4,4'-isopropylidenediphenol (bisphenol a) documents/p-i209/other/p-i209... ·...

4,4'-Isopropylidenediphenol (Bisphenol A)

EC Number: 201-245-8

CAS Number: 80-05-7

IUPAC Name: 2,2-bis(4-hydroxyphenyl)propane

Additional information on REACH registration dossier

The information compiled in this document is included in the REACH registration dossier for 4,4'-isopropylidenediphenol but is currently not disseminated on the ECHA website. However, the registrants believe that this information is required to comprehend the conclusions as derived in the REACH registration dossier for 4,4'-isopropylidenediphenol. Please note that at present two registration dossiers for 4,4'-isopropylidenediphenol are available on the ECHA website. This is due to the fact that one registrant chose to submit an individual registration dossier despite being aware that a joint registration dossier had already been submitted by the Bisphenol A REACH consortium. The extensive registration dossier of the Bisphenol A REACH consortium can easily be identified by the IUPAC name 2,2-bis(4-hydroxyphenyl)propane in section 1.1 of the disseminated registration dossier. The additional information compiled in this document does only refer to this registration dossier as submitted by the Bisphenol A REACH consortium.

EC number: 201-245-8 4,4'-isopropylidenediphenol CAS number: 80-05-7

2011-10-20 Additional information on REACH registration dossier 2

Table of Contents 1. PHYSICAL AND CHEMICAL PROPERTIES 3 2. ENVIRONMENTAL FATE PROPERTIES 4 3. ENVIRONMENTAL HAZARD ASSESSMENT 6

3.1 Aquatic toxicity 6 3.2 Sediment toxicity 13 3.3 Terrestrial toxicity 16 3.4 Environmental classification justification 19

4. HUMAN HEALTH HAZARD ASSESSMENT 20 4.1 Toxicokinetics, metabolism and distribution 20 4.2 Acute toxicity 21 4.3 Irritation / Corrosion 21 4.4 Sensitisation 22 4.5 Repeated dose toxicity 23 4.6 Genetic toxicity 23 4.7 Carcinogenicity 24 4.8 Toxicity to reproduction 25 4.9 Specific investigations 26 4.10 Derivation of DNEL(s) / DMEL(s) 26

5. PBT and vPvB ASSESSMENT 36 REFERENCES 38



1. PHYSICAL AND CHEMICAL PROPERTIES

Molecular Weight: 228.28 g/mol

Appearance/physical state/colour: Bisphenol A is a white solid at environmentally relevant temperatures

Melting Point: 155 °C

Boiling Point at 17 hPa: 250 - 252 °C (with potential decomposition), Boiling Point at 1013 hPa: 360 °C (with decomposition)

Density: 1.2 g/cm3 at 25 °C

Vapour Pressure: 4.12E-09 hPa at 25 °C

Log Kow: 3.4 at 21.5 °C and pH 6.4

Water Solubility: 300 mg/L at 25 °C

Flash point: 213 °C at 1013 hPa

Auto-ignition temperature: 510 °C at 1013 hPa (decomposition starts at lower temperatures)

Stability: Stable. Not self-heating and not corrosive to metals

pKa: 11.3

Bisphenol A is a white solid melting at 155 °C and boiling at 360 °C under decomposition. With a moderate water solubility of 300 mg/L and log Kow of 3.4, Bisphenol A is expected to adsorb somewhat to soil and sediment and not migrate through groundwater. With a pKa of 11.3, Bisphenol A is not expected to significantly dissociate in water at environmentally relevant pH values (5 to 8).



2. ENVIRONMENTAL FATE PROPERTIES

Stability (Phototransformation in air):

After exposure to air, Bisphenol A will be rapidly degraded by photochemical processes. The rate constant for indirect photolysis in air (OH-mediated photodegradation) is 80.6E-12 cm3 / (molecule*sec) and half-life is 0.13 days, assuming 12 hours daylight and a hydroxyl radical concentration of 1.5E6 molecules per cm3.

Stability (Hydrolysis):

Based on the molecular structure of Bisphenol A, hydrolysis is not expected under environmentally relevant conditions.

Stability (Phototransformation in water):

Bisphenol A has a half-life of 0.5 to 10 days for direct photolysis and 0.7 to 1.3 days for indirect photolysis.

Bisphenol A may undergo both direct and indirect photolysis in water. Direct photolysis half-lives of 0.5 to 10 days have been reported from laboratory experiments. Indirect photolysis (in the presence of various dissolved organic matter) half-lives of about 0.7 to 1.3 days have been reported from laboratory experiments.

Biodegradation (in water, screening tests):

Bisphenol A is readily biodegradable.

Across a number of tests using international guidelines for measuring aerobic biodegradation under stringent test conditions, Bisphenol A is shown to be readily biodegradable.

Biodegradation (in water and soil, simulation tests):

Bisphenol A was found to be rapidly biodegraded by microbial consortia found in many natural waters and sediments, with lag times and half-lives on the order of a few days.

