## Screening Assessment for the Challenge

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

Chemical Abstracts Service Registry Number
80-05-7

Environment Canada
Health Canada

October 2008Synopsis

Pursuant to section 74 of the Canadian Environmental Protection Act, 1999 (CEPA 1999), the Ministers of the Environment a nd 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 subs tance was identified as a result of the categorization of the Domestic Substances List as a high prior ity for action under the Ministerial Challenge, since it was considered to pose the greatest potential for exposure
to individuals in Canada and had been cl assified 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 persiste nce 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 in creased 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 manufactur ed, imported or in commerce in Canada during the calendar year 1986. In 2006, no bisphe nol A was manufactur ed in Canada at quantities equal to or greater than a repor ting threshold of 100 kg. However, bisphenol A was used in Canada in the range of 1 00 000 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 signi ficantly under aerobic conditions. However, the substance has been found not to degrade or to degrade only slowly under conditions of low or no oxygen. Bis phenol A has been detected in Canadian and U.S. surface waters, sediment, groundwater and soil, as well as in municipal and industrial waste treatment pr oducts. 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 bio accumulation potential and a capacity for metabolism in various species. Most meas ured 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. Ther e is evidence that low-level exposure to bisphenol A, particularly at se nsitive life cycle stages, may lead to permanent alterations in hormonal, developmental or reproductive cap acity. In laboratory testing, these effects have occurred within the range of concentrations measured in Canada, indicating that there is potential for adverse effects in populat ions, particularly close to point sources.

On the basis of expected continued or in creasing exposure of bi ota, and information indicating the potential for long-term adverse effects to organisms within the range of

iconcentrations currently measured in the e nvironment, it is considered appropriate to apply a precautionary approach when characte rizing 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, so il 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 μ g/kg body weight/day to 4.30 μ g/kg-bw per day. Specific exposur e estimates for the most highly exposed subpopulation (i.e., infants) range from an average of 0.50 μ g/kg body weight/day (maximum 4.30 μ g/kg-bw per day) for infants aged 0 to 1 month to an average of 0.27 μ g/kg body weight/day (maximum 1.75 μ g/kg body weight/day) for infants aged 12 to 18 months. A critical e ffect for characterizati on of risk to human health is reproductive and developmenta l toxicity. The neurodevelopmental and behavioural dataset for rodents, though highly uncertain, is su ggestive of potential effects at doses the same order of magnitude to 1- 2 orders of magnitude higher than exposures.
Given that toxicokinetics and metabolism da ta 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 he ightened susceptibility during stages of development in rodents, it is considered a ppropriate to apply a precautionary approach when characterizing risk. As such, it is conc luded that bisphenol A be considered as a substance that may be entering the environmen t 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 contro l 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.

iiTable of content

Introduction................................................................................................................... ...... 1 Substance Identity............................................................................................................. .. 3 Physical and Chemical Properties....................................................................................... 4

Sources................................................................................................................................ 5
Uses..................................................................................................................................... 6
Releases to the Environment............................................................................................... 7
Environmental Fate............................................................................................................. 9
Persistence and Bioaccumulation Potential ...................................................................... 11
Environmental Persistence............................................................................................ 11
Potential for Bioaccumulation ...................................................................................... 15 Potential to Cause Ecological Harm ................................................................................. 16 Ecological Effects Assessment ..................................................................................... 16 A - In the Aquatic Compartment .............................................................................. 16
B - In Other Environm ental Compartments.............................................................. 21 Ecological Exposure Assessment ................................................................................. 23 Characterization of Ecological Risk ............................................................................. 26 Uncertainties in Evaluation of Ecological Risk............................................................ 33
Potential to Cause Harm to Human Health....................................................................... 35 Exposure Assessment.................................................................................................... 35 Dietary Intake............................................................................................................ 35
Environmental Media................................................................................................ 44 Other Sources of Exposure ........................................................................................... 48 Aggregate Exposure from Dietar y and Non-dietary sources........................................ 52
Health Effects Assessment............................................................................................ 55
Toxicokinetics and metabolism ................................................................................ 56 Carcinogenicity......................................................................................................... 58 Genotoxicity.............................................................................................................. 59 Reproductive and Developmental Toxicity .............................................................. 60 Developmental Neurotoxicity................................................................................... 63 Weight of Evidence Summary of Deve lopmental Neurotoxicity Findings.............. 69 Epidemiology Investigations .................................................................................... 71
Characterization of Risk to Human Health................................................................... 72
Uncertainties in Evaluation of Risk to Human Health and............................................... 73
Identification of Research Needs...................................................................................... 73
Conclusion ........................................................................................................................ 76
References......................................................................................................................... 77 Appendix A

. Robust Study Summary – I nherent Toxicity............................................. 100 Appendix B

. Robust Study Summary – I nherent Toxicity............................................. 101 Appendix C

. Robust Study Summary – I nherent Toxicity............................................. 102 Appendix D

. Summary of Studies Inve stigating Neurobehavioural Effects in Rodents 103

.

iiiIntroduction

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 scr eening assessment, the Ministers can propose
to take no further action with re spect 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 Sc hedule 1 of the Act and, where applicable, the implementation of virtual elimination.

Based on the information obtained through th e categorization pro cess, the Ministers
identified a number of substa nces as high priorities for act ion. These include substances
that

• met all of the ecological categorizati on 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 grea test potential for exposure (GPE) or presented an intermediate potential for exposure (IPE), and had been identified as posing a high hazard to human health base d on classifications by other national or international agencies for carcinogenicity, genotoxicity, developm ental toxicity or reproductive toxicity.

The Ministers therefore publishe d a notice of intent in the Canada Gazette , Part I, on December 9, 2006 (Canada 2006), that cha llenged industry and other interested stakeholders to submit, within specified time lines, 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 s ubstances 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 cla ssified by other agencies on the basis of reproductive toxicity. The Cha llenge for bisphenol A was published in the
Canada Gazette on May 12, 2007. A substance profile wa s 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 th e ecological categorization crit erion for inherent toxicity to aquatic organisms, it did not meet the categorization criteria for potential for persistence or bioaccumulat ion. Therefore this assessm ent focuses principally on
information relevant to the evalua tion of risk to human health.

1Under CEPA 1999, screening assessments focus on information critical to determining
whether a substance meets the criteria for de fining 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-ter m 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 evid ence approach and precaution.

This screening assessment includes considera tion 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, stak eholder research reports and from recent literature searches, up to Se ptember 2008 for human health effects and exposure sections of the document. Key st udies 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 gene ral population, as well as information on health hazards (based principally on the weight of evidence assessments of other agencies that were used for prioritization of the substance). Decisions for human health are based
on the nature of the critical e ffect and/or margins between co nservative effect levels and estimates of exposure, taking into account confidence in the co mpleteness of the identified databases on both exposure and e ffects, 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 a nd 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 as sessment has undergone external written peer review/consultation. Comments on the tech nical portions releva nt to human health were received from Toxicology Excellence for Risk Assessment (TERA), Dr. Sergio Pellis of University of Lethbridge, Albe rta, and Dr. Alfonso Abizaid of Carleton University, Ottawa. While ex ternal comments were taken in to consideration, the final content and outcome of the screening risk asse ssment remain the res ponsibility of Health
Canada and Environment Canada. The cri tical information a nd considerations upon
which the assessment is based are summarized below.

2Substance Identity

For the purposes of this document, this subs tance will be referred to as bisphenol A.

Table 1. Substance identity

Chemical Abstracts Service Registry Number (CAS RN) 80-05-7

Name on Domestic Substances List (DSL) Phenol, 4,4' -(1-methylethylidene)bisInventory names 1 Phenol, 4,4'-(1-methylethylidene)bis- (TSCA, PICCS, ASIA-PAC) 4,4'-isopropylidenediphenol (EINECS, PICCS) 2,2-Bis(4'-hydroxyphenyl) propane (ENCS) Phenol, 4,4'-(1-methylethylidene)bis- (AICS, PICCS) 4,4'-(1-Methylethylidene)bisphenol (ECL) 4,4'-Bisphenol A (ECL) Phenol, 4,4'-(1-methylet hylidene)bis- (SWISS) Bisphenol A (SWISS, PICCS) p,p'-isopropylidene diphenol (PICCS) Diphenol methylethylidene (PICCS) bis[phenol], 4,4'-(1-methylethylidene)- (PICCS) Bisphenol-a (PICCS) Bisphenol, 4,4'-(1-methylethylidene)- (PICCS) 4,4-isopropylidene diphenyl (PICCS) 4,4'-dihydroxyphenyl-2,2-propane (PICCS) 2,2-di(4-hydroxyphenyl)propane (PICCS) 2,2-di(4-hydroxyphenyl) propane (PICCS) 2,2-bis-(4-hydroxy-phen yl)-propane (PICCS)

Other names Bisphenol A, Diphenylolpropane, BPA Chemical group Discrete organics Chemical subgroup Phenols

Chemical formula C 15 H 16 O 2

Chemical structure

Simplified Molecular Input Line Entry System (SMILES) Oc(ccc(c1)C(c(ccc(O)c2)c2)(C)C)c1

Molecular mass 228.29 g/mol

1 National Chemical Inventories (N CI). 2006: AICS (Australian Inve ntory of Chemical Substances); ASIA-PAC (Asia-Pacific

Substances List s); ECL (Korean Existing Chemicals List); EINECS (E uropean Inventory of Existing Commercial Chemical Substances); ENCS (Japanese Existing and New Chemical S ubstances); PICCS (Philippine Inventory of Chemicals and Chemical Substances); SWISS (Inventory of Newly Notified Substances and Giflist 1 – List of Toxic Substances) and TSCA (Toxic Substances Control Act Chemical Substance Inventory)

.

3Physical and Chemical Properties

Table 2 contains experimental and modelle d physical and chemical properties of bisphenol A that are relevant to its environmental fate.

Table 2. Physical and chemical properties for bisphenol A

Property Type Value Temperature (°C) Reference

Experimental 150-157 Howard 1989; Dow Europe 1993 Melting point (oC)

Modelled 132 MPBPWIN 2000

Experimental 220-398 Howard 1989; Dow Europe 1993 Boiling point (oC)

Modelled 364 MPBPWIN 2000 Density (kg/m 3 ) Experimental 1195 25 Sax and Lewis 1996

Experimental 5.3 x 10 -6 25 Bayer AG 1988 Vapour pressure (Pa)

Modelled 3.0 x 10 -5 (2.27 x 10 -7 mm Hg) 25 MPBPWIN 2000

Experimental 1.0 x 10 -6 (1.0 x 10 -11

atm·m 3 /mol) Hine and Mookerjee 1975 Henry’s Law

constant (Pa·m 3 /mol)

Modelled 9.3 x 10 -7 (9.16 x 10 -12

atm·m 3 /mol)

Water solubility 120 mg/L:

4.0 x 10 -5 (3.95 x 10 -10

atm·m 3 /mol)

Water solubility 257 mg/L:

4.7 x 10 -6 (4.7 x 10 -11

atm·m 3 /mol) 25 HENRYWIN 2000 (Bond estimation)

HENRYWIN 2000 (VP/Water solubility estimation)

Experimental 3.32 Howard 1989;
Hansch et al

. 1995 Log K
ow (Octanol-water partition coefficient) (dimensionless) Modelled 3.64 KOWWIN 2000

Experimental 2.53-2.85 (pH 4.5-5.9) Loffredo and Senesi 2006

Calculated 2.85 ECB 2003 Log K
oc(Organic carbonwater partition coefficient – L/kg) (dimensionless) Modelled 4.88 PCKOCWIN 2000

4Property Type Value Temperature (°C) Reference

Log K
oa(Organic carbonair partition coefficient) (dimensionless) Modelled 12.7 KOAWIN 2000

253-257 22-24 Lee and Peart 2000b

301 room temperature Bayer AG 1988 Experimental

120 25 Dorn et al

. 1987 Water solubility (mg/L)

Modelled 173 25 WSKOWWIN
2000

Other solubilities (g/L) Experimental (alcohol) soluble Lewis 2000

Experimental 9.59 – 11.30 Staples et al

. 1998 pK
a (Acid
dissociation constant) (dimensionless) Modelled 9.73-10.48 ACD 2005

Bisphenol A is characterized as having low vapour pressure, modera te water solubility and a moderate organic carbon-wa ter partition coefficient (log K
oc ). EPIWIN (2000) indicates that, based on the chemical structure of the substance, the K
oc of bisphenol A may be sensitive to and vary significantly with pH. The high pK
a range of 9.59 to 11.30 (see Table 2) indicates that bisphenol A is a very weak acid. While ionization of the substance may occur at high pH, there is unlikely to be appreciable ionization at environmental pH levels of 7 and lower (Cousins et al

. 2002). Cousins et al

. (2002) predict that chemicals such as bisphenol A, which have measurable vapour pressures, aqueous solubilities and octanol-w ater partition coefficients (K
ow ), will likely partition to some extent in all available environmental phases.

Sources

There is no reference in the published literatur e to the natural occu rrence of bisphenol A
in the environment.

Global production of bisphenol A was 4 billi on 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 secti on 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 (Environmen t Canada 2007a). The extent to which the reported values represent quantiti es of bisphenol A present in finished and semi-finished

5goods entering Canada from other parts of th e world is unknown as these uses would be unlikely to meet the reporti ng criteria for the survey.

Uses

Bisphenol A is primarily used as a monomer in the production of pol ycarbonates and as a precursor or a starting material for monome rs of certain epoxy resins (EFSA 2006). In 2003, consumption patterns of bisphenol A in the U.S. indicated that approximately 72% of bisphenol A was used to manufacture polycarbonates, 21% was used in epoxy resins and 6% was used in other applications (NTP 2007).

The type of mixture, product or manufactur ed item reported for 2006 in response to a survey of Canadian industry under sec tion 71 of CEPA 1999 in cluded resins, curing agents, epoxy curing agents, ha rdeners, plastic resin formulations, monomer, paperboard packaging, metal cans, phenolic resins, industr ial coatings, plasticizers, adhesives, two part epoxy adhesives, chain oil, brake fluid, heat transfer fl uid 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 adhesi ve, in custom colour powder coating and as a curing agent for resurfacing concrete.

The literature indicates that pol ycarbonate is used in the manufacture of compact discs, food and beverage contact containers (e .g., baby bottles, repeat use water bottles, pitchers, water carboys, tablewar e and storage contai ners), 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 h ousings, computer parts, household electric equipment, lamp fittings , power plugs) and the automotive industry (e.g., car headlight and rear light reflector s and coverings, bumpers, radiator and ventilation grills, safety gl azing, inside lights, motorcycle windshields and protective helmets) (NTP 2007; EFSA 2006). Medical de vices 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 the Food and Drugs Act (Health Canada, Medical Devices Bureau, Health Products and Food Branch, pers. comm., 2008 Feb 5, unreferenced).

Epoxy resins made with bisphenol A are common ly used in protective coatings, structural composites, electrical laminates, electri cal 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 bottl es, 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 Ca nada, Chemical Health Hazard Assessment Division, Health Products and Food Branch, pers. comm., 2008 Feb 26, unreferenced).

6Epoxy resins are used in automotive electr o-coating and in automotive and industrial primer and topcoat applications. Epoxy resi ns are also used in dustrially in powder coatings used on reinforcing bars for concrete , oil pipelines, wire ra cks 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 bisphenol A (Environment Canada 2007b).

Dental products containing bisphenol A-ba sed 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 Feb 5, unreferenced).

Bisphenol A is also used in the producti on of phenolplast, pheno lic and unsaturated polyester resins, thermal paper, polyvinylch loride (PVC), alkoxylat ed bisphenol A and polyols/polyurethane (NTP 2007). Additionally, it may be used in the manufacture of brake fluids, tires, modified polyamide, and tetrabromobisphenol A (ECB 2003). Other products of bisphenol A incl ude protective window glazing, building materials, optical lenses and development of dyes (NTP 2007).

The U.S. National Library of Medicine H ousehold Products Database reports that bisphenol A-based polymers are used in coati ngs, adhesives and putti es available to the general public for use in automobiles, home maintenance and repair, and hobbies (NTP 2007). Response to a survey conducted under section 71 of CEPA 1999 indicated that these products were also reported to be avai lable for use in Canada. Bisphenol A-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., 2008 March 18, unreferenced).

Releases to the Environment

There are no known natural sources of bi sphenol A and potential releases to the
environment are restricted to those asso ciated with anthropogenic activities.

Releases of bisphenol A may occur during pr oduction, 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 bisphenol A wa s 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 bisphenol A 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 a pplication materials such as polycarbonates and epoxy resins are th e most likely industria l sources of release

7of bisphenol A into the Canadian environment. Unintentional release of fugitive dust from closed systems during handling and transp ortation of the substance may also occur. Bisphenol A has low vapour pressure at ty pical 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.

Bisphenol A released to wastewater would lik ely 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 w ill be sequestered by sludge. However, the moderate water solubility of 120 to 300 mg/L indicates that a portion will likely remain in the water phase and will therefore be presen t in final effluents discharged to receiving
waters. Lee et al

. (2004) measured concentrations of 1.0 × 10 -5 to 1.73 × 10 -2 mg/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 wastewat er treatment plants. The re sults provide evidence that bisphenol A partitions into both the solid a nd liquid phases within treatment plants. A
more detailed review of concentrations meas ured in waste treatment products is provided in Table 9a. Bisphenol A present in sewage sludge could be released into the soil compartment through applicati on of sludge biosolids to ag ricultural and pasture lands. Lee and Peart (2000a) examined the efficiency of bisphenol A removal by the sewage treatment plant process. Overall reduction ra tes of < 1% to 99% were determined from influent/effluent sample pairs collected from 36 Canadian sewage trea tment plants, with a
median reduction value of 68%. Plants exhibi ting greater than 50% reduction rates were those employing secondary waste treatment.

The National Pollutant Release Inventory (NPR I) has tracked the release of bisphenol A
from industrial facilitie s in Canada since the 1990s. Tabl e 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 we re exclusively to the air compartment. In recent years,
total annual releases have varied between about 44 and 8770 kg, with no clear temporal trend apparent. Most recent reporting (2005) indi cates total on-site releases of 120 kg, a decrease from that seen in previous year s (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 tota l Canadian releases of bisphenol A. For example, all of the following reporting crit eria must have been met before NPRI reporting of bisphenol A was required for th e 2005 calendar year: (1) employees at the facility worked a total of 20 000 hours or more or the facility was us ed for an activity to which the 20 000-hour employee threshold does not apply; and (2) the facility manufactured, processed or otherwise used 10 tonnes (10 000 kg) or more of the substance in one year; and (3 ) the substance was manufactur ed, processed or otherwise used at a concentration greater than or e qual 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 re lating to the release and/or disposal of bisphenol A at their facility. While the number of such facilities is not known, any such releases would not be repor ted and would therefore be over and above those reported through the NPRI.

8Bisphenol A also appears on the United Stat es Environmental Protection Agency Toxics Release Inventory (TRI), a database of ch emicals managed through disposal and other releases, recycling, energy recovery and waste treatment. In 2005, a total of approximately 675 000 kg (1.5 × 10 6 pounds) of bisphenol A were reported to the TRI as disposed or released, with land disposal predominating (about 550 000 kg) and releases occurring to air (about 90 000 kg) and su rface waters (about 4 500 kg; USEPA 2007).

Table 3: NPRI release and disposal data (kg) for bisphenol A from 2001 to 2005

Year Releases to
air Releases to

water Releases to

land Releases to unspecified compartment Total

releases Total disposals

2001 2904
- - - 2904 4919

2002
- - - 44 44 5543

2003 8765
- - 5 8770 13 953

2004 4523
- - 2 4525 8512

2005 120
- - - 120 1154

Release over the service life of end produc ts may occur throu gh volatilization or leaching. The majority of bisphenol A appear s to be effectively retained within the polymer matrix of materials such as polycar bonates and therefore losses through leaching from the product surface are expected to be limited (ECB 2003). As well, the low vapour pressure suggests bisphenol A will have little tendency to volatilize from products at normal environmental temperatures. Some losse s could occur at elevated temperatures, for example, during heating of end produc ts. Losses may also take place through weathering and breakdown of e nd products, particularly those with outdoor applications.

Bisphenol A may enter the environment thr ough physical and chemical degradation of end products during disposal and recycling op erations. Releases w ould 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 circumstances (e.g., during rainfall events) and concentrat ions up to 1.41 × 10 -3 mg/L have been reported in groundwater samples collected in the vicinity of municipal landfills (Rudel et al

. 1998; see section on Ecological Exposure Assessment). Higher concen trations would be expected should the leachates have an elevated pH.

Environmental Fate

Based on its physical and chemical properties (T able 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 s ubstance with some water solubility, bisphenol A can be expected to partition to organic phases such as sediments and soils; however, an appreciable fraction will also likely be present in the di ssolved phase (Cousins et al

.

92002). This suggests that the medium of re lease may be particularly important in predicting the partitioning be haviour and fate of bisphenol A in the environment.

Level III fugacity modelling (Table 4) pred icts that when bisphenol A is released exclusively into water, the ma jority (96.9%) will remain w ithin 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 bisphenol A (78.7% to 99.3%) will partition to the soil compartment, with smalle r amounts expected to be present in water and sediment. Bisphenol A is not predicted to remain within air, even when released exclusively into that compartment.

