affected at the first breeding. Furthermore, oral exposures of pregnant and nursing Long-Evans rat dams from gestational day 12 to postnatal day 21 to 2.4 µg/kg-bw per day altered plasma luteinizing hormone levels and decreased testosterone levels in adulthood. This suggests that bisphenolA-dependent effects on androgen biosynthesis by adult leydig cells occurred as a result of targeting a sensitive period during perinatal development. Chronic exposure from weaning at postnatal day 21 to adulthood, representing adolescent exposure, also suppressed androgen biosynthesis by adult leydig cells at doses as low as 2.4 µg/kg-bw per day bisphenolA (Akingbemiet al. 2004). A LOAEL for the reproductive effects of bisphenolA was observed to be 2 µg/kg-bw per day in the aforementioned subset of studies. A non-monotonic (inverted-U) dose-response was observed in several instances.

Conflicting results have been described in various additional reports. Oral bisphenolA exposure to 0.2, 2, 20 or 200 µg/kg-bw per day from gestational day 11 to 17 did not produce reproductive toxicity in studies assessing outcomes in CF-1 mice (Cagen et al. 1999a; Ashby et al. 1999). Nor was there an effect on sperm density or male reproductive organ development following bisphenolA exposure of estrogen-sensitive C57Bl/6N mice during various life stages (Nagao et al. 2002). Notably, in the studies by Cagen et al. (1999a) and Ashby et al. (1999), the responses of the positive control did not differ from either the negative control or bisphenolA. There were no treatment-related effects on progeny following maternal drinking water exposure to 0.01, 0.1, 1.0 or 10 ppm bisphenolA among female Wistar albino rats. The authors conclude that bisphenolA should not be considered a selective reproductive or developmental toxicant (Cagen et al. 1999b). A 2-generation reproduction study using Crj: CD(SD) IGS rats administered bisphenolA orally at doses between 0.2 and 200 µg/kg-bw per day also found no significant adverse treatment-related changes in the multiple reproductive and developmental endpoints assessed (Ema et al. 2001); this study also lacked a positive control for interpretation of negative data. Most recently, assessing the impact of bisphenolA exposure in the male Long Evans hooded rat following oral dosing with 2, 20 or 200 µg/kg-bw per day from gestational day 7 to postnatal 18, Howdeshell et al.(2007) found no significant effect on any male reproductive endpoint investigated. These included male reproductive development, reproductive hormone levels or sperm production.

It is recognized that a number of experimental variables may account for divergent results in response to bisphenolA exposure at low doses: specifically, species and strain differences; variable responses depending on the tissues and/or endpoint(s) assessed; variability in feed with respect to amounts of estrogenic contaminants; inappropriate use or lack of positive controls; as well as the consideration of exposure-related effects that present in a non-monotonic dose-response curve (vom Saal and Hughes 2005; vom Saal et al. 2005a; Richter et al. 2007). Additionally, the dosing period or time of exposure with respect to critical developmental windows and the dosing regimen are important considerations, particularly when assessing alterations resulting from developmental exposures. Furthermore, the nature of the effects are such that it is difficult to characterize the degree to which they would be considered potentially "adverse" and, hence, form the basis of a human health risk assessment.

Recently, expert panel assessments of the current status of bisphenol A toxicity have identified neurodevelopmental and behavioural alterations as an endpoint of concern. The U.S. National Toxicology Program (NTP) Center for the Evaluation of Risks to Human Reproduction (CERHR) (NTP 2007) concluded there was "some" concern that pregnant women, fetuses, infants and children may be susceptible to bisphenol A disruption for neural and behavioural endpoints. "Some" concern for these effects was based on the finding that the bisphenol A literature regarding neural and behavioural effects consistently produced more "positive" studies, as compared to the reproductive and developmental studies, although the endpoints assessed differed between investigators. The level of concern was expressed relative to estimates of exposure for the general population in the United States (NTP 2007). Recently, the NTP finalized its evaluation of bisphenol A and confirmed some concerns for neurobehavior effects in fetuses, infants, and children at the current level of human exposure (NTP 2008) Additionally, based on existing evidence, the Chapel Hill bisphenol A expert panel states with confidence "that low doses of bisphenol A during development have persistent effects on brain structure, function and behaviour in rats and mice" (vom Saal et al.2007). Confidence was defined as research areas where the expert panel concluded that there was an agreement of findings from multiple papers from multiple labs. Given these expert panel conclusions, as part of the screening assessment it was considered appropriate to further consider the weight of evidence associated with neurodevelopmental and behavioural studies investigating effects of low dose bisphenol A exposure. These studies are presented in Appendix D and are discussed in further detail below.

