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 2008
Synopsis 

Pursuant to section 74 of the Canadian Environmental Protection Act, 1999 (CEPA 

1999), the Ministers of the Environment and of Health have conducted a screening 

assessment of phenol, 4,4' -(1-methylethylidene)bis- or bisphenol A, Chemical Abstracts 

Service Registry Number 80-05-7. This substance was identified as a result of the 

categorization of the Domestic Substances List as a high priority for action under the 

Ministerial Challenge, since it was considered to pose the greatest potential for exposure 

to individuals in Canada and had been classified by the European Commission on the 

basis of reproductive toxicity. Bisphenol A also met the ecological categorization 

criterion for inherent toxicity to aquatic organisms, but it did not meet the ecological 

categorization criteria for persistence or bioaccumulation potential. 

Bisphenol A is a high volume chemical, with global production at 4 billion kg in 2006. In 

the United States, production quantities increased from 521 million kg in 1990 to 736 

million kg in 1995. Estimated production in the U.S. in 2007 was 1 billion kg. Canadian 

market values may be lower than those for the U.S; however, approximately 12 million 

kg of bisphenol A were reported as manufactured, imported or in commerce in Canada 

during the calendar year 1986. In 2006, no bisphenol A was manufactured in Canada at 

quantities equal to or greater than a reporting threshold of 100 kg. However, bisphenol A 

was used in Canada in the range of 100 000 to 1 000 000 kg and approximately half a 

million kg was imported into Canada either alone, in a product, in a mixture or in a 

manufactured item. 

The available data indicate that bisphenol A does not persist significantly under aerobic 

conditions. However, the substance has been found not to degrade or to degrade only 

slowly under conditions of low or no oxygen. Bisphenol A has been detected in Canadian 

and U.S. surface waters, sediment, groundwater and soil, as well as in municipal and 

industrial waste treatment products. Studies from North America, Europe and Japan 

document detectable levels in several species of aquatic biota. The data demonstrate that 

bisphenol A is present in a wide range of environmental media. 

Most data indicate relatively low bioaccumulation potential and a capacity for 

metabolism in various species. Most measured bioaccumulation and bioconcentration 

factors range only up to about 150 L/kg, with one study reporting a bioaccumulation 

factor of 650 L/kg in lower trophic levels. These studies confirm that bisphenol A is 

bioavailable and can accumulate in tissues to some degree. Bisphenol A is acutely toxic 

to aquatic organisms and has been shown to adversely affect growth and development in 

both aquatic and terrestrial species. There is evidence that low-level exposure to 

bisphenol A, particularly at sensitive life cycle stages, may lead to permanent alterations 

in hormonal, developmental or reproductive capacity. In laboratory testing, these effects 

have occurred within the range of concentrations measured in Canada, indicating that 

there is potential for adverse effects in populations, particularly close to point sources. 

On the basis of expected continued or increasing exposure of biota, and information 

indicating the potential for long-term adverse effects to organisms within the range of 

i
concentrations currently measured in the environment, it is considered appropriate to 

apply a precautionary approach when characterizing risk. As such, it is concluded that 

bisphenol A is entering the environment in a quantity or concentration or under 

conditions that have or may have an immediate or long-term harmful effect on the 

environment or its biological diversity. 

Human exposure to bisphenol A in Canada can result from dietary intake (e.g., migration 

from food packaging, migration from repeat-use polycarbonate containers), from 

environmental media (i.e., ambient air, indoor air, drinking water, soil and dust), from use 

of consumer products and other sources. Dietary intake is the primary source of exposure. 

Exposure estimates for the general population of Canada range from 0.08 μg/kg body 

weight/day to 4.30 μg/kg-bw per day. Specific exposure estimates for the most highly 

exposed subpopulation (i.e., infants) range from an average of 0.50 μg/kg body 

weight/day (maximum 4.30 μg/kg-bw per day) for infants aged 0 to 1 month to an 

average of 0.27 μg/kg body weight/day (maximum 1.75 μg/kg body weight/day) for 

infants aged 12 to 18 months. A critical effect for characterization of risk to human 

health is reproductive and developmental toxicity. The neurodevelopmental and 

behavioural dataset for rodents, though highly uncertain, is suggestive of potential effects 

at doses the same order of magnitude to 1-2 orders of magnitude higher than exposures. 

Given that toxicokinetics and metabolism data from experimental animal and limited 

human studies indicate potential sensitivity to the maternal-fetal unit and infant, and that 

animal studies suggest a trend towards heightened susceptibility during stages of 

development in rodents, it is considered appropriate to apply a precautionary approach 

when characterizing risk. As such, it is concluded that bisphenol A be considered as a 

substance that may be entering the environment in a quantity or concentration or under 

conditions that constitute or may constitute a danger in Canada to human life or health. 

In addition and where relevant, research and monitoring will support verification of 

assumptions used during the screening assessment and, where appropriate, the 

performance of potential control measures identified during the risk management phase. 

Based on the information available, bisphenol A meets one or more of the criteria set out 

in section 64 of the Canadian Environmental Protection Act, 1999. 

ii
Table 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 Environmental 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 Dietary 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 Developmental 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 – Inherent Toxicity............................................. 100 

Appendix B. Robust Study Summary – Inherent Toxicity............................................. 101 

Appendix C. Robust Study Summary – Inherent Toxicity............................................. 102 

Appendix D. Summary of Studies Investigating Neurobehavioural Effects in Rodents 103 

.

iii
Introduction 

The Canadian Environmental Protection Act, 1999 (CEPA 1999) (Canada 1999) requires 

the Minister of the Environment and the Minister of Health to conduct screening 

assessments of substances that have met the categorization criteria set out in the Act to 

determine whether these substances present or may present a risk to the environment or 

human health. Based on the results of a screening assessment, the Ministers can propose 

to take no further action with respect to the substance, to add the substance to the Priority 

Substances List (PSL) for further assessment, or to recommend that the substance be 

added to the List of Toxic Substances in Schedule 1 of the Act and, where applicable, the 

implementation of virtual elimination. 

Based on the information obtained through the categorization process, the Ministers 

identified a number of substances as high priorities for action. These include substances 

that 

• met all of the ecological categorization criteria, including persistence (P), 

bioaccumulation potential (B) and inherent toxicity to aquatic organisms (iT), and 

were believed to be in commerce; and/or 

• met the categorization criteria for greatest potential for exposure (GPE) or 

presented an intermediate potential for exposure (IPE), and had been identified as 

posing a high hazard to human health based on classifications by other national or 

international agencies for carcinogenicity, genotoxicity, developmental toxicity or 

reproductive toxicity. 

 

The Ministers therefore published a notice of intent in the Canada Gazette, Part I, on 

December 9, 2006 (Canada 2006), that challenged industry and other interested 

stakeholders to submit, within specified timelines, specific information that may be used 

to inform risk assessment, and to develop and benchmark best practices for the risk 

management and product stewardship of those substances identified as high priorities. 

The substance, phenol, 4,4' -(1-methylethylidene)bis- or bisphenol A was identified as a 

high priority for assessment of human health risk because it was considered to present 

greatest potential for human exposure (GPE) and had been classified by other agencies on 

the basis of reproductive toxicity. The Challenge for bisphenol A was published in the 

Canada Gazette on May 12, 2007. A substance profile was released at the same time 

(Canada 2007). The substance profile presented the technical information available prior 

to December 2005 that formed the basis for categorization of this substance. As a result 

of the Challenge, submissions of information were received. 

Although bisphenol A was determined to be a high priority for assessment with respect to 

human health, and it also met the ecological categorization criterion for inherent toxicity 

to aquatic organisms, it did not meet the categorization criteria for potential for 

persistence or bioaccumulation. Therefore this assessment focuses principally on 

information relevant to the evaluation of risk to human health. 

1
Under CEPA 1999, screening assessments focus on information critical to determining 

whether a substance meets the criteria for defining a chemical as toxic as set out in 

section 64 of the Act, where 

“64. [...] a substance is toxic if it is entering or may enter the environment in a quantity or 

concentration or under conditions that 

(a) have or may have an immediate or long-term harmful effect on the environment or its 

biological diversity; 

(b) constitute or may constitute a danger to the environment on which life depends; or 

(c) constitute or may constitute a danger in Canada to human life or health.” 

Screening assessments examine scientific information and develop conclusions by 

incorporating a weight of evidence approach and precaution. 

This screening assessment includes consideration of information on chemical properties, 

hazards, uses and exposure, including the additional information submitted under the 

Challenge. Data relevant to the screening assessment of this substance were identified in 

original literature, review and assessment documents, stakeholder research reports and 

from recent literature searches, up to September 2008 for human health effects and 

exposure sections of the document. Key studies were critically evaluated; modelling 

results may have been used to reach conclusions. 

Evaluation of risk to human health involves consideration of data relevant to estimation 

of exposure (non-occupational) of the general population, as well as information on 

health hazards (based principally on the weight of evidence assessments of other agencies 

that were used for prioritization of the substance). Decisions for human health are based 

on the nature of the critical effect and/or margins between conservative effect levels and 

estimates of exposure, taking into account confidence in the completeness of the 

identified databases on both exposure and effects, within a screening context. The 

screening assessment does not represent an exhaustive or critical review of all available 

data. Rather, it presents a summary of the critical information upon which the conclusion 

is based. 

This screening assessment was prepared by staff in the Existing Substances Programs at 

Health Canada and Environment Canada and incorporates input from other programs 

within these departments. The draft of this screening assessment was subject to a 60-day 

public comment period. Additionally, this assessment has undergone external written 

peer review/consultation. Comments on the technical portions relevant to human health 

were received from Toxicology Excellence for Risk Assessment (TERA), Dr. Sergio 

Pellis of University of Lethbridge, Alberta, and Dr. Alfonso Abizaid of Carleton 

University, Ottawa. While external comments were taken into consideration, the final 

content and outcome of the screening risk assessment remain the responsibility of Health 

Canada and Environment Canada. The critical information and considerations upon 

which the assessment is based are summarized below. 

2
Substance Identity 

For the purposes of this document, this substance 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 names1

 

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-methylethylidene)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-phenyl)-propane (PICCS)

Other names Bisphenol A, Diphenylolpropane, BPA 

Chemical group Discrete organics 

Chemical subgroup Phenols 

Chemical formula C15H16O2

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 (NCI). 2006: AICS (Australian Inventory of Chemical Substances); ASIA-PAC (Asia-Pacific 

Substances Lists); ECL (Korean Existing Chemicals List); EINECS (European Inventory of Existing Commercial Chemical 

Substances); ENCS (Japanese Existing and New Chemical Substances); 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).

3
Physical and Chemical Properties 

Table 2 contains experimental and modelled 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/m3) Experimental 1195 25 Sax and Lewis 1996

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

(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·m3

/mol)

Hine and Mookerjee 1975

Henry’s Law 

constant 

(Pa·m3

/mol) 

Modelled

9.3 x 10-7

(9.16 x 10-12

atm·m3

/mol) 

Water solubility 

120 mg/L: 

4.0 x 10-5 

(3.95 x 10-10

atm·m3

/mol) 

Water solubility 

257 mg/L: 

4.7 x 10-6

(4.7 x 10-11

atm·m3

/mol) 

25

HENRYWIN 2000 

(Bond estimation) 

HENRYWIN 2000 

(VP/Water solubility 

estimation)

Experimental 3.32 Howard 1989; 

Hansch et al. 1995

Log Kow 

(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 Koc

(Organic carbonwater partition 

coefficient – L/kg) 

(dimensionless) Modelled 4.88 PCKOCWIN 2000

4
Property Type Value Temperature 

(°C) 

Reference 

Log Koa 

(Organic carbonair partition 

coefficient) 

(dimensionless) 

Modelled 12.7 KOAWIN 2000 

253-257 22-24 Lee and Peart 2000b 

301 room 

temperature Experimental Bayer AG 1988 

120 25 Dorn et al. 1987

Water solubility 

(mg/L) 

Modelled 173 25 WSKOWWIN 

2000

Other solubilities 

(g/L) 

Experimental 

(alcohol) soluble Lewis 2000 

pK Experimental 9.59 – 11.30 Staples et al. 1998 a (Acid 

dissociation 

constant) 

(dimensionless) 

Modelled 9.73-10.48 ACD 2005 

Bisphenol A is characterized as having low vapour pressure, moderate water solubility 

and a moderate organic carbon-water partition coefficient (log Koc). EPIWIN (2000) 

indicates that, based on the chemical structure of the substance, the Koc of bisphenol A 

may be sensitive to and vary significantly with pH. The high pKa 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-water partition coefficients (Kow), will likely partition to 

some extent in all available environmental phases. 

Sources 

There is no reference in the published literature to the natural occurrence of bisphenol A 

in the environment. 

Global production of bisphenol A was 4 billion kg in 2006. In the United States (U.S.), 

production quantities increased from 521 million kg in 1990 to 736 million kg in 1995. 

Estimated production in the U.S. in 2007 was 1 billion kg (SRI Consulting, 2007). 

Based on a survey conducted under section 71 of CEPA 1999, no bisphenol A was 

manufactured in Canada in 2006 at quantities greater than or equal to 100 kg. However, 

26 companies reported using bisphenol A in Canada in the range of 100 000 1 000 000 kg 

or importing approximately half a million kg into Canada either alone, in a product, in a 

mixture or in a manufactured item (Environment Canada 2007a). The extent to which the 

reported values represent quantities of bisphenol A present in finished and semi-finished 

5
goods entering Canada from other parts of the world is unknown as these uses would be 

unlikely to meet the reporting criteria for the survey. 

Uses 

Bisphenol A is primarily used as a monomer in the production of polycarbonates and as a 

precursor or a starting material for monomers of certain epoxy resins (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 manufactured item reported for 2006 in response to a 

survey of Canadian industry under section 71 of CEPA 1999 included resins, curing 

agents, epoxy curing agents, hardeners, plastic resin formulations, monomer, paperboard 

packaging, metal cans, phenolic resins, industrial coatings, plasticizers, adhesives, two 

part epoxy adhesives, chain oil, brake fluid, heat transfer fluid and lubricant formulations. 

Information voluntarily submitted in 2007 in response to the Challenge Questionnaire 

and other information submitted by industry additionally include use in epoxy polymer 

flooring, as a laminating adhesive, in custom colour powder coating and as a curing agent 

for resurfacing concrete. 

The literature indicates that polycarbonate is used in the manufacture of compact discs, 

food and beverage contact containers (e.g., baby bottles, repeat use water bottles, 

pitchers, water carboys, tableware and storage containers), water pipes, medical devices, 

and in glazing applications and film. Polycarbonate blends find application in the electric 

and electronics industry (e.g., alarm devices, mobile phone housings, computer parts, 

household electric equipment, lamp fittings, power plugs) and the automotive industry 

(e.g., car headlight and rear light reflectors and coverings, bumpers, radiator and 

ventilation grills, safety glazing, inside lights, motorcycle windshields and protective 

helmets) (NTP 2007; EFSA 2006). Medical devices which contain polycarbonate (e.g., 

hemodialyzers, hemofilters, blood oxygenators and respiratory devices) have been 

approved or licensed for use in Canada under the Medical Devices Regulations of 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 commonly used in protective coatings, structural 

composites, electrical laminates, electrical applications and adhesives (NTP 2007). 

Epoxy-phenolic resins are used as liners in metal cans for foods and beverages and as 

coatings on metal lids for glass jars and bottles, as well as surface coatings on residential 

drinking water storage tanks (EFSA 2006). Cans used for food and beverages, including 

ready-to-use, concentrate and powdered infant formula, that were evaluated by Health 

Canada are lined with epoxy resin (Health Canada, Chemical Health Hazard Assessment 

Division, Health Products and Food Branch, pers. comm., 2008 Feb 26, unreferenced). 

6
Epoxy resins are used in automotive electro-coating and in automotive and industrial 

primer and topcoat applications. Epoxy resins are also used industrially in powder 

coatings used on reinforcing bars for concrete, oil pipelines, wire racks and shelving, as 

automotive primers and as general industry primers and topcoats. Aerospace coatings 

(e.g., decals or logos on the sides or tails of planes) also contain bisphenol A 

(Environment Canada 2007b). 

Dental products containing bisphenol A-based polymers (e.g., resin-based composite 

restorative materials and sealants) have been approved or licensed for use in Canada 

under the Medical Devices Regulations of the Food and Drugs Act (Health Canada, 

Medical Devices Bureau, Health Products and Food Branch, pers. comm., 2008 Feb 5, 

unreferenced). 

Bisphenol A is also used in the production of phenolplast, phenolic and unsaturated 

polyester resins, thermal paper, polyvinylchloride (PVC), alkoxylated 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 include protective window glazing, building materials, optical 

lenses and development of dyes (NTP 2007). 

The U.S. National Library of Medicine Household Products Database reports that 

bisphenol A-based polymers are used in coatings, adhesives and putties available to the 

general public for use in automobiles, home maintenance and repair, and hobbies (NTP 

2007). Response to a survey conducted under section 71 of CEPA 1999 indicated that 

these products were also reported to be available 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 bisphenol A and potential releases to the 

environment are restricted to those associated with anthropogenic activities. 

Releases of bisphenol A may occur during production, processing, use or disposal of the 

substance or products containing it. Results from a section 71 Notice with respect to 

certain Batch 2 Challenge substances conducted for 2006 (Environment Canada 2007a) 

indicated that bisphenol A was not manufactured in Canada in that year in an amount 

equal to or greater than 100 kg, although it was imported into the country for use in 

processing. As production of 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 application materials 

such as polycarbonates and epoxy resins are the most likely industrial sources of release 

7
of bisphenol A into the Canadian environment. Unintentional release of fugitive dust 

from closed systems during handling and transportation of the substance may also occur. 

Bisphenol A has low vapour pressure at typical environmental temperatures; however, 

elevated temperatures occurring during some processing operations may increase the 

vapour pressure, resulting in formation and possible emission of gaseous bisphenol A 

from manufacturing facilities. 

Bisphenol A released to wastewater would likely be transported to a treatment facility. 

Once at the plant, the moderate organic carbon-water partition coefficient (log Koc 2.53) 

suggests that a fraction of the release will 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 present 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 wastewater treatment plants. The results provide evidence that 

bisphenol A partitions into both the solid and liquid phases within treatment plants. A 

more detailed review of concentrations measured in waste treatment products is provided 

in Table 9a. Bisphenol A present in sewage sludge could be released into the soil 

compartment through application of sludge biosolids to agricultural and pasture lands. 

Lee and Peart (2000a) examined the efficiency of bisphenol A removal by the sewage 

treatment plant process. Overall reduction rates of < 1% to 99% were determined from 

influent/effluent sample pairs collected from 36 Canadian sewage treatment plants, with a 

median reduction value of 68%. Plants exhibiting greater than 50% reduction rates were 

those employing secondary waste treatment. 

The National Pollutant Release Inventory (NPRI) has tracked the release of bisphenol A 

from industrial facilities in Canada since the 1990s. Table 3 provides a summary of NPRI 

on-site release and total disposal data over the period 2001 to 2005. Where information 

was provided, on-site releases were exclusively to the air compartment. In recent years, 

total annual releases have varied between about 44 and 8770 kg, with no clear temporal 

trend apparent. Most recent reporting (2005) indicates total on-site releases of 120 kg, a 

decrease from that seen in previous years (NPRI 2007). It should be noted that only 

facilities meeting established criteria are required to report to the NPRI, and therefore 

NPRI data are likely to underestimate total Canadian releases of bisphenol A. For 

example, all of the following reporting criteria must have been met before NPRI 

reporting of bisphenol A was required for the 2005 calendar year: (1) employees at the 

facility worked a total of 20 000 hours or more or the facility was used 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 manufactured, processed or otherwise 

used at a concentration greater than or equal to 1% by weight, with the exception of 

NPRI substances considered to be by-products. Facilities not meeting all of these criteria 

were not required to submit information relating to the release and/or disposal of 

bisphenol A at their facility. While the number of such facilities is not known, any such 

releases would not be reported and would therefore be over and above those reported 

through the NPRI. 

8
Bisphenol A also appears on the United States Environmental Protection Agency Toxics 

Release Inventory (TRI), a database of chemicals managed through disposal and other 

releases, recycling, energy recovery and waste treatment. In 2005, a total of 

approximately 675 000 kg (1.5 × 106

 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 surface 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 products may occur through volatilization or 

leaching. The majority of bisphenol A appears to be effectively retained within the 

polymer matrix of materials such as polycarbonates and therefore losses through leaching 

from the product surface are expected to be limited (ECB 2003). As well, the low vapour 

pressure suggests bisphenol A will have little tendency to volatilize from products at 

normal environmental temperatures. Some losses could occur at elevated temperatures, 

for example, during heating of end products. Losses may also take place through 

weathering and breakdown of end products, particularly those with outdoor applications. 

Bisphenol A may enter the environment through physical and chemical degradation of 

end products during disposal and recycling operations. Releases would be primarily to 

soil, and to a lesser extent, to water and air. Moderate water solubility suggests there is 

potential for leaching from disposal sites under some circumstances (e.g., during rainfall 

events) and concentrations 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 concentrations would be expected should the 

leachates have an elevated pH. 

Environmental Fate 

Based on its physical and chemical properties (Table 2) and compartments to which it is 

released, the results of Level III fugacity modelling (Table 4) suggest that bisphenol A is 

expected to partition predominantly to soil or water, depending on the compartment of 

release. 