Biodegradation simulation tests have measured the die-away of Bisphenol A using surface water and sediments collected from rivers, estuaries, and marine waters in North America, Europe, Japan, and Australia. Bisphenol A was found to be rapidly biodegraded by the microbial consortia found in the natural waters and sediments, with lag times and half-lives on the order of a few days. Biodegradation rates are expected to be slower and half-lives longer in tests conducted using lower temperatures and oxygen content. Microbial populations capable of rapidly biodegrading Bisphenol A appear to be ubiquitous in the environment. The biotransformation pathways for Bisphenol A have been defined in several studies.

Biodegradation (in soil):

Bisphenol A dissipates in soil in less than 3 days, with a major route being conversion to non-extractable bound residues.



Bioaccumulation:

Bisphenol A has low potential for bioaccumulation based on laboratory bioconcentration factors of less than or equal to 73 L/kg in fish.

Transport and distribution (Adsorption/desorption):

The soil sorption organic carbon normalised partition coefficient values (Koc) of Bisphenol A range from 251 to 1507 L/kg, averaging 750 +/- 348 L/kg.

Transport and distribution (Henry´s Law constant):

At ambient temperature and atmospheric pressure, Bisphenol A is not volatile, with a Henry's law constant calculated to be 3.12E-7 Pa*m3/mol from aqueous solubility and vapour pressure.

Transport and distribution (Distribution modelling):

Level III Mackay-type modelling shows that nearly all Bisphenol A goes into the water and soil compartments. The exact proportion in each depends on where emissions enter the environment.

Environmental Monitoring:

Concentrations of Bisphenol A have been taken from 89 reports and publications from 1997 to 2007 containing "reliable" or "very reliable" surface and/or sediment data from North America and Europe. The data were subjected to statistical analysis using the Kaplan-Meier method because the dataset contained numerous non-detected values. To assess their ecological significance, the exposure concentrations were compared with regulatory criteria. The results suggest that the frequency of location in which concentrations are likely to cause adverse effects in aquatic systems is low, with the exception of sediments collected from some highly urbanised and industrial locations.



3. ENVIRONMENTAL HAZARD ASSESSMENT

3.1 Aquatic toxicity

Short term toxicity to fish

There are four key studies that are fully reliable that characterise the acute toxicity of Bisphenol A to fresh and saltwater fish.

The 96-hour flow-through exposure to the freshwater fish, fathead minnow (Pimephales promelas), resulted in an LC50 of 4.6 mg/L and a NOEC of 2.28 mg/L based on survival (Alexander et al., 1985; also reported in Alexander et al., 1988). In a longer term exposure, a 14-day semistatic study with zebra fish (Brachydanio rerio) resulted in a LOEC of 10.15 mg/L and a NOEC of 3.2 mg/L (Bayer AG, 1998).

Two marine fish were tested in acute exposures. A 96-hour acute flow-through exposure to the sheepshead minnow (Cyprinodon variegatus) resulted in an LC50 of 11 mg/L (Sayers, 2009). Similarly, a 96-hour acute flow-through exposure to the marine fish Atlantic silversides (Menidia menidia), resulted in an LC50 of 9.4 mg/L (Springborn Bionomics, 1985a; also reported in Alexander et al., 1988). Other acute LC50s from studies deemed to be valid with restrictions exposing sheepshead minnow, rainbow trout, or medaka to Bisphenol A ranged from 3.0 to 8.3 mg/L.

Of these four key acute studies, the two most conservative LC50's, 4.6 mg/L for freshwater and 9.4 mg/L for marine water were chosen for the Chemical Safety Assessment.

The following information is taken into account for acute fish toxicity for the derivation of PNEC:

While there are four key studies with acute values that are very similar, the two most conservative LC50's were chosen for the CSA. Tests on acute toxicity of Bisphenol A towards freshwater fish resulted in an LC50 of 4.6 mg/L (Alexander et al., 1985, also reported in Alexander et al., 1988). Tests on acute toxicity of Bisphenol A towards marine water fish resulted in an LC50 of 9.4 mg/L (Springborn Bionomics, 1985a, also reported in Alexander et al., 1988).

Value used for CSA:

LC50 for freshwater fish: 4.6 mg/L

LC50 for marine water fish: 9.4 mg/L

Long term toxicity to fish

There are three key freshwater long-term toxicity studies, all with the fathead minnow, and one life-cycle study with the estuarine fish, the sheepshead minnow. Caunter et al. (1999) reported the results of a 36-day early life stage study with fathead minnow. The NOEC determined in that study based on hatchability, survival and growth was 640 µg/L. Caunter et al. (2000) reported on a multigeneration study exposing the fathead minnow (Pimephales promelas) for 444 days, through two generations, to a dilution water control and nominal Bisphenol A concentrations of 1.0, 16, 160, 640, and 1280 µg/L. Daily observations of mortality, behaviour and appearance were made and any abnormal effects recorded for the F0, F1, and F2 generation fish. Survival, growth, reproduction, gonadal size, vitellogenin, and gonadal histology were evaluated. Gonadal evaluations and vitellogenin concentrations were reported in the companion study (Sumpter et al., 2001) but they did not impact the most sensitive effect determined in the Caunter et al. (2000) study of F2 hatchability. Weaknesses in the spermatogenesis portion of the Sumpter et al. (2001) study (e. g. quality of histopathology sections/slides, number of fields counted, number of cell types assessed) result in not being able to use that data to set a NOEC based on the histology data alone. The NOEC for vitellogenin production in males was 16 µg/L. Overall, the NOEC established based on F2 hatchability in this multigenerational exposure was 16 µg/L. The other longer-term key study is a follow-on to the Caunter et al. (2000) and the Sumpter et al. (2001) studies to evaluate the possible effects of Bisphenol A on gonadal cell growth, as well as more population relevant endpoints (Rhodes, 2008). In this 164 day study, there were no statistically significant effects at any treatment level in males or females with respect to survival, growth, fecundity, hatchability, and gonadosomatic index, with the exception



of reduced survival in males at 640 µg/L. With respect to supplemental endpoints evaluated in this study, there was a statistically significant increase in vitellogenin at 64 µg/L and higher in both males and females, compared to controls. Gonadal histopathology showed a statistically significant increase in intravascular proteinaceous fluid in females, with minimal to mild changes at 640 µg/L, and in males with minimal to moderately-severe changes at 160 and 640 µg/L. A statistically significant shift towards less mature gametogenic cell types (relative cell frequency as compared to controls) was observed in females at 640 µg/L and in males at 160 µg/L and 640 µg/L. In summary, population relevant reproduction endpoints of survival, growth, fecundity and hatchability were not impacted at any concentration. Contrary to previous study results (Sumpter et al., 2001), changes to the distribution of testicular cell types only occurred at the highest treatment levels tested in this robust study. The ecologically relevant NOEC based on male survival was 160 µg/L.

There is one key estuarine/marine life-cycle study with the sheepshead minnow (York, 2010). The objective of this 116-day study was to evaluate the long-term (chronic) effects of exposure to Bisphenol A on the marine fish, sheepshead minnow (Cyprinodon variegatus). Data were compiled on the effects of exposure on hatching success, survival, growth (total length and wet weight) and reproductive success of first generation (F0) fish (eggs/female/day) and the hatching success, survival and growth (total length and wet weight) of their progeny (F1). No effects were noted in any of the parameters measured at the highest concentration tested except for reproductive success measured as eggs/female/day. Based on F0 reproductive success (eggs/female/day), the study NOEC was 66 µg a. i. /L and the study LOEC was 130 µg a. i. /L.

For Bisphenol A, the sheepshead minnow NOEC of 66 µg a. i. /L is chosen for the Chemical Safety Assessment for marine fish. The most sensitive effect endpoint that is relevant to fish populations is egg hatchability in the fathead minnow. The NOEC based on egg hatchability from a fathead minnow life-cycle study is 16 µg/L (Caunter et al., 2000).

The results from these robust, guideline studies are in direct contrast to a study by Lahnsteiner et al. (2005, Aquatic Toxicology 75:213 -224) that has been used in one geography as part of a risk assessment, although its use was rejected in the EU risk assessment as it was not deemed to be of sufficient quality. The reported apparent effects on gamete quality in wild-caught brown trout were not evaluated against any measures of reproductive success or population relevant endpoint. Due to lack of replication, lack of analytical confirmation of test concentrations, low numbers of fish, or examination of any primary endpoint related to survival, growth or reproductive success, this study is not suitable for risk assessment. A similar but more robust study was performed with brown trout (Bjerregaard et al., 2008, Ecotoxicology 17(4):252 -263) and no effects of Bisphenol A on sex hormone concentration or gonadal histology were observed. Although the study reported by Bjerregaard et al. (2008) used a more robust design, it still cannot be used for risk assessment because no endpoints related to survival, growth, and reproduction were measured. Thus, the F2 hatchability NOEC of 16 µg/L from the 2-generation fathead minnow study and the NOEC of 66 µg/L from the life-cycle sheepshead minnow study are the most conservative endpoints from fully valid studies and will be used in the deriving aquatic PNEC for use in risk assessment.

The following information is taken into account for long-term fish toxicity for the derivation of PNEC:

For Bisphenol A, the most sensitive effect endpoint that is relevant to fish populations is egg hatchability in the fathead minnow. The NOEC based on egg hatchability from a fathead minnow life-cycle study is 16 µg/L (Caunter et al., 2000). For marine species, a NOEC for reproductive success (egg/female/day) was observed in a life-cycle exposure to sheepshead minnow at 66 µg a. i. /L (York, 2010).