Neural development is a complex process beginning in early embryonic development and continuing through different life stages. Precise functioning of the nervous system is critical for all mental, sensory and motor activities, and regulates homeostasis through interaction with the endocrine system (Clancy et al. 2001). An array of behavioural tests have been conducted to evaluate the effect of low dose bisphenolA exposure. Some tests were among those recommended by the recently adopted OECD Developmental Neurotoxicity Study Guideline # 426 (OECD 2007). A number of developmental neurotoxicity endpoints were examined, including locomotory and exploratory activity, learning, social, sexual and maternal behaviour, as well as brain morphology, with a range of exposures from 0.2 to 400 000 µg/kg-bw per day [reviewed in NTP 2007; Richter et al. 2007; Willhite et al.2008; ECB 2008].

Evaluation of several pivotal studies on neural development and behaviour effects of bisphenolA in rodents showed that in utero, perinatal or postnatal exposure to bisphenolA, at levels below the established NOAELs of 50 mg/kg-bw per day for reproductive/developmental effects and 5 mg/kg-bw per day for systemic toxicity in rats and mice reported by Tyl et al.(2002; 2007), which were used in the most recent derivations of reference concentrations of 16-50 µg/kg-bw per day (EFSA 2006; Willhite et al. 2007), may produce some effects on neurobehavioral development in fetal or early postnatal life in rodents. The lowest observed effect in the neurodevelopmental literature was 0.02 µg/kg-bw per day based on changes in estrogen receptor mRNA expression, however, no effects were observed at 2 µg/kg-bw per day; this is suggestive of an inverted U shaped (non-monotonic) dose response phenomenon (Nishizawa et al.2003; 2005a; 2005b).

While this inverted non-monotonic dose-response relationship is considered to be mediated by both positive and negative feedback effects on estrogenic receptors (Welshons et al. 2003; 2006) or related to physiological responses to estrogenic activity of bisphenolA via the feedback mechanism of the endocrine system of the animals, the issue of non-monotonic dose response curves remains an area of ongoing research with efforts directed at the replication of perturbations at levels of exposure below those demonstrated to have no effects. Furthermore, it has been noted that the shape of the dose-response curves for effects of various estrogenic compounds differs depending on the endpoint assessed and the dosing regime applied (Melnick et al. 2002). As such, the significance of the LOEL of 0.02 µg/kg-bw per day for bisphenolA to human health risk assessment, however, is difficult to ascertain, particularly in light of the fact that it was the only dose level tested below the NOEL of 2 µg/kg-bw per day and some of the effects were not consistently observed throughout the study periods. In addition, since the studies (exposure period and endpoints measured) were carried out only up to certain embryo development stages, the relevance to the postnatal developmental period for newborns and infants is unknown. An additional study that examined substance-induced changes in brain receptors noted a fluctuating increase in the number of neurons expressing ERa and ERß after oral administration of 2 µg/kg-bw per day bisphenolA to ICR mice from gestational day 11 to 17. Increases were observed at 5 and 13 weeks, but not at 9 weeks (Kawai et al. 2007). Oral administration of Sprague Dawley rats with 40 µg/kg-bw per day bisphenolA from postnatal day 23-30, thereby targeting the period of juvenile development, resulted in a transient and gender-specific increase in ERa levels in the ventromedial nucleus of the hypothalamus. Changes were noted at postnatal day 37, but not at postnatal day 90, in females only (Ceccarelli et al. 2007). Although these effects may be considered biomarkers of exposure to bisphenolA, and potential precursor events to adverse effects, the biological relevance of these effects for purposes of human health risk assessment is not known.

Volumetric changes of sexually dimorphic regions of the brain have been used as a biomarker for developmental disruption by endocrine-active substances [reviewed in Patisaul and Polston 2008]. Following subcutaneous injection of 100 mg/kg-bw per day to Sprague-Dawley rat pups on postnatal day 1 and 2, no bisphenolA-induced changes in the volumes of the sexually dimorphic nucleus of the pre-optic area (SDN-POA) or the anterioventral periventricular nucleus of the hypothalamus was observed (Patisaul et al. 2006; 2007). Although bisphenolA was shown to disrupt neuronal phenotype by demasculinization of tyrosine hydroxylase (TH) immunoreactive cells and defeminization of ERa/TH double labelled cells in male and female rats, respectively, no functional changes were observed by examination of gonadotropin-releasing hormone (GnRH) neurons (Patisaul et al. 2006; 2007). Thus, neuroanatomical and expression changes detected by immunohistochemistry may not always translate into disruption of neuronal function [reviewed in Patisaul and Polston 2008]. Examination of GnRH neurons as an idex of feminization of the AVPV is an indirect means to determine bisphenolA-induced alterations in this sexually dimorphic region of the brain. Assessing the ability of exposed females to generate an LH surge would have been more informative regarding AVPV feminization. Oral administration of 12 - 60 mg/kg- bw per day bisphenolA to male pups from 5 days to 3 weeks of age led to reduced tyrosine hydroxylase immunoreactivity in the midbrain at 7 weeks of age. This occurred concomitantly with an increase in TUNEL-positive cells in the substantia nigra, possibly reflecting neurodegeneration, and reduced mesencephalic dopamine transporter expression (Ishido et al. 2007). Oral exposure to doses of bisphenolA ranging from 3.2 to 320 mg/kg-bw per day also failed to induce volumetric changes to the SDN-POA in female Sprague Dawley rats exposed during gestation and lactation (Kwon 2000). A reversal of gender differences in the volume of the locus coeruleus, a nucleus in the brainstem involved in stress and panic responses, was observed following administration of drinking water at 0.03 and 1.5 mg/kg-bw per day, respectively (Kubo et al. 2001; 2003). A non-monotonic dose response was reported as the extent of reversal was greater at 0.03 mg/kg-bw per day than at 0.3 mg/kg-bw per day (Kubo et al. 2003). Similarly, the normal gender difference in numbers of corticotrophin releasing hormone (CRH) neurons was disrupted in the bed nucleus of the stria terminalis and preoptic area in Wistar rats following gestational and lactational exposure in drinking water to 2.5 mg/kg-bw per day bisphenolA (Funabashi et al. 2004). hom*ologous sexually dimorphic regions exist in the human brain (Balthazart and Ball, 2007; MacLusky and Naftolin, 1981), however, the relevance to human health risk assessment is uncertain.