As a moderately hydrophobic substance with some water solubility, 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 dissolved phase (Cousins et al.

9
2002). This suggests that the medium of release may be particularly important in 

predicting the partitioning behaviour and fate of bisphenol A in the environment. 

Level III fugacity modelling (Table 4) predicts that when bisphenol A is released 

exclusively into water, the majority (96.9%) will remain within the water column with 

only a small fraction (3.1%) moving into sediment. When released to air or soil, or when 

equal proportions are released to all three media, most bisphenol A (78.7% to 99.3%) will 

partition to the soil compartment, with smaller 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 Level 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 sediment; partition coefficients 4.5 × 10-10 (dimensionless) for air-water, 9.7 L/kg for soil-water, 19.4 

L/kg for sediment-water, 97 L/kg for suspended particles-water, 5.5 L/kg for fish-water, 100 (dimensionless) for aerosol-water 

Eisenreich et al. (1981) predicted that bisphenol 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 fraction present as vapour would react with 

photochemically generated hydroxyl radicals (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 process for bisphenol A in most 

aquatic and terrestrial environments. Biodegradation half-lives of less than four days 

were measured in natural waters following an adaptation period (acclimation) from one to 

four days (Dorn et al. 1987). A slower rate of biodegradation is expected to occur in nonacclimated waters; other processes such as sorption to suspended solids and sediments 

and photolysis may also take place. Recent research (e.g., Chin et al. 2004; Peng et al. 

2006; Zhan et al. 2006) documents both direct and indirect photochemical transformation 

of bisphenol A in aquatic media. Chin et al. (2004) reported that the rate of direct 

photolysis was significantly slower than that of indirect photolysis, which includes the 

presence of dissolved organic matter (DOM), and hypothesized that indirect 

photodegradation may proceed via multiple pathways involving DOM-derived 

phototransient substances. Bisphenol A is not expected to undergo hydrolysis due to a 

lack of hydrolyzable functional groups (Lyman et al. 1982). The low Henry’s Law 

constant suggests that volatilization is unlikely to be a significant removal process. 

The moderate water solubility of bisphenol A implies that the substance will exhibit low 

to moderate mobility when released to soil (Dorn et al. 1987). However, Loffredo and 

10
Senesi (2006) examined the adsorption kinetics and adsorption/desorption isotherms of 

bisphenol A onto four soil samples collected from the surface (0-30 cm) and deep (30-90 

cm) horizons of two acidic sandy soils. The researchers concluded that soil adsorption of 

bisphenol A is generally reversible, with desorption occurring quickly and completely. 

Thus, the substance is expected to move down the soil profile and could possibly 

contaminate groundwater in acidic sandy soils (Loffredo and Senesi 2006). The 

movement of bisphenol A through the soil matrix and/or into groundwater will be 

influenced by the physical and chemical properties of the receiving soil and of water 

travelling through it, and in particular the pH of these media, as well as the nature and 

properties of organic material present in the soil and water. Quantities of bisphenol A 

available for transport will be determined by factors such as the loading rate of the 

substance into the receiving medium and degradation processes. Potential contamination 

of groundwater from soil-based bisphenol A contamination is thus site-specific and 

difficult to predict across the Canadian landscape. Rudel et al. (1998) reported 

concentrations of up to 1.41 × 10-3 mg/L in groundwater samples collected in the vicinity 

of municipal landfills and wastewater treatment plants. Bisphenol A was present in 

monitoring wells located downgradient of secondary wastewater treatment plant 

infiltration beds to which no wastewater had been discharged for six months, suggesting 

no or only slow degradation of the substance in subsurface waters. 

Persistence and Bioaccumulation Potential 

Environmental Persistence 

Data relevant to the evaluation of potential environmental persistence of bisphenol A are 

summarized in Tables 5a and 5b. 

Table 5a. Empirical data 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 Klečka et al. 2001 

Biodegradation 

(aerobic) 2-3 Half-life, days 

Water 

Biodegradation 

(anaerobic) < 10 10-day biodegradation, 

% 

Kang and Kondo 2002 

11
Medium 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, 

Ground- days 

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 dominant loss process for bisphenol A in most 

aquatic and terrestrial environments (Staples et al. 1998). Bisphenol A has demonstrated 

rapid biodegradation in standard aerobic 28-day ready biodegradability testing using 

12
OECD Test Guideline No. 301F (West et al. 2001), but is indicated to be nonbiodegradable using another aerobic OECD method, Test Guideline 301C (NITE 1977). 

Klečka et al. (2001) reported biodegradation half-lives of less than 3.5 days in river 

surface water samples, including those to which 0.05% surface sediment mixtures (top 1-

2 cm) were added. A lag phase of from two to eight days was required prior to the onset 

of degradation. 

A number of studies report resistance to biodegradation under conditions of low oxygen, 

suggesting there may be 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 conditions, but a less than 10% decrease in 

samples incubated anaerobically for an equivalent time period of 10 days. The 

researchers postulated that anaerobic bacteria may have little or no capability of 

degrading bisphenol A, although chemical degradation of the substance (for example, by 

lipoperoxide and sodium chloride) could still occur. Ying and Kookana (2005) 

documented rapid degradation of bisphenol A in soil within 7 days, but no degradation of 

the substance under anaerobic conditions in the soil. Bisphenol A was reported to persist 

in an anaerobic slurry of river bed sediment, with no degradation evident after 3 months 

of incubation (Ronen and Abeliovich 2000). The substance also remained stable in 

anoxic estuarine sediments, with no degradation observed after 120 days (Voordeckers et 

al. 2002). Concentrations of 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 soil types (Ying and Kookana 2005), and water 

and sediment collected from an aquifer (Ying et al. 2003). Little or no degradation was 

reported for periods of up to 4 months, suggesting that half-lives of bisphenol A in anoxic 

and anaerobic media could well exceed one year. 

Sarmah and Northcott (2008) examined degradation of bisphenol A in river watersediment and groundwater-aquifer test systems. A biphasic degradation pattern was 

observed under both aerobic and anaerobic testing conditions, with rapid initial 

degradation of greater than 90% occurring in the first four days and trace amounts still 

remaining at the end of the 70-day study period. The formation of degradation products 

was monitored for the duration of the study; however, no potential metabolites were 

detected. Measured data were fitted using a first-order double exponential decay model, 

with results being presented as dissipation times (i.e., losses from the test system) rather 

than degradation half-lives. Reported 50% dissipation times (DT50s) were 0.57 to 1.212 

days for aerobic conditions and 1.38 to 2.315 days for anaerobic conditions. Times for 

90% dissipation were 2.26 to 2.75 days and 4.901 to 306 days for aerobic and anaerobic 

conditions, respectively. The researchers noted the incongruency of their results with 

those of previous studies which found little or no biodegradation of bisphenol A under 

anaerobic conditions, and suggested that lower test concentrations used in the study and 

variations in experimental protocols may have contributed to the observed differences. In 

addition, losses of up to 40% of the initial amount applied occurred in the sterile (control) 

treatments; this was considered significantly high and attributed to the possible 

occurrence of abiotic degradation processes or irreversible sorption of bisphenol A to test 

vessels. 

13
The modelled data generally support the empirical evidence, with BIOWIN (2000) 

predicting a half-life for ultimate aerobic biodegradation of 37.5 days but no or very slow 

anaerobic biodegradation (Table 5b). The results from Table 5b show that the majority of 

the probability models suggest this substance does not biodegrade fast. However, only 

the MITI non-linear model produced a result less than 0.3, the cut-off suggested by 

Aronson et al. (2006) in identifying substances as having a half-life >60 days (based on 

the MITI probability models). CPOPs (2008), which uses the CATABOL model, 

predicts that bisphenol A will fail ready biodegradation testing based on OECD method 

301C, supporting the experimental results reported by NITE (1977). 

Table 5b. Modelled data for persistence 

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 Probability BIOWIN 2000, Linear 

Water Biodegradation 0.4653 Probability BIOWIN 2000, Non-linear 

Water Biodegradation 0.2956 Probability BIOWIN 2000, MITI linear 

Water Biodegradation 0.1559 Probability BIOWIN 2000, MITI non-linear 

Water Biodegradation -0.2593 Probability BIOWIN 2000, Anaerobic linear 

Water Biodegradation 11.71 BOD, % (301C) CPOPs 2008; Mekenyan et al. 

2005 

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

water2

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

water2

1

 In all domains of applicability (global parameter, structural and metabolic) 2

 Values were derived from the modelled half-life in water using the extrapolation factors of Boethling et 

al. (1995): t1/2 soil : t1/2 water : t1/2 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 substance will likely be 

rapidly oxidized in air. AOPWIN (2000) predicts reaction with atmospheric nitrate 

radicals may also occur. 

There is conflicting evidence on the aerobic biodegradation potential of bisphenol A. 

Laboratory tests show positive and negative results of ready biodegradation potential. 

However, there is insufficient detail for the only laboratory aerobic biodegradation test 

showing a negative ready biodegradation test (NITE 1977) to determine whether the lack 

of biodegradation was the result of a lag phase. This test received lower weighting 

overall. There is greater concordance with the remaining laboratory and field evidence 

showing a fairly rapid loss of bisphenol A under aerobic conditions. Field tests are 

expected to be more indicative of the biodegradation potential of bisphenol A under 

natural conditions than the laboratory studies or QSARs based on laboratory data. 

Therefore, field data received greater weight. Results for the majority of aerobic 

14
biodegradation models also suggest a fairly short half-life, but some models (e.g., CPOPS 

and MITI non-linear model) suggest greater persistence of bisphenol A. Both these 

models, however, are trained on the NITE ready biodegradation data which, as already 

suggested, may not adequately account for a potential lag phase of bisphenol A. 

Therefore, these predictions are subject to test error which will not be captured by 

bisphenol A’s structural characteristics alone and thus receive a lower weight. When 

combined laboratory, field and model evidence are considered, there is greater reliable 

evidence to suggest that bisphenol A does not meet persistence criterion for air (half-life 

in air ≥ 2 days), water and soil (half-life in soil and water ≥ 182 days) as set out in the 

Persistence and Bioaccumulation Regulations (Canada 2000). There are many studies 

showing that bisphenol A does not degrade or degrades only slowly under conditions of 

low or no oxygen and the measured presence in sediment, a medium to which there is no 

direct release, is further evidence of this slow degradation. Therefore, bisphenol A meets 

criteria for persistence in sediments (half-life in sediments ≥ 365 days) under these 

conditions. 

Potential for Bioaccumulation 

The empirical and modelled log Kow 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 bioaccumulation 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 bioaccumulation factors (BAF; i.e., 

incorporating consideration of uptake through food) of from 18 to 650 for periphyton 

(presence of detritus not reported, uncertain whether steady-state concentrations 

achieved) and from 8 to170 for benthos (procedures for sampling/collection, species not 

specified), and postulated that food may be an important uptake route for organisms in 

the aquatic environment. 

Table 6a. Empirical data for bioaccumulation 

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 occurring at the lowest test temperature of 2°C. This suggests 

that in addition to environmental exposure concentrations, body clearance rates and 

metabolism are important in determining final tissue levels of the substance. Metabolism 

15
of bisphenol A has been reported in fish (Lindholst et al. 2001), with the formation of 

glucuronides identified as the primary metabolic pathway. Kang et al (2006c) conducted 

a review of existing literature and also found that various organisms, including plants, 

fish, birds and mammals have capacity to metabolize bisphenol A. 

 

A range of QSAR modelled bioaccumulation and bioconcentration values are available, 

with BAF or BCF values of from 5.8 (Baseline BCF Model) to 955 (OASIS Forecast; 

Table 6b). Most modelled values support a conclusion of low bioaccumulation potential. 

Table 6b. Modelled data for bioaccumulation 

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 BAF 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 properties of the substance 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 concentrations (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) 

EC501

2.7, 3.12 

1.0, 1.83

Alexander et al. 1988 

Daphnid Acute (48 hours) EC50

LC504

 

LC50

10 

7.75 

12.8 

Alexander et al. 1988 

Brennan et al. 2006 

Hirano et al. 2004 

16
Test organism Type of test Endpoint Value (mg/L) Reference 

Chronic (7 days) IC255 

NOEC6

LOEC7

3.92 

0.94 

1.88 

Tatarazako et al. 2002 

Chronic (21 days) NOEC 

Threshold 

concentration8

≥ 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) 

LC50

NOEC 

LOEC 

NOEC 

LOEC 

NOEC 

LOEC 

NOEC 

LOEC 

LC50

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) LC50 

LC50

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 percent effect. A point estimate of the concentration of a test substance that causes 25% 

reduction in a quantitative biological measurement such as growth rate. 6

 No observed effect concentration; highest concentration in a toxicity test not causing a statistically significant effect in comparison to 

the controls. In this case, the study identified that the highest concentration tested did not result in statistically significant results. Since 

the NOEC could be higher, the NOEC is described as being greater than or equal to the highest test concentration. 7

 Low observed effect concentration; lowest concentration in a toxicity test that caused a statistically significant effect in comparison 

to the controls. 

8

 Minimum concentration in a toxicity test that just causes an observable effect 

Alexander et al. (1988) conducted a series of standard freshwater and saltwater toxicity 

tests on bisphenol A. Lowest endpoint values were a 96-hour EC50 of 1.0 mg/L for the 

marine diatom, Skeletonema costatum, based on cell counts, and a 96-hour LC50 of 1.1 

mg/L for the mysid shrimp, Mysidopsis bahia. The mysid 96-hour LC50 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 value was in the range of 

that reported by Alexander et al. (1988; see Table 7a). However, in chronic testing, 

higher mortality was observed in second generation daphnids compared with first 

generation animals exposed to the same test concentrations. The researchers postulated 

that elevated mortalities in the second generation may indicate weakening of the 

offspring relative to the first (parental) generation, leading to greater susceptibility to 

lethal effects in the second generation. This trend could indicate action at the cellular 

level of organisms, possibly DNA damage. 

The same study also examined potential reproductive effects. No significant impacts were 

observed in the fecundity of first and second generation daphnids exposed to 

concentrations up to 1.0 mg/L. Mu et al. (2005). However, reported reduced reproductive 

17
success for Daphnia magna at a threshold concentration of 1.30 mg/L and proposed that 

the substance could elicit chronic effects through interference with ecdysis (moulting). 

Seven-day NOEC and LOEC values of 0.94 and 1.88 mg/L, respectively, were reported 

for reproduction in the daphnid species, Ceriodaphnia dubia (Tatarazako et al. 2002); the 

IC25 for the study was 3.92 mg/L. These values are similar to results obtained using 

Daphnia magna and provide further evidence of possible chronic toxicity. 

Johnson et al. (2005) investigated potential impacts to the freshwater amphipod, 

Gammarus pulex. No statistically significant effects were evident on female moulting and 

juvenile production following a 14-day exposure period, although adult survival was 

reduced at the highest test concentration of 1.0 mg/L. 

Potential effects following long-term exposure (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 reproductive effects (i.e., induction of vitellogenin synthesis, 

alterations to sex cells) occurred at 0.016 mg/L. The researchers 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 spawning period, and investigated effects on 

maturation, quantity and quality of semen and eggs. Bisphenol A was added into the 

flow-through test system by means of an injection pump, with final reported test 

concentrations of 0.00175, 0.00240 and 0.005 mg/L calculated based on water flow and 

test chemical injection rates. Similar numbers 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 lower 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 control. The percentage of ovulated females 

was similar in the control group and two lowest exposure groups; however, females at the 

highest exposure group did not ovulate over the 103-day investigation period. Ovulation 

was delayed by approximately 2 weeks in the 0.00175 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 quality of eggs (egg mass, percentile mass 

increase during hardening, egg fertility). 

Acute aquatic toxicity estimates derived from QSAR models are in the range of 3 to 9 

mg/L (Table 7b), slightly above results obtained experimentally. However, the predicted 

values are sufficiently close to those of laboratory studies to support a decision of 

potential acute aquatic toxicity at low concentrations. The weight of experimental and 

modelled evidence indicate that bisphenol A can be considered highly hazardous to the 

18
aquatic environment (i.e., (L(E)C50 values at or approaching 1 mg/L and/or chronic 

NOEC equal to or less than 0.1 mg/L). 

Table 7b. Modelled data for aquatic toxicity 

Test organism Type of test Endpoint Value (mg/L) Reference 

Fish Acute 

(96 hours) 

LC501 3.3 ECOSAR 2004 

Fish Acute 

(96 hours) 

LC50 4.3 TOPKAT 2004 

Fish Acute 

(96 hours) 

LC50 4.9 ASTER 1999 

Fish Acute 

(96 hours) 

LC50 ≤6.33 (± 1.2)4 CPOPS 2005 

Fish Acute 

(96 hours) 

LC50 8.5 AIES 2003-2005 

Fish Chronic 

(30 days) 

MATC5 0.5 ECOSAR 2004 

Daphnid Acute 

(48 hours) 

LC50 2.6 ECOSAR 2004 

Daphnid Acute 

(48 hours) 

EC502 5.1(± 2.6)4 CPOPS 2005 

Daphnid Acute 

(48 hours) 

EC502 5.2 TOPKAT 2004 

Daphnid Chronic 

(21 days) 

MATC5 0.4 ECOSAR 2004 

Green Algae Acute 

(96 hours) 

EC50 4.0 ECOSAR 2004 

Green Algae Chronic 

(96 hours) 

MATC5 1.0 ECOSAR 2004 

1

 Median lethal concentration 

2

 Median effective concentration 

3

 CPOPS has identified bisphenol A as having an unspecified reactive model of toxicity in the fish 96hr LC50 QSAR. The ≤ denotes 

that the median lethal concentration is expected to be lower than the reported default value which is based on a baseline narcotic mode 

of action 

4

 ± 95% confidence interval 

5

 Maximum Allowable Toxicant Concentration, generally presented as the range between the No Observed Effect Concentration and 

Lowest Observed Effect Concentration or as the geometric mean of the two measures 

Evidence of disruption to reproductive and developmental processes following exposure 

to bisphenol A at concentrations below those causing acute effects has been reported in 

fish, aquatic invertebrates, amphibians and reptiles. A selection of reported endpoint 

values relevant to potential hormonal disruption is presented in Table 7c. While there is a 

widespread variation in reported values, many fall in the range of 0.001 to 1 mg/L. In 

addition, differing sensitivities are evident between groups of organisms, with endpoint 

values for fish generally higher than those for aquatic invertebrates. Considered together, 

the data provide strong evidence that bisphenol A is capable of eliciting adverse effects: 

(1) following prolonged exposure at levels below those usually seen to elicit effects in 

standard toxicity tests (i.e., tests based on recognized methods which evaluate endpoints 

such as survival, reproduction and growth); (2) following brief low-dose exposure, 

particularly at sensitive developmental stages, with effects apparent later in the life cycle; 

(3) on filial generations following parental exposure; and (4) using more than one mode 

of action. 

19
Table 7c. Selected endpoint values relating 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 0.0011 Mandich et al. 2007 

Turbot 21 Altered steroid hormone 

balance 

0.0592 Labadie and Budzinski 

2006 

Guppy 21 Reduced total sperm count 0.2741 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 

0.001751 Lahnsteiner et al. 2005 

Medaka 110 Altered gonad development 0.0061 Metcalfe et al. 2001 

Invertebrates: 

Copepod 21 Delayed development 0.000011 Marcial et al. 2003 

Mussel 21 Induction of vitellogeninlike proteins and spawning 

in both sexes 

0.0502 Aarab et al. 2006 

Mussel 21 Resorption of male and 

female gonads 

0.0502 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 (2nd

generation) 

mouthpart deformities 

0.078 

0.010 

Segner et al. 2003 

20
Test organism Duration of 

test (days)

Endpoint observed Lowest effect 

value (mg/L)

Reference

Amphibians: 

Frog < 1 Competitive binding to 

estrogen receptor 

0.1073 Suzuki et al. 2004 

Frog 9 Suppressed metamorphosis 0.2281 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 

1.4001 

(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 for 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) 

NOEC1

LOEC2

NOEC 

32 mg/kg dw 

100 mg/kg dw 

≥ 100 mg/kg dw3

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 

EC254

EC505

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) NOAEL6

LOAEL7

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 causing a statistically significant effect in comparison to 

the controls. 

2

 Low observed effect concentration; lowest concentration in a toxicity test that caused a statistically significant effect in comparison 

to the controls. 

3

 Study identified that the highest concentration tested did not result in statistically significant results. Since the NOEC could be 

higher, the NOEC is described as being greater than or equal to the highest test concentration. 4

 The concentration which causes an effect on 25% of the test organisms. 5

 Median effective concentration. 

21
6

 No observed adverse effect level; highest dose in a toxicity test not causing a statistically significant effect in comparison to the 

controls. 

7

 Lowest observed adverse effect level; lowest dose in a toxicity test that caused a statistically significant effect in comparison to the 

controls. 

No statistically significant effects on reproduction were evident in earthworm, Eisenia

sp., exposed for 56 days to concentrations of up to 100 mg/kg dw of soil; however, 

statistically significant and concentration-dependent 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, corn, oat, soybean, tomato and wheat) (Dow 

Chemical Canada Inc. 2007). Effects were evident at both concentrations tested. No 

information on statistical endpoints was provided in the summary report submitted to 

Environment Canada. 