Value used for CSA:

EC10/LC10 or NOEC for freshwater fish: 0.016 mg/L

EC10/LC10 or NOEC for marine water fish: 0.066 mg/L

Short-term toxicity to aquatic invertebrates

Many acute freshwater and marine water invertebrate studies are available for Bisphenol A, although often there has not been confirmation of dose during the study. Two key studies were identified that did follow appropriate guidelines with analytical confirmation of dose.



A 48-hr static exposure acute study with Daphnia magna was performed with Bisphenol A (Alexander et al., 1985; also reported in Alexander et al., 1988). Daphnia were exposed to nominal concentrations ranging from 0.93 to 20.0 mg/L and a control. These corresponded to measured concentrations ranging from 0.90 to 19.34 mg/L. The 24-hr and 48-hr EC50 values (and 95 % confidence intervals) were 16 (14 -17) and 10.2 (9.2 -11) mg/L, respectively.

A 96-hr acute flow-through study with the estuarine mysid shrimp (Mysidopsis bahia) was performed with Bisphenol A (Springborn, 1985b; also reported in Alexander et al., 1988). Mysidopsis bahia were exposed to nominal concentrations of 0, 0.89, 1.4, 2.1, 3.2, and 5.0 mg/L. These corresponded to measured concentrations of 0, 0.51, 0.86, 1.4, 1.9, and 3.3 mg/L, respectively. The 96-hr LC50 (95 % confidence interval) from this study was calculated to be 1.1 (0.92-1.2) mg/L.

These findings are in good agreement with the range of results for aquatic invertebrates, generally between 1.1 and 16 mg/L, published for Bisphenol A.

The following information is taken into account for short-term toxicity to aquatic invertebrates for the derivation of PNEC:

Tests to determine acute toxicity of Bisphenol A towards freshwater invertebrates resulted in an 48-hr EC50 of 10.2 mg/L (Alexander et al., 1985, also reported in Alexander et al., 1988). Tests on acute toxicity of Bisphenol A towards marine water invertebrates resulted in an 96-hr LC50 of 1.1 mg/L (Springborn, 1985b; also reported in Alexander et al., 1988).

Value used for CSA:

EC50/LC50 for freshwater invertebrates: 10.2 mg/L

EC50/LC50 for marine water invertebrates: 1.1 mg/L

Long-term toxicity to aquatic invertebrates

The existing literature for results of chronic invertebrate tests (Klimisch rating of 1) report NOECs ranging from 0.025 to 3.16 mg/L.

A standard, chronic Daphnia magna reproduction study as well as a study to assess the moulting behavior of Daphnia exposed to Bisphenol A were performed by Caspers (1998). No statistically significant difference in reproduction was observed in this standard OECD 211 study. The NOEC for reproduction was the highest dose tested of 3.16 ppm. Also, no statistically significant differences in the moulting frequency between the control and the two Bisphenol A concentrations (0.316 ppm and 3.16 ppm) were observed.

A second key study for chronic invertebrate exposure is a life-cycle rotifer study with Bisphenol A (Springborn Smithers, 2006a; also reported in Mihaich et al., 2009). The purpose of this study was to determine the chronic toxicity of Bisphenol A to the rotifer (Brachionus calyciflorus) under static test conditions. The intrinsic rate of increase (reproduction) was used to determine the NOEC and the LOEC. Based on mean measured concentrations and the intrinsic rate of increase, the NOEC was determined to be 1.8 mg/L. The LOEC was determined to be 3.6 mg /L.

In a 28-day life cycle study with the estuarine mysid shrimp (Americamysis bahia) (Lee, 2010), endpoints of F0 survival, growth (mean dry body weight and mean total body length) of both male and female mysids and reproduction (number of young released per female) were assessed. Based on the time-weighted average measured concentrations and the most sensitive endpoint analysed (reproduction), the Lowest-Observed-Effect Concentration (LOEC) was determined to be 370 µg a. i. /L. The No-Observed-Effect Concentration (NOEC) for Bisphenol A and mysids was determined to be 170 µg a. i. /L. Since no concentration tested resulted in ≥ 50% mortality, the 28 day LC50 value was empirically estimated to be > 370 µg a. i. /L, the highest time weighted average concentration tested.

Oehlmann and colleagues (Oehlmann et al., 2000, Ecotoxicology 9:383 -397; Schulte-Oehlmann et al., 2001, Umweltchem. Okotox. 13(6):319-333; Oehlmann et al., 2006, Env. Health Perspec. 114 (Supplement 1): 127-133) reported significant increases in egg production (referred to by the authors as superfeminasation) in giant ramshorn snails, Marisa cornuarietis exposed to Bisphenol A. The Oehlmann et al. (2000) experiments