Brain receptor expression and immunohistochemical investigations are key for providing mechanistic information, and are useful biomarkers, but relevance to human health risk assessment is uncertain. This is consistent with interpretations in the draft updated Risk Assessment of bisphenolA by the European Union (ECB 2008).

Behavioural testing was used to characterize the effect of bisphenolA exposure on general locomotory and exploratory activity, anxiety and learning and memory in rodents. At 10 µg/kg-bw per day, CD-1 mice exposed via drinking water during perinatal development showed altered maternal behaviour. Interestingly, the differences in nursing behaviour were observed following prenatal or adult exposure, but not following continuous exposure (Palanza et al. 2002). Reduced D-amphetamine-related reinforcing effects in females (Laviolaet al. 2005) and decreased or eliminated sex-related behavioural differences (Gioiosa et al. 2007), were also demonstrated in CD-1 mice exposed orally during gestation to 10 µg/kg-bw per day suggesting that exposure to bisphenolA at doses which are well below established NOAEL of 50 mg/kg-bw per day for reproductive/developmental effects may lead to organizational effects during critical periods of brain development in rodents. However, these studies, though, were limited to a single exposure level precluding the evaluation of dose-response. The lowest dose leading to bisphenolA-induced organizational effects in the brain was 10 µg/kg-bw per day in CD-1 mice. It should be recognized that studies were conducted using the same outbred strain of mice (CD-1), the same experimental dosing protocol (single dose, no positive control) and by the same group of researchers in one research institute.

Effects of gestational exposure to 40 µ/kg-bw per day bisphenolA have been noted in rats, including statistically significant gender-specific changes in sexual performance (Farabollini et al. 2002), effects on active and passive maternal behaviour of treated rats (Della Seta et al.2005), marginally diminished intergroup differences in pain response (Aloisi et al. 2002), altered novelty seeking and impulsive behaviour in females and males, respectively (Adrianiet al. 2003), as well as altered play behaviour in female rats (Porrini et al. 2005). In addition, in Sprague-Dawley rats, pubertal exposure to oral doses of 40 µg/kg-bw per day altered behaviour of male rats (Della Seta et al. 2006). These studies, again conducted by a common group of researchers, provide convincing evidence of bisphenolA-induced effects at 40 µg/kg-bw per day and illustrate the implementation of a comprehensive approach. Replication of the aforementioned results by independent researchers is needed.

The effect of bisphenolA exposure on open field behaviour (i.e., an assessment of general locomotory and exploratory activity, grooming and anxiety) (ECB 2008), was not consistent among the available studies. Kubo et al. (2001;2003) administered bisphenolA to Wistar rats during gestation and lactation at 0.03 or 0.3 mg/kg-bw per day in drinking water and noted slightly reduced anxiety-related behaviour in male rats. This behaviour was somewhat increased in females. Rearing behaviour, indicative of exploratory behaviour, observed by Fujimoto et al. (2006), also reflected opposing effects in male and female Wistar rats following 15 µg/kg-bw per day bisphenolA from gestational day 13 to birth. Increased duration of rearing was observed in males and was slightly reduced in females, suggesting that the sex difference in rearing activity had been abolished. No effects on open field behaviour were reported in additional studies assessing the effects of oral bisphenolA exposure during the prenatal and neonatal periods with doses ranging from 40 µg/kg-bw per day to 400 mg/kg-bw per day (Adriani et al. 2003; Negishi et al. 2003; 2004). Notably, the 2-generation study in Sprague Dawley rats by Ema et al. (2001), also reported no effects of oral bisphenolA exposure (0.2, 2, 20, 200 µg/kg-bw per day) on the behavioural endpoints tested in F1 Crj: CD(SD)IGS rats (in open field and water-filled T-maze). It was acknowledged in the report that differences in results to comparative studies may have been due to possible discrepancies in experimental conditions (Ema et al. 2001). However, this large study, though, did not provide data from behavioural measures preventing assessment of the relationship between exposed and control groups.