Ferrara et al. (2006) reported phytotoxicity and morphological anomalies in four plant 

species (broad bean, tomato, lettuce and durum wheat) grown hydroponically for 21 days 

at doses of 10 and 50 mg/L bisphenol A. 

ECB (2008) reported results from three unpublished studies conducted using springtail 

(Folsomia candida; ECT 2007a), enchytraeid worm (Enchytraeus crypticus; ECT 

2007b), and six terrestrial plant species (Springborn Smithers 2007). In springtail testing, 

no significant effects on survival occurred up to a highest exposure level of 1000 mg/kg 

dw; however, reproduction was significantly reduced at this concentration and the NOEC 

for this endpoint was therefore 500 mg/kg dw (EBC 2008). No significant effects on 

survival and reproduction were reported in enchytraeid worms up to a maximum 

exposure concentration of 100 mg/kg dw. Emergence in six species of terrestrial plants 

was significantly inhibited at a lowest exposure level of 320 mg/kg dw (cabbage, 

tomato), while the lowest effect level for growth was 50 mg/kg dw (tomato; ECB 2008). 

Furuya et al. (2006) reported on the development of male chicken phenotypes (comb, 

wattle and testes) in White Leghorn, Gallus domesticus, chicks administered oral doses of 

0.002 to 200 mg/kg-bw bisphenol A in corn oil/alcohol carrier every two days from 2 to 

25 weeks of age. Dose-dependent growth inhibition of all three endpoints occurred even 

at the minimum dosage, although recovery was evident at all but the highest dosage. This 

suggests development was delayed at lower doses and permanently altered at the highest 

dose of 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 rodents (see Health Effects Assessment section). Chronic oral 

exposure to low levels of contaminants, for example through ingestion in food, is more 

likely to be the scenario of concern for wildlife and there are some studies that, though 

limited when subject to a weight of evidence analysis, are suggestive of potential effects 

at doses well below the NOAELs of 5 and 50 mg/kg-bw per day established for bisphenol 

A in rodent species. There is also evidence that the pregnant female/fetus and developing 

offspring may be a sensitive subpopulation (see Health Effects Assessment section). A 

22
more detailed consideration of potential impacts 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 levels reported in water and wastewater 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 dw1

)

Number of 

samples 

Reference

Surface sediment Lake Erie, Canada; 2004 < 1.5×10-4 – 6.1×10-3 55 Chu et al. 2005 

MWWTP2 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° effluent3, 4

2° effluent5 

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 – 0.04996

0.104 – 36.7 

47 

51 

Lee and Peart 2000b 

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

23
Medium Location; year Concentration 

(mg/L or mg/kg dw1

)

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 

24
7

 Values estimated from graph 

8

 Median and maximum values are for different sample sets collected from two rivers in the same region 

The highest reported values for surface water and sediment were 0.012 mg/L and 0.0061 

mg/kg dw, respectively. Most surface water concentrations were much lower than the 

maximum reported value, however, with levels more in the range of 0.0001 to 0.001 

mg/L. In their study on sediment concentrations in Lake Erie, Chu et al. (2005a) reported 

that highest concentrations of bisphenol A (up to 0.0061 mg/kg dw) were measured in the 

western basin of the lake, likely reflecting the substantial loading and input from 

numerous wastewater treatment plants in the region into Lake Erie and the Detroit River, 

which empties into it. 

Rudel et al. (1998) reported a maximum concentration of 0.00141 mg/L in groundwater 

samples collected in the vicinity of municipal landfills, while up to 2.9×10-5 mg/L was 

detected in monitoring wells located downgradient of secondary wastewater treatment 

plant infiltration beds to which no wastewater had been discharged for six months. 

High levels have been measured in some industrial wastewaters, most notably those 

associated with paper and allied products (maximum 0.14923 mg/L; median 0.00872 

mg/L), chemicals and chemical products (maximum 0.09127 mg/L; median 0.00150 

mg/L) and commercial laundries (maximum 0.04345 mg/L; median 0.00656 mg/L) (Lee 

et al. 2002). 

Bisphenol A has also been detected in influent, effluent and sludge collected from 

municipal wastewater treatment plants across Canada. In general, effluent levels are 

lower than those found in influents from the same plant, indicating some removal from 

wastewater streams during processing. Lee and Peart (2000a) estimated a median 

reduction rate of 68% (range < 1-99%) based on concentration differences measured in 

36 influent/effluent sample pairs collected from Canadian municipal wastewater 

treatment plants over the period 1999-2000. The values incorporate consideration of 

losses through degradation and from partitioning to sludge. The presence of high 

concentrations in some sludge samples (up to around 40 mg/kg dw) confirms that 

bisphenol A partitions into this medium, even while some portion remains within the 

water phase. This supports the prediction made by Cousins et al. (2002; see 

Environmental Fate section) that based solely on properties of moderate hydrophobicity 

and water solubility, bisphenol A will partition into organic phases such as sediments and 

soils while an appreciable fraction remains in the dissolved phase. 

High levels of up to 4.6 mg/kg dw were measured in biosolid products produced by 

municipal wastewater treatment plants in seven different U.S. states (Kinney et al. 2006). 

The maximum concentration was found in a wet cake product intended for use in 

agriculture. 

Bisphenol A has been measured in surface waters, sediments, and wastewater treatment 

products in many European countries. These data are well summarized in ECB (2003). 

25
Limited information on measured levels in soil was found in the published literature, and 

this was more relevant to the assessment of human health (see Environmental Media 

section of Human Health Exposure Assessment). 

Only limited North American data were found for levels in biota; however, some 

European and Japanese data are available (see Table 9b). 

Table 9b. Concentrations of bisphenol A in biota 

Organism Location; year Concentration (mg/kg) Number of 

samples

Reference

Zooplankton Massachusetts, United 

States; 2003–2004 

< 4.0×10-6 – 0.0131 12 Stuart et al. 2005 

Clam Massachusetts, United 

States; 2003–2004 

0.0154, 0.02831 4 Stuart et al. 2005 

Fish The Netherlands; 1999 < 1.8×10-4 – 0.00262 6 Vethaak et al. 2005 

Fish: 

Freshwater 

Marine 

Norway; 2003 

0.001 – 0.01412

< 0.0017 – 0.061972

18 

6 

Fjeld et al. 2004 

Fish: 

Liver 

Muscle 

The Netherlands; 1999 

0.002 – 0.0751

0.001 – 0.0111

43

64

Belfroid et al. 2002 

Mussel The Netherlands; 1999 0.0101 1 Belfroid et al. 2002 

Snail Japan; 2002–2003 < 0.002–0.0111 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 screening assessment was to examine available 

scientific information and develop conclusions based on a weight-of-evidence approach 

and using a precautionary approach, as required under section 76.1 of CEPA 1999. 

Lines of evidence with respect to bisphenol A relate to persistence, bioaccumulation 

potential, toxicity, environmental occurrence, and trends in production and use. 

The available data indicate that bisphenol A does not persist under aerobic conditions. 

However, the substance has been found not to degrade or to degrade only slowly under 

conditions of low or no oxygen. This stability, combined with significant anthropogenic 

production and use, could lead to the formation of areas of accumulation or “sinks” in the 

environment, with the potential to act as future sources of exposure for organisms. 

The available data generally point to low bioaccumulation potential and a capacity for 

metabolism in various species. Most measured bioaccumulation and bioconcentration 

factors range upwards to approximately 147 L/kg; however, a lower trophic level study 

suggests a potential bioaccumulation factor of up to 650. These studies confirm that 

bisphenol A is bioavailable and can accumulate in tissues to some degree. 

26
Bisphenol A has been detected in surface waters, sediment, and groundwater over wide 

areas in Canada and the United States. It has also been measured in municipal and 

industrial wastewaters, sludge and biosolids. The published literature contains no 

reference to North American air concentrations; however, releases reported to the NPRI 

(see Table 3) were exclusively to the air compartment and therefore at least some 

presence in air can be expected from steady emission to the atmosphere. In addition, 

some limited data are available for concentrations in North American soils (see Exposure 

Assessment for Human Health). While few data exist for levels in North American biota, 

concentrations of up to 0.075 mg/kg dw are documented in fish and invertebrates from 

Europe and Japan. The data indicate that bisphenol A is present in a wide range of 

environmental media. 

Bisphenol A has shown high toxicity to aquatic organisms, with acute toxicity values 

falling below 13 mg/L and chronic values below 2 mg/L. The empirical and modelled 

data demonstrate that bisphenol A can be considered highly hazardous to the aquatic 

environment. 

A quantitative evaluation of exposure and ecological effects was conducted for bisphenol 

A as part of the weight of evidence evaluation of its potential to cause harm. Predicted 

Environmental Concentrations (PECs) were determined based on an analysis of exposure 

pathways, and these exposure pathways were used to identify sensitive receptor 

organisms. For bisphenol A, PECs were derived using Canadian environmental 

monitoring data, or if monitoring data were not available, the PECs were based on simple 

calculation procedures that took into account some degree of local environmental 

conditions, but largely relied on generic environmental parameters. A Predicted NoEffects Concentration (PNEC) was determined for each exposure pathway, by dividing a 

Critical Toxicity Value (CTV) by an application factor. CTVs typically represented the 

lowest ecotoxicity value (e.g., an EC10, LOEC, etc.) from an acceptable data set. A risk 

quotient analysis, integrating estimated potential exposures with potential adverse effects, 

was performed and the resulting risk quotient (PEC/PNEC) was used to estimate the 

potential for ecological risk. 

An effluent concentration of 0.0173 mg/L was reported for a municipal wastewater 

treatment plant in the Toronto area (Lee et al. 2004). This value is selected for derivation 

of the Predicted Exposure Concentration (PEC) for pelagic organisms as it represents the 

highest reliable Canadian effluent concentration from a municipal treatment plant 

providing both primary and secondary wastewater treatment and is therefore deemed both 

realistic and conservative. Applying a dilution factor of 10 (to account for exposures in 

the immediate mixing zone) yields a PEC value of 0.00173 mg/L. Lahnsteiner et al. 

(2005) reported significantly reduced semen quality and delayed ovulation in the brown 

trout, Salmo trutta f. fario, at a Lowest Observed Effect Concentration (LOEC) 

concentration of 0.00175 mg/L and this value is selected as the CTV. To derive a 

Predicted No-Effects Concentration (PNEC), the CTV is divided by an application factor 

of 10 in order to account for variability in interspecies and intraspecies sensitivity and 

extrapolation from laboratory to field conditions. The resulting PNEC for the pelagic 

compartment is 0.000175 mg/L and the risk quotient (PEC/PNEC) is 0.00173/0.000175 = 

27
9.9. The results indicate that bisphenol A concentrations in water have the potential to 

cause adverse effects on populations of pelagic organisms in Canada. 

For the sediment compartment, Johnson et al. (2005) reported a LOEC of 1.0 mg/L for 

reduced adult survival in the freshwater amphipod, Gammarus pulex, and this value is 

selected as the CTV. While this testing used water only as the exposure medium, the 

results reflect adverse effects on an epibenthic organism and are therefore considered 

appropriate for use in estimating potential risk to sediment species. A higher assessment 

factor of 100 is selected due to the paucity of data for this environmental compartment, 

and the resulting PNEC is 0.01 mg/L. 

While measured concentrations are available for sediment (e.g., Chu et al. 2005), the PEC 

value of 0.00173 mg/L derived for the pelagic compartment will also be used in the 

calculation of a risk quotient for the sediment compartment. This approach allows direct 

comparison with the PNEC value, which was determined in water-only testing, and is 

reasonable given that a significant exposure route for sediment species will be by means 

of interstitial water or water in the boundary layer directly above the sediment surface. 

The risk quotient for benthic species is then 0.00173 / 0.01 = 0.173. The risk quotient 

analysis results suggest there is currently a low risk of direct adverse effects to sediment 

organisms in Canada. 

The limited soil concentration data were deemed more relevant for the assessment of 

exposure to humans (see Exposure Assessment, Human Health); however, several studies 

provide Canadian data on measured levels of bisphenol A in sewage sludge. Application 

of sewage sludge to agricultural lands represents a direct pathway for bisphenol A into 

soil. A maximum concentration of 36.7 mg/kg dw was reported for digested sewage 

sludge collected from a municipal wastewater treatment plant in southern Ontario (Lee 

and Peart 2000a,b), and this value is selected for derivation of the PEC. The PEC is 

estimated for tilled agricultural soil and pastureland based on an approach described in 

Bonnell Environmental Consulting (2001). This approach uses the same basic approach 

as outlined in the EU’s Technical Guidance Document (European Communities 1994) 

but modified for Canadian conditions. The following assumptions were used: dry sludge 

is applied annually to soil at the rate of 0.5 kg/m2

; soil has a bulk density of 1700 kg/m3; 

sludge is mixed in soil to a depth of 0.2 m (depth of tillage) in agricultural soil and 0.1 m 

in pastureland (Bonnell Environmental Consulting 2001 and European Communities 

1994); no loss of the target substance due to erosion; no input from atmospheric 

deposition; and no background bisphenol A accumulations in the soil. It is also assumed 

that bisphenol A persists in the soil for a period of at least 56 days, the exposure period of 

the study used for the derivation of the PNEC value (see below). 

The PEC is then: PECagricultural soil = (36.7 x 0.5) / (0.2 x 1700) 

 = 0.054 mg/kg dw 

PECpastureland = (36.7 x 0.5) / (0.1 x 1700) 

 = 0.11 mg/kg dw 

28
A 56-d LOEC of 100 mg/kg dw was reported for significant mortality in the earthworm, 

Eisenia sp. (Johnson et al. 2005), and this value is selected as the CTV. A lower toxicity 

value of 10 mg/L was reported for significantly reduced growth in several terrestrial plant 

species (Ferrara et al. 2006), however, this value is not suitable for comparison with the 

PEC as the plant study was reported in terms of water concentrations only. An 

application factor of 100 is applied to the CTV in order to account for extrapolation from 

laboratory to field conditions and interspecies and intraspecies variations in sensitivity. 

The application factor is larger than that used for pelagic organisms as fewer data are 

available for soil species and therefore the uncertainty in extrapolating between species is 

greater. The resulting PNEC value is 1.0 mg/kg dw. No information is available on the 

organic carbon (OC) content of the sludge samples or the soil used in the earthworm 

study, therefore both will be assumed to have a standard OC content of 2%. The risk 

quotient, PEC/PNEC, for soil organisms is therefore 0.054/1.0 = 0.054 for tilled 

agricultural soil and 0.11/1.0 = 0.11 for pastureland. The risk quotient analysis results 

suggest there is currently a low risk of direct adverse effects to soil species in the 

Canadian environment. 

A PEC for mammalian wildlife was estimated based on a calculation of the total daily 

intake of the target substance by mink and otter. An energetics model based on the 

general exposure model for wildlife from the U.S. Environmental Protection Agency’s 

(EPA) Exposure Factors Handbook (USEPA 1993) was used: 

where, 

TDI = total daily intake (mg/kg-bw per day) 

FMR = normalized free metabolic rate of wildlife receptor of interest (236 kcal/kg-bw 

per day for mink and 183 kcal/kg-bw per day for river otter) 

Ci = concentration of contaminant in the ith prey species (mg/kg-bw) (see below) 

Pi = proportion of the ith prey species in the diet (unitless) (default = 35% for mink; 

100% for otter) 

GEi = gross energy of the ith prey species (default = 1240 kcal/kg-bw prey) 

AEi = assimilation efficiency of the ith prey species by the wildlife receptor (default = 

0.91) 

Pt = proportion of the time the receptor spends in the contaminated area (50% for 

mink and 50% for otter) 

The model incorporated the metabolic rate of the wildlife receptors of interest (mink and 

otter), the proportion of food uptake by the receptors and the amount of time the animals 

spend in the contaminated area, which is based on the typical habitat range of the wildlife 

receptors. 

29
As there are no Canadian data for bisphenol A concentrations in fish, concentrations of 

the target substance in fish (Ci) were estimated based on the highest PECwater and a BAF.

The BAF was estimated using the GOBAS BAF three trophic level model (Arnot and 

Gobas 2003; see Bioaccumulation section). The BAF represents a benthic/pelagic food 

chain and estimates the accumulation from all sources in a mid-trophic-level fish that 

would typically be eaten by a mammalian piscivore. 

Ci = PECwater · BAF 

where: 

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

PECwater = PEC of 0.00173 mg/L calculated above for the pelagic compartment 

BAF = bioaccumulation factor of 148 L/kg estimated for mid-trophic level fish based 

Arnot and Gobas (2003) 

Ci = 0.00173 · 148 = 0.256 mg/kg-bw 

The model estimated PECs of 0.009 mg/kg-bw per day and 0.021 mg/kg-bw per day for 

mink and otter, respectively. 

As identified in Characterization of Risk to Human Health, Health Effects Section, the 

dataset of neurodevelopmental and behavioural studies, though limited when subject to a 

weight of evidence analysis, is suggestive of potential effects at doses in the range of 0.01 

to 0.1 mg/kg-bw per day and higher (study details are provided in Appendix D). The 

range of 0.01 – 0.1 mg/kg-bw per day is selected as an estimate for CTVs in mammalian 

wildlife. The CTVs for predictive sentinel wildlife species (in this case, mink and otter) 

can then be calculated from the range of CTVs of the mammalian test species, using the 

formula (Sample et al. 1996): 

 CTVwildlife = CTVts · (bwts/bww)

1/4

where: 

CTVwildlife = critical toxicity value for wildlife 

CTVts = critical toxicity value for test species 

bwts = mean body weight of test species (0.35 kg) 

bww = mean body weight of predictive sentinel species (0.807 kg for mink, 6.01 kg for 

otter; Canadian Wildlife Service, Ontario Region, pers. comm., 2004, unreferenced). 

Therefore, the CTVwildlife ranges are then 11 to 112 mg/kg-bw per day for mink and 42 to 

420 mg/kg-bw per day for otter. Applying an application factor of 10 to account for 

extrapolation from laboratory to field conditions yields PNEC values of 0.0008 to 0.008 

mg/kg-bw per day and 0.0005 to 0.005 mg/kg-bw per day for mink and otter, 

respectively. The resulting risk quotients (PEC / PNEC) are 1.25 to 12.50 for mink and 

4.2 to 42.0 for otter. It is concluded that bisphenol concentrations in biota have the 

potential to cause adverse effects in populations of wildlife in Canada. 

30
A similar methodology was used to derive a risk quotient relevant to avian wildlife 

species. The total daily intake (TDI) was calculated for a sentinel avian wildlife species, 

the lesser scaup (Aythya affinis). The lesser scaup was selected as a suitable avian 

sentinel species as it is abundant through much of North America and consumes clams as 

part of its normal diet of aquatic invertebrates and vegetation (USEPA 1993). As clams 

are present in the diet of lesser scaup, the tissue concentration of 0.0283 mg/kg dw 

reported in clam collected off the Massachusetts coast (Stuart et al. 2005) is selected for 

use in the development of a PEC for avian wildlife. This concentration is converted to 

0.00535 mg/kg wet weight (ww) using a dry weight to wet weight ratio of 5.29 for clams 

(Vinogradov 1953). The total daily intake (TDI) for lesser scaup is then: 

where, 

TDI = total daily intake (mg/kg-bw per day) 

FMR = normalized free metabolic rate of wildlife receptor of interest (216 kcal/kg-bw 

per day for adult female scaup; USEPA 1993) 

Ci = concentration of contaminant in the ith prey species (mg/kg-bw) (0.00535 

mg/kg ww) 

Pi = proportion of the ith prey species in the diet (unitless) (19.2% for adults; Dirschl 

1969) 

GEi = gross energy of the ith prey species (default = 800 kcal/kg-bw prey) 

AEi = assimilation efficiency of the ith prey species by the wildlife receptor (default = 

0.77) 

Pt = proportion of the time the receptor spends in the contaminated area (assume 

50%) 

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

A LOAEL of 100 mg/kg-bw per day, based on permanently altered development in male 

White Leghorn chicks, Gallus domesticus, was reported by Furuya et al. (2006), and this 

value will be used as the CTV for the derivation of a PNEC for avian wildlife. The CTV 

for lesser scaup can then be calculated using the same formula as for mammalian wildlife 

(Sample et al. 1996). Sample et al. (1996) recommend using a scaling factor of 1 for 

interspecies extrapolation among birds; however, as the reported mean body weight for 

the lesser scaup considers the adult bird, while the toxicity value for Leghorn was derived 

using juveniles, an extrapolation which considers body weight was deemed appropriate in 

order to incorporate consideration of weight differences relating to life stage. 

The CTV for avian wildlife is then: 

31
 CTVwildlife = CTVts · (bwts/bww) 

where: 

CTVwildlife = critical toxicity value for wildlife 

CTVts = critical toxicity value for test species (100 mg/kg-bw per day) 

bwts = mean body weight of test species (1.3 kg; estimated from Leeson and Caston 

1991) 

bww = mean body weight of predictive sentinel species (0.77 kg; Nelson and Martin 

1953) 

The CTV for avian wildlife is therefore 169 mg/kg-bw per day. Applying an application 

factor of 100 to account for extrapolation from laboratory to field conditions and 

interspecies and intraspecies variations in sensitivity yields a PNEC value of 1.69 mg/kgbw per day and the risk quotient is then 0.00018/1.69 = 0.0001. The results suggest there 

is currently a low risk of direct adverse effects to avian wildlife species in Canada. 