Table 4: Results of the Leve l III fugacity modelling (EQC 2003) 1

Fraction of substance partitioning to each medium (%)

Substance Released to: Air Water Soil Sediment - Air (100%) 0.0 5.3 94.5 0.2
- Water (100%) 0.0 96.9 0.0 3.1 - Soil (100%) 0.0 0.7 99.3 0.0 - Air, water, soil (33% each) 0.0 20.7 78.7 0.6

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 se diment; 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 suspende d particles-water, 5.5 L/kg for fish-wat er, 100 (dimensionless) for aerosol-water

Eisenreich et al

. (1981) predicted that bi sphenol A 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 fracti on present as vapour would react with photochemically generated hydroxyl radica ls (half-life approximately 0.13 day; AOPWIN 2000) or undergo photolysis. Photodegradation products formed include phenol, 4-isopropylphenol, and a semiquinone derivative of bisphenol A (PhysProp 2006).

Biodegradation is expected to be the dominant loss proces s for bisphenol A in most aquatic and terrestrial envir onments. 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 biodegrad ation is expected to occur in nonacclimated waters; other processes such as sorption to suspended solids and sediments and photolysis may also take place. Recent re search (e.g., Chin et al. 2004; Peng et al. 2006; Zhan et al. 2006) documents both direct and indirect photoche mical 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 functi onal groups (Lyman et al

. 1982). The low Henry’s Law constant suggests that volatilization is un likely to be a significant removal process.

The moderate water solubility of bisphenol A implies that the substance will exhibit low to moderate mobility when re leased to soil (Dorn et al

. 1987). However, Loffredo and

10Senesi (2006) examined the adsorption kineti cs 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 rese archers concluded that soil adsorption of bisphenol A is generally reve rsible, with desorption occurr ing quickly and completely. Thus, the substance is exp ected 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 prop erties 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 th e 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 proces ses. 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 wa stewater 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 th e substance in subsurface waters.

Persistence and Bioaccumulation Potential

Environmental Persistence

Data relevant to the evaluati on of potential enviro nmental persistence of bisphenol A are
summarized in Tables 5a and 5b.

Table 5a. Empirical da ta for persistence

Medium Fate process Degradation value Degradation endpoint,
units Reference

Water Biodegradation (aerobic) 74.7-81.4 28-day biodegradation,
% West et al. 2001

Water Biodegradation (aerobic) 0 14-d biochemical oxygen demand, % NITE 1977

Water Biodegradation (aerobic) 0.5-3.4 (lag phase of 2-8 days required) Half-life, days Kle č ka et al. 2001

Biodegradation (aerobic) 2-3 Half-life, days

Water Biodegradation (anaerobic) < 10 10-day biodegradation,
% Kang and Kondo 2002

11Medium Fate process Degradation value Degradation endpoint,
units Reference

Sediment Biodegradation (anaerobic) 0 90-day (3-month) biodegradation, % Ronen and Abeliovich 2000

Sediment Biodegradation (anaerobic) 0 120-day biodegradation, % Voordeckers et al. 2002

Water,
Sediment Biodegradation (anaerobic) 0 70-day biodegradation,
% Ying and Kookana 2003

Biodegradation (aerobic) 1.212

1.38 50% dissipation time, days

90% dissipation time, days Water,
Sediment

Biodegradation (anaerobic) 2.75

4.901 50% dissipation time, days

90% dissipation time, days Sarmah and Northcott 2008

Soil Biodegradation (aerobic) 3 Dissipation half-life, days Fent et al. 2003

Biodegradation (aerobic) 7 Half-life, days
Soil

Biodegradation (anaerobic) 0 70-day biodegradation,
% Ying and Kookana 2005

Aquifer
water, sediment Biodegradation (anaerobic) 0 70-day biodegradation,
% Ying et al

. 2003

Biodegradation (aerobic) 0.57

2.26 50% dissipation time, days

90% dissipation time, days Ground-
water, Aquifer
Material

Biodegradation (anaerobic) 2.315

306.0 50% dissipation time, days

90% dissipation time, days Sarmah and Northcott 2008

Aerobic biodegradation is likely to be the dom inant loss process for bisphenol A in most aquatic and terrestrial envi ronments (Staples et al

. 1998). Bisphenol A has demonstrated rapid biodegradation in sta ndard aerobic 28-day ready bi odegradability testing using

12OECD Test Guideline No. 301F (West et al

. 2001), but is indicated to be nonbiodegradable using anothe r aerobic OECD method, Test Guideline 301C (NITE 1977).
Kle č ka et al

. (2001) reported biodegradation half- lives of less than 3.5 days in river surface water samples, including those to whic h 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 biodegradati on under conditions of low oxygen, suggesting there may be circumstances under which bisphenol A remains stable in the environment. Kang and Kondo (2002) measured rapid biodegradation in spiked river water samples incubated under aerobic condi tions, but a less than 10% decrease in samples incubated anaerobically for an equivalent time period of 10 days. The researchers postulated that anaerobic bacter ia may have little or no capability of degrading bisphenol A, although chemical degr adation of the substance (for example, by lipoperoxide and sodium chloride) coul d 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 th e soil. Bisphenol A was reported to persist in an anaerobic slurry of ri ver bed sediment, with no degrad ation evident after 3 months of incubation (Ronen and Abeliovich 2000). Th e substance also remained stable in anoxic estuarine sediments, with no degrada tion observed after 120 days (Voordeckers et
al

. 2002). Concentrations of bisphenol A remained unchanged in 70-day anaerobic biodegradation studies conducted using spiked samples of seawater and marine sediment (Ying and Kookana 2003), four different so il 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, suggest ing that half-lives of bisphenol A in anoxic
and anaerobic media could well exceed one year.

Sarmah and Northcott (2008) examined degr adation of bisphenol A in river watersediment and groundwater-aquifer test syst ems. 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 pe riod. The formation of degradation products
was monitored for the duration of the st udy; however, no potential metabolites were detected. Measured data were fitted using a first-order double exponential decay model, with results being pres ented 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 suggest ed that lower test concentr ations used in the study and variations in experimental pr otocols may have contributed to the observed differences. In addition, losses of up to 40% of the initial am ount applied occurred in the sterile (control) treatments; this was considered significan tly high and attributed to the possible occurrence of abiotic degradati on processes or irreversible so rption of bisphenol A to test
vessels.

13The modelled data generally support the empirical evidence, with BIOWIN (2000) predicting a half-life for ultimat e 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 substan ce 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 ha ving 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

Medium Fate process Degradation value Degradation endpoint, units Reference

Air Atm. oxidation 0.133 Half-life, days AOPWIN 2000

Air Ozone reaction No estimate.
Reaction with nitrate radicals may be important. Half-life, days AOPWIN 2000

Water Biodegradation 37.5 Half-life, days BIOWIN 2000, Ultimate survey

Water Biodegradation 0.6866 Pr obability BIOWIN 2000, Linear

Water Biodegradation 0.4653 Prob ability BIOWIN 2000, Non-linear

Water Biodegradation 0.2956 Prob ability BIOWIN 20 00, MITI linear

Water Biodegradation 0.1559 Probab ility BIOWIN 2000, MITI non-linear

Water Biodegradation -0.2593 Probab ility BIOWIN 2000, Anaerobic linear

Water Biodegradation 11.7 1 BOD, % (301C) CPOPs 2008; Mekenyan et al.

2005

Soil Biodegradation 37.5 Half-life, days Based on the modelled half-life in

water 2

Sediment Biodegradation 150 Half-life, days Based on the modelled half-life in

water 2

1 In all domains of applicability (globa l 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 bisphenol A due to reaction with the hydroxyl radical is 0.133 day (Table 5b) , indicating the subs tance will likely be rapidly oxidized in air. AOP WIN (2000) predicts reaction with atmospheric nitrate radicals may also occur.

There is conflicting evidence on the aerobic bi odegradation potential of bisphenol A. Laboratory tests show positive and negative results of ready biodegradation potential. However, there is insufficient detail for the only laboratory aerobi c biodegradation test showing a negative ready biodegr adation 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 QSAR s based on laboratory data. Therefore, field data recei ved greater weight. Results for the majority of aerobic

14biodegradation models also suggest a fairly short half-life, but so me models (e.g., CPOPS and MITI non-linear model) s uggest greater persistence of bisphenol A. Both these models, however, are trained on the NITE rea dy 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 bisphe nol 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 ev idence of this slow degradati on. Therefore, bisphenol A meets criteria for persistence in sediments (half-life in sediments ≥ 365 days) under these
conditions.

Potential for Bioaccumulation

The empirical and modelled log K
ow values of 3.32 and 3.64, respectively, suggest that bisphenol A may have some potential to accumulate in organisms. However, the
substance demonstrates low bioaccumula tion potential based on laboratory-derived bioconcentration factor (BCF) values obtained in testing with fish and other aquatic species (Table 6a). 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 bioaccum ulation factors (BAF; i.e., incorporating consideration of uptake through food) of fr om 18 to 650 for periphyton (presence of detritu s not reported, uncertain whethe r steady-state concentrations achieved) and from 8 to170 for benthos (pro cedures for sampling/collection, species not specified), and postulated that food may be an important up take route for organisms in the aquatic environment.

Table 6a. Empirical data for bioaccumulation
Test organism Endpoint Value wet weight (L/kg) Reference Periphyton
Benthos BAF 18-650
8-170 Takahashi et al

. 2003

Fish BCF 5.1-67.7 NITE 1977
Fish BCF 3.5-5.5 Lindholst et al

. 2001
Fish BCF 38±21 Lee et al

. 2004
Clam BCF 107-144 Heinonen et al

. 2002 Frog BCF 131-147 Koponen et al

. 2007

Heinonen et al

. (2002) reported very slow depuration of bisphenol A in freshwater clams,
Pisidium amnicum , exposed at temperatures ranging from 2°C to 12°C, with statistically insignificant depuration occurri ng at the lowest test temper ature of 2°C. This suggests
that in addition to environm ental exposure concentrations, body clearance rates and metabolism are important in determining final tissue levels of the substance. Metabolism

15of bisphenol A has been reported in fish (Lindholst et al

. 2001), with the formation of glucuronides identified as the primary meta bolic pathway. Kang et al (2006c) conducted a review of existing literature and also f ound that various organisms, including plants, fish, birds and mammals have capacity to metabolize bisphenol A.

A range of QSAR modelled bioaccumulation a nd bioconcentration valu es are available, with BAF or BCF values of from 5.8 (Bas eline BCF Model) to 955 (OASIS Forecast; Table 6b). Most modelled values support a co nclusion of low bioaccumulation potential.

Table 6b. Modelled data for bioaccumulation
Test organism Endpoint Value wet weight (L/kg) Reference
Fish BAF 107-170 GOBAS BAF three trophic level model (Arnot and Gobas 2003)

Fish BCF 104-170 GOBAS BCF three trophic level model (Arnot and Gobas 2003)

Fish BCF 955 OASIS Forecast 2005
Fish BCF 5.8 Baseline BCF model (all mitigating factors) (Dimitrov et al. 2005)

Fish BCF 71.8 BCFWIN 2000

Both the empirical and modeled BCF and BA F data are in good agreement and indicate that bisphenol A does not meet the bioaccumulation criterion (BCF, BAF > 5000) as set
out in the Persistence and Bioaccumulation Regulations (Canada 2000). This conclusion is consistent with the physical-chemical pr operties of the substa nce and evidence for
metabolism in biota.

Potential to Cause Ecological Harm

Ecological Effects Assessment

A - In the Aquatic Compartment

There is modelled and experimental evidence that bisphenol A causes harm to aquatic organisms at relatively low concen trations (Tables 7a, 7b and 7c)

Table 7a. Empirical data for aquatic toxicity
Test organism Type of test Endpoint Value (mg/L) Reference Algae
freshwater
marine Chronic (96 hours) EC 50 1
2.7, 3.1 2
1.0, 1.8 3 Alexander et al

. 1988

Daphnid Acute (48 hours) EC 50 LC 50 4

LC 50 10

7.75
12.8 Alexander et al

. 1988
Brennan et al

. 2006
Hirano et al

. 2004

16Test organism Type of test Endpoint Value (mg/L) Reference Chronic (7 days) IC 25 5

NOEC 6

LOEC 7 3.92
0.94
1.88 Tatarazako et al

. 2002

Chronic (21 days) NOEC
Threshold
concentration 8 ≥ 1.0

1.3 Brennan et al

. 2006

Mu et al

. 2005
Fish
freshwater

marine Acute (96 hours) Chronic (14 days)

Chronic (28 days)

Chronic (164 days) Chronic (103 days) Acute (96 hours) LC 50 NOEC
LOEC
NOEC
LOEC
NOEC
LOEC
NOEC
LOEC
LC 50 4.6
3.2
10.15
3.64
11.0
0.001-0.16
0.16-0.64
< 0.00175
0.00175
9.4 Alexander et al

. 1988 Bayer AG 1999a

Bayer AG 1999b

Sohoni et al

. 2001

Lahnsteiner et al. 2005

Alexander et al

. 1988 Mysid Acute (96 hours) LC 50 LC 50 1.1
1.03 Alexander et al

. 1988
Hirano et al

. 2004 Amphipod Acute (14 days) NOEC
LOEC 0.1
1.0 Johnson et al

. 2005

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 pe rcent effect. A point estimate of the con centration of a test substance that causes 25% reduction in a quantitative biological m easurement such as growth rate.

6 No observed effect concentration; highest concentration in a toxicity test not cau sing a statistically significant effect in c omparison to the controls. In this case, the study identified that the highest concentration tested did not result in statistically signific ant results. Since the NOEC could be higher, the NOEC is described as being greater than or equa l to the highest test concentration.

7 Low observed effect concentration; lowest concentration in a t oxicity test that caused a statistically significant effect in c omparison
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 sa ltwater toxicity tests on bisphenol A. Lowest e ndpoint values were a 96-hour EC 50 of 1.0 mg/L for the marine diatom, Skeletonema costatum , based on cell counts, and a 96-hour LC 50 of 1.1 mg/L for the mysid shrimp, Mysidopsis bahia. The mysid 96-hour LC 50 value 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 valu e was in the range of that reported by Alexander et al

. (1988; see Table 7a). Ho wever, 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) gene ration, leading to great er 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 reproduc tive 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, reporte d reduced reproductive

17success for Daphnia magna at a threshold concentration of 1.30 mg/L and proposed that
the substance could elicit chro nic effects through interferen ce with ecdysis (moulting).

Seven-day NOEC and LOEC values of 0.94 a nd 1.88 mg/L, respectively, were reported for reproduction in the daphnid species, Ceriodaphnia dubia (Tatarazako et al

. 2002); the

IC 25 for the study was 3.92 mg/L. These values are similar to resu lts obtained using Daphnia magna and provide further evidence of possible chronic toxicity.

Johnson et al

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

Potential effects following long-term exposur e (164 days) were studied in the fathead minnow, Pimephales promelas (Sohoni et 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 repr oductive effects (i.e., induction of vitellogenin synthesis, alterations to sex cells) occu rred at 0.016 mg/L. The research ers concluded that bisphenol A 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 spawni ng period, and investigated effects on maturation, quantity and quality of semen a nd eggs. Bisphenol A was added into the flow-through test system by m eans of an injection pump, with final reported test concentrations of 0.00175, 0.00240 and 0.005 mg/L calculated based on water flow and test chemical injection rates. Similar nu mbers of males (6-8) provided semen in the control group and at the two lowest concentrations, while only one male from the high test group gave semen. Semen quality was lo wer at estimated concentrations of 0.00175 and 0.00240 mg/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 semen was restricted to the end of the spawning season and delayed for approximately 4 weeks in comparison to the co ntrol. 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 th e 103-day investiga tion period. Ovulation was delayed by approximately 2 weeks in the 0.00175 mg/L exposure group and by approximately 3 weeks in the 0.00240 mg/L group. Therefore, the tested bisphenol A concentrations affected the percentage of ovulated females and the time point of ovulation. No effects were observed on the qua lity 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 9 mg/L (Table 7b), slightly a bove results obtained experiment ally. 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 bisphenol A can be consider ed highly hazardous to the

18aquatic environment (i.e., (L(E)C 50 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

Test organism Type of test Endpoint Value (mg/L) Reference
Fish Acute (96 hours) LC 50 1 3.3 ECOSAR 2004

Fish Acute (96 hours) LC 50 4.3 TOPKAT 2004

Fish Acute (96 hours) LC 50 4.9 ASTER 1999

Fish Acute (96 hours) LC 50 ≤ 6.3 3 (± 1.2) 4 CPOPS 2005

Fish Acute (96 hours) LC 50 8.5 AIES 2003-2005

Fish Chronic (30 days) MATC 5 0.5 ECOSAR 2004

Daphnid Acute (48 hours) LC 50 2.6 ECOSAR 2004

Daphnid Acute (48 hours) EC 50 2 5.1(± 2.6) 4 CPOPS 2005

Daphnid Acute (48 hours) EC 50 2 5.2 TOPKAT 2004

Daphnid Chronic (21 days) MATC 5 0.4 ECOSAR 2004

Green Algae Acute (96 hours) EC 50 4.0 ECOSAR 2004

Green Algae Chronic (96 hours) MATC 5 1.0 ECOSAR 2004

1 Median lethal concentration

2 Median effective concentration

3 CPOPS has identified bisphenol A as ha ving an unspecified reactive model of toxicity in the fish 96hr LC50 QSAR. The ≤ denotes
that the median lethal concentration is e xpected to be lower than the reported defa ult value which is based on a baseline narco tic mode
of action

4 ± 95% confidence interval

5 Maximum Allowable Toxicant Concentration, generally presented as the range between the No Obse rved Effect Concentration and Lowest Observed Effect Concentration or as the geometric mean of the two measures

Evidence of disruption to re productive and developmental processes following exposure to bisphenol A at concentra tions 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 disrupti on is presented in Table 7c. While there is a widespread variation in repor ted values, many fall in the range of 0.001 to 1 mg/L. In addition, differing sensitivities are evident be tween groups of organisms, with endpoint values for fish generally higher than those fo r aquatic invertebrates. Considered together, the data provide strong evidence that bisphenol A is capable of elici ting adverse effects: (1) following prolonged exposure at levels belo w those usually seen to elicit effects in standard toxicity tests (i.e ., tests based on recognized methods which evaluate endpoints such as survival, reproduc tion 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 follo wing parental exposure; and (4 ) using more than one mode
of action.

19Table 7c. Selected endpoint values relat ing to potential hormonal effects
Test organism Duration of test (days) Endpoint observed Lowest effect value (mg/L) Reference

Fish: Carp 4 Vitellogenin induction 22.8 Letcher et al

. 2005
Goldfish 8 Altered plasma calcium
homeostasis 0.228 Suzuki et al

. 2003

Rainbow trout 12 Vitellogenin induction 0.070 Lindholst et al

. 2000 Carp 14 Altered sex steroid levels 0.001 1 Mandich et al

. 2007
Turbot 21 Altered steroid hormone
balance 0.059 2 Labadie and Budzinski

2006 Guppy 21 Reduced total sperm count 0.274 1 Haubruge et al

. 2000
Fathead minnow 21 Reduced egg number at spawning 0.500 Brian et al

. 2007

Medaka 21 Vitellogenin induction 0.500 Tabata et al

. 2004
Medaka 21 Altered gonad development 1.720 Kang et al

. 2002
Medaka 60 Altered growth, sex ratio 1.820 Yokota et al

. 2000
Zebrafish Fertilization
to adult (65-75 days) Vitellogenin induction, altered gonad histology 0.375 Segner et al

. 2003

Brown trout 103 Reduced sperm quality and motility; delayed ovulation, reduced percent ovulation 0.00175 1 Lahnsteiner et al

. 2005

Medaka 110 Altered gonad development 0.006 1 Metcalfe et al

. 2001

Invertebrates: Copepod 21 Delayed development 0.00001 1 Marcial et al

. 2003
Mussel 21 Induction of vitellogeninlike proteins and spawning
in both sexes 0.050 2 Aarab et al

. 2006

Mussel 21 Resorption of male and female gonads 0.050 2 Ortiz-Zarragoitia and Cajaraville 2006

Mudsnail 56 Increased embryo production 0.001 (mg/kg) Duft et al

. 2003

Mudsnail 63 Increased embryo production 0.005 Jobling et al

. 2004

Ramshorn snail 180 Increased egg and clutch production 0.0000483 Oehlmann et al

. 2006

Chironomid 2 life cycles Delayed emergence (2 nd generation) mouthpart deformities 0.078

0.010 Segner et al

. 2003

20Test organism Duration of test (days) Endpoint observed Lowest effect value (mg/L) Reference

Amphibians: Frog < 1 Competitive binding to
estrogen receptor 0.107 3 Suzuki et al

. 2004

Frog 9 Suppressed metamorphosis 0.228 1 Goto et al

. 2006 Frog 75 Altered gonad development 0.228 Jagnytsch et al

. 2006 Frog 84 Feminized sex ratio 0.0228 Kloas et al

. 1999 Frog 90 No observable effect on larval growth, development
or sexual differentiation Highest: 0.500 Pickford et al. 2003

Frog 120 Feminized sex ratio at 0.0228 mg/L, no observable
effect at 0.00228 and 0.228 0.0228 Levy et al

. 2004

Reptiles:
Caiman 10 Reversed gonadal sex and altered gonad structure 1.400 1 (mg/kg egg) Stoker et al

. 2003

1 Significant effects occurred at the lowest test concentration

2 One test concentration

3 Test concentration was reported as 4.7 x 10 -7 M

B - In Other Environmental Compartments

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

Table 8. Empirical data for terrestrial toxicity
Test organism Type of test Endpoint Value Reference
Earthworm Acute (14 days)

Chronic (56 days) NOEC 1

LOEC 2

NOEC 32 mg/kg dw 100 mg/kg dw

≥ 100 mg/kg dw 3 Johnson et al

. 2005

Springtail Chronic (28 days) NOEC 500 mg/kg dw ECT 2007a Enchytraeid worm Chronic (28 days) NOEC > 100 mg/kg dw ECT 2007b Terrestrial plant (cabbage, corn, oat, soybean, tomato, wheat) Chronic (21 days) Emergence,
Growth Effects at 150 and 1000 mg/kg soil Dow Chemical
Canada Inc. 2007

Terrestrial plant (cabbage, corn, oat, soybean, tomato, wheat) Chronic (21 days) Lowest:
NOEC
EC 25 4

EC 50 5 20 mg/kg dw 19 mg/kg dw 67 mg/kg dw Springborn Smithers

2007

Terrestrial plant (bean, tomato, lettuce, wheat) Chronic (21 days) Growth, morphology 10 mg/L Ferrara et al

. 2006

Chicken Chronic (23 weeks) NOAEL 6

LOAEL 7 1 mg/kg-bw per day 100 mg/kg-bw per day Furuya et al. 2006

1 No observed effect concentration; highest concentration in a toxicity test not cau sing a statistically significant effect in c omparison to
the controls.