Some studies have investigated the effect of prenatal and neonatal bisphenolA exposure on anxiety-related behaviour using the hole board test, elevated maze test, forced swimming test and/or light-dark preference chamber (Farabollini, 1999; Negishiet al. 2004; Fujimoto et al. 2006; Ryan and Vanderbergh 2006). Farabollini et al. (1999) found that oral administration of 400 µg/kg-bw per day bisphenolA from gestational day 14 to postnatal day 6 increased anxiety-related behaviour in male and female rats using the hole board test. However, decreased anxiety-related behaviour was noted in males using the elevated plus maze. No effects were evident following administration of 40 µg/kg-bw per day. Ryan and Vanderbergh (2006) reported increased anxiety in females (males not tested) after administration of 200 µg/kg-bw per day bisphenolA from gestational day 3 to postnatal day 21, but not after 2 µg/kg-bw per day, in the light-dark preference chamber. No differences in spatial memory tests were observed. No changes were observed in additional studies assessing the same endpoints. Specifically, applying the elevated plus-maze test in rats, Fujimoto et al. (2006) observed no anxiety-related effects of bisphenolA following drinking water exposure to 15 µg/kg-bw per day from gestational day 13 to birth. Male and female offspring were tested. Negishi et al. (2004) also found no effects on female offspring exposed to 100 µg/kg-bw per day bisphenolA from gestational day 3 to postnatal day 20 (males not tested). Together, these findings illustrate the variable responses observed across gender and species in anxiety-related behavioural endpoints following a range of bisphenolA exposures.

Further research into the effects of bisphenolA exposure focused on disturbances in cognitive function. Of those studies examining bisphenolA-induced effects on learning and memory, no effects were observed by water-filled T-maze following gestational and lactational exposure to 0.2, 2, 20, 200 µg/kg-bw per day bisphenolA (Ema et al. 2001) (2-generation study in Sprague Dawley rats). Nor were effects observed using passive avoidance testing in Wistar rats after late gestational exposure via drinking water (15 µg/kg-bw per day) (Fujimoto et al.2006). Interestingly, in the same study, one week of bisphenolA exposure (gestational day 13 to postnatal day 0) altered the normal sex differences in struggling behaviour in the forced swimming test observed in control rats. Exposure increased the immobility of male rats in the forced swimming test, which is indicative of depression-like behaviour called behavioural despair. Male rats appeared to be more sensitive to the effects of bisphenolA on depressive response (Fujimoto et al.2006). Kubo et al. (2001), assessing the impact in rats of 1.5 mg/kg-bw per day bisphenolA, also administered via drinking water, found that the sexually dimorphic pattern of avoidance behaviour was abolished in exposed animals. Males showed enhanced avoidance memory while that of females was reduced. The authors therefore suggest that this exposureresults in demasculinization of males with a corresponding defeminization of females. Using similar tests, Negishi et al. (2003) found no consistent effects on active avoidance following exposure to 4, 40 or 400 mg/kg-bw per day bisphenolA from gestational day 10 to postnatal day 20. Responses were variable and did not follow a dose-related pattern. Further assessment of male F344 rats exposed to 100 µg/kg-bw per day from gestational day 3 to postnatal day 20 showed fewer avoidance responses in trials 1 to 3 of 4 in the active avoidance;. However, no differences were observed between bisphenolA-exposed and control groups was observed in passive avoidance (Negishiet al. 2004). Carr et al. (2003) exposed F344 rats to 100 or 250 µg/kg-bw per day bisphenolA from postnatal day1 to 14 and reported that the normal gender-dependent pattern of maze performance acquisition in the Morris water maze test was disrupted after the low dose only. Exposure to the higher dose significantly reduced spatial information retention in females.

Additionally, a number of studies have been conducted to investigate bisphenolA-induced behavioural alterations in response to pharmacological challenge; specifically, D-amphetamine (Laviola et al. 2005; Adriani et al. 2003), methamphetamine- (Suzuki et al. 2003) or morphine-induced (Mizuo et al. 2004a; 2004b; Narita et al. 2006; 2007) behavioural changes. These studies, using standard testing paradigms, provide evidence for altered stimulated responses following bisphenolA administration and are outlined in Appendix D. Prenatal exposure of CD-1 mice from gestational day 11 to 18 with 10 µg/kg-bw per day bisphenolA eliminated D-amphetamine-induced conditioned place preference in females suggesting exposure-related impairment of brain reward pathways; males were unaffected (Laviola et al. 2005). In rats, the D-amphetamine-induced increment activity in the open field test was significantly reduced in males exposed to 40 µg/kg-bw per day throughout gestation and lactation; in this case, females were unaffected (Adriani et al. 2003). Pharmacological challenge studies are suggestive of potential organizational alterations of the neural system following perinatal bisphenolA exposure. The details of additional studies are not elaborated on as the doses that resulted in significant effects on the neurochemical systems were above the established NOAEL of 50 mg/kg-bw per day for reproductive/developmental effects; altered behaviour was observed following administration of 250 or 400 mg/kg bw per day.