Considered together, the various lines of evidence indicate that the potential impacts of 

bisphenol A in the Canadian environment are of sufficient magnitude to warrant use of a 

precautionary approach in response to uncertainties in the evaluation of risk. 

There is evidence that bisphenol A is capable of altering hormonal, development or 

reproductive function: following prolonged exposure at levels below those usually seen 

to elicit effects in standard toxicity tests (i.e., tests based on recognized methods which 

evaluate endpoints such as survival, reproduction and growth); following brief low-level 

exposure, particularly when this occurs at sensitive developmental stages, with effects 

apparent later in the life cycle; in filial generations following parental exposure; and 

using more than one mode of action. Studies conducted using fish, aquatic invertebrates 

and amphibians demonstrate impacts in the range of 0.001 to 1 mg/L, levels that have 

been measured in Canadian and U.S. wastewaters, receiving waters and sediments. For 

this reason, potential exists for adverse effects in populations of aquatic organisms, 

particularly those residing in close proximity to treatment plant outfalls or other point 

sources. While the ability of 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 measured in the Canadian environment and the 

potential for adverse effects in organisms. 

 

Bisphenol A is one of the highest volume chemicals produced worldwide and global 

demand for the substance is predicted to continue increasing at a rate of 6 % to 10% 

annually. The total quantity of bisphenol A reported as imported or in use in Canada 

during the calendar year 2006 was approximately 0.5 million kg. Actual quantities 

available for release to the Canadian environment could be much higher, however, as 

these values likely do not fully capture amounts entering Canada in finished and semi32
finished products. This is particularly so, given the wide application of the substance and 

number of bisphenol A derivative products. 

In summary, the lines of evidence relating to persistence, bioaccumulation, toxicity, 

environmental occurrence and trends in production and use raise concerns regarding the 

presence of bisphenol A in the environment. The substance is predicted to degrade fairly 

rapidly under aerobic conditions (half-life in the order of days), suggesting that bisphenol 

A released into receiving waters should remain only a short time in this medium. 

However, it has been widely detected in surface waters. Possible explanations for this 

may be that because inputs to the environment are sufficiently large and continuous, 

measurable concentrations are always present and/or that the substance is not degrading 

as rapidly as laboratory testing would indicate. Measurable levels in sewage sludge and 

effluent are also suggestive of some level of stability of this substance. Further, bisphenol 

A is present in media to which there is no direct release, such as sediment and 

groundwater. This implies that the substance remains sufficiently long enough in the 

environment to move from its point of release into other environmental media. There is 

evidence that bisphenol A is stable in some environmental compartments, most notably 

under conditions of low oxygen, such as those found beneath the surface layers of 

sediment and soil and in marshes and bogs. Levels in the environment are expected to 

increase given that bisphenol A is a high production chemical and global demand for the 

substance is increasing. Bisphenol A has been detected in aquatic species, confirming that 

the substance can be taken up by organisms and stored in tissues. The potential for food 

web impacts is still undetermined, although bioaccumulation factors of 650 have been 

measured in lower trophic level species, suggesting that possible exposure of predator 

species could occur through consumption of prey. Bisphenol A is acutely toxic to aquatic 

organisms and adversely affects reproduction in earthworms, growth in terrestrial plants, 

and development in birds. Results from an analysis of risk quotients, which compare the 

potential for exposure with possible effects in organisms, indicate that bisphenol A 

concentrations in the Canadian environment have the potential to cause adverse effects in 

populations of pelagic organisms and mammalian wildlife but are unlikely to cause 

adverse effects in benthic and soil organisms and bird species in Canada. In addition to 

acute toxicity, bisphenol A can impact the normal development of individual organisms 

and influence the development of their offspring. In laboratory testing, these effects have 

occurred at concentrations below those shown to cause acute effects, and these effect 

concentrations have been measured in the Canadian environment. Considered together, 

the various lines of evidence indicate that the potential impacts of bisphenol A in the 

Canadian environment are of sufficient magnitude to warrant use of a precautionary 

approach in response to uncertainties in the evaluation of risk. 

Uncertainties in Evaluation of Ecological Risk 

There is uncertainty regarding sources of release and release quantities of bisphenol A 

into the Canadian environment. An industry survey was used to collect information on 

Canadian manufacture, import and use of bisphenol A for the calendar year 2006. 

However, it is not clear whether potential releases from partially-finished and finished 

33
goods being imported into Canada and from products currently in use in Canada have 

been adequately captured. This information is needed to more accurately estimate 

potential exposure concentrations for organisms in the Canadian environment. The 

presence of a recent database of measured environmental concentrations for some 

locations in Canada and the United States assisted with characterizing exposure potential 

in this assessment. However, quantitative exposure estimates were possible only for 

pelagic organisms. Increasing the size of this database by including a wider range of 

Canadian locations, more information on possible time trends, and more data on 

sediment, soil and surface water concentrations, as well as possible presence in leachates 

and groundwater, would increase the level of certainty in exposure estimation and allow 

estimation of potential exposure for benthic and terrestrial species. 

There is uncertainty about the measured presence and accumulation potential of 

bisphenol A in biota. The available information indicates that bisphenol A does not meet 

bioaccumulation criteria as specified in the Persistence and Bioaccumulation Regulations

of CEPA 1999. However, bioaccumulation factors of up to 650 have been determined for 

lower trophic level aquatic species, and this suggests there may be circumstances or 

conditions under which bisphenol A may accumulate within organisms. Lower trophic 

aquatic organisms form an important bridge between primary producers (i.e., 

phytoplankton) and higher trophic species such as fish. Therefore, while high tissue 

levels could indicate the presence of a site-specific condition, such as very high local 

concentrations, it is also possible that bioaccumulation is occurring with the subsequent 

potential for food chain transfer and secondary poisoning of predator species. An 

increased database of measured concentrations in Canadian biota, including trophic 

magnification studies, would provide greater clarity on the potential for accumulation 

within individual organisms and along food webs. 

There is uncertainty regarding the potential for adverse effects in Canadian biota resulting 

from single-dose or prolonged (i.e., lifetime and multigenerational) low-level exposure to 

bisphenol A, particularly during sensitive life cycle stages. This uncertainty reflects a 

need for greater clarification of minimum effects levels in pelagic, sediment, soil and 

wildlife species, and of exposure concentrations in the Canadian environment. Based on 

the weight of available data, the assessment indicates that there is a potential for adverse 

impacts in aquatic organisms, particularly those residing in the vicinity of treatment plant 

outfalls or other point sources. However, information is needed in order to more clearly 

delineate the relationship between exposure levels in the Canadian environment and 

possible effects in organisms. 

34
Potential to Cause Harm to Human Health 

Exposure Assessment 

Dietary Intake 

With respect to dietary intake, two sources of potential exposure are (i) migration into 

food and beverages from packaging materials that directly contact food, specifically food 

and beverage containers with internal epoxy resin coatings and (ii) migration from 

polycarbonate repeat use containers such as baby bottles and drinking bottles. 

(i) Migration from Food Packaging 

Epoxy resins are used as an interior protective lining for food and beverage cans and for 

metal closures for some jars and bottles. As a result of these food contact uses, very small 

quantities of the monomer bisphenol A can migrate into the food and beverage contents 

of the containers. 

Exposure to bisphenol A from food packaging uses has been characterized based on 

bisphenol A concentrations reported in a survey of canned foods on the U.S. market in 

2007 (EWG 2007). EWG (2007) tested foods and beverages from 97 cans purchased in 

grocery stores in three states. Twenty-eight different types of foods were analyzed for 

bisphenol A, including canned fruits, vegetables, pasta, beans, infant formula, meal 

replacement and canned milk. Results are presented in Table 10. Reported concentrations 

were considered to be representative of additional data available to Health Canada and of 

data reported in the literature (EFSA 2006). 

Table 10. Average 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 Nd2-385 

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 

35
1 Average is geometric mean 

2

 Not detected at the limit of detection 1 ppb 

Potential daily intakes from the use of epoxy resins as interior protective lining for food 

and beverage cans were estimated using average and maximum concentrations of 

bisphenol A, and daily intake values for the food commodities (Health Canada, Food 

Directorate, Health Products and Food Branch [HPFB], pers. comm., 2007 March 14, 

unreferenced, and are presented in Table 11. Leaching from metal closures was 

considered to contribute minimally to these exposures. 

Table 11. Potential daily intake of bisphenol A (μg/kg-bw per day) from use of epoxy 

resins as interior protective lining for food and beverage cans

1-4 year olds3 5-11 year olds3 12-19 year olds3 Adults3 Canned 

food 

Avg1 Max2 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 attributed this to additives in ink in paper 

used in the food packaging. Table 9b presents concentrations of bisphenol A found in 

aquatic biota. Seafood purchased in a Singapore market was analysed by Basheer et al. 

(2004) and mean concentrations (n=5 for each product) were reported as 13.3 ng/g in 

prawns, 213.1 ng/g in crabs, 56.5 ng/g in blood 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 3g/g in fresh market 

vegetables in Italy (8/14 detects). 

As there are no measured concentrations of bisphenol A in fish caught in Canada or 

North America, or of fish available on the Canadian market, the maximum concentration 

reported in fish from European countries (see Table 9b) was used to derive estimates of 

intake from fish consumption. Using a maximum concentration of 61.97 μg/kg as 

measured by Fjeld et al. (2004) and based on rates of consumption of fish for Canadians 

(Health Canada 1998), the estimated intake of bisphenol A from this source would range 

36
from 0.11 μg/kg-bw per day (for 12-19 year olds) to 0.24 μg/kg-bw per day (for 1-4 year 

olds). Studies have also indicated that concentrations of bisphenol A tend to be higher in 

fish liver and other organs (Belfroid et al. 2002) so consumption of these organs may 

slightly increase exposure. Confidence in these estimates is low as they are based on a 

number of assumptions; however it is likely that they are an overestimate of actual 

exposure from this source. 

As infant formula (ready-to-use, concentrate, powder) available on the Canadian market 

is packaged in cans, a separate estimate of the dietary intake of bisphenol A by infants 

was conducted. Due to the nature of the extracting medium, migration into liquid is 

considered to be greater than migration into powder and analysis therefore focussed on 

liquid formulations. Preliminary research conducted at Health Canada supports the 

conclusion that migration of BPA is significant only for liquid infant formulas (Health 

Canada, Food Directorate, HPFB, pers. comm., 2008 Sept 04, unreferenced). In 2007, 

Health Canada analyzed canned liquid infant formula to characterize residual levels of 

bisphenol A. Samples packaged in cans lined with epoxy-based coatings were collected 

in October, 2007 in Ottawa, Canada from 21 cans representing eight brands of ready-todrink or concentrated liquid infant formulas. Bisphenol A was detected and quantified in 

all samples, with concentrations ranging from 2.27 to 10.2 ppb (Cao et al. 2008). These 

data are consistent with other data reported in the literature. For instance, EWG (2007) 

reported average bisphenol A concentrations of 2.4 ppb (range from limit of detection of 

1 ppb to 17 ppb) in 6 cans of infant formula available on the US market. The Canadian 

data were considered to be the most relevant for use in the assessment. As a worst-case 

scenario it was assumed that liquid infant formula is the main source of food for infants 

in the first 18 months of their life, and that their formula intake is equal to their daily total 

fluid intake (i.e., formula, milk, water, juice). A dilution factor of 0.5 was applied to the 

concentrated liquid formulas as they would be diluted with water prior to the feeding (1 

part water to 1 part concentrated formula). No dilution factor was applied to the ready-touse infant formulas. Assumptions regarding formula intakes are based on 

recommendations of the Institute National de Santé Publique du Québec (INSPQ 2001). 

Potential daily intake of bisphenol A from liquid infant formula packaged in containers 

with epoxy resin interior protective lining is provided in Table 12. 

Table 12. Potential daily intake of bisphenol A (μg/kg-bw per day) from liquid 

infant formula packaged in containers with epoxy resin interior protective lining. 

Intake based on average 

bisphenol A concentration1

Intake based on maximum 

bisphenol A concentration Age of infants 1

Average 

formula intake

Maximum 

formula intake

Average 

formula 

intake

Maximum 

formula intake

0 to 1 month2 0.45 0.75 0.81 1.35 

2 to 3 months3 0.50 0.69 0.96 1.31 

4 to 7 months4 0.38 0.52 0.75 1.02 

8 to 12 months5 0.21 0.28 0.42 0.55 

12 to 18 months6 0.23 0.27 0.38 0.46 

37
1

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 according to the age of use intended by the manufacturer. 2

Based on average female infant weight of 3.9 kg (NCHS 2000) and an average formula intake of 644 

g/day and maximum formula intake of 1080 g/day (INSPQ 2001).The average and maximum 

concentrations of 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 bisphenol A through the consumption of human 

breastmilk, however no Canadian data on concentrations of bisphenol A in breastmilk 

were identified. In a survey of 20 U.S. lactating women with no known occupational 

exposure, Ye et al. (2006) identified concentrations of total bisphenol A (free + 

conjugate) ranging from the limit of detection (1 ng/ml) to 7.3 ng/mL (7.1 ppb – 

assuming breastmilk has a density of 1.03 g/mL (Health Canada 1998)), with a mean 

concentration of 1.9 ng/mL (1.8 ppb). Sun et al. (2004) reported a mean bisphenol 

concentration in 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 limit of detection (0.09 ng/g) to 0.70 ng/g (ppb) 

of bisphenol A in the breastmilk of three women. 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 (μg/kg-bw per day) from 

consumption of human breastmilk containing either average 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 month2 0.28 1.09 

2 to 3 months3 0.21 0.84 

4 to 7 months4,5 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). 

38
2

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 repeat-use containers 

such as baby bottles, drinking bottles, pitchers and carboys can contribute to oral 

exposure. Numerous studies have reported on migration of bisphenol A from 

polycarbonate bottles into water or milk simulant under different conditions (Brede et al. 

2003; Biedermann-Brem et al. 2007; Biles et al. 1997; Cao and Coriveau (2008); Ehlert 

et al. 2008; Kawamura et al. 1998; 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 conditions using food simulants such as water, 10% ethanol, 

50% ethanol and 3% acetic acid. In considering these results, it is important to 

differentiate between study designs intended to simulate realistic exposure scenarios 

versus those designed to characterize maximum concentration that may migrate from the 

product under extreme conditions or over the life of a container. 

Five studies of bottles obtained from the Canadian market are available. In 2000, Health 

Canada conducted a series of limited migration studies (Page et al. 2006). No bisphenol 

A was detected in milk simulant (50% ethanol/water) following storage in polycarbonate 

baby bottles for up to 7 days at 4 or 22o

C (detection limit of <0.1 μg/kg). In additional 

trials, bottles were subjected to the following treatments: (i) storage for 7 days at 4oC then 

storage for an additional 7 days at 22o

C, followed by 6 days at 70oC; (ii) storage for 7 

days at 22o

C then storage for 7 days at 4oC, followed by 6 days at 70oC C; (iii) storage for 

6 days at 70o

C. In these 3 trials, aliquots of simulant were collected after 1, 2, 3 and 6 

days at 70o

C and analyzed for bisphenol A. For all three trials, bisphenol A 

concentrations in the simulant on day 1 were 0.4 to 1 ppb, and concentrations increased 

on subsequent days of incubation at 70o

C. Sustained incubation at 70oC is not considered 

a meaningful study design for characterization of migration under realistic use conditions. 

A second study also measured the migration of bisphenol A from polycarbonate baby and 

drinking bottles into water. Cao and Corriveau (2008) tested five polycarbonate bottles, 

three baby bottles and two refillable drinking bottles, bought in Ottawa, Canada in 2007. 

The bottles were filled to capacity with boiling water and then left at room temperature 

for 24 hours. Bisphenol A concentrations in the water ranged from 1.7 to 4.1 μg/L (ppb). 

 

Preliminary raw data are available from a study conducted by the Health Canada in 2008 

(Health Canada, Safe Environments Programme, Healthy Environments and Consumer 

39
Safety Branch (HECSB), pers. comm., 2008 March 6, unreferenced). Migration was 

characterized by filling 14 brands of baby bottles available in Canada with water to 

simulate migration to aqueous and acidic foods, and a 50% ethanol/water solution to 

simulate migration to fatty foods such as infant formula. The bottles were incubated for 8, 

24 or 240 hours at 40°C to simulate use at room temperature, as per recommendations 

from the USFDA (2007). The 240-hour scenario (10 days) was included to estimate 

migration from both repetitive use and worst-case scenarios. Average migration for these 

tests ranged from 0.095 μg/L (ppb) in water incubated for 8 hrs, to 2.05 μg/L in the 50:50 

ethanol/water blend incubated for 240 hours. The scenario considered to represent the 

most realistic use consisted of filling bottles with the 50:50 ethanol/water blend and 

incubating them for 8 hours. This scenario resulted in average concentrations of 0.15 

μg/L. 

In another study of polycarbonate baby bottles on the Canadian market (Environmental 

Defence 2008), new baby bottles (n=9) were filled with water, sealed and allowed to sit 

for 24 hours; initially at room temperature and then in a subsequent trial at 80oC. The 

study investigators specify that the 80oC treatment was considered to simulate repeat 

washing of bottles (i.e., approximately 60-100 washes). As such, the 80o

C trials are not 

considered to represent a realistic migration from a single use. Results from the room 

temperature trials showed that bisphenol A levels ranged from below the limit of 

detection, i.e. 0.05 to 0.063 ng/mL (ppb). Results from the 80oC 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 37oC before the milk was tested for 

bisphenol A. The same bottles were then rewashed and placed under similar 

experimental conditions to test migration into apple juice; the only difference in protocol 

was that the juice was brought to room temperature. The concentration range of bisphenol 

A measured in milk migrating from new, hard polycarbonate bottles ranged from 0.44 to 

0.53 μg/L (n=10) and was reported to be 0.29 μg/L in used bottles (n=5). The migration 

concentrations measured in apple juice ranged from 0.29 to 0.46 μg/L in new bottles 

(n=10) and was 0.43 μg/L when measured in the used bottles (n=5). For both liquids, 

bisphenol A concentrations from the disposable plastic bags and glass bottles were below 

the detection limit (0.2 μg/L). 

Maragou et al. (2007) assessed migration potential from 31 new polycarbonate baby 

bottles (from 6 different brands) available on the Greek market under a variety of 

conditions, including those which study investigators considered to be representative of 

actual use. To determine the impact of the repeated cleaning of bottles, either with the 

dishwasher or with brushing, the bottles were cleaned using either method, rinsed and 

sterilized prior to being filled with water and incubated at 70o

C for 2 hours. This series of 

treatments did not lead to detectable bisphenol A migration with a reported detection 

limit of 2.4 ppb. Migration was only detected when bottles were filled with boiling water 

40
(100o

C). Specifically, 5 repeat cycles of cleaning the bottles by brushing with detergent, 

sterilizing in boiling water for 10 minutes, then filling with boiling water (100o

C) and 

leaving at ambient temperature for 45 minutes 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 μg/L (ppb) 

of bisphenol A in the food simulant contents. In a study of bottles filled with boiling 

water and kept overnight at room temperature Hanai (1997) measured bisphenol A 

migration varying from 3 to 55 μg/L (detection limit of 2 μg/L). Earls et al. (2000) 

detected bisphenol A in 5 out of 12 bottles filled with either boiling water or 3% acetic 

acid, placed in a refrigerator for 24 hours, and then heated to approximately 40°C by 

immersion in boiling water for approximately 2 minutes. Bisphenol A concentrations in 

detected samples ranged from 20 to 50 μg/L (detection limit of 10 μg/L). 

However, the Norwegian Food Safety Authority (Biedermann-Brem et al. 2007) 

investigated the effect of migration of bisphenol A from polycarbonate under extreme 

washing conditions such as washing using very strong alkali detergents (80oC for 1 hour, 

followed by drying of unrinsed bottles at 90oC for 30 minutes). They concluded that even 

under aggressive conditions, concentrations of bisphenol A in contents are unlikely to 

exceed 10 μg/L (ppb). 

A recent study investigated migration behaviour of bisphenol A into water during 

microwaving of polycarbonate baby bottles obtained from the European market (Ehlert et 

al. 2008). Polycarbonate bottles filled with water were heated to 100oC during 3 

microwave cycles. Migration of bisphenol A into water was shown to be in the range of 

<0.1 to 0.7 μg/L (ppb). 

Miyamoto and Kotake (2006) reported bisphenol 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 95o

C for 30 minutes. 

Polycarbonate carboys can be used to store drinking 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 DDH20 wash (2x20ml), with analysis by 

GC-MSD. The range of bisphenol A was 0.1 to 4.7 ng/L (0.0001 to 0.0047 ppb) 

(detection limit < 0.05 ng/L). 

Le et al. (2008) measured bisphenol A migration from polycarbonate drinking bottles 

used for consumption of water and other beverages. New and used polycarbonate bottles 

were filled with water and left for up to 7 days at room temperature (22o

C). Samples were 

collected on days 1, 3, 5 and 7. On day 1, the calculated mean concentration of bisphenol 

A detected in the water was 0.24 ng/mL (ppb). Mean concentrations across all sampling 

days, for both new and used polycarbonate containers, ranged from 0.08 to1.33 ng/mL 

(ppb). In general, the concentration of bisphenol A released from polycarbonate bottles 

41
into water at room temperature increased with time, but no significant difference was 

observed between the concentrations resulting from the use of new versus used bottles. 