2 Low observed effect concentration; lowest concentration in a t oxicity test that caused a statistically significant effect in c omparison
to the controls.

3 Study identified that the highest concentration tested did not result in statistically signifi cant results. Since the NOEC cou ld be higher, the NOEC is described as being greater th an or equal to the highest test concentration.

4 The concentration which causes an e ffect on 25% of the test organisms.

5 Median effective concentration.

216 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 compari son to the
controls.

No statistically significan t effects on reproduction were evident in earthworm, Eisenia

sp. , exposed for 56 days to concentrations of up to 100 mg/kg dw of soil; however, statistically significant and concentration-depe ndent mortality occurred at concentrations of 100 mg/kg dw of soil and above (up to 10 000 mg/kg dw; Johnson et al

. 2005).

A “limit test” using concentrations of 150 and 1000 mg/kg soil was performed using six species of terrestrial plants (cabbage, co rn, oat, soybean, tomato and wheat) (Dow Chemical Canada Inc. 2007). Effects were ev ident at both concen trations tested. No information on statistical endpoints was provi ded in the summary report submitted to
Environment Canada.

Ferrara et al

. (2006) reported phytotoxicity and mo rphological anomalies in four plant species (broad bean, tomato, lettuce and durum wheat) grown hydroponically for 21 days at doses of 10 and 50 mg/L bisphenol A.

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

Furuya et al. (2006) reported on the deve lopment of male chicken phenotypes (comb, wattle and testes) in White Leghorn, Gallus domesticus , chicks administered oral doses of 0.002 to 200 mg/kg-bw bisphenol A in corn oil/a lcohol carrier every two days from 2 to 25 weeks of age. Dose-dependent growth inhi bition of all three endpoints occurred even at the minimum dosage, although recovery was ev ident at all but the highest dosage. This suggests development was delayed at lower dos es and permanently altered at the highest dose of 200 mg/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 r odents (see Health Effects Asse ssment 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 wild life 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 bisphenol A in rodent species. There is also evidence th at the pregnant female/fetus and developing offspring may be a sensitive subpopulation (s ee Health Effects Assessment section). A

22more detailed consideration of potential imp acts in mammalian species is provided in the
Health Effect Assessment section of this document.

Ecological Exposure Assessment

Table 9a provides a summary of Canadian and North American water and sediment concentrations, as well as leve ls reported in water and wast ewater treatment products.

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

Medium Location; year Concentration (mg/L or mg/kg dw 1 ) Number of samples Reference

Surface sediment Lake Erie, Canada; 2004 < 1.5×10 -4 – 6.1×10 -3 55 Chu et al

. 2005

MWWTP 2 effluent Surface water (river) Drinking water Calgary, Canada; 2002–

2004 1.25×10 -4 , 1.95×10 -4

6.4×10 -7 – 1.88×10 -5

4.5×10 -7 , 7.6×10 -7 2
3

2 Chen et al

. 2006

Pulp & paper mill:
1° effluent 3, 4

2° effluent 5 Sludge Across Canada; 1998

2.6×10 -5 – 0.139
< 5.0×10 -6 – 4.06×10 -4

< 0.020–3.330 15
15
15 Lee and Peart 2000a

MWWTP effluent Surface water (rivers), upstream and
downstream of

MWWTPs Alberta, Canada; 2002 -

2003 <1.29×10 -6 – 1.95×10 -4

1.8×10 -6 – 1.53×10 -3 8

5 Sosiak and Hebbin (2005)

Industrial wastewaters discharged to MWWTP:

• Chemical & chemical products

• Commercial laundries

• Textile products & clothing industries

• Fabricated metal products industries

• Paper & allied products industries

• Miscellaneous
industries

• Plastic products Toronto, Canada; 1999–

2000
8.0×10 -5 – 0.09127 (median 1.5×10 -3 )

7.5×10 -4 – 0.04345 (median 6.56×10 -3 )

1.0×10 -4 – 4.8×10 -4 (median 2.1×10 -4 )

< 1.0×10 -5 – 6.51×10 -3

(median 7.0×10 -4 )

< 1.0×10 -5 – 0.14923 (median 8.72×10 -3 )

2.0×10 -4 –1.70×10 -3 (median 6.7×10 -4 )

5.0×10 -5 – 2.12×10 -3

(median 6.3×10 -4 ) 37

17

10

11

8

5

9 Lee et al

. 2002

Bleached Kraft Pulp Mill
Effluent Western Canada; 2002–

2005 1.1×10 -5 – 4.0×10 -5 (mean 2.1×10 -5 ) 6 Fernandez et al

. 2007b

MWWTP sludge: Raw sludge Digested sludge Across Canada; 1994–

2001 0.100 – 39.8
0.100 – 11.1 13

22 Lee and Peart 2002

MWWTP:
Influent & effluent Sludge Across Canada; 1998–

1999 3.1×10 -5 – 0.0499 6

0.104 – 36.7 47
51 Lee and Peart 2000b

MWWTP: Across Canada; 1999– Lee and Peart 2000a

23Medium Location; year Concentration (mg/L or mg/kg dw 1 ) Number of samples Reference

Influent
Effluent Sludge 2000 8.0×10 -5 – 4.98×10 -3

1.0×10 -5 – 1.08×10 -3

0.033 – 36.7 36
34
50

MWWTP:
Influent
Effluent Raw sludge Digested sludge Toronto, Canada; 1999–

2000 1.6×10 -4 – 0.0281
1.0×10 -5 – 0.0173
0.070– 10.6
0.120– 12.5 15
15
16
18 Lee et al

. 2004

MWWTP : Primary effluent
Final effluent British Columbia, Canada; 2002 – 2003 1.0×10 -5 – 4.5× 10 -5,7

1.0×10 -5 – 3.5×10 -5 8
8 Fernandez et al

. 2007a

MWWTP:
Influent
Effluent Western Canada; 2002 –

2005 < 2.1×10 -6 – 5.9×10 -4

< 2.1×10 -6 – 3.53×10 -4 11
29 Fernandez et al

. 2007b

MWWTP effluent Vancouver, Canada;

2003 5.8×10 -5 – 1.054×10 -3 16 Campbell et al

. 2006; Fernandez, pers. comm. 2007

MWWTP:
Influent
Effluent Vancouver, Canada;

2003 4.19×10 -5 – 7.18×10 -5

2.9×10 -6 – 7.64×10 -5 5

5 Nelson et al

. 2007

MWWTP:
Influent
Effluent Ontario, Canada; 2004

2.1×10 -4 – 2.4×10 -3

2.0×10 -5 – 4.5×10 -4 8
8 Lee et al

. 2005

MWWTP sludge Ontario, Canada; 2004 3.78×10 -3 – 0.07438 4 Chu et al

. 2005

Surface water:

Stream United States; 1996–

1997 < 0.001– 0.008 72 Staples et al

. 2000

Surface water:

Stream United States; 1999–

2000 1.4×10 -4 (median) 0.012 (maximum) 85 Kolpin et al

. 2002

Surface water:

Stream Iowa, United States;

2001 < 8.8×10 -5 – 7.4×10 -4 76 Kolpin et al

. 2004

Surface water:
River
Lake
Canal Bayou Louisiana, United States;

2003 6.0×10 -6 – 1.13×10 -4

< 1.0×10 -7 – 5.7×10 -5

< 1.0×10 -7 – 1.58×10 -4

< 1.0×10 -7 – 4.4×10 -5 8

7
14

5 Boyd et al

. 2004

Surface water :

River Missouri, United States;

2003–2004 1.8×10 -5 , 2.4×10 -5 (median) 8

7.2×10 -5 , 1.98×10 -4 (maximum) 31 Solís et al

. 2007

Surface water :
River Louisiana, United States;

2004 0 – 1.472×10 -4 6 Zhang et al

. 2007

Surface sediment Massachusetts, United States; 2003–2004 < 4.0×10 -6 – 0.005 15 Stuart et al

. 2005

Groundwater Massachusetts, United States; 1996–1997 < 3.0×10 -6 – 1.41×10 -3 9 Rudel et al

. 1998

MWWTP:
Biosolids United States; 2003–

2004 0.100 – 4.6 45 Kinney et al

. 2006

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

247 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 co ncentrations were much lower than the maximum reported value, however, with le vels 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 reflecti ng the substantial load ing 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 concentr ation 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 downgrad ient of secondary wastewater treatment plant infiltration beds to wh ich no wastewater had been discharged for six months.

High levels have been measur ed in some industrial wastew aters, most notably those associated with paper and allied pr oducts (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).

Bisphenol A has also been detected in infl uent, 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. Th e values incorporate consideration of losses through degradation and from part itioning to sludge. The presence of high concentrations in some sludge samples ( up to around 40 mg/kg dw) confirms that bisphenol A partitions into this medium, ev en while some portion remains within the water phase. This supports the prediction made by Cousins et al

. (2002; see Environmental Fate section) that based so lely on properties of moderate hydrophobicity and water solubility, bisphenol A will partition into organic ph ases 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 m easured 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.

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

25Limited information on measured levels in soil was found in the publ ished 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 bisphenol A in biota
Organism Location; year Concentration (mg/kg) Number of samples Reference

Zooplankton Massachusetts, United States; 2003–2004 < 4.0×10 -6 – 0.013 1 12 Stuart et al

. 2005

Clam Massachusetts, United States; 2003–2004 0.0154, 0.0283 1 4 Stuart et al

. 2005

Fish The Netherlands; 1999 < 1.8×10 -4 – 0.0026 2 6 Vethaak et al

. 2005

Fish:
Freshwater
Marine Norway; 2003
0.001 – 0.0141 2

< 0.0017 – 0.06197 2 18
6 Fjeld et al

. 2004

Fish:
Liver
Muscle The Netherlands; 1999

0.002 – 0.075 1

0.001 – 0.011 1 4 3

6 4 Belfroid et al

. 2002

Mussel The Netherlands; 1999 0.010 1 1 Belfroid et al

. 2002

Snail Japan; 2002–2003 < 0.002–0.011 1 12 Kang and Kondo 2006a

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

Characterization of Ecological Risk

The approach taken in this ecological scr eening assessment was to examine available scientific information and develop conclusi ons 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, environm ental occurrence, and trends in production and use.

The available data indicate that bisphenol A does not persist u nder aerobic conditions. However, the substance has been found not to degrade or to degr ade only slowly under conditions of low or no oxygen. This stabilit y, 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 lo w bioaccumulation potential and a capacity for metabolism in various species. Most meas ured 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.

26Bisphenol A has been detected in surface waters, sediment, and groundwater over wide
areas in Canada and the Unite d States. It has also been measured in municipal and industrial wastewaters, sludge and biosolid s. 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 st eady emission to the atmosphere. In addition,
some limited data are available for concentr ations 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 doc umented in fish and invertebrates from Europe and Japan. The data indicate that bi sphenol A is present in a wide range of
environmental media.

Bisphenol A has shown high toxicity to aquati c 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 eco logical effects was conducted for bisphenol A as part of the weight of evidence evaluati on of its potential to cause harm. Predicted
Environmental Concentrations (PECs) were determined based on an analysis of exposure pathways, and these exposure pathways we re used to identify sensitive receptor organisms. For bisphenol A, PECs were derived using Canadian environmental monitoring data, or if monitori ng data were not available, th e PECs were based on simple calculation procedures that took into acc ount some degree of local environmental conditions, but largely relied on generic environmental parameters. A Predicted NoEffects Concentration (PNEC) was determined for each exposure pathway, by dividing a Critical Toxicity Value (CTV) by an applica tion 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 quotie nt (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 (Lee et al

. 2004). This value is selected for derivation of the Predicted Exposure Con centration (PEC) for pelagic or ganisms as it represents the highest reliable Canadian e ffluent concentration from a municipal treatment plant providing both primary and secondary wastewat er treatment and is therefore deemed both realistic and conservative. Applying a dilution factor of 10 (to acc ount for exposures in the immediate mixing zone) yields a PEC value of 0.00173 mg/L. Lahnsteiner et al. (2005) reported significantly reduced semen quality and delayed ovulation in the brown
trout, Salmo trutta f. fario , at a Lowest Observed E ffect Concentration (LOEC) concentration of 0.00175 mg/L a nd this value is selected as the CTV. To derive a Predicted No-Effects Concentration (PNEC), th e 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 condi tions. The resulting PN EC for the pelagic compartment is 0.000175 mg/L and the ri sk quotient (PEC/PNEC) is 0.00173/0.000175 =

279.9. The results indicate that bisphenol A con centrations 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 LO EC 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 epibenth ic 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 availabl e 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 fo r 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 ro ute 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 deem ed 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 agricultura l 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 munici pal wastewater treatment plant in southern Ontario (Lee and Peart 2000a,b), and this value is select ed for derivation of the PEC. The PEC is estimated for tilled agricultural soil and past ureland based on an approach described in Bonnell Environmental Consulting (2001). This approach uses the same basic approach
as outlined in the EU’s Technical Guid ance 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/m 2 ; soil has a bulk density of 1700 kg/m 3 ; sludge is mixed in soil to a depth of 0.2 m (d epth 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 bisp henol A accumulations in the soil. It is also assumed that bisphenol A persists in the soil for a peri od 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 x 0.5) / (0.2 x 1700) = 0.054 mg/kg dw

PEC pastureland = (36.7 x 0.5) / (0.1 x 1700) = 0.11 mg/kg dw

28A 56-d LOEC of 100 mg/kg dw was reported fo r 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 te rms of water concentrations only. An application factor of 100 is a pplied to the CTV in order to account for extrapolation from laboratory to field conditions a nd interspecies and intraspecies variations in sensitivity. The application factor is larg er than that used for pelagi c organisms as fewer data are available for soil species and therefore the uncer tainty in extrapolating between species is greater. The resulting PNEC va lue is 1.0 mg/kg dw. No information is available on the organic carbon (OC) content of the sludge sa mples 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 0.1 1/1.0 = 0.11 for pastureland. The ri sk quotient analysis results suggest there is currently a lo w risk of direct adverse e ffects 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 th e U.S. Environmental Protection Agency’s (EPA) Exposure Factors Handbook (USEPA 1993) was used:

where,

TDI = total daily intake (mg/kg-bw per day)
FMR = normalized free metabolic rate of wild life receptor of inte rest (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) (d efault = 35% for mink; 100% for otter)

GE i = gross energy of the ith prey spec ies (default = 1240 kcal/kg-bw prey)

AE i = assimilation efficiency of the ith prey species by the wildli fe receptor (default = 0.91)

P t

= proportion of the time the receptor spe nds 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 upt ake 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.

29As there are no Canadian data for bisphenol A concentrations in fish, concentrations of the target substance in fish (Ci) we re estimated based on the highest PEC water and a BAF

. The BAF was estimated using the GOBAS BA F 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

where:

C i = concentration in a prey fish (mg/kg-bw)

PEC water = PEC of 0.00173 mg/L calculated a bove for the pelagic compartment BAF = bioaccumulation factor of 148 L/kg es timated 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 de tails are provided in Appendix D). The range of 0.01 – 0.1 mg/kg-bw per day is select ed as an estimate for CTVs in mammalian wildlife. The CTVs for predictive sentinel w ildlife 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

where:

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 sen tinel species (0.807 kg for mink, 6.01 kg for otter; Canadian Wildlife Service, Onta rio 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 c onditions 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 bi sphenol concentrations in biota have the potential to cause adverse effects in populations of wildlife in Canada.

30A similar methodology was used to derive a risk quotient relevant to avian wildlife species. The total daily intake (TDI) was calcu lated 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 a bundant through much of North Am erica and consumes clams as part of its normal diet of aquatic invertebrates and vege tation (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 w ildlife. This concentra tion is converted to 0.00535 mg/kg wet weight (ww) using a dry weig ht to wet weight ratio of 5.29 for clams (Vinogradov 1953). The total daily intake (TDI) for lesser scaup is then:

where,

TDI = total daily intake (mg/kg-bw per day)
FMR = normalized free metabolic rate of wild life receptor of inte rest (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 spec ies (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 deri vation of a PNEC for avian wildlife. The CTV for lesser scaup can then be cal culated 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; how ever, as the reported mean body weight for the lesser scaup considers the adult bird, while the toxicity value fo r Leghorn was derived using juveniles, an extrapolation which cons iders body weight was deemed appropriate in order to incorporate consider ation of weight differences relating to life stage.

The CTV for avian wildlife is then:

31CTV wildlife = CTV ts · (bw ts /bw

w )

where:

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 sen tinel 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 yiel ds a PNEC value of 1.69 mg/kgbw per day and the risk quotient is th en 0.00018/1.69 = 0.0001. The results suggest there is currently a low risk of direct adverse e ffects to avian wildlife species in Canada.

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

There is evidence that bisphenol A is capab le of altering hormona l, development or reproductive function: following prolonged exposur e at levels below those usually seen
to elicit effects in standard toxicity tests (i.e., tests ba sed 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 filia l generations following parental exposure; and using more than one mode of action. Studies conducted using fish, aquatic invertebrates and amphibians demonstrate impacts in the ra nge of 0.001 to 1 mg/L, levels that have
been measured in Canadian and U.S. wastew aters, receiving waters and sediments. For this reason, potential exists for adverse e ffects in populations of aquatic organisms, particularly those residing in close proximity to treatment plant outfalls or other point sources. While the ability of bisphenol A 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 measur ed in the Canadian environment and the potential for adverse effects in organisms.

Bisphenol A is one of the highest volum e chemicals produced worldwide and global demand for the substance is predicted to cont inue increasing at a rate of 6 % to 10% annually. The total quantity of bisphenol A re ported 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 envi ronment could be much higher, however, as these values likely do not fully capture amount s entering Canada in finished and semi32finished products. This is particularly so, gi ven the wide applicati on of the substance and number of bisphenol A derivative products.

In summary, the lines of evidence relating to persistence, bioa ccumulation, toxicity, environmental occurrence and trends in produc tion and use raise conc erns 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 su rface waters. Possible explanations for this may be that because inputs to the envir onment are sufficiently large and continuous, measurable concentrations are always present and/or that the substa nce is not degrading as rapidly as laboratory testi ng would indicate. Measurable levels in sewage sludge and effluent are also suggestive of some level of st ability of this substan ce. Further, bisphenol A is present in media to which there is no direct release, su ch as sediment and groundwater. This implies that the substa nce remains sufficiently long enough in the environment to move from its point of releas e into other environmental media. There is evidence that bisphenol A is stable in some environmental compartments, most notably under conditions of low oxygen, such as t hose found beneath the surface layers of sediment and soil and in marshes and bogs. Le vels in the environment are expected to increase given that bisphenol A is a high production chemical and global demand for the substance is increasing. Bisphenol A has been detected in aquatic sp ecies, confirming that the substance can be taken up by organisms a nd stored in tissues. The potential for food web impacts is still undetermined, although bi oaccumulation factors of 650 have been measured in lower trophic level species, s uggesting that possible exposure of predator species could occur through cons umption of prey. Bisphenol A is acutely toxic to aquatic organisms and adversely affect s reproduction in earthworms, gr owth 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 bisphenol A
concentrations in the Canadian environment ha ve the potential to cause adverse effects in populations of pelagic organisms and mamm alian wildlife but are unlikely to cause adverse effects in benthic and soil organisms and bird species in Canada. In addition to acute toxicity, bisphenol A can impact the no rmal development of individual organisms and influence the development of their offspr ing. In laboratory testi ng, these effects have
occurred at concentrations below those show n 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 bisphenol A in the Canadian environment are of sufficient magn itude to warrant use of a precautionary approach in response to uncertainti es in the evaluation of risk.