contained no replication, did not analytically confirm test concentrations, had varying density of organisms in the tanks, and employed incorrect statistics. Review of the scientific literature provided insufficient information on appropriate husbandry conditions, feeding habits, and breeding characteristics to design appropriate toxicity testing procedures with this species. A draft EU risk assessment called for the development of further testing to better assess impacts on snails. A colony of Marisa cornuarietis was established with adult snails collected from a remote lake in Puerto Rico located in the Gulf of Mexico. M. cornuarietis are typically found in tropical or subtropical freshwater bodies. Studies were conducted by scientists from three laboratories in the US (ABC Laboratories, Inc., Columbia, Missouri), the UK (Brixham Environmental Laboratory, Devon, England), and Denmark (Roskilde University). The results of a series of experiments on life history traits, appropriate food types, feeding frequencies, and the effects of temperature on Marisa snails were reported (Aufderheide et al., 2006; Selck et al., 2006). Using these data, Forbes et al. (2007a, b) established appropriate experimental design parameters with special focus on intra- and inter-laboratory variability in endpoints related to survival, growth, and reproduction of Marisa. Forbes et al. (2007a, b) also reported the results of preliminary toxicity testing in order to establish test design requirements that would provide sufficient statistical power to assess these endpoints.

Definitive toxicity tests with Marisa were conducted and reported by Warbritton et al. (2007a) and Forbes et al. (2008) who examined effects on adult fecundity, hatchability, and juvenile growth during a 328-day study with Bisphenol A. The study was run under flow-through conditions at 25oC, with nominal Bisphenol A concentrations of 0.1, 1.0, 25 and 640 µg/L, plus a control. For the adult fecundity trials, no effects of Bisphenol A were detected at any test concentration. Neither increased egg production nor superfeminization were observed. From these results, the NOEC for this trait was estimated to be > 640 μg/L. For the egg hatchability trial, no significant differences between the control and any Bisphenol A concentration were detected for percent hatch or for time to first- or 50% hatch. Thus the NOEC for hatchability traits was estimated to be > 640 μg/L. For the juvenile growth trial, significant impairments in female growth were detected at 640 μg/L. Though a statistically significant increase in male (but not female) juvenile growth was observed at 1 μg/L, the contributions of breeding pair and inter-juvenile variability contributed measurably more to the variance in juvenile growth rate than Bisphenol A treatment. Thus a NOEC for juvenile growth rate is estimated to be 25 μg/L based on the observed negative effects on female growth at 640 μg/L. Based on the combined results of the adult fecundity trial, the egg hatchability trial and the juvenile growth trial, a NOEC for Marisa cornuarietis of 25 μg/L was determined. This value is based on the highest concentration at which no statistically significant impairments in juvenile growth were observed (LOEC=640 μg/L). Warbritton et al. (2007b) and Forbes et al. (2008) conducted an additional 84 day fecundity trial at 22oC and found no differences in any endpoint similarly measured at 25oC. The work by Warbritton, Forbes, and colleagues did not confirm any of the findings reported by Oehlmann, thus the studies by Oehlmann are not considered suitable for use in risk assessment. The long term NOEC for Marisa cornuarietis of 25 μg/L, reported by Warbritton et al. (2007a) and Forbes et al. (2008), will be used to derive aquatic PNEC for use in risk assessment.

The following information is taken into account for long-term toxicity to aquatic invertebrates for the derivation of PNEC:

Studies considered to have a Klimisch rating of 1, with analytical confirmation of dose and following appropriate internationally recognised guidance, include a Daphnia magna reproduction study with a NOEC of 3.16 mg/L (Caspers, 1998), a rotifer life-cycle study with a NOEC of 1.8 mg/L (Springborn Smithers, 2006a; also reported in Mihaich et al., 2009), a mysid shrimp life-cycle study with a NOEC of 0.170 mg/L (Lee, 2010), and a Marisa cornuarietis snail reproduction study with a NOEC of 0.025 mg/L (Warbritton et al., 2007a; Forbes et al. 2008). The NOEC of 0.025 mg/L from the 328 day Marisa cornuarietis study based on a reduction in female growth was considered the key parameter for the long-term toxicity to aquatic invertebrates, although there are a number of chronic NOECs for invertebrates that are used in an SSD determination.

Value used for CSA:

EC10/LC10 or NOEC for freshwater invertebrates: 0.025 mg/L

EC10/LC10 or NOEC for marine water invertebrates: 0.17 mg/L



Toxicity to aquatic algae and cyanobacteria

There are two key algae studies for Bisphenol A, one freshwater and one marine.

The freshwater algae, Pseudokirchneriella subcapitata (formerly known as Selenastrum capricornutum), was exposed for 96-hr to nominal concentrations of Bisphenol A of 0.78, 1.30, 2.16, 3.6, 6.0, and 10.0 mg/L and a control (Alexander et al., 1985; also reported in Alexander et al., 1988). These corresponded to mean measured concentrations of 0.69, 1.17, 1.99, 3.41, 5,86, and 9.74 mg/L and non-detect, respectively. The 96-hr EC50 for cell count was 2.73 mg/L and the 96-hr EC50 for total cell volume was 3.10 mg/L. The EC10 for the most sensitive endpoint, cell count inhibition, was calculated by probit analysis to be 1.36 mg/L.