Repeating the study using boiling water (100o

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 appropriate to estimate bisphenol A intake from 

the use of polycarbonate baby bottles for both the recommended room temperature use 

and plausible high temperature use (boiling water) scenarios. Results calculated from Le 

et al (2008) indicate that filling polycarbonate bottles with room temperature water, could 

result in migration of approximately 0.24 ppb of bisphenol A into the contents of the 

bottle. This is supported by preliminary data from a migration study conducted by Health 

Canada indicating an average concentration of 0.15 ppb bisphenol A (Health Canada, 

Safe Environments Programme, HECSB, pers. comm., 2008 March 6, unreferenced). To 

estimate migration of bisphenol A from addition of boiling water [100°C] directly to the 

baby bottle, the results of Maragou et al. (2007), in which boiling water was added to 

polycarbonate bottles and cooled for 45 minutes, support a concentration of 10 ppb 

bisphenol A. As a worst-case scenario it was assumed that liquid infant formula is the 

main source of food for infants in the first 18 months of life, and so formula intake is 

considered to be equal to their daily total fluid intake (i.e., formula, milk, water, juice). 

Exposure estimates are provided in Table 14. 

Table 14. Contribution of use of polycarbonate baby bottles to potential daily intake 

(PDI) of bisphenol A (μg/kg-bw per day) for infants. 

PDI of bisphenol A (μg/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 month3 0.040 0.066 1.65 2.77 

2 to 3 months4 0.047 0.064 1.95 2.67 

4 to 7 months5 0.035 0.049 1.46 2.00 

8 to 12 months6 0.020 0.026 0.82 1.07 

12 to 18 months7 0.017 0.020 0.71 0.85 1

Based on an average concentration of bisphenol A detected in new and used polycarbonate bottles that 

were filled with room temperature liquid and left for 24 hours (Le et al. 2008). This value is also supported 

by an average concentration of 0.15 ppb reported from filling bottles with liquid at room temperature and 

42
leaving for 8 hours. (Health Canada, Safe Environments Programme, HECSB, pers. comm., 2008 March 

06, unreferenced). 

2

Based on an average concentration of bisphenol A resulting from filling polycarbonate baby bottles with 

boiling water and leaving at room temperature for 45 minutes (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 repeat use containers are available to help 

determine frequency and method of use of polycarbonate repeat-use bottles by other 

segments of the population (i.e., children aged 1-4 years, 5-11 years, 12-19 years and 

adults). Therefore exposure estimates were based on a migration of 0.24 ppb of bisphenol 

A as calculated from Le et al. (2008), resulting from the room temperature use of 

polycarbonate bottles, as it was assumed that contents would typically not be heated prior 

to serving for these age groups. This value is supported by preliminary results of a study 

conducted by Health Canada (Health Canada, Safe Environments Programme, HECSB, 

pers. comm., 2008 March 6, unreferenced), and represents an overestimate of 

concentrations reported in water carboys by Biles et al. (1997). It was considered, 

conservatively, that all drinking water consumed by these age groups was from 

polycarbonate containers (e.g., repeat use drinking bottles, pitchers and carboys). 

Exposure estimates for this use are presented in Table 15. 

Table 15. Contribution of use of polycarbonate drinking containers to potential 

daily intake of bisphenol A (μg/kg-bw per day) for children (aged 1+ years) and 

adults 

Age group Potential daily intake of bisphenol A 

(μg/kg-bw per day)1

1–4 years2 0.012

5–11 years3 0.010 

12–19 

years4

0.005 

Adults5 0.006 1

Based on a calculated average bisphenol A concentration of 0.24 ppb (Le et al. 2008), assuming contents 

would not be heated prior to serving. 

2

assumed to weigh 14.4 kg (Nutrition Canada 1980) and to drink 0.7 L of water per day (Health Canada 

1998) 

3

assumed to weigh 26.4 kg (Nutrition Canada 1980) and to drink 1.l L of water per day (Health Canada 

1998) 

4

assumed to weigh 53.8 kg (Nutrition Canada 1980) and to drink 1.2 L of water per day (Health Canada 

1998) 

5

assumed to weigh 60.0 kg (Nutrition Canada 1980) and to drink 1.5 L of water per day (Health Canada 

1998) 

43
Another potential source of exposure to bisphenol A is through the use of polycarbonate 

tableware and storage containers. 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 minutes (n=9). Under both sets of conditions, 

migration was below the limit of quantification (0.6 ppb) in 7 of 9 samples. In the 4 

samples with detectable concentrations of bisphenol A, concentrations ranged from 1.7 to 

4.5 ppb. Due to limited data on migration from polycarbonate tableware and storage 

containers, the lack of information on availability of these products on the Canadian 

market, and use patterns by the general population, exposure estimates cannot be derived. 

However, based on available information, under normal use conditions exposure from 

this source is expected to be limited. 

Environmental Media 

Many recent reviews have summarized bisphenol 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 Canadian waters are summarized in Table 9a. 

The only published Canadian drinking water values were measured by Chen et al. (2006) 

in Alberta at the Bearspaw (on the Bow River) and the Glenmore (on the Elbow River) 

water treatment plants with bisphenol A concentrations of 0.00045 and 0.00076 μg/L

respectively (no detection limit presented). These values fall within the range reported by 

other countries, with concentrations between 300 pg/L to 2 ng/L (0.0003 to 0.002μg/L) 

(Kuch and Ballschmiter 2001) in Germany, and from below detection (0.01μg/L) to 

0.06μg/L in Japan (Nakanishi et al. 2007). Boyd et al. (2003) was the only other 

Canadian study identified and although bisphenol A was detected (LOD 0.1 ng/L), it was 

not quantifiable, making this study unsuitable for deriving estimates of bisphenol A 

intake for the Canadian population. 

No Canadian data were identified for ambient or indoor air. Although bisphenol A is 

released to air, due to its low vapour pressure it is more likely to partition to soil or water 

rather than remain in the air, where it has a half life of a few hours (see Environmental 

Fate section of Ecological Exposure Assessment). Concentrations in ambient air have 

been measured in two pilot studies conducted in 1997 (Wilson et al. 2001, 2003) and in a 

2000-2001 survey (Wilson et al. 2007) of outside residences and daycare centres in North 

Carolina and Ohio. In the 2000-2001 survey, bisphenol A was detected in ambient air 

outside residences in 31% of samples in North Carolina, ranging from below the limit of 

detection (0.9 ng/m3

) to 44.6 ng/m3 (50th percentile 0.9 ng/m3), and 34% of samples in 

Ohio, ranging from the limit of detection to 19.0 ng/m3 (50th percentile 0.920 ng/m3) 

(Wilson et al. 2007). Concentrations of bisphenol A measured in ambient air outside 

daycare centres, ranged from below the limit of detection (0.01 ng/m3

) to 51.5 ng/m3

(50th percentile below detection) in North Carolina (detected in 38% of samples), and 

below detection to 6.94 ng/m3

 (50th percentile below detection) in Ohio (detected in 44% 

of samples). In deriving intake estimates for the Canadian population, it was considered 

44
appropriate to use the maximum and fiftieth percentile ambient air concentration in North 

Carolina (44.6 ng/m3

, 0.9 ng/m3), taking into consideration the residential location and 

larger sample size (n=127). 

Concentrations of bisphenol A in indoor air were measured by Wilson et al. (2001, 2003, 

2007) in 42 daycare centres and 153 residences across North Carolina and Ohio in 

2000-2001. In Ohio residences, bisphenol A was detected in 63% of the samples within a 

range of concentration from below the limit of detection (0.01 ng/m3) to 30.3 ng/m3 (50th

percentile 0.980ng/m3

). In North Carolina, this range was extended to include a 

maximum concentration of 193 ng/m3

 (50th percentile 1.82 ng/m3) with bisphenol A 

detected in 68% of samples. For daycare centres, indoor air concentrations ranged from 

below the detection limit (0.01 ng/m3

) to 7.42 ng/m3 in Ohio, and 8.99 ng/m3 in North 

Carolina, with bisphenol A detected in 73% and 45% of the samples, respectively. In 

another survey of 120 homes in Cape Cod, Massachusetts, conducted by Rudel et al. 

(2003), concentrations in indoor air were below the reporting limit (18 ng/m3

). In an 

earlier study, Rudel et al. (2001) measured bisphenol A in air in the range of 0.002 to 

0.003 μg/m3 (2-3 ng/m3) in offices and residential locations (n=5). The maximum and 

fiftieth percentile indoor air concentrations in North Carolina (193 ng/m3

, 1.82 ng/m3) 

were considered the most appropriate for use in deriving estimates of intake based on the 

residential location and sample size (n=127). 

No Canadian data were identified for soil. One study in the United States analyzed soil in 

four daycare facility play areas, and reported 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 μg/g)(Wilson et al. 2003). The maximum and mean 

concentrations in soil (0.007 μg/g; 0.006 μg/g) from residential locations were considered 

most appropriate for use in deriving estimates of intake. 

Preliminary raw data are available for Phase 1 of the Canadian House Dust Study. Based 

on air-dried vacuum samples collected from 260 Canadian homes, median concentrations 

of bisphenol A in dust were reported at 1.60 μg/g dust (ranged from less than the method 

reporting limit [0.17 μg/g] to 23.84 μg/g). Bisphenol A was detected in 99% of homes 

(Health Canada, Environmental Health Science and Research Bureau, pers. comm., 2008 

Feb 7, unreferenced). These results are consistent with concentrations reported in dust 

from homes in other countries. In a survey of 118 homes in Cape Cod, Massachusetts, 

Rudel et al. (2003) identified a maximum value of 17.6 μg/g (mean of 1.65 μ/g) with 

86% of samples above the reporting limit. Wilson et al. (2007) also reported dust levels 

in residences in North Carolina and Ohio. In North Carolina, 25% of homes had 

detectable levels, with a maximum value of 707 ng/g (50th percentile was reported as the 

limit of detection). In Ohio, 72% of homes had detectable levels, with a maximum value 

of 589 ng/g (50th percentile is below detection limit). The maximum and median 

concentrations (23.84 μg/g; 1.60 μg/g) identified in dust measured in homes in Canada 

were considered most appropriate for use in deriving estimates of intake. 

 

45
Tables 16 and 17 present mean and upper bounding estimates of intake for each age 

group in the general population of Canada based upon concentrations of bisphenol A 

identified in environmental media. Dust and soil concentrations have been summed for 

the purposes of the intake table. 

Table 16. Average estimates of potential daily intake of bisphenol A by the general 

population in Canada from environmental media. 

Estimated intake (μg/kg-bw per day) of bisphenol A by various age 

groups 

Route of 

exposure 

0–7 

months1

8–12 

months2 

1–4 

years3

5–11 

years4

12–19 

years5

20+ 

years6

Ambient 

air7 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 air8 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 

water9 NA10 NA10 3.7 x 10-5 3.2 x 10-5 1.7 x 10-5 1.9 x 10-5

Soil and 

dust11,12,13

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 m3 of air per day and ingest 30 mg of soil per 

day (Health Canada 1998). 2 Assumed to weigh 9.0 kg (NCHS 2000), to breathe 9.3 m3

 of air per day and to ingest 100 mg of soil 

per day (Health Canada 1998). 3

 Assumed to weigh 14.4 kg (Nutrition Canada 1980), to breathe 9.3 m3 of air per day, to drink 0.7 L of 

water per day and to ingest 100 mg of soil per day (Health Canada 1998). 4

 Assumed to weigh 26.4 kg (Nutrition Canada 1980), to breathe 14.5 m3 of air per day, to drink 1.1 L of 

water per day and to ingest 65 mg of soil per day (Health Canada 1998). 5

 Assumed to weigh 53.8 kg (Nutrition Canada 1980), to breathe 15.8 m3 of air per day, to drink 1.2 L of 

water per day and to ingest 30 mg of soil per day (Health Canada 1998). 6

 Assumed to weigh 60.0 kg (Nutrition Canada 1980) , to breathe 16.2 m3 of air per day, to drink 1.5 L 

of water per day and to ingest 30 mg of soil per day (Health Canada 1998). 7

 Wilson et al. (2007) reported that 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/m3

 (detected in 31% of samples, n=127). Canadians are assumed to spend 3 hours outdoors each 

day (Health Canada 1998). 8

 Wilson et al. (2007) reported that the fiftieth percentile indoor air concentration as measured in 

samples inside residences in North Carolina, U.S.A, was 0.00182 μg/m3

 (detected in 68% of samples, 

n=127). Canadians are assumed to spend 21 hours indoors each day (Health Canada 1998). 9

 The only published Canadian values for 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.. 

46
12 The average concentration in soil of 7 μg/kg is based on a survey of 9 residences in North Carolina, 

U.S.A as part of a study by Wilson et al. (2003) which also included day care play facilities (detection 

limit of 1 μg/kg). 13 The median concentration in dust of 1600 μg/kg was identified in the preliminary data of the Phase 1 

Canadian House Dust Study of 260 homes (Health Canada, Environmental Health, Science and 

Research Bureau, pers. comm., 2008 Feb, unreferenced). 

Table 17. Upper-bounding estimates of potential daily intake of bisphenol A by the 

general population in Canada from environmental media. 

Estimated intake (μg/kg-bw per day) of bisphenol A by various age 

groups 

Route of 

exposure 

0–7 

months1

8-12 

months2 

1–4 

years3

5–11 

years4

12–19 

years5

20+ 

years6

Ambient 

air7 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 air8 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 

water9 NA10 NA10 3.7 x 10-5 3.2 x 10-5 1.7 x 10-5 1.9 x 10-5

Soil and 

dust11,12,13

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 m3 of air per day and to ingest 30 mg of soil 

per day (Health Canada 1998). 2 Assumed to weigh 9.0 kg (NCHS 2000), to breathe 9.3 m3

 of air per day and to ingest 100 mg of soil 

per day (Health Canada 1998). 3

 Assumed to weigh 14.4 kg (Nutrition Canada 1980), to breathe 9.3 m3 of air per day, to drink 0.7 L of 

water per day and to ingest 100 mg of soil per day (Health Canada 1998). 4

 Assumed to weigh 26.4 kg (Nutrition Canada 1980), to breathe 14.5 m3 of air per day, to drink 1.1 L of 

water per day and to ingest 65 mg of soil per day (Health Canada 1998). 5

 Assumed to weigh 53.8 kg (Nutrition Canada 1980), to breathe 15.8 m3 of air per day, to drink 1.2 L of 

water per day and to ingest 30 mg of soil per day (Health Canada 1998). 6

 Assumed to weigh 60.0 kg (Nutrition Canada 1980), to breathe 16.2 m3 of air per day, to drink 1.5 L of 

water per day and to ingest 30 mg of soil per day (Health Canada 1998). 7

 Wilson et al. (2007) reported the maximum ambient air concentration as measured in samples taken 

outside residences in North Carolina, U.S.A., as 0.0446 μg/m3

 (detected in 31% of samples, n=127). 

Canadians are assumed to spend 3 hours outdoors each day (Health Canada 1998). 8

 Wilson et al. (2007) reported the maximum indoor air concentration as measured in samples inside 

residences in North Carolina, U.S.A., as 0.193 μg/m3

 (detected in 68% of samples, n=127). Canadians 

are assumed to spend 21 hours indoors each day (Health Canada 1998). 9

 The only published Canadian values for 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). 

47
13 The maximum concentration in dust of 23 840 μg/kg was identified in the preliminary data of the 

Phase 1 Canadian House Dust Study of 260 homes (Health Canada, Environmental Health, Science 

and Research Bureau, pers.comm., 2008 February, unreferenced). 

Other Sources of Exposure 

Dental materials manufactured from bisphenol 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 permanent first and second molars that erupt into 

the mouth at ages six and twelve respectively (Health Canada, Medical Devices Bureau, 

pers. comm., 2008 February 27, unreferenced) 

While current Canadian data on the prevalence of dental decay in children are not 

available, it is known through studies of individual populations that it is a common 

childhood disorder, particularly in certain at-risk populations. Leake and Main (1996) 

found that children aged five in Ontario had 1.22 decayed, missing or filled teeth. 

Peressini et al. (2004) found that children aged seven in an Ontario First Nations 

Community had 6.20 decayed, missing or filled teeth. Leake et al. (2001) in a sample 

survey of the oral health of 3657 Toronto children aged five, seven or thirteen, found that 

30.0% of five year olds, 41.3% of seven year olds and 39.3% of thirteen year olds 

children had one or more cavities. 

The restoration of decayed teeth, both primary and permanent, may be accomplished 

using a variety of materials including dental amalgam and composite resin. While current 

data on the prevalence of usage of composite resins versus other restorative materials are 

not available, anecdotal information suggests that the use of composite resin is increasing 

and is currently used more frequently than other materials to restore primary and 

permanent teeth in children. 

Santerre (2007) has stated the three sources of bisphenol A derived from dental materials 

are unpolymerized monomer on the surface of the restoration or sealant; hydrolyzed 

polymers of bisphenol-derived monomers in the material and fine particulate polymer 

produced by abrasion of the restoration. Santerre (2007) further states that BisGMA, one 

of the most commonly used bisphenol A-derived monomers used in the manufacture of 

dental restorative materials and sealants, shows no capacity to be hydrolyzed back into 

bisphenol A under normal oral conditions. 

A number of investigators have shown that placement of composite resin restorations and 

sealants can result in acute, transient exposures to bisphenol A (Fung et al. 2000; Joskow 

et al. 2006; Sasaki et al. 2005) 

Fung et al. (2000) collected saliva following the application of 8 mg (applied to one 

surface) and 32 mg (8 mg applied to four surfaces) of a dental sealant to a group of 18 

and 22 adults, respectively. Concentrations of bisphenol A in saliva ranged from 5.8 to 

105.6 ppb (detection limit of 5 ppb) at 1 hr and 3 hrs following treatment but was not 

48
detected immediately after treatment or at other time intervals (24 hr, 72 hrs and 120 hrs) 

post-treatment. 

More recently, Joskow et al. (2006) measured concentrations of bisphenol A in saliva of 

14 adults before and after the application of between 39.0 to 42.5 mg of two dental 

sealants. Mean concentrations of bisphenol A before treatment, immediately following 

treatment, and an hour post-treatment were reported as 0.22 ng/ml, 0.54 ng/ml, and 

0.24ng/ml in five patients treated with one sealant and 0.30 ng/ml, 26.5 ng/ml and 5.12 

ng/ml in nine patients treated with the other sealant. 

Sasaki et al. (2005) measured 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) labelled A (n=4) measured in the range of 

0.3 to 2.0 ng/ml (mean 0.87ng/ml) prior to treatment, 21.0 to 60.1 ng/ml (mean 32.1 

ng/ml) following treatment and 1.6 to 4.7 ng/ml (mean 3.1 ng/ml) after gargling for 30 

seconds with water at 37o

C. Patients treated with the composite labelled F (n=3) had the 

highest concentration of bisphenol 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 the use of the other composites labelled C (n=2), G (n=2), 

H (n=2), and I (n=2) ranged from 10 to 45 ng/ml. 

Azarpazhooh and Main (2008) observed that no dental sealants that carried the Seal of 

the American Dental Association [product recognition program, now discontinued] in 

2007 released detectable amounts of BPA and continued with recommendations for 

simple clinical procedures [commonly followed] to reduce the possibility of 

unpolymerized BPA-derived monomer remaining on the surface of the tooth following 

sealant placement. 

A method to extrapolate from concentrations in saliva to daily estimates of exposure was 

not identified. However, consistent with the data presented above, other international 

assessments have characterized exposure to 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 infants from contact with consumer products 

such as toys. The contact time was established using a survey of mothers who observed 

mouthing behaviour in a total of 50 infants (Tanimura 1999) and video recordings of 25 

infants in which time spent mouthing toys was monitored (Tanimura 2000). A minimum 

migration rate (of 0 μg/cm2

/min (for products that do not contain bisphenol A) and a 

maximum migration rate 0.0162 μg/cm2

/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 

49
are expected to be made of bisphenol A-based polymers and, additionally, these hard 

plastics tend to have a lower migration rate than would be expected from soft plastic toys 

(Health Canada, Consumer Product Safety Bureau, pers. comm., 2008 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 may be present in the adhesive products at 

concentrations ranging from 0.1%– 100%. However, only trace quantities 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 bisphenol A from epoxy adhesives predicts air 

concentrations during use to range from 3.82 x 10-5 μg/m3 to 6.08 x 10-5 μg/m3 (Table 18). 

The chosen scenarios correspond to gluing a handle on a coffee mug or gluing a large 

vase. However it is expected that the majority of exposures will occur at the lower end of 

this range. Dermal exposure may also result from the use of this product and was 

estimated at 0.0167 μg/kg-bw per use (Table 18). Confidence in these estimates is low as 

they are based on a number of assumptions; however it is likely that they overestimate 

actual exposures from this source. 

Table 18. Upper-bounding estimates of potential exposure to 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 increasing 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 cm2

, and an application duration of 5 

minutes 

- Assume a room volume of 20 m3

, exposure duration of 

240 minutes, a ventilation rate of 0.6 times/hr, a mass 

transfer rate based on Langmuirs method and a molecular 

weight matrix of 3000 g/mol (RIVM 2007). 