Uncertainties in Evaluation of Ecological Risk

There is uncertainty regardi ng sources of release and rel ease 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 whet her potential releases from pa rtially-finished and finished

33goods being imported into Canada and from pr oducts currently in use in Canada have been adequately captured. This informati on is needed to more accurately estimate potential exposure concentrations for orga nisms in the Canadian environment. The presence of a recent database of measured environmental concentrations for some
locations in Canada and the United States as sisted with characterizing exposure potential in this assessment. However, quantitativ e 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 exposur e estimation and allow estimation of potential exposure for be nthic and terrestrial species.

There is uncertainty about the measured presence and accumulation potential of bisphenol A in biota. The avai lable information indicates th at bisphenol A does not meet bioaccumulation criteria as specified in the Persistence and Bioaccumulation Regulations of CEPA 1999. However, bioaccumulation fact ors of up to 650 have been determined for lower trophic level aquatic species, and this suggests there may be circumstances or conditions under which bisphenol A may accu mulate within organisms. Lower trophic aquatic organisms form an important bridge between primary producers (i.e., phytoplankton) and higher trophic species su ch as fish. Therefore, while high tissue levels could indicate the pres ence of a site-specific condi tion, such as very high local concentrations, it is also possible that bio accumulation is occurring with the subsequent potential for food chain transfer and secondary poisoning of predator species. An
increased database of measured concentra tions 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 a nd multigenerational) low-level exposure to bisphenol A, particularly duri ng sensitive life cycle stages. This uncertainty reflects a need for greater clarification of minimum ef fects levels in pelagic, sediment, soil and wildlife species, and of expos ure concentrations in the Ca nadian environment. Based on the weight of available data, the assessment indicates that there is a potential for adverse impacts in aquatic organisms, particularly thos e residing in the vicinity of treatment plant outfalls or other point sources. However, information is needed in or der to more clearly delineate the relationship between exposure levels in the Canadian environment and possible effects in organisms.

34Potential to Cause Harm to Human Health

Exposure Assessment

Dietary Intake

With respect to dietary inta ke, two sources of potential e xposure are (i) mi gration into food and beverages from packaging materials that directly contac t food, specifically food and beverage containers with internal e poxy 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 bisphenol A can mi grate into the food and beverage contents
of the containers.

Exposure to bisphenol A from food packagi ng uses has been characterized based on bisphenol A concentrations reported in a surv ey of canned foods on the U.S. market in 2007 (EWG 2007). EWG (2007) tested foods a nd beverages from 97 cans purchased in grocery stores in three states. Twenty-eight different types of foods were analyzed for bisphenol A, including canned fruits, vegetables, pasta, be ans, infant formula, meal replacement and canned milk. Results are pres ented in Table 10. Reported concentrations were considered to be represen tative of additional data availa ble to Health Canada and of data reported in the literature (EFSA 2006).

Table 10. Average bisphenol A concentrations for various canned food types

Canned foods Number of cans

tested Average bisphenol A (ppb) 1 Bisphenol A
concentration

range (ppb)

All foods 97 7.9 Nd 2 -385

Pasta 6 63.5 Nd-247

Soup 19 57.6 Nd-385
Beans (baked) 6 9.7 Nd-38

Tuna 6 9.6 Nd-108
Vegetable 17 7.8 Nd-330
Meal replacement 5 4.2 Nd-66
Evaporated milk 3 3.5 Nd-9

Infant formula 6 2.4 Nd-17

Fruit 17 2.3 Nd-27

Soft drink 12 1.7 Nd-8

351 Average is geometric mean

2 Not detected at the limit of detection 1 ppb

Potential daily intakes from the use of epoxy re sins as interior prot ective lining for food and beverage cans were estimated using average and maximum concentrations of bisphenol A, and daily intake values fo r the food commodities (Health Canada, Food Directorate, Health Products and Food Branch [HPFB], pers. comm., 2007 March 14, unreferenced, and are presented in Tabl e 11. Leaching from metal closures was considered to contribute minimally to these exposures.

Table 11. Potential daily intake of bisphenol A ( μ g/kg-bw per day) from use of epoxy resins as interior protective lin ing for food and beverage cans 1-4 year olds 3 5-11 year olds 3 12-19 year olds 3 Adults 3 Canned
food Avg 1 Max 2 Avg Max Avg Max Avg Max

Soft drinks 0.00189 0.00889 0.00216 0.0102 0.00253 0.0119 0.00181 0.00854
Evaporated
milk 0.00248 0.00637 0.00106 0.00273 0.00056 0.00143 0.00057 0.00145

Soups 0.187 1.25 0.104 0.694 0.0596 0.398 0.0587 0.392

Tuna 2.7x10 -5 3.0x10 -4 4.2x10 -4 4.7x10 -3 2.7x10 -4 3.0x10 -3 2.0x10 -4 2.3x10 -3

Pastas 0.0494 0.330 0.0304 0.203 0.0189 0.126 0.00804 0.0537
Vegetables 0.00847 0.358 0.00814 0.345 0.00418 0.177 0.00309 0.131

Fruits 0.00260 0.0306 0.00132 0.0156 0.00034 0.00402 0.00056 0.00657

TOTAL 0.252 1.98 0.147 1.28 0.0864 0.731 0.0730 0.596

1 Based on the average concentration of bisphenol A found in foods as reported in Table 10.

2 Based on the maximum concentration of bisphenol A 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 bisphenol A in other food items are limited. Miyamoto and Kotake (2006) identified bisphenol A in noncanned foods, but at tributed this to additives in ink in paper used in the food packaging. Table 9b presen ts concentrations of bisphenol A found in aquatic biota. Seafood purchased in a Singapor e 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 cockle, 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 bisphenol A in the range of 0.25 to 1.0 3 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 μ g/kg as measured by Fjeld et al. (2004) and based on ra tes of consumption of fish for Canadians (Health Canada 1998), the estimated intake of bisphenol A from this source would range

36from 0.11 μ g/kg-bw per day (for 12-19 year olds) to 0.24 μ g/kg-bw per day (for 1-4 year olds). Studies have also indicat ed that concentrations of bi sphenol A tend to be higher in fish liver and other organs (Belfroid et al. 2002) so cons umption of these organs may slightly increase exposure. C onfidence in these estimates is low as they are based on a number of assumptions; however it is likely that they are an ove restimate 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 ex tracting medium, migra tion into liquid is considered to be greater than migration in to powder and analysis therefore focussed on liquid formulations. Preliminary research conducted at Health Canada supports the conclusion that migration of BPA is significan t only for liquid infant formulas (Health Canada, Food Directorate, HPFB, pe rs. 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 can s representing eight brands of ready-todrink or concentrated liquid infant formulas . Bisphenol A was detected and quantified in all samples, with concentra tions ranging from 2.27 to 10.2 ppb (Cao et al. 2008). These data are consistent with other data reporte d 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 formul a available on the US market. The Canadian
data were considered to be th e most relevant for use in the assessment. As a worst-case scenario it was assumed that liquid infant fo rmula is the main source of food for infants in the first 18 months of their life, and that th eir formula intake is equal to their daily total fluid intake (i.e., formula, milk, water, jui ce). 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-touse infant formulas. Assumptions rega rding formula intakes are based on
recommendations of the Institute National de Santé Publique du Québec (INSPQ 2001). Potential daily intake of bisphenol A from li quid infant formula packaged in containers with epoxy resin interior protective lining is provided in Table 12.

Table 12. Potential daily intake of bisphenol A ( μ g/kg-bw per day) from liquid infant formula packaged in containers with epoxy resin interior protective lining. Intake based on average bisphenol A concentration 1 Intake based on maximum bisphenol A concentration 1 Age of infants

Average
formula intake Maximum
formula intake Average
formula

intake Maximum
formula intake

0 to 1 month 2 0.45 0.75 0.81 1.35

2 to 3 months 3 0.50 0.69 0.96 1.31

4 to 7 months 4 0.38 0.52 0.75 1.02

8 to 12 months 5 0.21 0.28 0.42 0.55

12 to 18 months 6 0.23 0.27 0.38 0.46

371 Based on the average and maximum concentration of bisphenol A 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 accord ing 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 bisphenol A 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 bisphenol A 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 bisphenol A 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 bisphenol A 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 bisphenol A 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 bisphe nol A through the consumption of human breastmilk, however no Canadian data on con centrations of bisphe nol A in breastmilk were identified. In a survey of 20 U.S. lactating women with no known occupational exposure, Ye et al. (2006) identified con centrations of total bisphenol A (free + conjugate) ranging from the limit of dete ction (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 breastmilk 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, Otaka et al. (2003) quantified levels ranging from the lim it of detection (0.09 ng/ g) to 0.70 ng/g (ppb) of bisphenol A in the breastmilk of three wo men. Based on the U.S. data (Ye et el. 2006), potential daily intakes for infants exposed to bisphenol A through breastmilk are provided
in Table 13.

Table 13. Potential daily intake of bisphenol A ( μ g/kg-bw per day) from consumption of human breast milk containing either aver age or maximum bisphenol
A concentrations.

Infant age group Average Concentration of bisphenol A (1.8 ppb) 1 Maximum concentration of bisphenol A (7.1 ppb) 1

0 to 1 month 2 0.28 1.09

2 to 3 months 3 0.21 0.84

4 to 7 months 4,5 0.19 0.73

1 Average and maximum concentrations of bisphenol A as measured in breastmilk from 20 lactating women
in the United States (Ye et al. 2006).

382 Based on average female infant weight of 3.9 kg (NCHS 2000) and an estimated breast milk intake of 596 g/day (Arcus-Arth et 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. Therefore intake from this source was not estimated past seven months of age.

ii) Repeat-use Containers

Migration of small amounts of bisphenol A from polycarbonate rep eat-use containers such as baby bottles, drinking bottles, p itchers and carboys can contribute to oral exposure. Numerous studies have repor ted on migration of bisphenol A from polycarbonate bottles into water or milk simula nt 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; Le et al . 2008; Maragou et al. 2007; Mountfort et 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 conditi ons using food simulants such as water, 10% ethanol,
50% ethanol and 3% acetic acid. In consid ering these results, it is important to differentiate between study designs intended to simulate realistic exposure scenarios versus those designed to characterize maximu m concentration that may migrate from the product under extreme conditions or ove r the life of a container.

Five studies of bottles obtained from the Ca nadian market are available. In 2000, Health Canada conducted a series of limited migrati on studies (Page et al . 2006). No bisphenol A was detected in milk simulant (50% etha nol/water) following stor age in polycarbonate baby bottles for up to 7 days at 4 or 22 o C (detection limit of <0.1 μ g/kg). In additional trials, bottles were subjected to the followi ng treatments: (i) storage for 7 days at 4 o C then storage for an additional 7 days at 22 o C, followed by 6 days at 70 o C; (ii) storage for 7 days at 22 o C then storage for 7 days at 4 o C, followed by 6 days at 70 o C C; (iii) storage for 6 days at 70 o C. In these 3 trials, aliquots of simu lant were collected after 1, 2, 3 and 6 days at 70 o C and analyzed for bisphenol A. For all three trials, bisphenol A concentrations in the simulant on day 1 we re 0.4 to 1 ppb, and concentrations increased on subsequent days of incubation at 70 o C. Sustained incubation at 70 o C is not considered a meaningful study design for characterization of migration under real istic use conditions.

A second study also measured the migration of bisphenol A from polycarbonate baby and drinking bottles into water. Cao and Corriv eau (2008) tested five polycarbonate bottles, three baby bottles and two refillable drinki ng bottles, bought in Ottawa, Canada in 2007. The bottles were filled to capaci ty 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 μ g/L (ppb).

Preliminary raw data are avai lable from a study conducted by the Health Canada in 2008 (Health Canada, Safe Environments Program me, Healthy Environments and Consumer

39Safety Branch (HECSB), pers. comm., 2008 March 6, unreferenced). Migration was characterized by filling 14 brands of baby bottl es 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 infa nt formula. The bottles were incubated for 8,
24 or 240 hours at 40°C to simulate use at r oom temperature, as per recommendations from the USFDA (2007). The 240-hour scenario (10 days) was included to estimate migration from both repetitive use and worst-ca se scenarios. Average migration for these tests ranged from 0.095 μ g/L (ppb) in water incubated for 8 hrs, to 2.05 μ g/L in the 50:50
ethanol/water blend incubated for 240 hours. The scenario co nsidered to represent the most realistic use consisted of filling bottl es with the 50:50 ethanol/water blend and incubating them for 8 hours. This scenario resulted in average concentrations of 0.15

μ g/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 temperatur e and then in a subsequent trial at 80 o C. The study investigators specify that the 80 o C treatment was considered to simulate repeat washing of bottles (i.e., approximate ly 60-100 washes). As such, the 80 o C trials are not considered to represent a realistic migrati on 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 o 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 o C 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 temper ature. The concentration range of bisphenol A measured in milk migrating from new, ha rd polycarbonate bottles ranged from 0.44 to

0.53 μ g/L (n=10) and was reported to be 0.29 μ g/L in used bottles (n=5). The migration concentrations measured in apple juice ranged from 0.29 to 0.46 μ g/L in new bottles (n=10) and was 0.43 μ g/L when measured in the used bottles (n=5). For both liquids, bisphenol A concentrations from the disposable plastic bags and gla ss bottles were below the detection limit (0.2 μ g/L).

Maragou et al. (2007) assesse d migration potential from 31 new polycarbonate baby bottles (from 6 different brands) availabl e 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 we re cleaned using either method, rinsed and sterilized prior to be ing filled with water and incubated at 70 o C for 2 hours. This series of
treatments did not lead to detectable bis phenol A migration with a reported detection limit of 2.4 ppb. Migration was only detected when bottles were filled with boiling water

40(100 o C). Specifically, 5 repeat cycles of clean ing the bottles by brushing with detergent, sterilizing in boiling wate r for 10 minutes, then fill ing with boiling water (100 o C) and leaving at ambient temperature for 45 mi nutes resulted in an average bisphenol A 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, which measured concentrations of up to approximately 50 μ g/L (ppb) of bisphenol A in the food simulant conten ts. In a study of bottles filled with boiling water and kept overnight at room temper ature Hanai (1997) m easured bisphenol A migration varying from 3 to 55 μ g/L (detection limit of 2 μ g/L). Earls et al. (2000) detected bisphenol A in 5 out of 12 bottles f illed 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 μ g/L (detection limit of 10 μ g/L).

However, the Norwegian Food Safety Au thority (Biedermann-Brem et al. 2007) investigated the effect of migration of bisphenol A from polycarbonate under extreme washing conditions such as washing us ing very strong alka li detergents (80 o C for 1 hour, followed by drying of unrinsed bottles at 90 o C for 30 minutes). They concluded that even under aggressive conditions, concentrations of bisphenol A in contents are unlikely to
exceed 10 μ g/L (ppb).

A recent study investigated migration beha viour of bisphenol A into water during microwaving of polycarbonate baby bottles obtai ned from the European market (Ehlert et al. 2008). Polycarbonate bottl es filled with wate r were heated to 100 o C during 3 microwave cycles. Migration of bisphenol A in to water was shown to be in the range of

<0.1 to 0.7 μ g/L (ppb).

Miyamoto and Kotake (2006) reported bis phenol A 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 o C for 30 minutes.

Polycarbonate carboys can be used to store dr inking water. Biles et 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 DDH 2 0 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 be verages. New and used polycarbonate bottles were filled with water and left for up to 7 days at room temperature (22 o C). Samples were collected on days 1, 3, 5 and 7. On day 1, the ca lculated 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 releas ed from polycarbonate bottles

41into water at room temperature increased w ith time, but no significant difference was
observed between the concentrations resulti ng from the use of new versus used bottles. Repeating the study using boiling water (100 o C) 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 appropr iate to estimate bisphenol A intake from the use of polycarbonate baby bottles for bot h the recommended room temperature use and plausible high temperature use (boiling wate r) scenarios. Results calculated from Le et al (2008) indicate that f illing polycarbonate bottles with ro om 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 concentra tion of 0.15 ppb bisphenol A (Health Canada, Safe Environments Programme, HECSB, pe rs. comm., 2008 March 6, unreferenced). To estimate migration of bisphenol A from additi on 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 cool ed 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 fl uid intake (i.e., formula, milk, water, juice). Exposure estimates are provided in Table 14.

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

Concentration of bisphenol A resulting from filling bottles with

room temperature water (0.24 ppb) 1 Concentration of bisphenol A resulting from filling bottles with boiling water (10.0 ppb) 2 Infant age group

Average
formula intake Maximum
formula intake Average
formula intake Maximum
formula intake

0 to 1 month 3 0.040 0.066 1.65 2.77

2 to 3 months 4 0.047 0.064 1.95 2.67

4 to 7 months 5 0.035 0.049 1.46 2.00

8 to 12 months 6 0.020 0.026 0.82 1.07

12 to 18 months 7 0.017 0.020 0.71 0.85

1 Based on an average concentration of bisphenol A de tected 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 fro m filling bottles with liquid at room temperature and

42leaving 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 (Maragou et 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 repe at use containers are available to help determine frequency and method of use of polycarbonate repeat-u se 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), re sulting from the room temperature use of polycarbonate bottles, as it was a ssumed that contents would t ypically not be heated prior to serving for these age groups. This value is supported by preliminar y 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 c onsumed 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 Table 15.

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

adults Age group Potential daily in take of bisphenol A (μg/kg-bw per day) 1
1–4 years 2 0.012 5–11 years 3 0.010

12–19

years 4 0.005

Adults 5 0.006

1 Based on a calculated average bisphenol A concentratio n 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)

43Another potential source of e xposure to bisphenol A is thr ough the use of polycarbonate tableware and storage containe rs. Kawamura et al. (1998) measured the migration of bisphenol A 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 minut es (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 bisphenol A, concentrations ranged from 1.7 to 4.5 ppb. Due to limited data on migration fr om polycarbonate tabl eware and storage containers, the lack of information on avai lability of these products on the Canadian market, and use patterns by the general popul ation, exposure estimates cannot be derived. However, based on available information, under normal use conditions exposure from this source is expected to be limited.

Environmental Media

Many recent reviews have summarized bis phenol A concentrations in environmental media (Willhite et al. 2008; Vandenburg et al. 2007; NTP 2007; Kang et al. 2006b).

Data on concentrations of bisphenol A in Ca nadian waters are summarized in Table 9a. The only published Canadian drinking water va lues were measured by Chen et al. (2006) in Alberta at the Bearspaw (on the Bow Ri ver) and the Glenmore (on the Elbow River) water treatment plants with bispheno l A concentrations of 0.00045 and 0.00076 μg/L respectively (no detection limit presented). Thes e values fall within the range reported by other countries, with con centrations between 300 pg/L to 2 ng/L (0.0003 to 0.002μg/L) (Kuch and Ballschmiter 2001) in Germany, a nd 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 bis phenol A was detected (LOD 0.1 ng/L), it was not quantifiable, making this study unsuitabl e for deriving estimates of bisphenol A intake for the Canadian population.

No Canadian data were identified for am bient or indoor air. Although bisphenol A is released to air, due to its lo w 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 E xposure Assessment). Concentrations in ambient air have been measured in two pilot studies conduc ted 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, bisphenol A was detected in ambient air outside residences in 31% of samples in No rth Carolina, ranging from below the limit of detection (0.9 ng/m 3 ) to 44.6 ng/m 3 (50th percentile 0.9 ng/m 3 ), and 34% of samples in Ohio, ranging from the limit of detection to 19.0 ng/m 3 (50th percentile 0.920 ng/m 3 ) (Wilson et al. 2007). Concentrations of bis phenol A measured in ambient air outside daycare centres, ranged from belo w the limit of detection (0.01 ng/m 3 ) to 51.5 ng/m 3 (50th percentile below detecti on) in North Carolina (detected in 38% of samples), and below detection to 6.94 ng/m 3 (50th percentile below detectio n) in Ohio (detected in 44% of samples). In deriving inta ke estimates for the Canadian population, it was considered

44appropriate to use the maximum and fiftieth percentile ambient air concentration in North Carolina (44.6 ng/m 3 , 0.9 ng/m 3 ), taking into consideration the residential location and larger sample size (n=127).

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

No Canadian data were identified for soil. On e study in the United States analyzed soil in four daycare facility play areas, and reporte d bisphenol A 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 μ g/g)(Wilson et al. 2003). The maximum and mean concentrations in soil (0.007 μ g/g; 0.006 μ g/g) from residential locations were considered most appropriate for use in deriving estimates of intake.

Preliminary raw data are available for Phas e 1 of the Canadian House Dust Study. Based on air-dried vacuum samples collected from 260 Canadian homes, median concentrations of bisphenol A in dust were reported at 1.60 μ g/g dust (ranged from less than the method reporting limit [0.17 μ g/g] to 23.84 μ g/g). Bisphenol A was detected in 99% of homes (Health Canada, Environmental Health Scie nce and Research Bureau, pers. comm., 2008 Feb 7, unreferenced). These results are consiste nt with concentrati ons reported in dust
from homes in other countries. In a surv ey 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 Oh io. 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 ha d detectable levels, with a maximum value of 589 ng/g (50 th percentile is below detecti on limit). The maximum and median concentrations (23.84 μ g/g; 1.60 μ g/g) identified in dust measured in homes in Canada were considered most appropriate for use in deriving estimates of intake.

45Tables 16 and 17 present mean and upper bounding estimates of intake for each age group in the general pop ulation of Canada based upon c oncentrations of bisphenol A
identified in environmental media. Dust a nd soil concentrations have been summed for the purposes of the intake table.