Several organizations have recently evaluated the developmental neurotoxicity dataset (ECB 2008; EFSA 2006; NTP 2007; NTP 2008; USFDA 2008; Willhite et al. 2008). Specifically, the EU in the draft Updated Risk Assessment of bisphenol A concluded that "overall, taking together the low confidence in the reliability of the developmental neurotoxicity studies and the lack of consistency in the results of behavioural testing, no conclusions can be drawn from these studies" (ECB 2008). Other jurisdictions concluded that the scientific evidence cited supports the conclusion of some concern for neural and behavioural effects of bisphenol A in potentially vulnerable populations (NTP 2007; NTP 2008).

It was considered appropriate to characterize the weight of evidence supporting neurobehavioral effects in rodents following low-dose exposures to bisphenol A from the following perspectives: rigour, power, corroboration/consistency, biological plausibility/coherence.

The rigour of the neurobehavioral dataset was assessed based on use of suitable test methodology, endpoints monitored, and analysis used to report results. A majority of the studies assessing the effect of bisphenolA on neurobehavioral endpoints used acceptable methodology. All studies were published in the peer-reviewed literature. One study (Ema et al, 2001) declared compliance with good laboratory practices (GLP). However, the rigor of several studies was limited by small sample size, incomplete reporting and inadequate statistical analysis. Additionally, only a few studies reported use of concurrent positive controls. There were few common endpoints across studies and it was difficult to evaluate whether related endpoints that are indicative of anxiety, hyperactivity, avoidance, and so forth, respond in a similar manner. As such, the overall rigour of the neurobehavioral dataset is considered limited.

The statistical power of the studies comprising the dataset was considered. Ema et al. (2001) conducted a 2-generation study in rats using the largest number of animals per exposure group. This study may have had sufficient power to detect subtle changes in the endpoints evaluated, however, no effects were observed between exposed and control groups. The majority of other studies used fewer animals per exposure group which may not be sufficient to detect the neurobehavioral alterations associated with exposure to bisphenol A. Diverse study designs and lack of reporting of data, in many cases where similar study designs were followed, precluded meta-analysis methodology. As such, the overall power of the neurobehavioral dataset is considered limited.

With respect to corroboration/consistency (i.e, the replication of findings among similar studies and observation of similar effects under relevant conditions), the limited data preclude characterization of consistency with regard to the qualitative direction of neurobehavioral changes regardless of the statistical significance of reported alterations. Furthermore, apparent inconsistencies in results may be a function of design and/or the behavioural bioassays used to evaluate the potential developmental alterations (e.g., different tests of the same underlying functions, age of testing). For instance, inconsistencies in the effects of bisphenolA on anxiety-related behaviours should consider that holeboard, light-dark test (emergence test) and elevated-plus mazes all provide measures of anxiety but are also differentially sensitive to other concurrent neurobehavioural effects such as motor effect. The same applies to active and passive avoidance tests (active avoidance is more susceptible to motor effects). Thus, the apparent absence of a consistent pattern of effects may indicate that other factors (such as motor effects) may play a role in the discrepant results. As such, a lack of consistent effects when comparing these measures does not indicate that bisphenolA does not affect anxiety per se. In addition, not only does the comparison between hole board and elevated plus maze appear to include differences in the actual test, but the doses and dosing regimes used in these tests were different: doses of 15, 100 or 400 µg/kg per day bisphenolA with exposure periods from gestational day 13 to birth, gestational day 3 to postnatal day 21 or gestational day 14 to postnatal day 6, respectively. Variations in the dosing period, especially perinatal dosing, can itself have a substantial impact on subsequent effects. As such, there are sufficient methodological differences between these studies to preclude drawing firm conclusions regarding lack of effects of bisphenolA. A consistent finding among studies was that exposure during gestation and/or during early postnatal life consistently represented periods of sensitivity to bisphenolA-induced alterations. Additionally, a consistency in the reported effects regarding alterations of sexually dimorphic behaviours appears to be emerging in rodents. From the current database, it seems likely that the more stongly sexually dimorphic traits are more likely to be disupted by bisphenolA, and moreover, are those most likely to be detected given the measurements used. The degree of corroboration demonstrated in the dataset is considered limited.