Mean event 

concentration 

= 0.0000382 

μg/m3 

Two 

component 

epoxy 

Inhalation 

- Used ConsExpo model version 4.1, exposure to vapour: 

evaporation from an increasing area as mode of release 

Mean event 

concentration 

= 0.0000608 

50
Consumer 

product 

scenarios 

Assumptions Estimated 

exposure

adhesive–

gluing a 

large vase1

(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 cm2, an application duration of 30 

minutes, a room volume of 20 m3

, exposure duration of 

240 minutes, a ventilation rate of 0.6 times/hr, a mass 

transfer rate based on Langmuirs method and a molecular 

weight matrix of 3000g/mol (RIVM 2007). 

μg/m3 

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 cm2, and an 

applied amount of 0.1 grams of product is applied (RIVM 

2007). 

- Assume adult exposed weighs 60.0 kg (Nutrition Canada 

1980). 

- Assume 100% uptake.

Acute dose 

per event = 

0.00167 μg/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 are not known but are 

expected to be low and are not expected to contribute significantly to the overall exposure 

to the general population. 

51
Aggregate Exposure from Dietary 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 (μg/kg-bw per day).

Total estimated intake 

(dietary intake + media) 

Age group Intake from 

infant 

formula1,2 

Intake from 

environmental 

media6 

Intake from 

migration from 

polycarbonate 

bottles3,4

Total estimated 

dietary intake4,5 

(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 compared to boiling water and are not presented here 

(see Table 14). 

4

 Based on the average concentration of bisphenol A measured in water, 10 ppb, that resulted from filling 

polycarbonate bottles with boiling water (Maragou et al. 2007) (see Table 14). 5

 Heath Canada recommends that solid food be introduced to infants after 6 months of age. For this 

assessment it was assumed that an infants main source of food is infant formula, however if infants 

consume other canned foods this may increase their total dietary exposure to bisphenol A. 6

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

52
Table 19b. Aggregated maximum estimates of potential exposure for formula fed 

infants aged 0 to 18 months (μg/kg-bw per day).

Total estimated intake2

(dietary intake + media) 

Age group Intake from 

infant 

formula1,2 

Intake from 

environmental 

media6 

Intake from 

migration from 

polycarbonate 

bottles3,4

Total estimated 

dietary intake4,5 

(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 compared to boiling water and are not presented here 

(see Table 14). 

4

 Based on the average concentration of bisphenol A measured in water, 10 ppb, that resulted from filling 

polycarbonate bottles with boiling water (Maragou et al. 2007) (see Table 14). 5

 Heath Canada recommends that solid food be introduced to infants after 6 months of age. For this 

assessment it was assumed that an infants main source of food is infant formula, however if infants 

consume other canned foods this may increase their total dietary exposure to bisphenol A. 6

 Based on maximum concentrations measured in various environmental media (see Table 17). 

For exposure among breastfed infants aged 0 to7 months, exposures from multiple 

sources (breast milk and through environmental media) were aggregated and are 

presented in Table 20. 

Table 20. Aggregate estimate of potential exposure for breastfed infants aged 0 to 7 

months (μg/kg-bw per day)1

.

Intake from 

consumption of 

breastmilk2

Overall intake from 

environmental media3

Age group Total estimated intake 

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. 

53
2

Based on the average and maximum concentrations of bisphenol A identified in breastmilk by Ye et al. 

(2006) (see Table 13). 

3

Based on the average and maximum concentrations of bisphenol A identified in various environmental 

media (see Tables 16 and 17)

For exposure among children and adults (aged 1+ years), exposure from multiple sources 

(food packaging of canned food and beverages, migration from polycarbonate drinking 

bottles and through environmental media) were aggregated and are presented in Table 21. 

Table 21. Aggregate estimate of potential exposure for children (aged 1+ years) and 

adults (μg/kg-bw per day).

Intake from 

consumption of 

canned food1

Intake from 

migration 

from 

polycarbonate 

bottles2

Overall intake from 

environmental 

media3

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 bisphenol A detected in water, 0.24 ppb, that resulted 

from the room temperature use of polycarbonate bottles (Le et al. 2008) (see Table 15). 

3

Based on the average and maximum concentrations identified in various environmental media (see Tables 

16 and 17).

In a limited study, Wilson et al. (2003) reported aggregate exposure estimates for nine 

children aged 2-5 year olds from concentrations of bisphenol A reported in indoor and 

ambient air, floor dust, play area soil and solid and liquid food collected at two daycare 

centres and nine households in North Carolina. Mean exposure was 42.981 ng/kg-bw per 

day and maximum exposure was 71.124 ng/kg-bw per day. The aggregate exposure was 

based on the following assumptions: 100% absorption, that a child spends 12 hours a day 

at the daycare centre, ingests soil and dust at a rate of 100 mg/day (Stanek and Calabrese 

1995a,b), and inhales air at a rate of 8.3 m3

/day (EFH 2000). In an expanded study by 

Wilson et al. (2007), 257 children aged 1.5 to 5 years old were surveyed in daycare 

centres and homes located within North Carolina and Ohio using the approach described 

above. A 50th percentile aggregate exposure of 71.4 ng/kg-bw per day and a maximum of 

1570 ng/kg-bw per day was reported in North Carolina. A 50th percentile aggregate 

exposure of 60.8 ng/kg-bw per day and maximum of 775 ng/kg-bw per day was reported 

in Ohio (an assumption of 50% absorption was made). 

54
Concentrations in biological tissues and fluid have been reported by a number of 

investigators and are summarized in Vandenburg et al (2007), Dash et al. (2006) NTP 

(2007) and Dekant and Volkel (2008). Tissues and fluid analyzed include semen, 

amniotic fluid, placental tissue and serum. Urine is considered the most appropriate body 

fluid for quantifying exposure to bisphenol A. Several studies have analyzed bisphenol A 

in urine. Most recently, bisphenol A (free and conjugated) was an analyte of the 

2003-2004 National Health and Nutrition Examination Survey (NHANES) .The 

NHANES results are reported by Calafat et al. (2008) and present results of bisphenol A 

analysis (free and conjugated) in 2517 individuals aged 6 years and older in the United 

States. Using these data, Lakind and Naiman (2008) have derived estimates of daily 

intake of bisphenol A, assuming steady state excretion. Fiftieth and ninety-fifth 

percentiles for the following subpopulations of the U.S. (6–11 years, 12–19 years, 20–39 

years, 40–59 years and 60+ years) have been derived based on standard values for urinary 

output and adjusted for individual bodyweights. For all individuals, the 50th percentile 

exposure estimate is 50.5 ng/kg-bw per day (the 95th percentile exposure estimate is 

274.2 ng/kg-bw per day). For children aged 6 to 11, the 50th percentile exposure estimate 

is 67.4 ng/kg-bw per day (the 95th percentile exposure estimate is 310.5 ng/kg-bw per 

day). 

Confidence in the exposure assessment is considered to be moderate. Canadian 

multimedia data are very limited. Biomonitoring data indicates that exposures may be 

less than those derived from aggregating exposure estimates based on concentrations in 

different media, but the absence of biomonitoring data for a key segment of the 

population (infants) precludes use of biomonitoring data to characterize risk. 

Additionally, applying precaution, derivation of exposure estimates based on 

concentrations in media captures exposures in the most exposed individuals. 

Furthermore, there are uncertainties associated with the use of spot urine samples from 

one point in time to predict daily exposure estimates. 

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 considered to have low acute oral toxicity in animals (ECB 2003). 

The European Union reported that bisphenol A is not a skin irritant, but may cause eye or 

respiratory irritation in experimental animals (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 literature. 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. 

55
Toxicokinetics 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 levels of free bisphenol A. Following oral 

exposure, the glucuronidation of bisphenol A by hepatic uridine diphosphate 

glucuronyltransferase (UDPGT), appears to be a major in vivo metabolism pathway in 

rats (specifically the isoform, UGT2B1) (Pottenger et al. 2000; Yokota et al. 1999), nonhuman primates (Kurebayashi et al. 2002), and humans (Volkel et al. 2002). The UGT 

isoform(s) which may be involved in metabolism of bisphenol A in humans is/are not 

known (NTP 2008). In a recent draft, the European Food Safety Authority (EFSA) 

acknowledged glucuronidation as the major pathway of bisphenol A metabolism in adult 

humans and animals while sulfation was suggested as a minor pathway in humans and 

rats (EFSA 2008). The United States Food and Drug Administration (USFDA) also 

recognized glucuronidation as a major metabolic pathway in rodents, monkeys and 

humans, but the possibility of bisphenol A sulfation during metabolism was not 

sufficiently reviewed (USFDA 2008). 

In humans and other non-human primates, bisphenol A administered orally enters firstpass metabolism in the gut wall and the liver and is quickly metabolized 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 bisphenol 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 also undergoes glucuronidation, but the resulting 

bisphenol A-glucuronide is excreted from liver 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 may cause slow elimination of 

bisphenol A in rodents (Pottenger et al. 2000; Snyder et al. 2000). An overview of 

toxicokinetics or pharmacokinetics of bisphenol A by the EFSA and USFDA elaborated 

the similar stages of bisphenol A metabolism in rodents, non-human primates and 

humans (EFSA 2008; USFDA 2008). 

Additionally, the toxicokinetics and metabolism of bisphenol A in pregnant animals, 

fetuses or neonates may be different to that of non-pregnant adults. The absorption and 

distribution of bisphenol A was rapid in maternal F344 rat organs following a single oral 

exposure to a high dose of 1 g/kg, reaching the fetus by crossing the placenta within 20 

minutes of exposure; the fetal bisphenol A concentration exceeded that of maternal blood 

40 minutes after exposure (Takahashi and Oishi 2000). Also, in humans, bisphenol A was 

present in serum and follicular fluid of pregnant women, as well as fetal serum and 

amniotic fluid, demonstrating placental transfer. Bisphenol A (presumably bound and 

free) concentrations in amniotic fluid, mid-term, were 5-fold higher as compared to 

56
serum and follicular fluid (Ikezuki et al. 2002). These data indicate that repeated maternal 

exposures could lead to an accumulation of fetal circulating levels of free bisphenol A 

and correspondingly elevated in 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 (Schönfelder et al. 2002). In women, bisphenol A 

has been detected in biological fluids/tissue including breastmilk, colostrum and adipose 

tissue in parts per billion concentrations (Fernandez et al 2007; Kuruto-Niwa et al 2007; 

Sun et al 2004). A careful evaluation of biologically active or free bisphenol A 

concentration in various tissues of a sensitive population was previously identified as a 

research need and has also recently been identified as being of critical importance by the 

USFDA (2008). 

A reduced UGT2B1 activity was seen in pregnant rats as compared to non-pregnant 

females. The UGT2B1 enzyme was not present in the fetus, its activity was slow in 

neonates, and reached the adult level of activity within three weeks after birth 

(Matsumoto et al. 2002). It has been suggested that an increase in glucuronidation 

capacity during pregnancy may result in negligible fetal exposure to xenobiotics (EFSA 

2008). Human data show that glucuronide enzyme activity may be induced during 

pregnancy as it was evidenced by an increased clearance of orally administered 

lorazepam, paracetamol and lamotrigine in pregnant women (Miners et al 1986; Papini et 

al 2006; Pennell et al 2008). It is important to note that based on the limited evidence of 

selective glucuronidation and considering other factors including individual variation, 

tissue localization, and substrate-specificity of these enzymes in animals, it cannot be 

assumed that glucuronidation activity would be increased in pregnant women following 

exposure to bisphenol A. 

Limited activity of glucuronidation enzymes in the human fetal liver has also been 

demonstrated (Ring et al. 1999). In human liver microsomes, most of the UGT isozymes 

were not expressed prenatally. A full complement was present at three months after birth; 

however, levels were substantially reduced compared to the adult (Coughtrie et al. 1988). 

Studies indicate that, in humans, sulfotransferase (SULT) enzymes are fully functional at 

birth and it has been suggested that sulfoconjugation of bisphenol A may overcome the 

absence of or very low glucuronidation activity in newborns (Benedetti and Baltes 2003; 

EFSA 2008; Richard et al. 2001). 

Some in vitro data showed that SULT enzymes may have the ability to metabolize some 

phenolic or estrogen-like compounds including bisphenol A in cell cultures (Nishiyama 

et al 2002; Shimizu et al 2002; Suiko et al 2000). This limited data needs to be 

interpreted carefully as different liver cells were studied under controlled conditions 

using different protocols. Whether the sulfotransferase activity reported in these studies 

would be sufficient to efficiently metabolize the bisphenol A in in vivo is not known at 

this time. A study conducted using post-mortem human (fetal, term and adults) liver 

57
microsomes, in general, showed reduced UDPGT activity towards some phenolic 

compounds, which were structurally similar to bisphenol A, in fetal and term samples as 

compared to adults. Also, the human fetal liver sulphotransferase activity was found to 

be less than 3% of that present in non-human primate fetal liver (unpublished work – 

Leakey). The authors suggested that glucuronic acid or sulphate conjugation pathway 

may not be efficient in human perinate against phenolic compounds (Leakey et al 1987). 

It is therefore not clear, based on the data of Leaky et al. (1987) whether sulfotransferases 

would act as a major metabolic enzyme system in human neonates following exposure to 

bisphenol A. Together, the 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 (Welshons et al. 2006) and their 

inability to efficiently glucuronidate bisphenol A (Ring et al. 1999; Coughtrie et al. 

1988). 

Conversely, some evidence suggests that there were no significant differences between 

pregnant (at different stages of gestation) and non-pregnant Sprague-Dawley rats in the 

toxicokinetics and metabolism of orally administered bisphenol A (10 mg/kg single 

dose), and fetal bisphenol A elimination half-life appeared to be dependent on maternal 

bisphenol A elimination (Domoradzki et al. 2003). It was reported that, in developing 

Sprague-Dawley rats, dose-dependent metabolism of 1 or 10 mg/kg (gavage) bisphenol A 

to its glucuronide conjugate occurred as early as postnatal day 4 through postnatal day 21 

and no significant sex-related differences were noted in the metabolism and 

pharmacokinetics in neonatal to weanling rats (Domoradzki et al. 2004). 

In the absence of fully developed physiologically-based pharmacokinetic models to 

quantitatively characterize absorption, distribution, metabolism and excretion of 

bisphenol A and its metabolites in test animals and humans via different routes and at 

different life stages, it is considered appropriate to consider the pregnant woman/fetus 

and infant as potentially sensitive subpopulations. 

Carcinogenicity 

A lifetime (103-week) bisphenol A chronic toxicity/carcinogenicity study in male (0, 74 

or 148 mg/kg/d) and female (0, 74 or 135 mg/kg per day) F344 rats reported reduced 

body weight gain in all exposed animals which was correlated with reduced feed 

consumption. A marginal increase in the incidence of leukemia and testicular interstitial 

tumors in male rats was considered statistically non-significant due to failure to meet the 

Bonferoni inequality criterion; the incidence of both tumor types was more common in 

the high-dose groups. Male and female B6C3F1 mice were administered doses of 0, 120 

or 600 mg/kg-bw per day or 0, 650 or 1300 mg/kg-bw per day, respectively. Reduced 

body weight gain was observed in male and female B6C3F1 mice that received bisphenol 

A doses of 600 mg/kg per day or 650 or 1300 mg/kg per day respectively for 103 weeks. 

There was an increase in the combined incidence of leukemia and lymphoma in male 

mice, which was not considered dose-related. Overall, it was concluded that there was no 

58
convincing evidence that dietary exposure to bisphenol A causes carcinogenicity in rats 

or mice of either sex (NTP 1982). Some debate exists regarding the biological 

significance of the NTP (1982) findings. Specifically, it has been argued that the 

marginal increases in leukemias of rats and the combined incidence of lymphomas and 

leukemias in male mice should be considered to be bisphenol A related. It has been 

proposed as well that the observed testicular effects should also be considered as some 

evidence for carcinogenic activity (Huff 2001; 2002). However, others maintain that the 

2-year study correctly concluded that the results demonstrate equivocal evidence of 

carcinogenicity in male rats and no evidence of carcinogenicity in female rats and mice of 

both genders (Haighton et al. 2002; Munro et al. 2002a; 2002b). It was noted that the 

incidence of malignancies observed in the bisphenol A exposed groups of rats and mice 

were within the historical control ranges for the strains and species tested (Haighton et al. 

2002). In support of the NTP (1982) conclusions, Tyl et al. (2002) observed no increase 

in cancer incidence in a three-generation assessment of bisphenol A toxicity in male and 

female Sprague-Dawley rats exposed to doses ranging from 0.001 to 500 mg/kg-bw per 

day (Tyl et al. 2002). The available evidence suggests that bisphenol A is not a 

carcinogen following adult exposure (Keri et al. 2007). 

Recent reports have suggested that prenatal and/or neonatal exposure to bisphenol A in 

rats at doses of 25 or 250 ng/kg-bw (via osmotic pumps implanted into pregnant dams) 

and ranging from 2.5 μg/kg-bw per day to 1.0 mg/kg-bw per day (administered via 

maternal subcutaneous injection) during development, may increase susceptibility to 

neoplastic transformation in the prostate and mammary gland in adult rats (Ho et al. 

2006; Murray et al. 2007; Durando et al. 2007; Markey et al. 2003, 2005; Munoz-de-Toro 

et al. 2005). However, the limited evidence is insufficient to demonstrate that early 

bisphenol A exposure, acting independently, could lead to neoplastic events (Newbold et 

al. 2007; Prins et al. 2007; Ho et al. 2006; Keri et al. 2007; Ichihara et al. 2003). Further 

research is needed on the role of early life exposures to bisphenol A, in particular via 

routes most relevant to human exposure, in the process of carcinogenesis (Keri et al. 2007 

Genotoxicity 

In vitro genotoxicity studies demonstrated a lack of mutagenic activity; however, it was 

noted that bisphenol A has aneugenic potential and induced morphological 

transformation and micronuclei formation in the absence of metabolic activation in 

cultured Syrian hamster embryo cells and Chinese hamster V79 cells, respectively (NTP 

2007; ECB 2003; Haighton et al. 2002). In contrast, a large number of standard 

genotoxicity assays, including the in vivo mammalian dominant lethal assay and the 

mouse micronucleus assay, provided no evidence of mutagenic or clastogenic potential of 

bisphenol A (ECB 2003; Haighton et al. 2002). Using the 32P-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 lack of significant tumour induction in animal 

studies, the European Union concluded that “it does not appear that bisphenol A has 

significant mutagenic potential in vivo” (NTP 2007; ECB 2003). 

59
The aneugenic potential of bisphenol A was further investigated following the 

observation of a spontaneous increase in oocyte 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 oogenesis. 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% and 10.9% for 0, 20, 40, and 100 μg/kg-bw per 

day, respectively) and a dose-related increase in meiotic abnormalities (Hunt et al. 2003). 

Subsequent examination of the effects of bisphenol A on female and male germ cells and 

somatic cells did not confirm aneugenic potential. There was no evidence of increased 

hyperploidy or polyploidy, alterations on the timing of meiotic progression or 

micronuclei induction following bisphenol A exposure via drinking water (7 daily doses 

of 0.04 mg/kg or for 7 weeks at a concentration of 0.5 mg/L) to C57Bl/6 female mice or 

via gavage (0, 0.002, 0.02, 0.2 mg/kg) to (102/ElxC3H/E1) F1 male mice (Pacchierotti et 

al. 2008). Contrary to the induction of aneuploidy, it has been suggested that oocytes 

from 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; Nakanishi 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 and 500 mg/kg-bw per day) and mice (0, 

0.003, 0.03, 0.3, 5, 50, 600 mg/kg-bw per day) (Tyl et al. 2002; 2007). NOAELs of 5 

mg/kg-bw per day and 50 mg/kg-bw per day were identified for adult systemic effects 

and reproductive and developmental toxicity in rats and mice, respectively. The NOAEL 

of 5 mg/kg-bw per day for adult systemic effects was based on significantly reduced body 

weight and reduced body weight gain, reduced absolute and increased relative organ 

weights in rats, and an increased incidence of minimal to mild hepatocyte hypertrophy in 

adult male and female mice. The NOAEL of 50 mg/kg-bw per day for reproductive and 

developmental toxicity was based on effects on the offspring at the highest doses tested. 

These effects included reduced weanling body weight, exposure-related changes in 

absolute and relative organ weights in male and female rats, and hypoplasia of 

seminiferous tubules and delayed acquisition of preputial separation, as well as a slight 

increase in the incidence of undescended testes in weanling male mice. No changes of 

toxicologic significance were seen within the dose range of 0.001-5.0 or 0.003-5.0 

mg/kg-bw per day in adult or developing rats and mice respectively (Tyl et al. 2002; 

2007). The three-generation reproductive toxicity study by Tyl et al. (2002) in rats did not 

include an estrogenic reference control against which the negative results could be 

60
evaluated. However, the most recent multigeneration report in mice incorporated a 

dietary 17β-estradiol reference group (Tyl et al. 2007). Developmental toxicity in CD-1 

mice and CD rats has been demonstrated to occur following prenatal exposure during 

major periods of organogenesis (gestational days 6-15), through gavage, at high doses of 

bisphenol A. Measures of developmental toxicity included percentage resorptions, 

number of live fetuses per ratio, sex ratio, average fetal body weight per litter and 

percentage of malformed fetuses. No developmental effects were observed following 

administration of 640 mg/kg-bw per day (highest dose reported due to severe maternal 

toxicity at 1280 mg/kg-bw per day) or 1000 mg/kg-bw per day in rats and mice, 

respectively. Bisphenol A was not teratogenic at any of the administered doses 

(Morrissey et al. 1987). 