The marine algae, Skeletonema costatum, was exposed for 96-hr to nominal concentrations of Bisphenol A of 0, 0.72, 1.1, 1.9, 3.2, 5.4, 9.0, and 15 mg/L (Springborn Bionomics, 1985c). These corresponded to analysed test concentrations of 0, 0.66, 1.1, 1.6, 2.8, 5.2, 8.8, and 15 mg/L, respectively. At the end of the exposure, chlorophyll A concentration and cell counts were significantly reduced at all Bisphenol A test concentrations. Based on the analysed concentration and the observed effects, the EC50s for cell count inhibition and chlorophyll A inhibition,calculated by probit analysis, were 1.1 mg/L and 1.4 mg/L, respectively. The 96-hr EC10 for the most sensitive endpoint, cell count, was calculated by probit analysis to be 0.4 mg/L.

The following information is taken into account for effects on algae / cyanobacteria for the derivation of PNEC:

Alexander et al. (1985) reported the 96-hr EC50 value for Pseudokirchneriella subcapitata (formerly known as Selenastrum capricornutum) of 2.73 mg/L, based on cell count. Using probit analysis, an EC10 of 1.36 mg/L was derived. Springborn Bionomics (1985c) reported the results of a 96-hr study with the marine algae, Skeletonema costatum. The EC50 reported in the study using non-linear interpolation was 1.0 mg/L based on cell count. These original data have been analysed by the UK rapporteur in the EU Risk Assessment Update in 2008 (EC, 2008a, p. 63) using probit analysis in accordance with the OECD Guideline. The resulting EC50 for cell count is 1.1 mg/l, and that for chlorophyll A content is 1.4 mg/l. In order to calculate an EC10 value, probit analysis was used. The resulting 96-hr EC10, based on the most sensitive endpoint, cell count, was determined to be 0.4 mg/L. Both EC10s of 1.36 mg/L for freshwater algae and 0.4 mg/l for marine algae are acceptable to be used in PNEC derivation.

Value used for CSA:

EC50/LC50 for freshwater algae: 2.73 mg/L

EC50/LC50 for marine water algae: 1.1 mg/L

EC10/LC10 or NOEC for freshwater algae: 1.36 mg/L

EC10/LC10 or NOEC for marine water algae: 0.4 mg/L

Toxicity to aquatic plants other than algae

The freshwater vascular plant, Lemna gibba, strain G3, was tested in a 7-day chronic study with Bisphenol A (Putt, 2003; Mihaich et al., 2009). Nominal concentrations of 0, 1.3, 3.2, 8.0, 20, and 50 mg/L corresponded to measured concentrations of 0, 1.0, 3.0, 7.8, 20, and 54 mg/L, respectively. The 7-day NOEC and LOEC for frond density were determined to be 7.8 mg/L and 20 mg/L, respectively. The 7-day EC50 value for frond density was calculated to be 20 mg/L (95% CI 19-21 mg/L). The 7-day NOEC and LOEC for frond growth rate were determined to be 7.8 mg/L and 20 mg/L, respectively. The 7-day EC50 for frond growth rate was 32 mg/L (95% CI of 31 to 32 mg/L). The 7-day NOEC and LOEC for frond biomass were determined to be 7.8 mg/L and 20 mg/L, respectively. The 7-day EC50 value for frond biomass is 22 mg/L (95% CI 18 to 26 mg/L).

The following information is taken into account for effects on aquatic plants other than algae for the derivation of PNEC:

A 7-day EC50 of 20 mg/L and a 7-day NOEC of 7.8 mg/L were determined for Bisphenol A in a chronic study with the aquatic plant Lemna gibba (Putt, 2003; Mihaich et al., 2009).



Value used for CSA:

EC50/LC50 for freshwater plants: 20 mg/L

EC10/LC10 or NOEC for freshwater plants: 7.8 mg/L

Toxicity to microorganisms

Studies with micro-organisms can be used to determine the PNEC for waste water treatment plants (WWTP). Two weight of evidence studies are available to provide information for the toxicity to microorganisms endpoint. Stone and Watkinson (1983) studied the biodegradability of Bisphenol A in a number of tests. One of them evaluated the inhibition of growth of the bacteria, Pseudomonas fluorescens. In this microbial inhibition test, the IC50 for the inhibition of growth of Pseudomonas by Bisphenol A was determined to be 54.5 mg/L. However, P. fluorescens uses glucose as a substrate which, according to the technical guidance document (TGD) is not appropriate for determining a PNEC for WWTPs.

A study with Pseudomonas putida on an agar-solidified medium by Fabig (1988) reported an EC10 of > 320 mg/L, the highest dose tested in the study. While these studies are not the same as the activated sludge respiration test that would typically be performed to satisfy this data need, Bisphenol A is readily biodegradable and does not show high toxicity to microbial populations. The Weight of Evidence suggests that no additional studies are necessary for the assessment of microbial populations exposed to Bisphenol A. The EC10 of > 320 mg/L, can be considered the PNEC for WWTP.