Table 16. Average estimates of potential daily intake of bisphenol A by the general population in Canada from environmental media. Estimated intake ( μ g/kg-bw per day) of bisphenol A by various age

groups Route of

exposure

0–7

months 1 8–12

months 2 1–4

years 3 5–11

years 4 12–19

years 5 20+

years 6

Ambient
air 7 4 x 10 -5 1.2 x 10 -4 7.3 x 10 -5 6.2 x 10 -5 3.3 x 10 -5 3.0 x 10 -5

Indoor air 8 5.6 x 10 -4 1.6 x 10 -3 1.0 x 10 -3 8.7 x 10 -4 4.7 x 10 -4 4.3 x 10 -4

Drinking

water 9 NA 10 NA 10 3.7 x 10 -5 3.2 x 10 -5 1.7 x 10 -5 1.9 x 10 -5

Soil and
dust 11,12,13 8.1 x 10 -3 1.8 x 10 -2 1.1 x 10 -2 4.0 x 10 -3 9.0 x 10 -4 8.0 x 10 -4

Total
intake 8.7 x 10 -3 2.0 x 10 -2 1.2 x 10 -2 4.9 x 10 -3 1.4 x 10 -3 1.3 x 10 -3

1 Assumed to weigh 5.9 kg (NCHS 2000), to breathe 2.1 m 3 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 m 3 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 m 3 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 m 3 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 m 3 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 m 3 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 the 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/m 3 (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 the fiftieth percentile indoor air concentration as measured in samples inside residences in North Carolina, U.S.A, was 0.00182 μg/m 3 (detected in 68% of samples, n=127). Canadians are assumed to spend 21 h ours indoors each day (Heal th Canada 1998).

9 The only published Canadian values for bisphenol A 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..

4612 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) wh ich also included day care play facilities (detection limit of 1 μg/kg).

13 The median concentration in dust of 1600 μg/kg wa s 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 bisphenol A by the general population in Canada from environmental media. Estimated intake ( μ g/kg-bw per day) of bisphenol A by various age

groups Route of

exposure

0–7

months 1 8-12

months 2 1–4

years 3 5–11

years 4 12–19

years 5 20+

years 6

Ambient
air 7 2.0 x 10 -3 5.8 x 10 -3 3.6 x 10 -3 3.1 x 10 -3 1.6 x 10 -3 1.5 x 10 -3

Indoor air 8 6.0 x 10 -2 1.8 x 10 -1 1.1 x 10 -1 9.3 x 10 -2 5.0 x 10 -2 4.6 x 10 -2

Drinking

water 9 NA 10 NA 10 3.7 x 10 -5 3.2 x 10 -5 1.7 x 10 -5 1.9 x 10 -5

Soil and
dust 11,12,13 1.2 x 10 -1 2.6 x 10 -1 1.7 x 10 -1 5.9 x 10 -2 1.3 x 10 -2 1.2 x 10 -2

Total
intake 1.8 x 10 -1 4.4 x 10 -1 2.8 x 10 -1 1.6 x 10 -1 6.4 x 10 -2 5.9 x 10 -2

1 Assumed to weigh 5.9 kg (NCHS 2000), to breathe 2.1 m 3 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 m 3 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 m 3 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 m 3 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 m 3 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 m 3 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 the maximum ambien t air concentration as measured in samples taken outside residences in North Carolina, U.S.A., as 0.0446 μg/m 3 (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 the maximum indoor air concentration as measured in samples inside residences in North Carolina, U.S.A., as 0.193 μg/m 3 (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 bisphenol A 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

. 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).

4713 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).

Other Sources of Exposure

Dental materials manufactured from bisphe nol A 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 perm anent first and second molars that erupt into the mouth at ages six and twelve respectiv ely (Health Canada, Medical Devices Bureau, pers. comm., 2008 February 27, unreferenced)

While current Canadian data on the prevalen ce of dental decay in children are not available, it is known through studies of individual popula tions that it is a common childhood disorder, particularly in certain at-risk populations. Leake and Main (1996) found that children aged five in Ontari o had 1.22 decayed, missing or filled teeth. Peressini et al. (2004) found th at 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 h ealth 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 co mposite resin. While current data on the prevalence of usage of composite resins versus ot her restorative materials are not available, anecdotal information suggests th at the use of composite resin is increasing and is currently used more frequently th an other materials to restore primary and permanent teeth in children.

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

A number of investigators have shown that pl acement of composite re sin restorations and sealants can result in acute, transient expos ures to bisphenol A (F ung et al. 2000; Joskow et al. 2006; Sasaki et al. 2005)

Fung et al. (2000) collected saliva following the a pplication of 8 mg (applied to one surface) and 32 mg (8 mg applied to four su rfaces) of a dental sealant to a group of 18 and 22 adults, respectively. Concentrations of bisphenol A in saliva ranged from 5.8 to 105.6 ppb (detection limit of 5 ppb) at 1 hr a nd 3 hrs following treatment but was not

48detected immediately after treatment or at ot her time intervals (24 hr, 72 hrs and 120 hrs)
post-treatment.

More recently, Joskow et al. (2006) measured c oncentrations of bisphenol A in saliva of 14 adults before and after the applicati on of between 39.0 to 42.5 mg of two dental
sealants. Mean concentrations of bisphenol A before treat ment, 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 treate d with the other sealant.

Sasaki et al. (2005) measured bisphenol A concentration in saliva collected from 21 patients before and after dental treatments using 9 different commercial composite resins. The concentration of bisphenol A in saliva for patients treated with bis-GMA monomerbased composite resins (and bonding agents) la belled 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 o C. Patients treated with the composite labelled F (n=3) had the highest concentration of bis phenol A 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 th e 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 de ntal sealants that carried the Seal of the American Dental Association [product recognition program, now discontinued] in
2007 released detectable amounts of BPA a nd continued with recommendations for simple clinical procedures [commonly followed] to reduce the possibility of unpolymerized BPA-derived monomer remainin g on the surface of the tooth following sealant placement.

A method to extrapolate from concentrations in saliva to daily estim ates of exposure was not identified. However, consistent with th e data presented above , other international assessments have characterized exposure to bisphenol A 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 bisphenol A for in fants from contact with consumer products such as toys. The contact time was establis hed using a survey of mothers who observed mouthing behaviour in a total of 50 infants (Tanimura 1999) and video recordings of 25 infants in which time spent mouthing toys was monitored (Tanimura 2000). A minimum migration rate (of 0 μg/cm 2 /min (for products that do not contain bisphenol A) and a maximum migration rate 0.0162 μg/cm 2 /min identified in the Japanese literature [Kanagawa 2000] were used. The estimated intake of bisphenol A 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 bisphenol A-based polymers and, as such, these are likely to be overestimates of exposure . Similarly, few toys on the Canadian market

49are expected to be made of bisphenol A- based polymers and, additionally, these hard plastics tend to have a lower migration rate than would be e xpected from soft plastic toys (Health Canada, Consumer Product Safe ty Bureau, pers. comm., 2008 March 17, unreferenced) and therefore contribution to exposure among infants and children is expected to be limited.

The Household Products Database (National Institutes of Health 2007) indicates that bisphenol A may be found as a residual monomer in epoxy resins used in the manufacture of epoxy adhesives. The resins ma y be present in the adhesive products at concentrations ranging from 0.1%– 100%. Howeve r, only trace quantiti es of bisphenol A 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 bisphe nol A from epoxy adhesives predicts air concentrations during us e to range from 3.82 x 10 -5 μ g/m 3 to 6.08 x 10 -5 μ g/m 3 (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 occu r at the lower end of this range. Dermal exposure may also result from the use of this product and was
estimated at 0.0167 μ g/kg-bw per use (Table 18). Confid ence in these estimates is low as they are based on a number of assumptions; how ever it is likely that they overestimate actual exposures from this source.

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

Consumer product
scenarios Assumptions Estimated

exposure

Two
component

epoxy adhesive– gluing the
handle of a
coffee

mug 1 Inhalation

- Used ConsExpo model version 4.1, exposure to vapour: evaporation from an increasi ng area as mode of release (RIVM 2006) - Assume saturation conditions (i.e. select “limit air concentration to vapour pressure of pure substance” check box) - Based on a reported weight fraction of 0.1 for epoxy resins in epoxy adhesives (Emerson and Cuming 2004) and assuming 0.0001% residual bisphenol A monomer in epoxy resins (ECB 2003) - Assume amount of product used is 0.5 g/event to cover a
surface area of 2 cm 2 , and an application duration of 5
minutes
- Assume a room volume of 20 m 3 , exposure duration of 240 minutes, a ventilation ra te of 0.6 times/hr, a mass transfer rate based on Lang muirs method and a molecular weight matrix of 3000 g/mol (RIVM 2007). Mean event
concentration

= 0.0000382

μ g/m 3

Two
component

epoxy Inhalation

- Used ConsExpo model version 4.1, exposure to vapour: evaporation from an increasi ng area as mode of release Mean event
concentration

= 0.0000608

50Consumer product
scenarios Assumptions Estimated

exposure

adhesive– gluing a large vase 1 (RIVM 2006). - Assume saturation conditions (i.e. select “limit air concentration to vapour pressure of pure substance” check box). - Based on a reported weight fraction of 0.1 for epoxy resins in epoxy adhesives (Emerson and Cuming 2004) and assuming 0.0001% residual bisphenol A monomer in epoxy resins (ECB 2003). - Assume amount of product used is 20 g/event to cover a
surface area of 500 cm 2 , an application duration of 30 minutes, a room volume of 20 m 3 , exposure duration of 240 minutes, a ventilation ra te of 0.6 times/hr, a mass transfer rate based on Lang muirs method and a molecular weight matrix of 3000g/mol (RIVM 2007). μ g/m 3

Dermal - Used ConsExpo model version 4.1, direct dermal contact with product: instant application as mode of release (RIVM 2006). - Based on a reported weight fraction of 0.1 for epoxy resins in epoxy adhesives (Emerson and Cuming 2004) and assuming 0.0001% residual bisphenol A monomer in epoxy resins (ECB 2003). - Assume the exposed area of skin is 43 cm 2 , and an applied amount of 0.1 grams of product is applied (RIVM 2007). - Assume adult exposed wei ghs 60.0 kg (Nutrition Canada 1980). - Assume 100% uptake. Acute dose
per event =
0.00167 μ g/kg

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 bisphenol A 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 March 18, unreferenced). Residual bisphenol A in these products ar e not known but are expected to be low and are not expected to co ntribute significantly to the overall exposure to the general population.

51Aggregate Exposure from Diet ary and Non-dietary sources

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 ( μ g/kg-bw per day).

Total estimated intake (dietary intake + media) Age group Intake from
infant
formula 1,2 Intake from
environmental
media 6 Intake from migration from polycarbonate
bottles 3,4 Total estimated dietary intake 4,5 (bottles + formula) Boiling

water Room
temp.

water

0 to 1 month 0.45 0.009 1.65 2.1 2.11 0.50

2 to 3 months 0.50 0.009 1.95 2.450 2.46 0.56

4 to 7 months 0.38 0.009 1.46 1.840 1.85 0.42

8 to 12 months 0.21 0.02 0.82 1.03 1.05 0.25

12 to 18 months 0.23 0.02 0.71 0.94 0.96 0.27

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 comp ared to boiling water and are not presented here (see Table 14).

4 Based on the average concentration of bisphenol A meas ured in water, 10 ppb, that resulted from filling polycarbonate bottles with boiling water (M aragou et al. 2007) (see Table 14).

5 Heath Canada recommends that solid food be introdu ced 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 th eir total dietary exposure to bisphenol A.

6 Based on average concentrations measured in various environmental media (see Table 16).

52Table 19b. Aggregated maximum estimates of potential exposure for formula fed infants aged 0 to 18 months ( μ g/kg-bw per day).

Total estimated intake 2 (dietary intake + media) Age group Intake from
infant
formula 1,2 Intake from
environmental
media 6 Intake from migration from polycarbonate
bottles 3,4 Total estimated dietary intake 4,5 (bottles + formula) Boiling

Water Room temp.

water
0 to 1 month 1.35 0.18 2.77 4.12 4.30 1.60

2 to 3 months 1.31 0.18 2.67 3.98 4.16 1.55

4 to 7 months 1.02 0.18 2.00 3.02 3.20 1.25

8 to 12 months 0.55 0.44 1.07 1.62 2.06 1.02

12 to 18 months 0.46 0.44 0.85 1.31 1.75 0.92

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 comp ared to boiling water and are not presented here (see Table 14).

4 Based on the average concentration of bisphenol A meas ured in water, 10 ppb, that resulted from filling polycarbonate bottles with boiling water (M aragou et al. 2007) (see Table 14).

5 Heath Canada recommends that solid food be introdu ced 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 th eir 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 environm ental media) were aggregated and are presented in Table 20.

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

. Intake from consumption of

breastmilk 2 Overall intake from
environmental media 3 Total estimated intake Age group

Average Maximum Average Maximum Average Maximum

0 to 1
month 0.28 1.09 0.009 0.18 0.29 1.27

2 to 3
months 0.21 0.84 0.009 0.18 0.22 1.02

4 to 7
months 0.19 0.73 0.009 0.18 0.20 0.91

1 If infants are fed pumped breastmilk via polycarbonate bottles, this may add to their overall exposure to bisphenol A.

532 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 estimate of potential exposure for children (aged 1+ years) and adults ( μ g/kg-bw per day).
Intake from consumption of

canned food 1 Intake from migration
from polycarbonate
bottles 2 Overall intake from
environmental
media 3 Total estimated
intake Age

group

Average Maximum Average Average Maximum Average Maximum

1 to 4

years 0.25 1.98 0.012 0.012 0.28 0.27 2.27

5 to11

years 0.15 1.28 0.010 0.005 0.16 0.17 1.45

12 to 19

years 0.09 0.73 0.005 0.001 0.06 0.10 0.80

Adults 0.07 0.60 0.006 0.001 0.06 0.08 0.67

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 bisp henol 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 iden tified in various environmental media (see Tables 16 and 17).

In a limited study, Wilson et al. (2003) repor ted aggregate exposure estimates for nine children aged 2-5 year olds from concentra tions of bisphenol A reported in indoor and ambient air, floor dust, play area soil and so lid and liquid food collected at two daycare
centres and nine households in North Caro lina. 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% abso rption, 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 m 3 /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 Caro lina 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 maximu m of 775 ng/kg-bw per day was reported in Ohio (an assumption of 50% absorption was made).

54Concentrations in biological tissues and fluid have been reported by a number of investigators and are summarized in Vandenbur g et al (2007), Dash et al. (2006) NTP (2007) and Dekant and Volkel (2008). Tissues and flui d analyzed include semen, amniotic fluid, placental tissue and serum. Ur ine is considered the most appropriate body fluid for quantifying exposure to bisphenol A. Several studies have analyzed bisphenol A in urine. Most recently, bi sphenol A (free and conjugate d) 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 bisphenol A analysis (free and conjugated) in 2517 individuals aged 6 ye ars and older in the United States. Using these data, Lakind and Naiman (2008) have derived estimates of daily intake of bisphenol A, assuming steady st ate 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 de rived 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 (t he 95th percentile exposure estimate is 274.2 ng/kg-bw per day). For children aged 6 to 11, the 50th percentil e exposure estimate is 67.4 ng/kg-bw per day (the 95th percen tile 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 expos ure estimates based on concentrations in different media, but the absence of biom onitoring data for a key segment of the population (infants) precludes use of bi omonitoring data to characterize risk. Additionally, applying precaution, deriva tion of exposure estimates based on concentrations in media captures exposu res 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.

Health Effects Assessment

The European Chemicals Bureau (ECB) has classified bisphenol A 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 (ECB 2003). In general, bisphenol A is consid ered to have low acute oral toxicity in animals (ECB 2003). The European Union reported that bisphenol A is not a skin i rritant, but may cause eye or respiratory irritation in experimental an imals (ECB 2003). Extensive literature is available regarding effects on various organ systems in adult or developing animals; encompassing acute, short-term and long-term durations of exposure and different routes of exposure. A critical review of all hazard data on bisphenol A is beyond the scope of this screening assessment. Key toxicokinetics and metabolism data are summarized; as is the carcinogenicity and genotoxicity literatur e. Analysis of the recent literature, and recent assessments by other organizations, confirmed that reproductive and developmental endpoints were the critical effects, and the focus of the screening
assessment is therefore on these effects.

55Toxicokinetics and metabolism

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 le vels of free bisphenol A. Following oral exposure, the glucuronidation of bisp henol A by hepatic uridine diphosphate glucuronyltransferase (UDPGT), appears to be a major in vivo metabolism pathway in rats (specifically the isoform, UGT2B1) (Po ttenger et al. 2000; Yokota et al. 1999), nonhuman primates (Kurebayashi et al. 2002), and humans (Volkel et al. 2002). The UGT isoform(s) which may be involved in metabo lism 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 pa thway of bisphenol A metabolism in adult humans and animals while sulfation was sugge sted as a minor pathway in humans and rats (EFSA 2008). The United States F ood and Drug Administration (USFDA) also recognized glucuronidation as a major meta bolic 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, bis phenol A administered orally enters firstpass metabolism in the gut wall and the liver and is quickly metabol ized to bisphenol Amonoglucuronide, 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 bi sphenol A, it is expected that very low concentrations of free bisphenol A would be available for receptor binding in humans (EFSA 2006).

In rats, bisphenol A administered orally al so undergoes glucuronida tion, but the resulting bisphenol A-glucuronide is excreted from liv er into bile. In the gut, the bisphenol Aglucuronide is then cleaved into bisphenol A and aleuronic acid and bisphenol A is
reabsorbed into the blood. This enterohepatic recirculation ma y cause slow elimination of bisphenol A in rodents (Potte nger et al. 2000; Snyder et al. 2000). An overview of toxicokinetics or pharmacokinetics of bisphe nol A by the EFSA and USFDA elaborated the similar stages of bisphenol A metabo lism in rodents, non-human primates and humans (EFSA 2008; USFDA 2008).

Additionally, the toxicokinetics and metabolis m 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 or gans following a single oral exposure to a high dose of 1 g/kg, reaching th e fetus by crossing the placenta within 20 minutes of exposure; the fetal bisphenol A concentration exce eded that of maternal blood 40 minutes after exposure (Takahashi and Oi shi 2000). Also, in humans, bisphenol A was present in serum and follicular fluid of pr egnant women, as well as fetal serum and amniotic fluid, demonstrating placental tr ansfer. Bisphenol A (presumably bound and free) concentrations in amni otic fluid, mid-term, were 5- fold higher as compared to

56serum and follicular fluid (Ikezuki et al. 2002). These data indicate that repeated maternal exposures could lead to an accumulation of fe tal circulating levels of free bisphenol A and correspondingly elevated in utero exposures (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 (Schönf elder et al. 2002). In women, bisphenol A has been detected in biological fluids/tissu e including breastmilk, colostrum and adipose tissue in parts per bill ion 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 sensit ive population was previ ously identified as a research need and has also recently been id entified as being of cr itical 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 suggest ed that an increase in glucuronidation capacity during pregnancy may result in neglig ible fetal exposure to xenobiotics (EFSA 2008). Human data show that glucuronide enzyme activity may be induced during pregnancy as it was evidenced by an increa sed clearance of orally administered lorazepam, paracetamol and lamotrigine in pr egnant women (Miners et al 1986; Papini et al 2006; Pennell et al 2008). It is important to note that ba sed on the limited evidence of selective glucuronidation and c onsidering other factors incl uding individual variation, tissue localization, and substrat e-specificity of these enzyme s 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 liv er microsomes, most of the UGT isozymes were not expressed prenatally. A full compleme nt was present at three months after birth; however, levels were substantially reduced co mpared to the adult (Coughtrie et al. 1988). Studies indicate that, in humans , sulfotransferase (SULT) en zymes are fully functional at birth and it has been suggested that sulf oconjugation of bisphenol A may overcome the absence of or very low glucuronidation activ ity in newborns (Ben edetti 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 com pounds 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. Wh ether the sulfotransferase activ ity 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-mor tem human (fetal, term and adults) liver

57microsomes, 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 comp ounds (Leakey et al 1987).

It is therefore not clear, based on the data of Leaky et al. (1 987) whether sulfotransferases would act as a major metabolic enzyme syst em in human neonates following exposure to bisphenol A. Together, the toxicokinetics 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 (Wel shons et al. 2006) and their inability to efficiently gluc uronidate bisphenol A (Ring et al. 1999; Coughtrie et al. 1988).

Conversely, some evidence suggests that ther e were no significant differences between pregnant (at different stages of gestation) and non-pregnant Sprague-Dawley rats in the toxicokinetics and metabolism of orally ad ministered 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 metabolis m 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 pharmacokinetics in neonatal to weanli ng rats (Domoradzki et al. 2004).

In the absence of fully developed physio logically-based pharmacokinetic models to quantitatively characterize absorption, di stribution, metabolism and excretion of bisphenol A and its metabolites in test anim als and humans via different routes and at different life stages, it is cons idered appropriate to consid er the pregnant woman/fetus and infant as potentially sensitive subpopulations.