Recent studies have shown that chemicals that alter endocrine function, including 

bisphenol A, can alter the cell function at levels below NOAELs identified in these 

standard toxicology studies (reviewed in Welshons et al. 2003; vom Saal and Welshons 

2006; Welshons et al. 2006). These studies have been described as ‘low dose studies’. 

Effects observed are not necessarily manifested following a linear dose-response 

relationship and, in several instances, have been found to follow a non-monotonic 

response curve. Bisphenol A has been considered as a weak environmental estrogen, 

based on traditional bioassays, as it binds to the estrogen receptor alpha and beta (ERα

and ERβ) (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 17β-estradiol (Vandenberg et al. 

2007). A mode-of-action analysis is beyond the scope of this screening assessment; 

however, exposure to low doses of bisphenol A has been demonstrated to produce 

disruptive effects in endocrine organs including the androgen or estrogen responsive 

tissues, immune system, thyroid hormone function, and developing nervous system 

(reviewed in Richter et al. 2007; Vandenberg et al. 2007; Wetherill et al. 2007). Evidence 

suggests that bisphenol A can stimulate the aforementioned cellular responses at very low 

concentrations through genomic (nuclear estrogen receptor) or non-genomic (membraneassociated or intracellular transduction) mechanisms, with effects on cellular function at 

doses as low as 1 pM or 0.23 pg/ml (Anway et al. 2005, 2006; Wozniak et al. 2005, 

Zoeller et al. 2005) (reviewed in Wetherill et al. 2007). Specifically, it has been 

demonstrated that some of the effects of bisphenol A may be mediated through the cell 

surface estrogen receptor (GPR30) being equipotent with 17β-estradiol and 

diethylstilbestrol (Alonso-Magdelena et al. 2005). 

There are many studies in the peer-reviewed literature that describe effects in 

experimental animals resulting from exposure to bisphenol A administered at doses 

below the NOAELS established in standard toxicological studies for systemic and 

reproductive/developmental endpoints. Effects observed include changes in prostate 

growth and development, mammary gland organization, protein induction in the uterus, 

organization of sexually dimorphic circuits in the hypothalamus, onset of estrus cyclicity, 

early puberty, body weight, genital malformations and others (reviewed in Richter et al. 

2007; Wetherill et al. 2007). Thorough overviews of these studies have been summarized 

in several recent reviews (NTP 2007; Richter et al. 2007; Willhite et al. 2008). For 

purposes of the screening assessment, the primary focus will be placed on identification 

61
of alterations considered potentially ‘adverse’ for the purposes of human health risk 

assessment. 

It should be noted that, due to the finding by Taylor et al. (2008) that the differences 

between oral and subcutaneous injection toxicokinetics in adults are not likely conserved 

in neonates, those studies delivering bisphenol A during fetal development and/or early 

postnatal periods via the subcutaneous (sc) injection route were evaluated for the current 

assessment (Kato et al. 2003; 2006; Ho et al. 2006; Honma et al. 2002; Nakamura et al. 

2006; Suzuki et al. 2002). The neonatal data, using sc injection permits the evaluation of 

possible effects of ecposure during the early stages of development. While informative, 

these studies were not the basis for hazard characterization of bisphenol A for human 

health. 

A number of reproductive and developmental studies provide evidence that in utero

and/or neonatal exposure to doses below the established reproductive/developmental 

NOAEL of 50 mg/kg-bw per day may result in altered reproductive function and 

developmental endpoints. Prenatal exposure from gestational day 11 to 17, of CF-1 mice 

to 0, 2 or 20 μg/kg-bw per day bisphenol A administered orally, resulted in an increase in 

prostate size in male offspring beginning at 2 μg/kg-bw per day (Nagel et al. 1997), 

reduced epididymal weight and decreased daily sperm production at doses of 2 and 20 

μg/kg –bw per day, respectively (vom Saal et al. 1998). It also resulted in increased 

postnatal growth in both males and females and early onset of sexual maturation in 

female offspring, observed following exposure to 2.4 μg/kg-bw per day (Howdeshell et 

al. 1999). Additionally, CD-1 mice exposed to 10 μg/kg-bw per day, from gestational day 

14 to 18, displayed disrupted development of the prostate as shown by an increase in the 

number and size of the prostate ducts and an increase in the overall volume of the 

prostate; a malformation of the urethra was also observed (Timms 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 was not affected at the first breeding. 

Furthermore, oral exposures of pregnant and nursing Long-Evans rat dams from 

gestational day 12 to postnatal day 21 to 2.4 μg/kg-bw per day altered plasma luteinizing 

hormone levels and decreased testosterone levels in adulthood. This suggests that 

bisphenol A-dependent effects on androgen biosynthesis by adult leydig cells occurred as 

a result of targeting a sensitive period during perinatal development. Chronic exposure 

from weaning at postnatal day 21 to adulthood, representing adolescent exposure, also 

suppressed androgen biosynthesis by adult leydig cells at doses as low as 2.4 μg/kg-bw 

per day bisphenol A (Akingbemi et al. 2004). A LOAEL for the reproductive 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 observed in several instances. 

Conflicting results have been described in various additional reports. Oral bisphenol A 

exposure to 0.2, 2, 20 or 200 μg/kg-bw per day from gestational day 11 to 17 did not 

produce reproductive toxicity in studies assessing outcomes in CF-1 mice (Cagen et al. 

62
1999a; Ashby et al. 1999). Nor was there an effect on sperm density or male reproductive 

organ development following bisphenol A exposure of estrogen-sensitive C57Bl/6N mice 

during various life stages (Nagao et al. 2002). Notably, in the studies by Cagen et al. 

(1999a) and Ashby et al. (1999), the responses of the positive control did not differ from 

either the negative control or 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 reproductive or developmental toxicant (Cagen et al. 

1999b). A 2-generation reproduction 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 changes in the multiple reproductive and 

developmental endpoints assessed (Ema et al. 2001); this study also lacked a positive 

control for interpretation of negative data. Most recently, assessing the impact of 

bisphenol A exposure in the male Long Evans hooded rat following oral dosing with 2, 

20 or 200 μg/kg-bw per day from gestational day 7 to postnatal 18, Howdeshell et al. 

(2007) found no significant effect on any male reproductive endpoint investigated. These 

included male reproductive development, reproductive hormone levels or sperm 

production. 

It is recognized that a number of experimental variables may account for divergent results 

in response to bisphenol A exposure at low doses: specifically, species and strain 

differences; variable responses depending on the tissues and/or endpoint(s) assessed; 

variability in feed with respect to amounts of estrogenic contaminants; inappropriate use 

or lack of positive controls; as well as the consideration of exposure-related effects that 

present in a non-monotonic dose-response curve (vom Saal and Hughes 2005; vom Saal 

et al. 2005a; Richter et al. 2007). Additionally, the dosing period or time of exposure with 

respect to critical developmental windows and the dosing regimen are important 

considerations, particularly when assessing alterations resulting from developmental 

exposures. Furthermore, the nature of the effects are such that it is difficult to 

characterize the degree to which they would be considered potentially “adverse” and, 

hence, form the basis of a human health risk assessment. 

Developmental Neurotoxicity 

Recently, expert panel assessments of the current status of bisphenol A toxicity have 

identified neurodevelopmental and behavioural alterations as an endpoint of concern. The 

U.S. National Toxicology Program (NTP) Center for the Evaluation of Risks to Human 

Reproduction (CERHR) (NTP 2007) concluded there was “some” concern that pregnant 

women, fetuses, infants and children may be susceptible to bisphenol A disruption for 

neural and behavioural endpoints. “Some” concern for these effects was based on the 

finding that the bisphenol A literature regarding neural and behavioural effects 

consistently produced more “positive” studies, as compared to the reproductive and 

developmental studies, although the endpoints assessed differed between investigators. 

The level of concern was expressed relative to estimates of exposure for the general 

population in the United States (NTP 2007). Recently, the NTP finalized its evaluation 

63
of bisphenol A and confirmed some concerns for neurobehavior effects in fetuses, 

infants, and children at the current level of human exposure (NTP 2008) Additionally, 

based on existing evidence, the Chapel Hill bisphenol A expert panel states with 

confidence “that low doses of bisphenol A during development have persistent effects on 

brain structure, function and behaviour in rats and mice” (vom Saal et al. 2007). 

Confidence was defined as research areas where the expert panel concluded that there 

was an agreement of findings from multiple papers from multiple labs. Given these 

expert panel conclusions, as part of the screening assessment it was considered 

appropriate to further consider the weight of evidence associated with 

neurodevelopmental and behavioural studies investigating effects of low dose bisphenol 

A exposure. These studies are presented in Appendix D and are discussed in further detail 

below. 

Neural development is a complex process beginning in early embryonic development and 

continuing through different life stages. Precise functioning of the nervous system is 

critical for all mental, sensory and motor activities, and regulates homeostasis through 

interaction with the endocrine system (Clancy et al. 2001). An array of behavioural tests 

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 (OECD 2007). A number of developmental 

neurotoxicity endpoints were examined, including locomotor and exploratory activity, 

learning, social, sexual and maternal behaviour, as well as brain morphology, with a 

range of exposures from 0.2 to 400 000 μg/kg-bw per day [reviewed in NTP 2007; 

Richter et al. 2007; Willhite et al. 2008; ECB 2008]. 

Evaluation of several pivotal studies on neural development and behaviour effects of 

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 μg/kg-bw per day (EFSA 2006; Willhite 

et al. 2007), may produce some effects on neurobehavioral development in fetal or early 

postnatal life in rodents. The lowest observed effect in the neurodevelopmental literature 

was 0.02 μg/kg-bw per day based on changes in estrogen receptor mRNA expression, 

however, no effects were observed at 2 μg/kg-bw per day; this is suggestive of an 

inverted U shaped (non-monotonic) dose response phenomenon (Nishizawa et al. 2003; 

2005a; 2005b). 

While this inverted non-monotonic dose-response relationship is considered to be 

mediated by both positive and negative feedback effects on estrogenic receptors 

(Welshons et al. 2003; 2006) or related to physiological responses to estrogenic activity 

of 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 replication of perturbations at levels of exposure below those 

demonstrated to have no effects. Furthermore, it has been noted that the shape of the 

dose-response curves for effects of various estrogenic compounds differs depending on 

64
the endpoint assessed and the dosing regime applied (Melnick et al. 2002). As such, the 

significance of the LOEL of 0.02 μg/kg-bw per day for bisphenol A to human health risk 

assessment, however, is difficult to ascertain, particularly in light of the fact that it was 

the only dose level tested below the NOEL of 2 μg/kg-bw per day and some of the effects 

were not consistently observed throughout the study periods. In addition, since the studies 

(exposure period and endpoints measured) were carried out only up to certain embryo 

development stages, the relevance to the postnatal developmental period for newborns 

and infants is unknown. An additional study that examined substance-induced changes in 

brain receptors noted a fluctuating increase in the number of neurons expressing ERα and 

ERβ after oral administration of 2 μg/kg-bw per day bisphenol A to ICR mice from 

gestational day 11 to 17. Increases were observed at 5 and 13 weeks, but not at 9 weeks 

(Kawai et al. 2007). Oral administration of Sprague Dawley rats with 40 μg/kg-bw per 

day bisphenol A from postnatal day 23–30, thereby targeting the period of juvenile 

development, resulted in a transient and gender-specific increase in ERα levels in the 

ventromedial nucleus of the hypothalamus. Changes were noted at postnatal day 37, but 

not at postnatal day 90, in females only (Ceccarelli et al. 2007). Although these effects 

may be considered biomarkers of exposure to bisphenol A, and potential precursor events 

to adverse effects, the biological relevance of these effects for purposes of human health 

risk assessment is not known. 

Volumetric changes of sexually dimorphic regions of the brain have been used as a 

biomarker for developmental disruption by endocrine-active substances [reviewed in 

Patisaul and Polston 2008]. Following subcutaneous injection of 100 mg/kg-bw per day 

to Sprague-Dawley rat pups on postnatal day 1 and 2, no 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 was observed (Patisaul et al. 

2006; 2007). Although bisphenol A was shown to disrupt neuronal phenotype by 

demasculinization of tyrosine hydroxylase (TH) immunoreactive cells and defeminization 

of ERα/TH double labelled cells in male and female rats, respectively, no functional 

changes were observed by examination of gonadotropin-releasing hormone (GnRH) 

neurons (Patisaul et al. 2006; 2007). Thus, neuroanatomical and expression changes 

detected by immunohistochemistry may not always translate into disruption of neuronal 

function [reviewed in Patisaul and Polston 2008]. Examination of GnRH neurons as an 

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 would 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 mesencephalic 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 changes to the SDN-POA in female Sprague 

Dawley rats exposed during gestation and lactation (Kwon 2000). A reversal of gender 

differences in the volume of the locus coeruleus, a nucleus in the brainstem involved in 

stress and panic responses, was observed following administration of drinking water at 

65
0.03 and 1.5 mg/kg-bw per day, respectively (Kubo et al. 2001; 2003). A non-monotonic 

dose response was reported as the extent of reversal was greater at 0.03 mg/kg-bw per 

day than at 0.3 mg/kg-bw per day (Kubo et al. 2003). Similarly, the normal gender 

difference in numbers of corticotrophin releasing hormone (CRH) neurons was disrupted 

in the bed nucleus of the stria terminalis and preoptic area in Wistar rats following 

gestational and lactational exposure in drinking water to 2.5 mg/kg–bw per day bisphenol 

A (Funabashi et al. 2004). Homologous sexually dimorphic regions exist in the human 

brain (Balthazart and Ball, 2007; MacLusky and Naftolin, 1981), however, the relevance 

to human health risk assessment is uncertain. 

Brain receptor expression and immunohistochemical investigations are key for providing 

mechanistic information, and are useful biomarkers, but relevance to human health risk 

assessment is uncertain. This is consistent with interpretations in the draft updated Risk 

Assessment of bisphenol A by the European Union (ECB 2008). 

Behavioural testing was used to characterize the effect of bisphenol A exposure on 

general locomotor and exploratory activity, anxiety and learning and memory in rodents. 

At 10 μg/kg-bw per day, CD-1 mice exposed via drinking water during perinatal 

development showed altered maternal behaviour. Interestingly, the differences in nursing 

behaviour were observed following prenatal or adult exposure, but not following 

continuous exposure (Palanza et al. 2002). Reduced D-amphetamine-related reinforcing 

effects in females (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 effects during critical 

periods of brain development in rodents. However, these studies, though, were limited to 

a single exposure level precluding the evaluation of dose-response. The lowest dose 

leading to 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 conducted using the same 

outbred strain of mice (CD-1), the same experimental dosing protocol (single dose, no 

positive control) and by the same group of researchers in one research institute. 

Effects of gestational exposure to 40 μ/kg-bw per day 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 maternal behaviour of treated rats 

(Della Seta et al. 2005), marginally diminished intergroup differences in pain response 

(Aloisi et al. 2002), altered novelty seeking and impulsive behaviour in females and 

males, respectively (Adriani et al. 2003), as well as altered play behaviour in female rats 

(Porrini et al. 2005). In addition, in Sprague-Dawley rats, pubertal exposure to oral doses 

of 40 μg/kg-bw per day altered behaviour of male rats (Della Seta et al. 2006). These 

studies, again conducted by a common group of researchers, provide convincing evidence 

of bisphenol A-induced effects at 40 μg/kg-bw per day and illustrate the implementation 

of a comprehensive approach. Replication of the aforementioned results by independent 

researchers is needed. 

66
The 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 gestation and lactation at 0.03 or 0.3 mg/kg-bw per day in drinking 

water and noted slightly reduced anxiety-related behaviour in male rats. This behaviour 

was somewhat increased in females. Rearing behaviour, indicative of exploratory 

behaviour, observed by Fujimoto et al. (2006), also reflected opposing effects in male and 

female Wistar rats following 15 μg/kg-bw per day bisphenol A from gestational day 13 to 

birth. Increased duration of rearing was observed in males and was slightly reduced in 

females, suggesting that the sex difference in rearing activity had been abolished. No 

effects on open field behaviour were reported in additional studies assessing the effects of 

oral bisphenol A exposure during the prenatal and neonatal periods with doses ranging 

from 40 μg/kg-bw per day to 400 mg/kg-bw per day (Adriani et al. 2003; Negishi et al. 

2003; 2004). Notably, the 2-generation study in Sprague Dawley rats by Ema et al. 

(2001), also reported no effects of oral bisphenol A exposure (0.2, 2, 20, 200 μg/kg-bw 

per day) on the behavioural endpoints tested in F1 Crj: CD(SD)IGS rats (in open field 

and water-filled T-maze). It was acknowledged in the report that differences in results to 

comparative studies may have been due to possible discrepancies in experimental 

conditions (Ema et al. 2001). However, this large study, though, did not provide data 

from behavioural measures preventing assessment of the relationship between exposed 

and control groups. 

Some studies have investigated the effect of prenatal and neonatal 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 Vanderbergh 2006). Farabollini et al. (1999) found 

that oral administration of 400 μg/kg-bw per day bisphenol A from gestational day 14 to 

postnatal day 6 increased anxiety-related behaviour in male and female rats using the 

hole board test. However, decreased anxiety-related behaviour was noted in males using 

the elevated plus maze. No effects were evident following administration of 40 μg/kg-bw 

per day. Ryan and Vanderbergh (2006) reported increased anxiety in females (males not 

tested) after administration of 200 μg/kg-bw per day bisphenol A from gestational day 3 

to postnatal day 21, but not after 2 μg/kg-bw per day, in the light-dark preference 

chamber. No differences in spatial memory tests were observed. No changes were 

observed in additional studies assessing the same endpoints. Specifically, applying the 

elevated plus-maze test in rats, Fujimoto et al. (2006) observed no anxiety-related effects 

of bisphenol A following drinking water exposure to 15 μg/kg-bw per day from 

gestational day 13 to birth. Male and female offspring were tested. Negishi et al. (2004) 

also found no effects on female offspring exposed to 100 μg/kg-bw per day bisphenol A 

from gestational day 3 to postnatal day 20 (males not tested). Together, these findings 

illustrate the variable responses observed across gender and species in anxiety-related 

behavioural endpoints following a range of bisphenol A exposures. 

Further research into the effects of bisphenol A exposure focused on disturbances in 

cognitive function. Of those studies examining bisphenol A-induced effects on learning 

and memory, no effects were observed by water-filled T-maze following gestational and 

67
lactational 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). Interestingly, in the same study, one week of 

bisphenol A exposure (gestational day 13 to postnatal day 0) altered the normal sex 

differences in struggling behaviour in the forced swimming test observed in control rats. 

Exposure increased the immobility of male rats in the forced swimming test, which is 

indicative of depression-like behaviour called behavioural despair. Male rats appeared to 

be more sensitive to the effects of 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. Responses were variable and did not follow a 

dose-related pattern. Further assessment of male F344 rats exposed to 100 μg/kg-bw per 

day from gestational day 3 to postnatal day 20 showed fewer avoidance responses in 

trials 1 to 3 of 4 in the active avoidance;. However, no differences were observed 

between bisphenol A-exposed and control groups 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; Adriani et al. 2003), methamphetamine- (Suzuki et 

al. 2003) or morphine-induced (Mizuo et al. 2004a; 2004b; Narita et al. 2006; 2007) 

behavioural changes. These studies, using standard testing paradigms, provide evidence 

for altered stimulated responses following bisphenol A administration and are outlined in 

Appendix D. Prenatal exposure of CD-1 mice from gestational day 11 to 18 with 10 

μg/kg-bw per day bisphenol A eliminated D-amphetamine-induced conditioned place 

preference in females suggesting exposure-related impairment of brain reward pathways; 

males were unaffected (Laviola et al. 2005). In rats, the D-amphetamine-induced 

increment activity in the open field test was significantly reduced in males exposed to 40 

μg/kg-bw per day throughout gestation and lactation; in this case, females were 

unaffected (Adriani et al. 2003). Pharmacological challenge studies are suggestive of 

potential organizational alterations of the neural system following perinatal bisphenol A 

exposure. The details of additional studies are not elaborated on as the doses that resulted 

in significant effects on the neurochemical systems were above the established NOAEL 

of 50 mg/kg-bw per day for reproductive/developmental effects; altered behaviour was 

observed following administration of 250 or 400 mg/kg bw per day. 

68
Weight of Evidence Summary of Developmental Neurotoxicity Findings 

Several organizations have recently evaluated the developmental neurotoxicity dataset 

(ECB 2008; EFSA 2006; NTP 2007; NTP 2008; USFDA 2008; Willhite et al. 2008). 

Specifically, the EU in the draft Updated Risk Assessment of bisphenol A concluded that 

“overall, taking together the low confidence in the reliability of the developmental 

neurotoxicity studies and the lack of consistency in the results of behavioural testing, no 

conclusions can be drawn from these studies” (ECB 2008). Other jurisdictions concluded 

that the scientific evidence cited supports the conclusion of some concern for neural and 

behavioural effects of bisphenol A in potentially vulnerable populations (NTP 2007; NTP 

2008). 