The following information is taken into account for effects on microorganisms for the derivation of PNEC:

Value used for CSA:

EC10/LC10 or NOEC for aquatic microorganisms: 320 mg/L

Toxicity to amphibians

African clawed frog (Xenopus laevis) larvae at a developmental stage between 43 and 45, were exposed to nominal Bisphenol A concentrations ranging from 1 to 500 µg/L, in flow-through conditions until they reached complete metamorphosis (Pickford et al., 2000, 2001, 2003). During the study, nominal test concentrations of 0, 1.0, 2.3, 10, 23, 100, and 500 µg/L corresponded to mean measured test concentrations of <0.5, 0.83, 2.1, 9.5, 23.8, 100, and 497 µg/L, respectively. Results are reported based on nominal concentrations. After 90 days, larvae were sacrificed, and total length, snout-vent length and wet weight were measured. Sexual differentiation in larvae at developmental stages 58-66 was evaluated by inspection of gross gonadal morphology. The NOEC for survival, sex ratio, length and weight was determined to be 500 µg/L. LOEC for survival, sex ratio, length and weight was ≥ 500 µg/L. There was no significant difference in mean percentage survival between Bisphenol A, positive control (17-B estradiol), and dilution water control but there was a significantly lower survival rate at 500 µg/L than 100 µg/L of Bisphenol A. Compared to the dilution control, there was no significant difference between growth, development, and time to metamorphosis in the Bisphenol A treatment groups while there was a significant difference in the positive control with regards to mean time to metamorphosis. There was also a significant increase in total length, snout-length, and wet weight in males between positive control and the dilution control but no noticeable difference in Bisphenol A froglets.

The Pickford et al. (2000, also reported in Pickford et al. 2001, 2003) study was run under full GLP, was thoroughly reported, and was specifically designed with sufficient replication and appropriate statistics to determine a no-effect level for a range of effects. Earlier studies by Kloas et al. (1999, The Science of the Total Environment 225: 59-68) and a follow-up study reported by Levy et al. (2004, Env. Research, 94:102-111) resulted in inconclusive findings due to experimental design flaws or assessing endpoints not appropriate for risk assessment. For example, the Kloas et al. (1999) and Levy et al. (2004) studies had insufficient evidence of a concentration-response, utilised insufficient replication in the test design, used inappropriate statistics, reported alteration of sex ratio but no histopathological evidence of gonadal abnormalities, and no or inadequate analytical confirmation of test concentrations. Therefore, the studies on Xenopus laevis by Kloas, Levy and colleagues are not suitable for use in risk assessment. The study by Pickford et al. (2000, also reported in Pickford et al., 2003) will be used in deriving aquatic PNEC for use in risk assessment.



The following information is taken into account for any hazard / risk assessment:

In a 90-day study, the NOEC for Xenopus laevis based on larval survival, adult growth and sex ratio was determined to be 500 µg/L, the highest dose of Bisphenol A tested (Pickford et al., 2000, 2001, 2003). No impact on sex ratio was observed in the treatment or control test concentrations, although there was a significant feminisation observed in the positive control group.

PNEC Assessment factor

Remarks/Justification

PNEC aqua (freshwater): 0.018 mg/L

1 Extrapolation method: statistical extrapolation

Studies used to calculate a statistically-derived PNECfreshwater were as discussed in Staples et al. (2008; Hum. Ecol. Risk Assess., 14(3):455-478). For Bisphenol A, there are at least 19 valid chronic freshwater studies covering 14 different species of organisms from 10 unique families. These include four species of fish, one amphibian, two crustaceans, one insect, one mollusk, a rotifer, an hydra, a sponge, a green algae, and an aquatic macrophyte. They include a 28-day juvenile growth test with rainbow trout, a multi-generation test with medaka, a multi-generation study with fathead minnow, a 30-day mortality study with guppy, a 90-day growth and sexual differentiation study with the amphibian Xenopus, a 21-day life-cycle study with Daphnia magna, a 42-day growth and reproduction study with the amphipod Hyalella, a life-cycle study with rotifers, a 90-day hatching and juvenile growth study with the snail Marisa, a 181-day reproduction and growth study with Marisa, a mortality and development study with hydra, a 9-day growth study with the sponge Heteromyenia, a 4-day chronic study with the green algae Selenastrum, and a 7-day chronic study with the duckweed, Lemna. One of the statistical methods used in Staples et al. (2008) in developing PNEC values for Bisphenol A was the Hazard Concentration (HC5) approach developed by The Netherlands National Institute of Public Health and the Environment (RIVM).The HC5 approach, representing the lower-bound 5th percentile of NOEC values of the distribution of species-specific toxicity results is robust and includes statistical tests on the normality of the dataset. Given the extent and quality of the dataset available for Bisphenol A no assessment factor is needed. The PNECwater for Bisphenol A, calculated using the HC5 approach, is 18 µg/L.