Carcinogenicity

A lifetime (103-week) bisphenol A chronic t oxicity/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 wh ich was correlated with reduced feed consumption. A marginal increase in the incide nce of leukemia and te sticular interstitial
tumors in male rats was considered statistical ly non-significant due to failure to meet the Bonferoni inequality criterio n; the incidence of both tumor types was more common in the high-dose groups. Male and female B6C3 F1 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 fe male 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 in cidence of leukemia and lymphoma in male mice, which was not considered dose-related. Ov erall, it was concl uded that there was no

58convincing evidence that dietary exposure to bisphenol A causes car cinogenicity in rats or mice of either sex (NTP 1982). Some de bate exists regarding the biological significance of the NTP (1982) findings. Sp ecifically, it has been argued that the marginal increases in leukemias of rats a nd the combined incide nce of lymphomas and
leukemias in male mice should be considered to be bisphenol A related. It has been proposed as well that the observe d testicular effects should al so be considered as some evidence for carcinogenic activity (Huff 2001; 20 02). However, others maintain that the 2-year study correctly concl uded that the results demons trate 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 bi sphenol 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 assess ment 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 eviden ce suggests that bisphenol A is not a carcinogen following adult expos ure (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 os motic 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 de velopment, may increase susceptibility to neoplastic transformation in the prostate a nd 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 dem onstrate that early bisphenol A exposure, acting in dependently, 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 carcinogene sis (Keri et al. 2007

Genotoxicity

In vitro genotoxicity studies demonstrated a l ack of mutagenic activity; however, it was noted that bisphenol A has aneugeni c potential and i nduced morphological
transformation and micronuclei formation in the absence of metabolic activation in cultured Syrian hamster embryo cells and Ch inese hamster V79 cells, respectively (NTP 2007; ECB 2003; Haighton et al. 2002). In c ontrast, 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 clas togenic potential of bisphenol A (ECB 2003; Haight on et al. 2002). Using the 32 P-postlabeling technique, bisphenol A 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 la ck of significant tumour induction in animal studies, the European Union concluded that “it does not appear th at bisphenol A has significant mutagenic potential in vivo ” (NTP 2007; ECB 2003).

59The aneugenic potential of bisphenol A was further investigated following the observation of a spontaneous increase in oocyt e meiotic abnormalities in mice attributed to inadvertent exposure to bisphenol A from damage to caging materials (Hunt et al. 2003). The authors subsequently investigated the impact of short-term bisphenol A exposure on meiotic processes during oogenesi s. Oral bisphenol A 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% a nd 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 bi sphenol A on female and male germ cells and somatic cells did not confirm aneugenic poten tial. There was no evidence of increased hyperploidy or polyploidy, a lterations on the timing of meiotic progression or micronuclei induction following bisphenol A e xposure via drinking water (7 daily doses of 0.04 mg/kg or for 7 weeks at a concentra tion of 0.5 mg/L) to C 57Bl/6 female mice or via gavage (0, 0.002, 0.02, 0.2 mg/kg) to (102/Elx C3H/E1) F1 male mice (Pacchierotti et al. 2008). Contrary to the induction of aneupl oidy, it has been suggested that oocytes from bisphenol A 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 bisphenol A has no significant aneugenic potential (ECB 2003).

Reproductive and Developmental Toxicity

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; Nakani shi et al. 2007; EFSA 2006; ECB 2003). Reproductive and developmental endpoints were examined following dietary exposure to bisphenol A in rats (0.001, 0.02, 0.30, 5, 50 a nd 500 mg/kg-bw per day) and mice (0, 0.003, 0.03, 0.3, 5, 50, 600 mg/kg-bw per day) (T yl et al. 2002; 2007). NOAELs of 5 mg/kg-bw per day and 50 mg/kg-bw per day we re 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 ef fects was based on significantly reduced body weight and reduced body weight gain, reduced absolute and increas ed relative organ weights in rats, and an increas ed incidence of minimal to m ild 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 dela yed acquisition of preputial se paration, as well as a slight
increase in the incidence of undescended testes in weanling male mice. No changes of toxicologic significance were seen with in the dose range of 0.001-5.0 or 0.003-5.0 mg/kg-bw per day in adult or developing ra ts and mice respectively (Tyl et al. 2002; 2007). The three-generation reprodu ctive toxicity study by Tyl et al. (2002) in rats did not include an estrogenic reference control ag ainst which the negative results could be

60evaluated. However, the most recent multig eneration report in mice incorporated a dietary 17 β -estradiol reference group (Tyl et al. 2007). Developm ental toxicity in CD-1
mice and CD rats has been demonstrated to occur following pren atal exposure during major periods of organogenesis (gestational da ys 6-15), through gava ge, at high doses of bisphenol A. Measures of developmental toxicity included percentage resorptions, number of live fetuses per ratio, sex rati o, average fetal body we ight per litter and percentage of malformed fetuses. No de velopmental effects were observed following administration of 640 mg/kg-bw per day (highe st 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. Bisphenol A wa s not teratogenic at any of the administered doses (Morrissey et al. 1987).

Recent studies have shown that chemical s that alter endocrine function, including bisphenol A, can alter the cell function at levels below NOAELs identified in these standard toxicology studies (reviewed in We lshons 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 mani fested following a linear dose-response relationship and, in several instances, have been found to follow a non-monotonic response curve. Bisphenol A has been cons idered as a weak environmental estrogen, based on traditional bioassays, as it binds to the estrogen receptor alpha and beta (ER α and ER β ) (Gould et al. 1998; Kuiper et al. 1998; Pennie et al. 1998) with affinity, which
is about 10 000 to 100 000-fold weaker than that of 17 β -estradiol (Vandenberg et al. 2007). A mode-of-action analysis is beyond th e scope of this screening assessment; however, exposure to low doses of bisphe nol A has been demonstrated to produce disruptive effects in endocrine organs in cluding 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 bisphenol A can stimulate the af orementioned cellular responses at very lowconcentrations through genomic (nuclear estrogen receptor) or non-genomic (membraneassociated or intracellular transduction) mech anisms, with effects on cellular function at doses as low as 1 pM or 0.23 pg/ml (A nway et al. 2005, 2006; Wozniak et al. 2005, Zoeller et al. 2005) (reviewed in Wetheri ll et al. 2007). Specifically, it has been
demonstrated that some of the effects of bisphenol A may be mediated through the cell surface estrogen receptor (GPR30) being equipotent with 17 β -estradiol and diethylstilbestrol (Alons o-Magdelena et al. 2005).

There are many studies in the peer-reviewe d literature that describe effects in experimental animals resulting from exposur e to bisphenol A administered at doses
below the NOAELS established in standard toxicological studies for systemic and reproductive/developmental endpoints. Effect s observed include changes in prostate growth and development, mammary gland or ganization, protein induc tion 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; Richte r et al. 2007; Willhite et al. 2008). For purposes of the screening assessment, the pr imary focus will be placed on identification

61of alterations considered potentially ‘advers e’ 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 toxic okinetics in adults are not likely conserved in neonates, those studies de livering bisphenol A during feta l development and/or early postnatal periods via the subcutaneous (sc) in jection route were evaluated for the current assessment (Kato et al. 2003; 2006; Ho et al . 2006; Honma et al. 2002; Nakamura et al. 2006; Suzuki et al. 2002). The neonatal data, us ing sc injection perm its the evaluation of possible effects of ecposure duri ng the early stages of develo pment. While informative,
these studies were not the basis for hazard characterization of bi sphenol A for human
health.

A number of reproductive and developm ental 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 repro ductive function and developmental endpoints. Prenatal exposure from gestational day 11 to 17, of CF-1 mice to 0, 2 or 20 μg/kg-bw per day bisphenol A admi nistered 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 re sulted in increased postnatal growth in both males and female s and early onset of sexual maturation in female offspring, observed following exposur e 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 th e 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 th e urethra was also observed (T imms et al. 2005). Honma et al. (2002), using subcutaneous injections of 0, 2 or 20 μg/kg-bw per day bisphenol A 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 vaginal opening at 20 μg/kg-bw per day. However, female reproductive function wa s not affected at the first breeding. Furthermore, oral exposures of pregna nt 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 bisphenol A-dependent effects on androgen biosynthesis by adu lt leydig cells occurred as a result of targeting a sensitive period during perinatal development. Chronic exposure from weaning at postnatal day 21 to adu lthood, representing adolescent exposure, also suppressed androgen biosynthesis by adult leydi g cells at doses as low as 2.4 μg/kg-bw per day bisphenol A (Akingbemi et al. 2004) . A LOAEL for the re productive effects of bisphenol A was observed to be 2 μg/kg-bw per day in the aforementioned subset of studies. A non-monotonic (inverted-U) dose-response was observe d in several instances.

Conflicting results have been described in various additional repor ts. Oral bisphenol A exposure to 0.2, 2, 20 or 200 μg/kg-bw per da y from gestational day 11 to 17 did not produce reproductive toxicity in studies assess ing outcomes in CF-1 mice (Cagen et al.

621999a; Ashby et al. 1999). Nor was there an effect on sperm density or male reproductive organ development following bisphenol A expos ure 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), th e responses of the positive control did not differ from either the negative control or bisphenol A. There were no treatment-related effects on progeny following maternal drinking water exposure to 0.01, 0.1, 1.0 or 10 ppm bisphenol A among female Wistar albino rats . The authors conclude that bisphenol A should not be considered a selective reproducti ve or developmental toxicant (Cagen et al. 1999b). A 2-generation reproductio n study using Crj: CD(SD) IGS rats administered bisphenol A orally at doses between 0.2 and 200 μg/kg-bw per day also found no significant adverse treatment-related ch anges in the multiple reproductive and developmental endpoints assessed (Ema et al. 2001); this study al so lacked a positive control for interpretation of negative data . Most recently, assessing the impact of bisphenol A exposure in the male Long Evan s 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 numb er of experimental variables may account for divergent results in response to bisphenol A exposure at lo w doses: specifically, species and strain differences; variable responses depending on the tissues and/or endpoint(s) assessed; variability in feed with resp ect to amounts of estrogenic co ntaminants; inappropriate use or lack of positive controls; as well as the consideration of exposure-related effects that present in a non-monotonic dose-response cu rve (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 window s 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.

Developmental Neurotoxicity

Recently, expert panel assessments of the curr ent 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 th ere was “some” concern that pregnant women, fetuses, infants and ch ildren 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 re garding neural and behavioural effects consistently produced more “ positive” studies, as compar ed to the reproductive and developmental studies, although the endpoints assessed differed betw een investigators. The level of concern was expr essed relative to estimates of exposure for the general population in the United States (NTP 2007). Recently, the NTP finalized its evaluation

63of 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 dur ing 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 wh ere 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 th e screening assessment it was considered appropriate to further consider the weight of evidence associated with neurodevelopmental and behaviour al studies investigating e ffects 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 be ginning in early embryonic development and continuing through different life stages. Preci se 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
has been conducted to evaluate the effect of low dose bisphenol A exposure. Some tests were among those recommended by the recently adopted OECD Developmental Neurotoxicity Study Guideline # 426 (OEC D 2007). A number of developmental neurotoxicity endpoints were examined, incl uding locomotor and exploratory activity, learning, social, sexual and maternal beha viour, as well as brain morphology, with a range of exposures from 0.2 to 400 000 μ g/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 developmen t and behaviour effects of bisphenol A in rodents showed that in utero, perinatal or postnatal exposure to bisphenol A, 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 μ g/kg-bw per day (EFSA 2006; Willhite et al. 2007), may produce some effects on neur obehavioral development in fetal or early postnatal life in rodents. The lowest observe d effect in the neurodevelopmental literature

was 0.02 μ g/kg-bw per day based on changes in estrogen receptor mRNA expression, however, no effects were observed at 2 μ g/kg-bw per day; this is suggestive of an inverted U shaped (non-monotonic) dose res ponse phenomenon (Nishizawa et al. 2003; 2005a; 2005b).

While this inverted non-monotonic dose-res ponse relationship is considered to be mediated by both positive and negative f eedback effects on estrogenic receptors (Welshons et al. 2003; 2006) or related to physiological responses to estrogenic activity of bisphenol A 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 replic ation of perturbations at le vels of exposure below those
demonstrated to have no effects. Furthermor e, it has been noted that the shape of the dose-response curves for effects of vari ous estrogenic compounds differs depending on

64the endpoint assessed and the dosing regime a pplied (Melnick et al. 2002). As such, the significance of the LOEL of 0.02 μg/kg-bw per day for bisphenol A to human health risk assessment, however, is difficult to ascertain, par ticularly in light of the fact that it was the only dose level tested below the NOEL of 2 μ g/kg-bw per day and some of the effects were not consistently observ ed 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 pos tnatal developmental period for newborns and infants is unknown. An additional study that examined substance-induced changes in brain receptors noted a fluctuating increas e in the number of neurons expressing ER α and

ER β after oral administration of 2 μg/kg-bw per day bisphenol A to ICR mice from gestational day 11 to 17. Increases were obser ved 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 bisphenol A from postnatal day 23–30, th ereby targeting the period of juvenile development, resulted in a transien t and gender-specific increase in ER α levels in the ventromedial nucleus of the hypothalamus. Ch anges were noted at postnatal day 37, but not at postnatal day 90, in females only (C eccarelli et al. 2007). Although these effects may be considered biomarkers of exposure to bisphenol A, and potenti al 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 e ndocrine-active substances [reviewed in Patisaul and Polston 2008]. Following subcutan eous injection of 100 mg/kg-bw per day to Sprague-Dawley rat pups on postnatal day 1 and 2, no bisphenol A-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 wa s observed (Patisaul et al. 2006; 2007). Although bisphenol A was show n to disrupt neuronal phenotype by demasculinization of tyrosine hydroxylase (TH) immunoreactive cells and defeminization
of ER α /TH double labelled cells in male and fe male 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 alwa ys translate into disruption of neuronal function [reviewed in Patisaul and Polston 2008]. Examination of GnRH neurons as an
index of feminization of the AVPV is an indirect means to determine bisphenol Ainduced alterations in this sexually dimorphic region of the brain. Assessing the ability of exposed females to generate an LH surge w ould have been more informative regarding
AVPV feminization. Oral administration of 12 – 60 mg/kg- bw per day bisphenol A 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 mesencephali c dopamine transporter expression (Ishido et al. 2007). Oral exposure to doses of bisphenol A ranging from 3.2 to 320 mg/kg-bw per day also failed to induce volumetric ch anges to the SDN-POA in female Sprague Dawley rats exposed during gestation and la ctation (Kwon 2000). A reversal of gender
differences in the volume of the locus coerul eus, a nucleus in the brainstem involved in stress and panic responses, was observed fo llowing administration of drinking water at

650.03 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 cort icotrophin 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 dri nking water to 2.5 mg/kg–bw per day bisphenol A (Funabashi et al. 2004). Homologous sexuall y 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 immunohistoche mical investigations are key for providing mechanistic information, and are useful biomar kers, but relevance to human health risk
assessment is uncertain. This is consistent wi th interpretations in the draft updated Risk Assessment of bisphenol A by the European Union (ECB 2008).

Behavioural testing was used to character ize the effect of bisphenol A exposure on general locomotor and exploratory activity, an xiety and learning and memory in rodents. At 10 μg/kg-bw per day, CD-1 mice expos ed 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 (Laviola et 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 bisphenol A at doses which are well below established NOAEL of 50 mg/kg-bw per day for reproductive/developmental effects may lead to organizational e ffects during critical periods of brain development in rodents. Ho wever, these studies, though, were limited to a single exposure level precluding the eval uation of dose-respons e. The lowest dose leading to bisphenol A-induced organizational effects in the brain was 10 μg/kg-bw per day in CD-1 mice. It should be recognized that studies were c onducted using the same outbred strain of mice (CD-1), the same e xperimental dosing prot ocol (single dose, no positive control) and by the same group of re searchers in one research institute.

Effects of gestational exposure to 40 μ/kg-bw per day bisphenol A have been noted in rats, including statistically significant gender-specific changes in sexual performance (Farabollini et al. 2002), effects on active and passive mate rnal behaviour of treated rats (Della Seta et al. 2005), marginally diminish ed intergroup differences in pain response (Aloisi et al. 2002), altered novelty seeking and impulsive behavi our in females and males, respectively (Adriani et al. 2003), as we ll as altered play behaviour in female rats (Porrini et al. 2005). In addition, in Sprague-D awley rats, pubertal exposure to oral doses
of 40 μ g/kg-bw per day altered behaviour of male rats (Della Seta et al. 2006). These studies, again conducted by a co mmon group of researchers, provide convincing evidence of bisphenol A-induced effects at 40 μ g/kg-bw per day and illust rate the implementation of a comprehensive approach. Replication of the aforementioned results by independent
researchers is needed.

66The effect of bisphenol A exposure on open field behaviour (i.e., an assessment of general locomotor and exploratory activity, grooming and anxiety) (ECB 2008), was not consistent among the available studies. Kubo et al. (2001;2003) administered bisphenol A to Wistar rats during gestati on 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. Rear ing behaviour, indica tive of exploratory behaviour, observed by Fujimoto et al. (2006), also reflected opposing effects in male and female Wistar rats following 15 μ g/kg-bw per day bisphenol A from gestational day 13 to
birth. Increased duration of r earing 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 asse ssing the effects of oral bisphenol A exposure duri ng the prenatal and neonatal periods with doses ranging
from 40 μ g/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 or al bisphenol A exposure (0.2, 2, 20, 200 μ g/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, th is large study, though, di d not provide data from behavioural measures preventing asse ssment of the relationship between exposed and control groups.

Some studies have investigated the effect of prenatal and neonatal bisphenol A exposure on anxiety-related behaviour using the hole board test, elevated maze test, forced swimming test and/or light-dark preference chamber (Farabollini, 1999; Negishi et al. 2004; Fujimoto et al. 2006; Ryan and Vande rbergh 2006). Farabollini et al. (1999) found

that oral administration of 400 μ g/kg-bw per day bisphenol A from gestational day 14 to postnatal day 6 increased anxiety-related beha viour in male and female rats using the hole board test. However, decr eased anxiety-related behaviou r was noted in males using the elevated plus maze. No effects we re evident following administration of 40 μ g/kg-bw per day. Ryan and Vanderbergh (2006) reported increased anxiety in females (males not tested) after admi nistration of 200 μ g/kg-bw per day bisphenol A from gestational day 3 to postnatal day 21, but not after 2 μ g/kg-bw per day, in the light-dark preference chamber. No differences in spatial memo ry 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 bisphenol A following drinking water exposure to 15 μ g/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 μ g/kg-bw per day bisphenol A from gestational day 3 to postnatal day 20 (m ales not tested). Together, these findings illustrate the variable responses observed ac ross gender and species in anxiety-related behavioural endpoints following a ra nge of bisphenol A exposures.

Further research into the e ffects of bisphenol A exposure focused on disturbances in cognitive function. Of those studies examin ing bisphenol A-induced effects on learning and memory, no effects were observed by wa ter-filled T-maze following gestational and

67lactational exposure to 0.2, 2, 20, 200 μg/kg-bw per day bisphenol A (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). In terestingly, in the same study, one week of bisphenol A exposure (gestati onal day 13 to postnatal day 0) altered the normal sex differences in struggling behavi our in the forced swimming te st observed in control rats. Exposure increased the immobility of male rats in the for ced swimming test, which is indicative of depression-like behaviour called behavioural despair. Male rats appeared to be more sensitive to the effects of bisphenol A on depressive response (Fujimoto et al. 2006). Kubo et al. (2001), assessing the impact in rats of 1.5 mg/kg-bw per day bisphenol A, 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 exposure results 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 bisphenol A from gestational day 10 to postnatal day 20. Res ponses were variable and did not follow a dose-related pattern. Further assessment of ma le F344 rats exposed to 100 μg/kg-bw per day from gestational day 3 to postnatal da y 20 showed fewer avoidance responses in
trials 1 to 3 of 4 in the active avoidan ce;. However, no differences were observed between bisphenol A-exposed and control group s in passive avoidance (Negishi et al. 2004). Carr et al. (2003) exposed F344 rats to 100 or 250 μg /kg-bw per day bisphenol A 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 bisphenol Ainduced behavioural alterations in response to pharmacological challenge; specifically, D-amphetamine (Laviola et al. 2005; Adrian i et al. 2003), methamphetamine- (Suzuki et al. 2003) or morphine-induced (Mizuo et al. 2004a; 2004b; Narita et al. 2006; 2007) behavioural changes. These stud ies, using standard testing paradigms, provide evidence for altered stimulated responses following bi sphenol A administration and are outlined in Appendix D. Prenatal exposure of CD-1 mi ce from gestational day 11 to 18 with 10 μg/kg-bw per day bisphenol A eliminated D-amphetamine-induced conditioned place preference in females suggesting exposure-relat ed 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). Pharmacol ogical challenge studies are suggestive of potential organizational alterations of the ne ural system following perinatal bisphenol A exposure. The details of additional studies are not elaborated on as th e doses that resulted in significant effects on the neurochemical systems were above the established NOAEL of 50 mg/kg-bw per day for reproductive/deve lopmental effects; altered behaviour was observed following administration of 250 or 400 mg/kg bw per day.

68Weight of Evidence Summary of De velopmental Neurotoxicity Findings

Several organizations have recently evaluate d 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 la ck of consistency in the resu lts of behavioural testing, no
conclusions can be drawn from these studies ” (ECB 2008). Other jurisdictions concluded that the scientific evidence cited supports th e conclusion of some concern for neural and behavioural effects of bisphe nol A in potentially vulnerabl e populations (NTP 2007; NTP 2008).