It was considered appropriate to characterize the weight of evidence supporting 

neurobehavioral effects in rodents following low-dose exposures to bisphenol A from the 

following perspectives: rigour, power, corroboration/consistency, biological 

plausibility/coherence. 

The rigour of the neurobehavioral dataset was assessed based on use of suitable test 

methodology, endpoints monitored, and analysis used to report results. A majority of the 

studies assessing the effect of bisphenol A on neurobehavioral endpoints used acceptable 

methodology. All studies were published in the peer-reviewed literature. One study (Ema 

et al, 2001) declared compliance with good laboratory practices (GLP). However, the 

rigor of several studies was limited by small sample size, incomplete reporting and 

inadequate statistical analysis. Additionally, only a few studies reported use of concurrent 

positive controls. There were few common endpoints across studies and it was difficult to 

evaluate whether related endpoints that are indicative of anxiety, hyperactivity, 

avoidance, and so forth, respond in a similar manner. As such, the overall rigour of the 

neurobehavioral dataset is considered limited. 

The statistical power of the studies comprising the dataset was considered. Ema et al. 

(2001) conducted a 2-generation study in rats using the largest number of animals per 

exposure group. This study may have had sufficient power to detect subtle changes in the 

endpoints evaluated, however, no effects were observed between exposed and control 

groups. The majority of other studies used fewer animals per exposure group which may 

not be sufficient to detect the neurobehavioral alterations associated with exposure to 

bisphenol A. Diverse study designs and lack of reporting of data, in many cases where 

similar study designs were followed, precluded meta-analysis methodology. As such, the 

overall power of the neurobehavioral dataset is considered limited. 

With respect to corroboration/consistency (i.e., the replication of findings among similar 

studies and observation of similar effects under relevant conditions), the limited data 

preclude characterization of consistency with regard to the qualitative direction of 

neurobehavioral changes regardless of the statistical significance of reported alterations. 

Furthermore, apparent inconsistencies in results may be a function of design and/or the 

behavioural bioassays used to evaluate the potential developmental alterations (e.g., 

different tests of the same underlying functions, age of testing). For instance, 

69
inconsistencies in the effects of bisphenol A on anxiety-related behaviours should 

consider that holeboard, light-dark test (emergence test) and elevated-plus mazes all 

provide measures of anxiety but are also differentially sensitive to other concurrent 

neurobehavioural effects such as motor effect. The same applies to active and passive 

avoidance tests (active avoidance is more susceptible to motor effects). Thus, the 

apparent absence of a consistent pattern of effects may indicate that other factors (such as 

motor effects) may play a role in the discrepant results. As such, a lack of consistent 

effects when comparing these measures does not indicate that bisphenol A does not affect 

anxiety per se. In addition, not only does the comparison between hole board and 

elevated plus maze appear to include differences in the actual test, but the doses and 

dosing regimes used in these tests were different: doses of 15, 100 or 400 μg/kg per day 

bisphenol A with exposure periods from gestational day 13 to birth, gestational day 3 to 

postnatal day 21 or gestational day 14 to postnatal day 6, respectively. Variations in the 

dosing period, especially perinatal dosing, can itself have a substantial impact on 

subsequent effects. As such, there are sufficient methodological differences between 

these studies to preclude drawing firm conclusions regarding lack of effects of bisphenol 

A. A consistent finding among studies was that exposure during gestation and/or during 

early postnatal life consistently represented periods of sensitivity to bisphenol A-induced 

alterations. Additionally, a consistency in the reported effects regarding alterations of 

sexually dimorphic behaviours appears to be emerging in rodents. From the current 

database, it seems likely that the more stongly sexually dimorphic traits are more likely to 

be disrupted by bisphenol A, and moreover, are those most likely to be detected given the 

measurements used. The degree of corroboration demonstrated in the dataset is 

considered limited. 

Finally, biological plausibility is an important consideration when considering the weight 

of evidence. Biological plausibility is informed by investigations of mechanism(s) of 

action. Bisphenol A produces estrogenic effects via interactions with estrogen receptors 

(ER). The estrogen receptor family includes the classic ERα, ERβ, and their splicevariants, extranuclear estrogen receptors associated with cell membranes and neuronal 

synapses (Woolley et al. 2007). Bisphenol A was found to have limited binding affinity 

with ERα, but it showed a nearly tenfold higher affinity for ERβ. Specifically, the relative 

binding affinity (RBA) of bisphenol A for ERα and ERβ was 0.05 and 0.33 respectively, 

whereas RBA for 17b-estradiol was 100 (Kuiper et al., 1997). Strikingly, these chemicals 

are equally potent at inducing effects through cell membrane ER (Quesada et al. 2002). It 

is possible that bisphenol A preferentially affects ERβ-containing tissues including the 

ovary, cardiovascular system and brain (Harris 2007) and those with cell membrane 

receptors. Importantly, estrogen is a key regulator of organizational and activational 

processes in the brain of both males and females. Note that the relative contribution of 

estrogen and androgen to sexually dimorphic behaviours has been demonstrated to differ 

between species. Specifically, it is generally held that the estrogen-signalling pathway 

predominates in rodents, while the androgen-signalling pathway plays a more 

predominate role in humans and non-human primates (Li et al. 2008). Thus, the relative 

significance of the estrogenic mode of action of bisphenol A in rodents with respect to 

human relevance requires further investigation. 

70
Typically, demonstration of a dose-response relationship supports biological plausibility. 

The dose-response relationship following exposure to bisphenol A at doses below the 

established NOAEL of 50 mg/kg-bw per day for reproductive/developmental effects is 

complex and not yet completely understood. This is in large part due to a lack of studies 

consistently assessing multiple endpoints following administration of 3 or more 

incremental doses of bisphenol A. Regarding the neurobehavioral dataset, a very limited 

number of studies assessed more than one dose. Therefore, a dose-response relationship 

could not be established. An attempt was made to compare and evaluate bisphenol Arelated changes at similar doses between similar experiments. However, this was not 

possible due to a number of the aforementioned design limitations. In a few studies which 

included multiple doses, bisphenol A-related effects were seen at some doses, but not at 

others (Kubo et al. 2003; Narita et al. 2006; Tando et al. 2007). Further characterization 

of the dose-response and examination of the significance and relevance of non-linear 

dose-response relationships are needed for a comprehensive understanding of 

neurodevelopmental effects associated with bisphenol A exposures in the μg/kg-bw per 

day range. The weight of evidence for biological plausibility is considered limited. 

Overall, taking into consideration rigour, power, corroboration/consistency, and 

biological plausibility/coherence, the weight of evidence supporting neurobehavioral 

effects in rodents following exposures to bisphenol A at exposures below established 

NOAELs for reproductive/developmental toxicity is limited. 

An emerging trend from the developmental neurotoxicity studies is that prenatal and/or 

neonatal bisphenol A exposure may cause organizational effects in the brain leading to 

disruptions in the sexually dimorphic behaviours in mice and rats. Although the mode of 

action, and thus biological plausibility, of the reported effects of bisphenol A is not clear, 

it is not deemed necessary for risk assessment if the evidence indicates an adverse effect 

(Li et al. 2008). Detailed mechanistic information can reduce uncertainty regarding the 

relevance of rodent findings to humans, but is not crucial for general toxicity screening. 

Thus, as highlighted by the NTP (2007), it is not clear whether the reported changes 

constitute critical adverse toxicological responses, however, the uncertainties associated 

with the results from neurodevelopmental and behavioural testing warrant concern with 

respect to human health, in particular during certain sensitive life stages. 

Epidemiology Investigations 

Epidemiology studies have investigated cohorts exposed to bisphenol A and these are 

summarized in Vandenberg et al. (2007). Although concentrations of bisphenol A in 

blood have been associated with a variety of conditions in women including obesity, 

endometrial hyperplasia, recurrent miscarriages, and polycystic ovarian syndrome, as 

well as elevated androgen hormone levels, these studies have many limitations including 

small sample size, potential exposure misclassification, and lack of control of potential 

confounding factors. These studies did not investigate neurodevelopmental and 

behavioural effects in the potentially sensitive subpopulations (the pregnant woman/fetus 

and infant). 

71
Confidence in the toxicological database, overall, is considered moderate. Data for acute 

toxicity, repeated dose toxicity, genotoxicity, carcinogenicity and reproductive and 

developmental toxicity are available. Although there is an extensive database from which 

to evaluate the effects of bisphenol A exposure, inconsistencies and experimental 

limitations exist among studies evaluating reproductive and developmental endpoints. 

Additionally, several of the studies administered a single dose precluding characterization 

of dose response. 

Characterization of Risk to Human Health 

Consistent with the assessment of the European Commission (ECB 2003), which 

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, a critical effect for characterization of risk to human health is 

reproductive and developmental toxicity. 

Exposure estimates for the general population of Canada range from 0.08 μg/kg-bw per 

day to 4.30 μg/kg-bw per day. Specific exposure estimates for the most highly exposed 

subpopulation (i.e., infants) range from an average of 0.27 μg/kg-bw per day (maximum 

1.75 μg/kg-bw per day) for infants aged 12 to 18 months to an average of 0.50 μg/kg-bw 

per day (maximum 4.30 μg/kg-bw per day) for infants aged 0 to 1 month. The NOAELs 

from the multigeneration reproductive toxicity studies in Sprague- Dawley rats and CD-1 

mice of 5 mg/kg-bw per day for systemic effects (reduced body weight gain in rats and 

minimal to mild hepatocyte hypertrophy in adult male and female mice) and 50 mg/kgbw per day for reproductive and developmental toxicity (Tyl et al. 2002; 2007) are 

considered an appropriate departure point for characterizing risk to human health from 

exposure to bisphenol A. Based on the NOAEL of 5 mg/kg-bw per day, margins of 

exposure for infants exposed to bisphenol A from consumption of infant formula, use of 

polycarbonate bottles and via environmental media range (Tables 19a and 19b) from 

1160 to 18 520. For infants 0-7 months, exposed via the consumption of breastmilk and 

environmental media (Table 20), margins of exposure to the NOAEL of 5 mg/kg-bw per 

day range from 3940 to 25 000. For other segments of the population, exposed via the 

consumption of canned food, use of polycarbonate bottles and environmental media 

(Table 21), margins of exposure to the NOAEL of 5 mg/kg-bw per day range from 2200 

to 62 500. Margins of exposure to the NOAEL of 50 mg/kg-bw per day would be 10-fold 

higher than those presented above. Although these margins are considered adequate to 

address interspecies and intraspecies differences with respect to the adverse effects on 

which the NOAELs are based, the dataset of neurodevelopmental and behavioural 

studies, though limited when subject to a weight of evidence analysis, is suggestive of 

potential effects at doses well below these NOAELs. 

Specifically, a small number of studies in CD-1 mice reported alterations in behaviour 

following oral administration of 10 μg/kg-bw per day. These include altered maternal 

behaviour in dams exposed either as fetuses or during adulthood and decreased or 

72
eliminated sex-related behavioural differences following gestational or gestational and 

lactational exposures (Palanza et al. 2002; Laviola et al. 2005; Gioiosa et al. 2007). At 40 

μg/kg-bw per day, a small number of studies in rats have reported behavioural effects 

including gender-specific changes in sexual performance, effects on active and passive 

maternal behaviour, and altered novelty seeking and impulsive behaviour in both sexes in 

adults whose mothers were administered bisphenol A during gestation and lactation 

(Farabollini et al. 2002; Della Seta et al. 2005; Adriani et al. 2003). Further supporting 

the effect of bisphenol A exposure on altered behaviour, oral administration of 100 

μg/kg-bw per day in rats during gestation and lactation reduced the number of correct 

avoidance responses (Negishi et al. 2004), while early postnatal exposure disrupted the 

normal gender-dependent pattern of maze performance acquisition in the Morris water 

maze (Carr et al. 2003). 

While collectively these studies provide evidence that exposure to 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 limitations such as conduct of behavioural 

assessments at a single time point); power (e.g., limited number of animals per test 

group), corroboration/consistency ( limited consistency of studies) and biological 

plausibility (e.g., certain studies involve use of a single dose, lack of dose response 

relationship). These limitations make it difficult to determine actual significance of 

findings to human health risk assessment

The neurodevelopmental and behavioural dataset in rodents, though highly uncertain, is 

suggestive of potential effects at doses at the same order of magnitude to 1-2 orders of 

magnitude higher than exposures. Given that toxicokinetics and metabolism data indicate 

potential sensitivity to the pregnant woman/fetus and infant; and that animal studies 

suggest a trend towards heightened susceptibility during stages of development in 

rodents, it is considered appropriate to apply a precautionary approach when 

characterizing risk. As such, it is concluded that bisphenol A be considered as a substance 

that may be entering the environment in a quantity or concentration or under conditions 

that constitute or may constitute a danger in Canada to human life or health. 

Uncertainties in Evaluation of Risk to Human Health and 

Identification of Research Needs 

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

including: 

• Uncertainty in the estimation of exposure to all age groups from the use of 

polycarbonate repeat-use containers due to absence of Canadian use pattern 

information (e.g., frequency of use, method of use) as well as the availability of 

studies representative of actual use conditions. 

73
• fants from the consumption of 

breastmilk due to lack of surveys measuring concentrations in breastmilk from 

 of 

 or 

t 

• vels of bisphenol A in polymers used to 

manufacture consumer products, such as epoxy adhesives, toys and cosmetics, and 

• intake 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. Rodents clear bisphenol A in the faeces via 

 to 

•

itive populations. Actual levels of free 

s with 

Uncertainty in the estimation of exposures to in

Canadian women, or surveys involving larger sample sizes. Further exploration

exposures (e.g., food, environmental media) contributing to levels in breastmilk

other biomonitoring would assist in determining which sources of exposure are mos

important for risk management, if required. 

Uncertainty associated with actual residual le

uncertainty associated with the assumptions 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 reprodu

uncertainty. Limitations with respect to rigour, power, corroboration/consistency

biological plausibility/coherence exist. 

Interspecies variation. Uncertainties exis

biliary excretion of the glucuronide (EFSA 2006). Humans, on the other hand, clear

bisphenol A primarily via urine after glucuronidation (Snyder et al. 2000; Volkel et 

al. 2002). It has been suggested that the fetus from rats and humans have little 

glucuronidation activity (Matsumoto et al. 2002; 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 uncertainty in extrapolating from rodents to 

humans. Uncertainties also exist regarding relevance to humans of potential effects of 

bisphenol A on brain function in rodents. Species differences in brain development 

are important (Seegal 2001). The brain of a newborn rat pup is comparable to that of 

a human fetus nearing the third trimester of pregnancy. Conversely, the human 

newborn might be compared with a 2- to 3-week-old rat pup (Morreale et al. 2004). 

The mechanism(s) by which bisphenol A exerts estrogenic effects depends on many 

factors, including species, and this is an additional uncertainty with respect to 

interspecies extrapolation. 

With regards to intraspecies variation, available data indicate that the pregnant 

female/fetus and infant are potentially sens

bisphenol A in the fetus however remains an uncertainty. Uncertainty also exist

respect to critical windows of susceptibility during in utero and postnatal 

development and potential subsequent long-term effects. 

74
esearch needs include the following: 

rtain environmental media (e.g., air, soil, drinking 

water) are lacking or limited. Such data would decrease uncertainty regarding the 

• nant 

women and children younger than 6 years old) would decrease uncertainty 

• ration from polycarbonate 

would inform risk management. To date, studies in this area report inconsistent 

ase 

• establish relationships between bisphenol 

A exposure and health outcomes for sensitive subpopulations. 

•

may induce alterations in the developing 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 population of Canada (including preg

regarding the exposure assessment. Investigations into accuracy of spot urine 

samples in predicting daily exposures are needed. 

Studies investigating the underlying process of mig

results, with some study investigators indicating that migration rates will decre

with use and others indicating that migration 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 mechanisms 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 environmentally 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 the sexual differentiation of the brain a

neurobehavioural endpoints conducted in acco

developmental neurotoxicity. 

Investigation of the effects of bisphenol A exposure during precise age window

that correspond with specific a

of the neuraxis that are the most critical to the development of neurobehavioural 

features. For example, sensory and motor differences are generally brainstem, 

anxiety-related activity involves mid-brain and lower telencephalon, speciestypical behaviour patterns are most likely lower telencephalon and learning and 

memory differences involve the cortex. 

75
• ed changes on the reproductive and 

nervous system during early and later stages of development including 

• identify the doses and ages or windows of sensitivity 

that are the most effective in producing developmental disturbances. 

• exposures 

via oral or parenteral routes 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-induc

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 bisphenol A exposures is required. 

this screening

bi

the environment in a quantity or concentration 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 this screening assessment, and applying a 

precautionary approach, it is concluded that bisphenol A be considered as a substance 

that may be entering the environment in a quantity or concentration or under 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 criteria in paragraphs 64(a) and 

64(c) of CEPA 1999. 

76
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99
Appendix A. Robust Study Summary – Inherent Toxicity

Item Yes No 

Reference: Sohoni et al. 2001. Reproductive effects of long-term exposure 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 substance (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 (laboratory or field) specified? X 

System type (static, semi-static, flow through)? X 

Negative or positive controls : Negative X 

Number of replicates (including controls) and concentrations X 

Exposure pathways (food, water, both) X 

Exposure duration X 

*Measured concentrations reported? X 

Exposure media conditions (temperature, pH, electrical 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 (LC50, EC50, or IC50- 164 d LOEC (lowest-observed effect concentration – somatic growth in adults, egg hatchability) = 

0.64 mg/L; 164 d LOEC (vitellogenin [VTG] synthesis, sex cell type proportion) = 0.016 mg/L; 164 d LOEC (egg production) = 1.28 mg/L 

Other endpoints reported – e.g., bioconcentration factor / bioaccumulation factor (BCF/BAF): 

*Was toxicity value below the chemical’s water solubility? X 

Other adverse effects (carcinogenicity, 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 high performance liquid chromatography (HPLC) analysis criteria are provided and this is considered to 

satisfactorily confirm the identity of the test material. 

100
Appendix 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 substance (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 (laboratory or field) specified? X 

System type (static, semi-static, flow through)? X 

Negative or positive controls : Solvent control X 

Number of replicates (including controls) and concentrations X 

Exposure pathways (food, water, both) X 

Exposure duration X 

*Measured concentrations reported? X

Exposure media conditions (temperature, pH, electrical 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 (LC50, EC50, or IC50 –: 14 d LOEC (lowest-observed-effect concentration – 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 for endocrine 

disruption 

*Was toxicity value below the chemical’s water solubility? X 

Other adverse effects (carcinogenicity, 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 substance and standard testing methods used are not included in the 

report, it is likely that these met acceptability criteria given that the study is part of a larger government-sponsored program and the results 

were submitted to the United Kingdom Department of Environment, Food and Rural Affairs. The description of test methods includes 

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 terms of nominal concentration values. Similarly, procedures for measurement of 

water quality parameters are described, but the numerical values are not provided in the report. Consider data generated from the study to be 

of acceptable quality, even though the resulting report does not fully describe all results. 

101
Appendix C. 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 substance (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 (laboratory or field) specified? X 

System type (static, semi-static, flow through)? X 

Negative or positive controls: Solvent control X 

Number of replicates (including controls) and concentrations X 

Exposure pathways (food, water, both) X 

Exposure duration X 

*Measured concentrations reported? X

Exposure media conditions (temperature, pH, electrical 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 (LC50, EC50, or IC50 –: 14 d LOEC (lowest-observed-effect concentration – 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 (carcinogenicity, 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 substance and standard testing methods used are not included in the 

report, it is likely that these met acceptability criteria given that the study is part of a larger government-sponsored program and the results 

were submitted to the United Kingdom Department of Environment, Food and Rural Affairs. The description of test methods includes 

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 terms of nominal concentration values. Similarly, procedures for measurement of 

water quality parameters are described, but the numerical values are not provided in the report. Consider data generated from the study to be 

of acceptable quality, even though the resulting report does not fully describe all results. 

102
Appendix D. Summary of Studies Investigating 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. 

103
Author 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 males 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 17β-oestradiol and 

testosterone level 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 (17β-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 avoidance responses in 

active avoidance test trials 1 to 3 of 4 

(suggests slower learning). 

104
Author 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 females (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 ERα and ERβ on 

male mice. Immunostaining to 

ERα and ERβ, serotonin and 

serotonin transporter at 4–5, 8–9 

or 12–13 weeks of age. 

Increased number of neurons 

expressing ERα and ERβ 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 juvenile BPA exposure 

in brain development. ERα levels 

assessed in three sexually 

dimorphic regions of the 

hypothalamus (PND 37 and 90) 

Increased ERα 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 (sst2) 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 sst2 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 sst3 

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 αGABAA receptors 

enhanced responses. 

Central Dopaminergic System 

105
Author 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, 

[

35S]GTPγS binding assay in 

response to morphine. 

Increased place preference (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, 

[

35S]GTPγS 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. 

106
Author 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 Immunohistochemical 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 ERα

and/or tyrosine hydroxylase (TH). 

TH positive cells increased in males 

(demasculinization); decreased TH/ 

ERα 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