Finally, biological plausibility is an important consideration when considering the weight of evidence. Biological plausibility is informed by investigations of mechanism(s) of action. BisphenolA produces estrogenic effects via interactions with estrogen receptors (ER). The estrogen receptor family includes the classic ERa, ERß, and their splice-variants, extranuclear estrogen receptors associated with cell membranes and neuronal synapses (Woolley et al. 2007). BisphenolA was found to have limited binding affinity with ERa, but it showed a nearly tenfold higher affinity for ERß. Specifically, the relative binding affinity (RBA) of bisphenolA for ERa and ERß was 0.05 and 0.33 respectively, whereas RBA for 17b-estradiol was 100 (Kuiper et al., 1997). Strikingly, these chemicals are equally potent at inducing effects through cell membrane ER (Quesada et al.2002). It is possible that bisphenolA preferentially affects ERß-containing tissues including the ovary, cardiovascular system and brain (Harris 2007) and those with cell membrane receptors. Importantly, estrogen is a key regulator of organizational and activational processes in the brain of both males and females. Note that the relative contribution of estrogen and androgen to sexually dimorphic behaviours has been demonstrated to differ between species. Specifically, it is generally held that the estrogen-signalling pathway predominates in rodents, while the androgen-signalling pathway plays a more predominate role in humans and non-human primates (Li et al. 2008). Thus, the relative significance of the estrogenic mode of action of bisphenol A in rodents with respect to human relevance requires further investigation.

Typically, demonstration of a dose-response relationship supports biological plausibility. The dose-response relationship following exposure to bisphenolA at doses below the established NOAEL of 50 mg/kg-bw per day for reproductive/developmental effects is complex and not yet completely understood. This is in large part due to a lack of studies consistently assessing multiple endpoints following administration of 3 or more incremental doses of bisphenolA. Regarding the neurobehavioral dataset, a very limited number of studies assessed more than one dose. Therefore, a dose-response relationship could not be established. An attempt was made to compare and evaluate bisphenolA-related changes at similar doses between similar experiments. However, this was not possible due to a number of the aforementioned design limitations. In a few studies which included multiple doses, bisphenolA-related effects were seen at some doses, but not at others (Kubo et al. 2003; Narita et al. 2006; Tando et al. 2007). Further characterization of the dose-response and examination of the significance and relevance of non-linear dose-response relationships are needed for a comprehensive understanding of neurodevelopmental effects associated with bisphenolA exposures in the µg/kg-bw per day range. The weight ofevidence for biological plausibility is considered limited.

Overall, taking into consideration rigour, power, corroboration/consistency, and biological plausibility/coherence, the weight-of-evidence supporting neurobehavioral effects in rodents following exposures to bisphenol A at exposures below established NOAELs for reproductive/developmental toxicity is limited.

An emerging trend from the developmental neurotoxicity studies is that prenatal and/or neonatal bisphenol A exposure may cause organizational effects in the brain leading to disruptions in the sexually dimorphic behaviours in mice and rats. Although the mode of action, and thus biological plausibility, of the reported effects of bisphenol A is not clear, it is not deemed necessary for risk assessment if the evidence indicates an adverse effect (Liet al. 2008). Detailed mechanistic information can reduce uncertainty regarding the relevance of rodent findings to humans, but is not crucial for general toxicity screening. Thus, as highlighted by the NTP (2007), it is not clear whether the reported changes constitute critical adverse toxicological responses, however, the uncertainties associated with the results from neurodevelopmental and behavioural testing warrant concern with respect to human health, in particular during certain sensitive life stages.

Epidemiology studies have investigated cohorts exposed to bisphenol A and these are summarized in Vandenberg et al.(2007). Although concentrations of bisphenol A in blood have been associated with a variety of conditions in women including obesity, endometrial hyperplasia, recurrent miscarriages, and polycystic ovarian syndrome, as well as elevated androgen hormone levels, these studies have many limitations including small sample size, potential exposure misclassification, and lack of control of potential confounding factors. These studies did not investigate neurodevelopmental and behavioural effects in the potentially sensitive subpopulations (the pregnant woman/fetus and infant).

Confidence in the toxicological database, overall, is considered moderate. Data for acute toxicity, repeated dose toxicity, genotoxicity, carcinogenicity and reproductive and developmental toxicity are available. Although there is an extensive database from which to evaluate the effects of bisphenolA exposure, inconsistencies and experimental limitations exist among studies evaluating reproductive and developmental endpoints. Additionally, several of the studies administered a single dose precluding characterization of dose response.

Consistent with the assessment of the European Commission (ECB 2003), which classified bisphenolA as a Category 3 reproductive toxicant, that is a substance which causes concern for human fertility based on sufficient evidence of reproductive toxicity in experimental animals, a critical effect for characterization of risk to human health is reproductive and developmental toxicity.