It was considered appropriate to characterize the weight of evidence supporting
neurobehavioral effects in r odents following low-dose exposures to bisphenol A from the following perspectives: rigour, power, co rroboration/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 bisphenol A on neurobehavi oral endpoints used acceptable methodology. All studies were published in the peer-reviewed literat ure. One study (Ema et al, 2001) declared compliance with good la boratory practices (GLP). However, the rigor of several studies was limited by sma ll sample size, incomplete reporting and inadequate statistical analysis. Additionally, on ly a few studies reporte d use of concurrent positive controls. There were few common endpoi nts across studies and it was difficult to evaluate whether related endpoints that are indicative of anxiety, hyperactivity, avoidance, and so forth, respond in a simila r manner. As such, the overall rigour of the
neurobehavioral dataset is considered limited.

The statistical power of the studies compri sing 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 sufficien t power to detect su btle 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 neurobehavio ral 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, preclud ed 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 unde r relevant conditions), the limited data preclude characterization of consistency with regard to the qualit ative direction of neurobehavioral changes regardless of the stat istical significance of reported alterations. Furthermore, apparent inconsistencies in re sults may be a function of design and/or the behavioural bioassays used to evaluate th e potential developmen tal alterations (e.g., different tests of the same underlying f unctions, age of testing). For instance,

69inconsistencies in the effects of bisphe nol A on anxiety-related behaviours should consider that holeboard, light-dark test (e mergence test) and elevated-plus mazes all provide measures of anxiety but are also differentially sensitive to other concurrent
neurobehavioural effects such as motor eff ect. The same applies to active and passive avoidance tests (active avoida nce is more susceptible to motor effects). Thus, the apparent absence of a consistent pattern of e ffects may indicate that other factors (such as motor effects) may play a role in the discre pant results. As such, a lack of consistent effects when comparing these measures does not indicate that bisphe nol A does not affect anxiety per se. In addition, not only does the comparison between hole board and elevated plus maze appear to include differe nces in the actual test, but the doses and dosing regimes used in these tests were different: doses of 15, 100 or 400 μ g/kg per day bisphenol A with exposure peri ods from gestational day 13 to birth, gestational day 3 to postnatal day 21 or gestational day 14 to postn atal day 6, respectively. Variations in the dosing period, especially pe rinatal dosing, can itself have a substantial impact on subsequent effects. As such, there are sufficient methodological differences between these studies to preclude draw ing firm conclusions regarding lack of effects of bisphenol A. A consistent finding among studies was th at exposure during gest ation and/or during early postnatal life consistently represented pe riods of sensitivity to bisphenol A-induced alterations. Additionally, a cons istency in the reported effe cts 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 disrupted by bisphenol A, and moreover, are those most likely to be detected given the measurements used. The degree of corrobor ation demonstrated in the dataset is
considered limited.

Finally, biological plausibility is an important consideration when considering the weight of evidence. Biological plausibi lity is informed by investig ations of mechanism(s) of action. Bisphenol A produces estrogenic effect s via interactions with estrogen receptors (ER). The estrogen receptor family includes the classic ER α , ER β , and their splicevariants, extranuclear estrogen receptors a ssociated with cell membranes and neuronal synapses (Woolley et al. 2007). Bisphenol A was found to have limited binding affinity
with ER α , but it showed a nearly tenfold higher affinity for ER β . Specifically, the relative binding affinity (RBA) of bisphenol A for ER α and ER β 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 ce ll membrane ER (Quesada et al. 2002). It is possible that bisphenol A preferentially affects ER β -containing tissues including the ovary, cardiovascular system and brain (H arris 2007) and those with cell membrane receptors. Importantly, estrogen is a key re gulator of organizational and activational processes in the brain of both males and females

. Note that the relative co ntribution of estrogen and androgen to sexuall y dimorphic behaviours has b een demonstrated to differ between species. Specifically, it is genera lly held that the estrogen-signalling pathway predominates in rodents, while the an drogen-signalling pathway plays a more predominate role in humans and non-human prim ates (Li et al. 2008). Thus, the relative significance of the estrogenic m ode of action of bisphenol A in rodents with respect to human relevance requires further investigation.

70Typically, demonstration of a dose-response re lationship supports bi ological plausibility. The dose-response relationship following expos ure to bisphenol A at doses below the established NOAEL of 50 mg/kg-bw per day fo r 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 bisphenol A. Regarding the neurobehavioral dataset, a very limited
number of studies assessed more than one dose. Therefore, a dose -response relationshipcould not be established. An attempt was made to compare and evaluate bisphenol Arelated changes at similar doses between si milar experiments. However, this was not possible due to a number of the aforementioned design limitations. In a few studies which included multiple doses, bisphenol A-related eff ects 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 asso ciated with bisphenol A expos ures in the μg/kg-bw per day range. The weight of evidence for biol ogical plausibility is considered limited.

Overall, taking into consideration rig our, power, corroboratio n/consistency, and biological plausibility/coherence, the wei ght of evidence supporting neurobehavioral effects in rodents following exposures to bi sphenol A at exposures below established NOAELs for reproductive/developm ental toxicity is limited.

An emerging trend from the developmental neur otoxicity studies is th at prenatal and/or neonatal bisphenol A exposure may cause organi zational effects in the brain leading to disruptions in the sexually dimorphic behavi ours 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 indica tes an adverse effect (Li et al. 2008). Detailed mechanistic info rmation can reduce uncertainty regarding the relevance of rodent findings to humans, but is not crucial for genera l toxicity screening. Thus, as highlighted by the NTP (2007), it is not clear whether the reported changes
constitute critical adverse t oxicological responses, however, the uncertainties associated with the results from neurodevelopmental a nd behavioural testing warrant concern with respect to human health, in particular during certain sensitive life stages.

Epidemiology Investigations

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

71Confidence in the toxicological database, overall, is considered moderate. Data for acute toxicity, repeated dose toxicity, genotoxi city, carcinogenicity and reproductive and developmental toxicity are ava ilable. Although there is an ex tensive database from which to evaluate the effects of bisphenol A e xposure, inconsistencies and experimental limitations exist among studies evaluating re productive and developmental endpoints. Additionally, several of the st udies administered a single dos e precluding characterization of dose response.

Characterization of Ri sk to Human Health

Consistent with the assessment of th e European Commission (ECB 2003), which classified bisphenol A as a Category 3 reproduc tive 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 populati on of Canada range from 0.08 μg/kg-bw per day to 4.30 μg/kg-bw per day. Specific exposur e estimates for the most highly exposed subpopulation (i.e., infants) ra nge from an average of 0.27 μ g/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) fo r infants aged 0 to 1 month. The NOAELs from the multigeneration reproduc tive toxicity studies in Sprague- Dawley rats and CD-1 mice of 5 mg/kg-bw per day for systemic eff ects (reduced body weight gain in rats andminimal to mild hepatocyte hypertrophy in adult male and female mice) and 50 mg/kgbw per day for reproductive and developmen tal toxicity (Tyl et al. 2002; 2007) are considered an appropriate depa rture point for characterizing risk to human health from exposure to bisphenol A. Based on the NOA EL of 5 mg/kg-bw per day, margins of exposure for infants exposed to bisphenol A fr om 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 consumpti on 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 segmen ts of the populati on, exposed via the consumption of canned food, use of poly carbonate bottles and environmental media (Table 21), margins of exposure to the NOA EL 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 th ese margins are considered adequate to address interspecies and intraspecies differe nces 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 wei ght of evidence analys is, is suggestive of potential effects at doses well below these NOAELs.

Specifically, a small number of studies in CD -1 mice reported altera tions in behaviour following oral administration of 10 μg/kg-bw per day. These include altered maternal behaviour in dams exposed either as fetu ses or during adulthood and decreased or

72eliminated sex-related behavioural differen ces following gestational or gestational and lactational exposures (Palanza et al. 2002; Lavi ola et al. 2005; Gioi osa et al. 2007). At 40 μg/kg-bw per day, a small number of studies in rats have reported behavioural effects including gender-specific cha nges in sexual performance, effects on active and passive maternal behaviour, and altered novelty seeki ng and impulsive behaviour in both sexes in
adults whose mothers were administered bisphenol A during gest ation and lactation (Farabollini et al. 2002 ; Della Seta et al. 2005; Adrian i et al. 2003). Further supporting the effect of bisphenol A exposure on alte red behaviour, oral administration of 100 μg/kg-bw per day in rats dur ing gestation and lactation re duced the number of correct avoidance responses (Negishi et al. 2004), wh ile early postnatal e xposure disrupted the normal gender-dependent pattern of maze pe rformance acquisition in the Morris water maze (Carr et al. 2003).

While collectively these studies provide evid ence that exposure to bisphenol A 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 lim itations 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 studi es involve use of a single dose, lack of dose response relationship). These limitations make it diffi cult to determine actual significance of findings to human health risk assessment

The neurodevelopmental and behavioural data set 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 toxicokinetics and metabolism data indicate potential sensitivity to the pregnant woman/ fetus and infant; and that animal studies suggest a trend towards heightened suscepti bility during stages of development in rodents, it is considered appropriate to apply a precautionary approach when characterizing risk. As such, it is concluded th at bisphenol A be cons idered as a substance that may be entering the environment in a qua ntity or concentrati on or under conditions that constitute or may constitute a danger in Canada to human life or health.

Uncertainties in Evaluation of Risk to Human Health and

Identification of Research Needs

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

• Uncertainty in the estimation of exposu re to all age groups from the use of polycarbonate repeat-use cont ainers 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.

73• fants from the consumption of breastmilk due to lack of surveys measur ing concentrations in breastmilk from
of

or

t

• vels of bisphenol A in polymers used to manufacture consumer products, such as e poxy adhesives, toys and cosmetics, and

• inta ke of bisphenol A from the consumption
of fresh fish and other fresh foods due to the lack of surveys measuring

• ctive/developmental effects below the established NOAEL of 50 mg/kg-bw per day for these effects in rodents is an , and

• t regarding toxicokinetics and metabolism
differences between rodents and humans. R odents clear bisphenol A in the faeces via

to

• itive populations. Actual levels of free
s with Uncertainty in the estimation of exposure s to in

Canadian women, or surveys involving larger sample sizes. Further exploration exposures (e.g., food, environmental media) c ontributing to levels in breastmilk other biomonitoring would assist in determ ining which sources of exposure are mos important for risk management, if required.

Uncertainty associated with actual residual le

uncertainty associated with the assumpti ons made to calculate general population exposure from use of these products.

Uncertainty associated with estimating

concentrations in Canada or elsewhere.

The adequacy of the database on repr odu

uncertainty. Limitations with respect to ri gour, power, corrobora tion/consistency biological plausibility/coherence exist.

Interspecies variation. Uncertainties exis

biliary excretion of the glucuronide (E FSA 2006). Humans, on the other hand, clear bisphenol A primarily via urine after gluc uronidation (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; Ring et al. 1999; Coughtrie et al. 1988). Evidence suggests that the human fetus metabolizes bisphenol A less extensively as compared to the rat fetus. Thus, the human fetus may be more prone effects of bisphenol A (Elsby et al. 2001). These important differences in
toxicokinetics and metabolism introduce uncerta inty in extrapolati ng from rodents to humans. Uncertainties also exist regarding re levance to humans of potential effects of bisphenol A on brain function in rodents. Species differe nces 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 trimes ter of pregnancy. Conversely, the human newborn might be compared with a 2- to 3-week-old rat pup (M orreale et al. 2004). The mechanism(s) by which bisphenol A ex erts estrogenic eff ects depends on many factors, including species, a nd this is an additional uncertainty with respect to interspecies extrapolation.

With regards to intraspecies variation, av ailable data indicate that the pregnant female/fetus and infant are potentially sens bisphenol A in the fetus however remains an uncertainty. Uncertainty also exist respect to critical windows of suscep tibility during in utero and postnatal development and potential subsequent long-term effects.

74esearch needs include the following:

rtain envi ronmental media (e.g., air, soil, drinking water) are lacking or limited. Such data would decrease uncerta inty regarding the

• nant women and children younger than 6 year s old) would decrease uncertainty

• ration from polycarbonate would inform risk management. To date, st udies in this area report inconsistent

ase

• es tablish relationships between bisphenol A exposure and health outcomes for sensitive subpopulations.

• may induce alterations in the developi ng reproductive system in males and

• thyroid hormone homeostasis, as well as the impact

•

• nd rdance to OECD guidelines for

• s spects of in utero and early postnatal development R

• Canadian monitoring data for ce

contribution of these media to exposure among the general population.

Biomonitoring data for the general popul ation of Canada (including preg

regarding the exposure assessm ent. Investigations into accuracy of spot urine samples in predicting dail y exposures are needed.

Studies investigating the underlying pro cess of mig

results, with some study investigators indi cating that migration rates will decre with use and others indicating that migr ation rates remain relatively constant throughout the life of the container.

Epidemiological studies are needed to

Further exploration of mechanisms by which low dose exposure to bisphenol A

females.

Further exploration of molecular mechan isms by which low doses of bisphenol A may disrupt the estrogen and on the central nervous system during critical windows of development.

Dose response studies using environmenta lly relevant levels of bisphenol A 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 bisphenol A on th e sexual differentiation of the brain a neurobehavioural endpoints conducted in acco developmental neurotoxicity.

Investigation of the effects of bisphe nol A exposure during precise age window that correspond with sp ecific a
of the neuraxis that are the most critical to the development of neurobehavioural features. For example, sensory and moto r differences are generally brainstem, anxiety-related activity involves mid-br ain and lower telencephalon, speciestypical behaviour patterns are most like ly lower telencephalon and learning and memory differences involve the cortex.

75• ed changes on the reproductive and nervous system during early and late r stages of development including

• identify the doses and ages or windows of sensitivity that are the most effective in pr oducing developmental disturbances.

• exposures via oral or parenteral rout es to characterize the actual versus administered dose.

• carcinogenicity from exposure during sensitive stages of development.

• ect to

Conclusion

Based on the information presented in assessment, and applying a precautionary approach, it is concluded that sphenol A is entering or may be entering ay have

itions Further assessment of bisphenol A-i nduc

adolescence and adulthood.

Further research is needed to

Assessment of tissue levels of bisphenol A following repeat low dose

Comprehensive evaluation of the potential of bisphenol A to initiate or promote

An emphasis on research to establish more precisely the magnitude of eff better understand the impact of bisp henol A exposures is required.

th is screening
bi the environment in a quantity or concentrati on or under conditions that have or m an immediate or long-term harmful effect on the environment or its biological diversity.

Based on the information presented in th is screening assessment, and applying a

recautionary approach, it is concluded that bisphenol A be considered as a substance p that may be entering the environment in a qua ntity or concentrati on or under cond that constitute or may constitute a danger in Canada to human life or health.

It is therefore concluded that bisphenol A does meet the crite ria in paragraphs 64(a) and

4(c) of CEPA 1999. 6

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Yokota H, Iwano H, Endo M, Kobayashi T, Inoue H, Ikushiro S, Yuasa A. 1999. Glucuronidation of the environmental oestrogen bisphenol A by an isoform of UDP-glucuronosyltransferase, UGT2B1, in the rat liver. Biochem J. 340 ( Pt 2):405-409.

Yokota H, Tsuruda Y, Maeda M, Oshima Y, Tadokor o H, Nakazono A, Honjo T, Kobayashi K. 2000. Effect of bisphenol A on the early life stage in Japanese medaka ( Oryzias latipes ). Environ Toxicol Chem 19(7): 1925-1930.

Zhan M, Yang X, Xian Q, Kong L. 2006. Photosensitized degradation of bisphenol A involving reactive oxygen species in the presence of humic substances. Chemosphere 63: 378-386.

Zhang S, Zhang Q, Darisaw S, Ehie O, Wang G. 20 07. Simultaneous quantification of polycyclic aromatic hydrocarbons (PAHs), polychlorinated biphenyls (PCBs ), and pharmaceuticals an d personal care products (PPCPs) in Mississippi river water, in New Orleans, Louisiana, USA. Chemosphere 66:1057-1069.

Zoeller RT. 2005. Bisphenol-A, an environmental contaminant that acts as thyroid hormone receptor antagonist in vitro , increases serum thyroxine, and alters RC3/Neu rogranin expression in the developing rat brain. Endocrinology 146(2):607-612.

99Appendix A. Robust Study Summary – Inherent Toxicity

Item Yes No

Reference : Sohoni et al

. 2001. Reproductive effects of long-term exposur e to bisphenol A in the fathead minnow ( Pimephales promelas ).
Environ Sci Technol 35: 2917–2925.

Test Substance : 80-05-7 (Bisphenol A)

*Chemical composition of the substa nce (including purity, by-products) X Persistence/stability of test substance in test system X

Method

References X *OECD, EU, national, or other standard method? X Justification of the method/protocol if a non-standard method was used X
*GLP (good laboratory practice) X Test organisms : Fathead minnow ( Pimephales promelas )

Latin or both Latin and common names reported? X Life cycle age/stage of test organism X

Sex X Length and weight of test organisms X Number of test organisms per replicate X Food type / feeding periods (acclimation / during test) X Test design/conditions

Test type – acute or chronic: Chronic Experiment type (laborat ory or field) specified? X System type (static, semi-static, flow through)? X Negative or positive controls : Negative X Number of replicates (includi ng controls) and concentrations X Exposure pathways (food, water, both) X Exposure duration X *Measured concentrations reported? X Exposure media conditions (temperature, pH, elect rical conductivity, hardness, total organic carbon, dissolved organic carbon, dissolved oxygen, major cations and anions; other) X

Was pH within 6–9 range? X Was temperature within 5–28°C range? X Photoperiod and light intensity X Stock and test solution preparation X Use of emulgators/solubilizers (especially for poorly soluble / unstable substances) Not required Analytical monitoring intervals X

Statistical methods used X

Results

Toxicity values (LC 50 , EC 50 , or IC 50 - 164 d LOEC (lowest-observed effect concentration – somatic growth in adults, egg hatchability) = 0.64 mg/L; 164 d LOEC (vitellogenin [VTG] sy nthesis, sex cell type proportion) = 0. 016 mg/L; 164 d LOEC (egg production) = 1.28 mg/L Other endpoints reported – e.g., bioconcentration factor / bioaccu mulation factor (BCF/BAF): *Was toxicity value below the chemical’s water solubility? X Other adverse effects (carcinogeni city, mutagenicity, etc. X

Score : Major items – 4/5; overall score – 24/25 (96%) EC Reliability code : 1

Reliability category (high, satisfactory, low) : High

Comments: With respect to the chemical composition of the test substance, the source and purity of bisphenol A (BPA) were not reported; however detailed description of hi gh performance liquid chromatography (HPLC) anal ysis criteria are provided and this is consid ered to satisfactorily confirm the iden tity of the test material.

100Appendix B. Robust Study Summary – Inherent Toxicity

Item Yes No

Reference : Johnson et al. 2005. Endocrine disruption in aquatic and terrestrial invertebrates. Final report produced by WRc NSF Ltd., Marlow, Buckinghamshire for the United Kingdom Department of Environment, Food and Rural Affairs (DEFRA). March 2005.

Test Substance : 80-05-7 (Bisphenol A)

*Chemical composition of the substa nce (including purity, by-products) X Persistence/stability of test substance in test system X

Method

References X *OECD, EU, national, or other standard method ? X Justification of the method/protocol if a non-standard method was used X

*GLP (good laboratory practice) X
Test organisms : Amphipod ( Gammarus pulex )

Latin or both Latin and common names reported? X Life cycle age/stage of test organism X

Sex X Length and weight of test organisms n/a Number of test organisms per replicate X Food type / feeding periods (acclimation / during test) X
Test design/conditions

Test type – acute or chronic: Chronic Experiment type (laborat ory or field) specified? X System type (static, semi-static, flow through)? X Negative or positive contro ls : Solvent control X Number of replicates (includi ng controls) and concentrations X Exposure pathways (food, water, both) X Exposure duration X *Measured concentrations reported? X Exposure media conditions (temperature, pH, elect rical conductivity, hardness, total organic carbon, dissolved organic carbon, dissolved oxygen, major cations and anions; other) X

Was pH within 6–9 range? X Was temperature within 5–28°C range? X Photoperiod and light intensity X Stock and test solution preparation X Use of emulgators/solubilizers (especially for poorly soluble / unstable substances) X Analytical monitoring intervals X

Statistical methods used X

Results

Toxicity values (LC 50 , EC 50 , or IC 50 –: 14 d LOEC (lowest-observed-effect concentrati on – survival) = 1.0 mg/L; 14 d NOEC (no-observedeffect concentration – survival) = 0.1 mg/L Other endpoints reported – e.g., bioconcentration factor / bioaccumulation factor (BCF /BAF): Reproduction, moulting, markers fo r endocrine disruption *Was toxicity value below the chemical’s water solubility? X Other adverse effects (carcinogeni city, mutagenicity, etc. X

Score : Major items – 2/5; overall score – 21/25 (84%)

EC Reliability code : 2

Reliability category (high, satisfactory, low) : Satisfactory

Comments: While details relating to the composition/purity of the test subs tance and standard testing methods used are not included in th e report, it is likely that these met acceptability criteria given th at the study is part of a larger government-sponsored progra m and the results were submitted to the United Kingdom Departme nt of Environment, Food and Rural Affairs. The description of test methods include s discussion of procedures for collection and analysis of chemical subsamples for the assessment of chemical stability; however, measured concentrations are not reported and results are presented in term s of nominal concentration values. Similarly, procedures for m easurement of water quality parameters are described, but the numerical values are not provided in the report. C onsider data generated from t he study to be of acceptable quality, even though the resulting re port does not fully describe all results.

101Appendix C. Robust Study Summary – Inherent Toxicity

Item Yes No

Reference : Johnson et al

. 2005. Endocrine disruption in aquatic and terrestrial invertebrates. Fi nal report produced by WRc NSF Ltd., Marlow, Buckinghamshire for the United Kingdom Department of Environment, Food and Rural Affairs (DEFRA). March 2005.

Test Substance : 80-05-7 (Bisphenol A)

*Chemical composition of the substa nce (including purity, by-products) X Persistence/stability of test substance in test system X

Method

References X *OECD, EU, national, or other standard method ? X Justification of the method/protocol if a non-standard method was used X

*GLP (good laboratory practice) X
Test organisms (specify common and Latin names): Earthworm ( Eisenia sp. )

Latin or both Latin and common names reported? X Life cycle age/stage of test organism X

Sex X Length and weight of test organisms n/a Number of test organisms per replicate X Food type / feeding periods (acclimation / during test) X
Test design/conditions

Test type – acute or chronic: Chronic Experiment type (laborat ory or field) specified? X System type (static, semi-static, flow through)? X Negative or positive controls: Solvent control X Number of replicates (includi ng controls) and concentrations X Exposure pathways (food, water, both) X Exposure duration X *Measured concentrations reported? X Exposure media conditions (temperature, pH, elect rical conductivity, hardness, total organic carbon, dissolved organic carbon, dissolved oxygen, major cations and anions; other) X

Was pH within 6–9 range? X Was temperature within 5–28°C range? X Photoperiod and light intensity X Stock and test solution preparation X Use of emulgators/solubilizers (especially for poorly soluble / unstable substances) X Analytical monitoring intervals X

Statistical methods used X

Results

Toxicity values (LC 50 , EC 50 , or IC 50 –: 14 d LOEC (lowest-observed-effect concentra tion – survival) = 100 mg/kg soil dw; 14 d NOEC (noobserved-effect concentration – survival) = 32 mg/kg soil dw Other endpoints reported – e.g., bioconentration factor / bioaccumulation factor (BCF /BAF): Reproduction, markers for endocrine disruption *Was toxicity value below the chemical’s water solubility? n/a Other adverse effects (carcinogeni city, mutagenicity, etc. X

Score : Major items – 1/4; overall score – 20/24 (83%)

EC Reliability code : 2

Reliability category (high, satisfactory, low) : Satisfactory

Comments: While details relating to the composition/purity of the test subs tance and standard testing methods used are not included in th e report, it is likely that these met acceptability criteria given th at the study is part of a larger government-sponsored progra m and the results were submitted to the United Kingdom Departme nt of Environment, Food and Rural Affairs. The description of test methods include s discussion of procedures for collection and analysis of chemical subsamples for the assessment of chemical stability; however, measured concentrations are not reported and results are presented in term s of nominal concentration values. Similarly, procedures for m easurement of water quality parameters are described, but the numerical values are not provided in the report. C onsider data generated from t he study to be of acceptable quality, even though the resulting re port does not fully describe all results.

102Appendix D. Summary of Studies Invest igating Neurobehavioural Effects in

Rodents

Author Species, route, dose(s) Treatment length Developmental neurotoxicity endpoints assessed Critical effects

Behaviour and Receptor Expression

Ema et al.

2001 SD rats N=25 (F0) Oral, gavage 0, 0.2, 2, 20 or 200 μg/kg per day Two-generation study
Gestation and
lactation Conventional study including behavioural endpoints: open field,
water-filled T-maze. No effects noted by authors.

Nishizawa et
al. 2003 ICR mice
Oral 0, 0.002 mg/kg per day Days post
coitum:

6.5–11.5
6.5–13.5
6.5–15.5
6.5–17.5 Effects of prenatal exposure on expression of retinoic acid receptors in mouse embryos. Variable changes in retinoic acid retinoid X receptors α mRNA expression in brain, ovary and testis. Changes varied according to sex, tissue and dosing period.

Nishizawa et
al. 2005 a ICR mice
Oral 0, 00002, 0.002, 0.20, 20 mg/kg per day Days post
coitum:

6.5–13.5
6.5–17.5 Effects of exposure on expression of mRNA for arylhydrocarbon and retinoid receptors in mouse embryos. Numerous changes; varied as above. Increased mRNA expression for the arylhydrocarbon receptors in brain, testis and ovary; retinoic acid

α receptor in brain, ovary and testis;
retinoid X α receptor in testis and

ovary.

Nishizawa et
al. 2005b ICR mice
Oral 0, 00002, 0.002, 0.20, 20 mg/kg per day Days post
coitum:

6.5–13.5
6.5–17.5 Effects of exposure on expression of arylhydrocarbon receptor, related factors and metabolizing enzymes in mouse embryos. Increased mRNA expression for arylhydrocarbon receptor, arylhydrocarbon receptor repressor, arylhydrocarbon receptor nuclear translocator in brain, testis and ovary.

Kawai et al.

2003 CD-1 mice

N=7–9 Oral, micropipette 0, 2, 20 μg/kg per day Gestational day (GD) 11–18 Aggressive behaviour in male mice at 8, 12 and 16 weeks of age: aggression scores determined by contact times with opponent. Increased aggression scores at 8 weeks (2, 20 μg/kg per day); no effect was
observed at 12 and 16 weeks.

Palanza et al.

2002 CD-1 mice N=10–12 (F0) Oral, micropipette 10 μg/kg-bw per day GD 14–18 (F0) GD 14–18 (F1) Maternal nursing behaviour of F1 monitored from postnatal day (PND) 2–15. Reflex responses examined in F2 offspring. Differences in nursing behaviour
observed when F1 females were exposed to BPA only prenatally or as an adult; not during both periods. No effects on F2 postnatal reflex development.

Laviola et al.

2005 CD-1 mice

N=10–12 Oral, feeding from syringe 10 μg/kg-bw per day GD 11–18 At PND 60 males and females were subjected to conditioned place preference testing following a d-amphetamine injection (d-amphetamine-reinforcing effects). Conditioned place preference was not observed in treated females; males
were unaffected.

Gioiosa et al.

2007 CD-1 mice

N=14 Oral (trained to spontaneously drink) 10 μg/kg-bw per day GD 11 to PND 8 Observed males and females at different ages to investigate explorative and emotional behaviours: novelty test (adolescents), open field and elevated plus-maze (adults). BPA reduced or eliminated sex-related behavioural differences in novelty test, open field behaviour and elevated plus-maze.

103Author Species, route, dose(s) Treatment length Developmental neurotoxicity endpoints assessed Critical effects

Fujimoto et
al. 2006 Wistar rats

N=6 Drinking water 15 μg/kg per day GD 13 to PND 0 Behavioural testing in offspring: open field, elevated plus-maze, passive avoidance, forced swimming tests In open field testing, rearing was
increased in male s and slightly
decreased in females. Differences were observed in the forced swimming test
for males and females.
No effects observed in elevated plus-maze and passive avoidance tests.

Farabollini et
al. 2002 SD rats

N=7 Oral, micropipette 40 μg/kg-bw per day Gestation or
lactation.
Cross-fostered to establish prenatal and postnatal

groups. Social and sexual behaviour assessed beginning at ~14 weeks of age; intruder test. Statistically significant changes in sexual performance:
males – reduced females – slight increase in sexual
behaviour.

Aloisi et al.

2002 SD rats

N=7 Oral, micropipette 40 μg/kg-bw per day Gestation or
lactation.
Cross-fostered to establish prenatal and postnatal

groups. Behavioural response to pain (licking, flexing and jerking of the paw). Response to formalin injection at 22 weeks of age. No BPA-related effect on open field
behaviours. Slight inter-group differences in pain response (marginally diminished).

Adriani et al.

2003 SD rats
N=9
Oral 40 μg/kg-bw per day GD 0 to PND 23 (behaviour
examined at PND 35–45; adolescence) Novelty preference (PND 35–45; adolescence), impulsivity testing (adults), open field testing with and without amphetamine challenge. Decreased novelty seeking in females; decreased impulsive behaviour in
males. The amphetamine-induced increment activity was significantly less marked in BPA-treated male rats compared
with controls. No open field differences without challenge.

Porrini et al.

2005 SD rats, females

N=12 Oral, micropipette 40 μg/kg-bw per day GD 0 to PND 21
Cross-fostered

pups Play behaviour of female offspring (PND 35, 45, 55).
Identified six behaviours (principle component analysis); significant differences noted for
three. Social and non-social exploration increased (day 35 and 45); Play with males decreased (45 days); Duration of grooming behaviour decreased (day 45).

Della Seta et
al. 2005 SD rats, females

N=17 Oral, micropipette 40 μg/kg-bw per day Gestation and lactation (day after mating through lactation) Assessed maternal behaviour: the frequency, duration and latency of
each element of maternal behaviour was analyzed with software (retrieving, ano-genital licking, licking-grooming, arched-back posture, lactation, on nest, nest building). Overall significant reduction in licking-grooming behaviour. Marginal reduction in frequency of ano-genital licking and duration of arched-back posture; altered behaviour was not dependent on pup gender.

Della Seta et
al. 2006 SD rats, males

N=26 Oral, micropipette 40 μg/kg-bw per day PND 23–30 Socio-sexual behaviour of juvenile males (examined on PND 45 and at PND >90).
Plasma 17 β -oestradiol and
testosterone leve l were measured
on PND 37 and 105. Biting/sniffing/climbing behaviours (sexual activity) (directed at a PVC tube) were lower in BPA group (slight modifications compared to those observed with ethinylestradiol).
Sexual behaviour assessment: short-term effects; latency to first
intromission was reduced.
Plasma testosterone levels were significantly lower (17 β -oestradiol was not affected). Negishi et al.

2004 F344 rats
N=8–10 Oral, gavage 0, 100 μg/kg per day GD 3 to PND 20 Behaviour of males: open field, spontaneous motor activity, passive avoidance, elevated plus-maze, active
avoidance and monoamine-disruption tests Fewer correct avoida nce responses in
active avoidance test trials 1 to 3 of 4 (suggests slower learning).

104Author Species, route, dose(s) Treatment length Developmental neurotoxicity endpoints assessed Critical effects

Ryan and Vandenbergh

2006 C57/B1/6 mice

N=14–16 Oral, gavage 0, 2, 200 μg/kg per day GD 3 to PND 21 Effects on non-reproductive sexually dimorphic behaviour: ovariectomized female offspring tested in elevated plus-maze, light-dark preference chamber,
radial arm maze and modified
barnes maze. Earlier onset of puberty (200 μg/kg per day). Increased anxiety (light-dark preference chamber) at 200 μg/kg per day. No differences in spatial memory tests.

Carr et al.

2003 F344 rats N=10 pups/sex Oral, gavage 0, 0.1, 0.25 mg/kg per day PND 1–14 Neonatal exposure on Morris water maze performance (tested on PND 33) Reduced memory retention at 0.25 mg/kg per day in males (not significant) and fema les (significant).

Nigishi et al.

2003 F344 rats
N=8–9 Oral, gavage 0, 4, 40, 400 mg/kg per day GD 10 to PND
20 Behavioural evaluations: spontaneous motor activity, active avoidance test, open field
behaviour Variable responses; did not follow a dose-related pattern. No effect in the open field test.

Farabollini et
al. 1999 SD rats

N=11 Oral, micropipette 1) 40 μg/kg-bw per day 2) 400 μg/kg-bw per day 1) 10 days prior to mating – PND

21 2) GD 14 to PND

21 Conducted between PND 85 and

87. Holeboard box, elevated plus-maze (tests to provide measure of anxiety and locomotion). Increased anxiety-related behaviour (males and females in holeboard test) – 400 μg/kg-bw per day. Decreased anxiety-related behaviour in males in maze test (in conflict with holeboard test).

DessiFulgheri et
al. 2002 SD rats
N=11 Oral, micropipette 1) 40 μg/kg-bw per day 2) 400 μg/kg-bw per day 1) 10 days prior to mating – PND

21 2) GD 14 to PND

6 Play behaviour observations at PND 35, 45, 55 (groups were pooled for analysis). Identified eight behaviours; significant differences noted for
four. At 40 μg/kg-bw per day: play directed to females by females increased; frequency of social interest
behaviour increased in males. At 400 μg/kg-bw per day: frequency of social interest behaviour
was decreased in both males and females; sociosexual exploration decreased in males (both test groups)
and females.

Kawai et al.

2007 ICR mice

N=18 Oral, micropipette 0, 2 μg/kg per day GD 11–17 Expression of ER α and ER β on male mice. Immunostaining to

ER α and ER β , serotonin and serotonin transporter at 4–5, 8–9 or 12–13 weeks of age. Increased number of neurons expressing ER α and ER β at 5 and 13 weeks, but not at 9 weeks.

Ceccarelli et
al. 2007 SD rats
N=14 Oral, micropipette 40 μg/kg-bw per day PND 23–30 Effects of j uvenile BPA exposure in brain development. ER α levels assessed in three sexually dimorphic regions of the hypothalamus (PND 37 and 90) Increased ER α levels in ventromedial nucleus in females at PND 37; not at
PND 90. No differences in males.

Facciolo et
al. 2002 SD rats N=unspecified
Oral 40 μg/kg-bw per day 400 μg/kg-bw per day Pre-mating to

PND 23
Cross-fostered

pups Effect on the somatostatin receptor subtype 2 (sst 2 ) in limbic regions of the brain (PND 10 and 23). Binding activity was examined. The higher dose proved to be the more effective one; the most dramatic effects
of BPA on sst 2 levels occurred at the low-affinity states.

Facciolo et
al. 2005 SD rats
N =12
Oral 40 μg/kg-bw per day 400 μg/kg-bw per day 8 days premating to PND

23
Cross-fostered

pups Effect on the expression of sst 3 mRNA receptors in the limbic regions of the brain in females. At 400 μg/kg-bw per day, mRNA
levels were increased in some areas
and decreased in others. Agonists for α GABA A receptors enhanced responses.

Central Dopaminergic System

105Author Species, route, dose(s) Treatment length Developmental neurotoxicity endpoints assessed Critical effects

Suzuki et al.

2003 ddY mice

N=? Dietary 0, 2.5, 60, 250 mg/kg per day Gestation and
lactation In male offspring: place preference conditioning test conducted using methamphetamine; locomotor activity; G-protein activation in the limbic forebrain; expression of dopamine D1 receptor mRNA. Methamphetamine-induced place preference (dose-dependent); hyperlocomotion by methamphetamine was dramatically potentiated(250 mg/kg per day)); increased Gprotein activation and upregulated dopamine D1 receptor expression at the high dose.

Mizuo et al.

2004a ddY mice

N=? Dietary 0, 2.5, 60, 250 mg/kg per day Gestation and
lactation In male offspring: rewarding effects, locomotion activity and receptor activity induced by morphine; G-protein activation and μ-opioid receptor mRNA expression. Morphine-induced place preference (dose-dependent; significant at 60 and 250 mg/kg per day.) and hyperlocomotion (250 mg/kg per day.). G-protein activation and μ-opioid receptor mRNA was not changed by

BPA.

Mizuo et al.
2004b ddY mice

N=? Dietary 0, 250 mg/kg per day (estimated dose levels) Gestation and
lactation Functional changes in dopamine D3 receptors in limbic forebrain(expression measured in an RTPCR assay) Reduced G-protein activation by the dopamine D3 receptor agonist; decreased receptor density in brain following BPA exposure.

Narita et al.

2006 ddY mice

N=? Dietary 0, 0.006, 0.06, 0.6, 100, 400 mg/kg per day Gestation and
lactation Effects on central dopaminergic system in mouse limbic area – place conditioning, locomotion, [ 35 S]GTP γ S binding assay in response to morphine. Increased place pref erence (0.6, 100, 400 mg/kg per day.); hyperlocomotion (0.6 and 400 mg/kg per day.); increased dopamine-induced binding (0.006, 0.6 or 400 mg/kg per day.).[BPA may potentiate dopaminereceptor-dependent neurotransmission]

Narita et al.

2007 ddY mice

N=? Dietary 0, 400 mg/kg per day GD 0–7
GD 7–14
GD 14–20
PND 0–20 Investigate importance of exposure period with respect to morphine-induced behaviours. Male offspring: place conditioning, locomotion, [ 35 S]GTP γ S binding assay in response to morphine. Morphine associated place preference, hyperlocomotion and potentiated dopamine-induced binding at GD 7–14 and PND 0–20. Suggests organogenesis and lactation as sensitive periods for BPA-induced alternation of the dopaminergic
system. Brain Structure

Nakamura et
al. 2006 ICR/Jcl mice subcutaneous (s.c.) injection 0, 20 μg/kg per day GD 0–16 Brain neocortex development: embryonic forebrains assessed after intraperitoneal (i.p.) injection of BrdU on either GD 10, 12, 14 or 16 (immunohistochemistry) RT-PCR analysis of various

genes. Results suggest BPA-induced
acceleration of neuronal differentiation/migration. The expression of various genes was upregulated at E14.5 in the BPA-treated group.

Kubo et al.

2003 Wistar rats

N=5–6 Drinking water 0, 0.03, 0.3 mg/kg per day (estimated intakes) Gestation and
lactation Open field testing, sexual behaviour; sexually dimorphic nucleus of the preoptic area (SDN-POA) and locus coeruleus volumes; number of neurons in
locus coeruleus Open field behaviour (anxiety-related) slightly reduced in males and slightly
increased in females.
Gender difference in locus coeruleus volume was reversed; extent of reversal was greater at 0.03 mg/kg per day.

Kubo et al.

2001 Wistar rats

N=5 Drinking water 0, 1.5 mg/kg per day (approx) Gestation and
lactation Behaviour and brain development: open field behaviour, passive avoidance and
SDN-POA and locus coeruleus
volumes Gender-related differences in open field and passive avoidance tests
diminished. Gender difference in locus
coeruleus volume was reversed.

106Author Species, route, dose(s) Treatment length Developmental neurotoxicity endpoints assessed Critical effects

Funabashi et
al. 2004 Wistar rat

N=8–11 Drinking water 0, 2.5 mg/kg per day Gestation to

PND 21 Numbers of corticotrophin releasing hormone (CRH)
neurones in the brain: immunohistochemistry of neurons
in the bed nucleus of the stria terminalis and preoptic area. Normal gender difference in numbers of CRH neurons not maintained; due to
increase in males and decrease in
females.

Tando et al.

2007 ddY mice Dietary 0, 4.5 or 1200 mg/kg per day GD 0 to PND 21 Immunohi stochemical assessment at 8–11 weeks of age: tyrosine hydroxylase, calbindin D-28 K, calretinin, parvalbumin. Cell death was assessed using TUNEL staining. Reduced density and volume of tyrosine hydroxylase positive nuclei and fibers in substantia nigra in females at 4.5 kg per day, not at 1200 kg per day. TUNEL did not identify the presence
of cell death.

Honma et al.

2006 SD rats
N=6 Oral, gavage 0, 4, 40 mg/kg per day GD 6 to PND 20 HPLC analysis of brain neurotransmitter levels at 1, 3, 6, or 9 weeks of age. No discernable variations in neurotransmitter patterns.

Kwon et al.

2000 SD rats

N=8 Oral, gavage 0, 3.2, 32, 320 mg/kg per day GD 13 to PND

21 Sexually dimorphic nucleus of the preoptic area (SDN-POA) volume
in females (immunohistochemistry) No effects.

Ishido et al.

2007 Wistar rats N=10 (dams) Oral, 0, 600 μg/pup/day [authors indicated dose equivalent to 12–60 mg/kg per day] Daily
administration
from PND 5 to 3 weeks of age Measurements in males: Spontaneous motor activity (Supermex system) (4–5 weeks); immunohistochemistry of tyrosine hydroxylase and TUNEL; RT-PCR of dopamine transporter (7 weeks) Increased motor activity (hyperactivity) through nocturnal phase (12-hr dark period); 1.3 times as
active as controls.

Tyrosine hydroxylase immunoreactivity in substantia nigra largely reduced (not quantified). Increase in TUNEL positive cells with nuclear condensation (apoptotic) in substantia nigra. Gene expression of dopamine transporter was completely inhibited.

[BPA thought to contribute to the degeneration of dopaminergic neurons]

Patisaul et al.

2006 SD rats N=5–8 pups s.c. injection 100 mg/kg per day PND 1 and 2 Development of the anterioventral periventricular nucleus of the hypothalamus (AVPV) (assessed on PND 19). Immunohistochemistry for ER α and/or tyrosine hydroxylase (TH). TH positive cells increased in males (demasculinization); decreased TH/

ER α double-labelled cells in AVPV in females (defeminization – depleted to male-typical levels).
SDN or CALB-SDN volumes unaffected; AVPV volume unaffected.

Patisaul et al.

2007 SD rats N=5–8 pups s.c. injection 100 mg/kg per day PND 1 and 2 Volumes of AVPV and

SDN-POA (immunohistochemistry), calbindin, GnRH and fos immunoreactivity; assessed in
adults. Increased SDN-POA calbindin-positive nuclei (considered “hypermasculinization” of males). No
other BPA-induced effects.

107