Exposure estimates for the general population of Canada range from 0.08 µg/kg-bw per day to 4.30 µg/kg-bw per day. Specific exposure estimates for the most highly exposed subpopulation (i.e., infants) range from an average of 0.27 µg/kg-bw per day (maximum 1.75 µg/kg-bw per day) for infants aged 12 to 18 months to an average of 0.50 µg/kg-bw per day (maximum 4.30 µg/kg-bw per day) for infants aged 0 to 1 month. The NOAELs from the multigeneration reproductive toxicity studies in Sprague- Dawley rats and CD-1 mice of 5 mg/kg-bw per day for systemic effects (reduced body weight gain in rats and minimal to mild hepatocyte hypertrophy in adult male and female mice) and 50 mg/kg-bw per day for reproductive and developmental toxicity (Tyl et al. 2002; 2007) are considered an appropriate departure point for characterizing risk to human health from exposure to bisphenol A. Based on the NOAEL of 5 mg/kg-bw per day, margins of exposure for infants exposed to bisphenol A from consumption of infant formula, use of polycarbonate bottles and via environmental media range (Tables 19a and 19b) from 1160 to 18 520. For infants 0-7 months, exposed via the consumption of breastmilk and environmental media (Table 20), margins of exposure to the NOAEL of 5 mg/kg-bw per day range from 3940 to 25 000. For other segments of the population, exposed via the consumption of canned food, use of polycarbonate bottles and environmental media (Table 21), margins of exposure to the NOAEL of 5 mg/kg-bw per day range from 2200 to 62 500. Margins of exposure to the NOAEL of 50 mg/kg-bw per day would be 10-fold higher than those presented above. Although these margins are considered adequate to address interspecies and intraspecies differences with respect to the adverse effects on which the NOAELs are based, the dataset of neurodevelopmental and behavioural studies, though limited when subject to a weight of evidence analysis, is suggestive of potential effects at doses well below these NOAELs.

Specifically, a small number of studies in CD-1 mice reported alterations in behaviour following oral administration of 10 µg/kg-bw per day. These include altered maternal behaviour in dams exposed either as fetuses or during adulthood and decreased or eliminated sex-related behavioural differences following gestational or gestational and lactational exposures (Palanzaet al. 2002; Laviola et al. 2005; Gioiosa et al. 2007). At 40 µg/kg-bw per day, a small number of studies in rats have reported behavioural effects including gender-specific changes in sexual performance, effects on active and passive maternal behaviour, and altered novelty seeking and impulsive behaviour in both sexes in adults whose mothers were administered bisphenolA during gestation and lactation (Farabollini et al. 2002; Della Seta et al. 2005; Adriani et al. 2003). Further supporting the effect of bisphenolA exposure on altered behaviour, oral administration of 100 µg/kg-bw per day in rats during gestation and lactation reduced the number of correct avoidance responses (Negishi et al. 2004), while early postnatal exposure disrupted the normal gender-dependent pattern of maze performance acquisition in the Morris water maze (Carr et al. 2003).

While collectively these studies provide evidence that exposure to bisphenolA during gestation and early postnatal life may be affecting neural development and some aspects of behaviour in rodents, the overall weight of evidence was considered limited from the perspective of rigour (e.g., study design limitations such as conduct of behavioural assessments at a single time point); power (e.g., limited number of animals per test group), corroboration/consistency (limited consistency of studies) and biological plausibility (e.g., certain studies involve use of a single dose, lack of dose response relationship). These limitations make it difficult to determine actual significance of findings to human health risk assessment

The neurodevelopmental and behavioural dataset in rodents, though highly uncertain, is suggestive of potential effects at doses at the same order of magnitude to 1-2 orders of magnitude higher than exposures. Given that toxico*kinetic and metabolism data indicate potential sensitivity to the pregnant woman/fetus and infant; and that animal studies suggest a trend towards heightened susceptibility during stages of development in rodents, it is considered appropriate to apply a precautionary approach when characterizing risk. As such, it is concluded that bisphenolA be considered as a substance that may be entering the environment in a quantity or concentration or under conditions that constitute or may constitute a danger in Canada to human life or health.

There are a number of uncertainties related to evaluation of risk to human health, including:

• Uncertainty in the estimation of exposure to all age groups from the use of polycarbonate repeat-use containers due to absence of Canadian use pattern information (e.g., frequency of use, method of use) as well as the availability of studies representative of actual use conditions.
• .Uncertainty in the estimation of exposures to infants from the consumption of breastmilk due to lack of surveys measuring concentrations in breastmilk from Canadian women, or surveys involving larger sample sizes. Further exploration of exposures (e.g., food, environmental media) are contributing to levels in breastmilk or other biomonitoring would assist in determining which sources of exposure are most important for risk management, if required.
• Uncertainty associated with actual residual levels of bisphenol A in polymers used to manufacture consumer products, such as epoxy adhesives, toys and cosmetics, and uncertainty associated with the assumptions made to calculate general population exposure from use of these products.
• Uncertainty associated with estimating intake of bisphenolA from the consumption of fresh fish and other fresh foods due to the lack of surveys measuring concentrations in Canada or elsewhere.
• The adequacy of the database on reproductive/developmental effects below the established NOAEL of 50 mg/kg-bw per day for these effects in rodents is an uncertainty. Limitations with respect to rigour, power, corroboration/consistency, and biological plausibility/coherence exist.
• Interspecies variation. Uncertainties exist regarding toxico*kinetic and metabolism differences between rodents and humans. Rodents clear bisphenolA in the faeces via biliary excretion of the glucuronide (EFSA 2006). Humans, on the other hand, clear bisphenolA primarily via urine after glucuronidation (Snyder et al. 2000; Volkel et al. 2002). It has been suggested that the fetus from rats and humans have little glucuronidation activity (Matsumoto et al. 2002; Ringet al. 1999; Coughtrie et al. 1988). Evidence suggests that the human fetus metabolizes bisphenolA less extensively as compared to the rat fetus. Thus, the human fetus may be more prone to effects of bisphenolA (Elsby et al.2001). These important differences in toxico*kinetics and metabolism introduce uncertainty in extrapolating from rodents to humans. Uncertainties also exist regarding relevance to humans of potential effects of bisphenolA on brain function in rodents. Species differences in brain development are important (Seegal 2001). The brain of a newborn rat pup is comparable to that of a human fetus nearing the third trimester of pregnancy. Conversely, the human newborn might be compared with a 2- to 3-week-old rat pup (Morrealeet al. 2004). The mechanism(s) by which bisphenolA exerts estrogenic effects depends on many factors, including species, and this is an additional uncertainty with respect to interspecies extrapolation.
• With regards to intraspecies variation, available data indicate that the pregnant female/fetus and infant are potentially sensitive populations. Actual levels of free bisphenolA in the fetus however remains an uncertainty. Uncertainty also exists with respect to critical windows of susceptibility during in utero and postnatal development and potential subsequent long-term effects.

Research needs include the following:

• Canadian monitoring data for certain environmental media (e.g., air, soil, drinking water) are lacking or limited. Such data would decrease uncertainty regarding the contribution of these media to exposure among the general population.
• Biomonitoring data for the general population of Canada (including pregnant women and children younger than 6 years old) would decrease uncertainty regarding the exposure assessment. Investigations into accuracy of spot urine samples in predicting daily exposures are needed.
• Studies investigating the underlying process of migration from polycarbonate would inform risk management. To date, studies in this area report inconsistent results, with some study investigators indicating that migration rates will decrease with use and others indicating that migration rates remain relatively constant throughout the life of the container.
• Epidemiological studies are needed to establish relationships between bisphenolA exposure and health outcomes for sensitive subpopulations.
• Further exploration of mechanisms by which low dose exposure to bisphenolA may induce alterations in the developing reproductive system in males and females.
• Further exploration of molecular mechanisms by which low doses of bisphenolA may disrupt the estrogen and thyroid hormone homeostasis, as well as the impact on the central nervous system during critical windows of development.
• Dose response studies using environmentally relevant levels of bisphenolA and investigation of mechanisms responsible for various dose-response curves (linear versus non-monotonic dose response relationships) and its relevance to human health risk assessment.
• Studies on the effects of bisphenolA on the sexual differentiation of the brain and neurobehavioural endpoints conducted in accordance to OECD guidelines for developmental neurotoxicity.
• Investigation of the effects of bisphenolA exposure during precise age windows that correspond with specific aspects of in utero and early postnatal development of the neuraxis that are the most critical to the development of neurobehavioural features. For example, sensory and motor differences are generally brainstem, anxiety-related activity involves mid-brain and lower telencephalon, species-typical behaviour patterns are most likely lower telencephalon and learning and memory differences involve the cortex.
• Further assessment of bisphenolA-induced changes on the reproductive and nervous system during early and later stages of development including adolescence and adulthood.
• Further research is needed to identify the doses and ages or windows of sensitivity that are the most effective in producing developmental disturbances.
• Assessment of tissue levels of bisphenolA following repeat low dose exposures via oral or parenteral routes to characterize the actual versus administered dose.
• Comprehensive evaluation of the potential of bisphenolA to initiate or promote carcinogenicity from exposure during sensitive stages of development.
• An emphasis on research to establish more precisely the magnitude of effect to better understand the impact of bisphenolA exposures is required.

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Conclusion

Based on the information presented in this screening assessment, and applying a precautionary approach, it is concluded that bisphenol A is entering or may be entering the environment in a quantity or concentration or under conditions that have or may have an immediate or long-term harmful effect on the environment or its biological diversity.

Based on the information presented in this screening assessment, and applying a precautionary approach, it is concluded that bisphenol A be considered as a substance that may be entering the environment in a quantity or concentration or under conditions that constitute or may constitute a danger in Canada to human life or health.

It is therefore concluded that bisphenol A does meet the criteria in paragraphs 64(a) and 64(c) of CEPA 1999.

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Appendix A: Robust Study Summary- Inherent Toxicity

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Appendix B: Robust Study Summary - Inherent Toxicity

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Appendix C: Robust Study Summary - Inherent Toxicity

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Appendix D: Summary of Studies Investigating Neurobehavioural Effects in Rodents