PNEC aqua (marine water): 0.016 mg/L

1 Extrapolation method: statistical extrapolation

Studies used to calculate a statistically-derived PNECfreshwater were as discussed in Staples et al. (2008; Hum. Ecol. Risk Assess., 14(3):455-478) with the addition of two chronic marine studies. In developing the PNECmarine, ECHA guidance allows the use of a combination of freshwater and marine species. For Bisphenol A, there are at least 21 valid chronic aquatic studies covering 16 species of organisms from 10 unique families. There are five species of fish, one amphibian, three crustaceans, one insect, one mollusk, a rotifer, a hydra, a sponge, a green algae, and an aquatic macrophyte. They include a 28-day rainbow trout juvenile growth test, multi-generation tests with medaka and fathead minnow, a 116-day sheepshead minnow life-cycle test, a 30-day mortality study with guppy, a 90-day growth and sexual differentiation study with the amphibian Xenopus, a 21-day Daphnia magna life-cycle study, a 28-day mysid shrimp life-cycle test, a 42-d growth and reproduction study with the amphipod Hyalella, a life-cycle study with rotifers, a 90-day hatching and juvenile growth study with the snail Marisa, a 181-day reproduction and growth study with Marisa, a hydra mortality and development study, a 9-day growth study with the sponge Heteromyenia, a 4-day chronic study with the green algae Selenastrum, and a 7-day chronic study with the duckweed, Lemna. One of the



statistical methods used in Staples et al. (2008) in developing PNEC values for Bisphenol A was the Hazard Concentration (HC5) approach developed by The Netherlands National Institute of Public Health and the Environment (RIVM).The HC5 approach, representing the lower-bound 5th percentile of NOEC values of the distribution of species-specific toxicity results is robust and includes statistical tests on the normality of the dataset. Given the extent and quality of the dataset for Bisphenol A no assessment factor is needed. The PNECmarine for Bisphenol A, calculated using the HC5 approach, is 16 µg/L.

PNEC aqua (intermittent releases): 0.01 mg/L

100 Extrapolation method: assessment factor

For short-term, intermittent release of compounds into a waterbody, a PNECwater, intermittent is calculated. The PNECwater, intermittent is normally derived by application of an assessment factor of 100 to the lowest L(E)C50 of at least three short-term tests from three trophic levels. For Bisphenol A, there are many acute studies with a variety of species. Considering the key studies, there are 4 acute fish studies, 2 acute invertebrate studies, and 2 algae studies which can be considered for use in the derivation of the PNECwater, intermittent. Of the four fish studies, the lowest LC50 of 4.6 mg/L is with the fathead minnow. The lowest invertebrate acute value is an EC50 of 1.1 mg/L for the mysid shrimp. The lowest algae LC50 is for the marine algae, Skeletonema, of 1.1 mg/L. Using an assessment factor of 100, the PNECwater, intermittent is calculated to be 10 µg/L.

PNEC STP: 320 mg/L


No inhibition of cell growth was observed in the test with P. putida in a liquid broth medium so the EC10 of > 320 mg/L, the highest dose tested in the study (Fabig, 1988), can be considered the PNEC for WWTPs.

3.2 Sediment toxicity

Six fully valid freshwater sediment studies exist with exposure to Bisphenol A, although in two of them the Bisphenol A concentration was measured in the water, not the sediment.

A 10-day Corophium volutator mortality and sublethal effect study was performed with Bisphenol A-dosed sediment according to OSPAR (1995) guidance for the performance of sediment bioassays (Whale et al., 1999). The nominal concentrations were 0, 1, 3.2, 10, 32, 100, 320, 1000, and 3200 mg/kg dry weight. The mean measured concentrations were approximately 86% of nominal so the test concentrations were reported as the nominal dose range. The 10-day LC50 values calculated for the Bisphenol A/acetone and directly spiked tests based on bulk sediment concentrations were 46 and 40 mg/kg, respectively. The corresponding lowest 10-day EC50 values based on mortality and failure to burrow in the Bisphenol A/acetone and directly spiked tests were 31 and 36 mg/kg, respectively. When the endpoints of the toxicity tests are based on predicted interstitial water concentrations the 10-day LC50 and 10-day EC50 values are between 1.4 and 1.5 and 1.1 and 1.3 mg/L, respectively. As such, these values are similar to other acute and longer term toxicity values for other aquatic invertebrates exposed to Bisphenol A.

A 28-day chronic freshwater oligochaete study was performed exposing Lumbriculus variegatus to nominal Bisphenol A concentrations in artificial sediment of2.0, 5.1, 13, 32 and 80 mg a. i. /kg under static renewal conditions (Picard, 2010a). The results of the analysis of sediment exposure concentrations established mean measured concentrations of 1.6, 3.9, 9.9, 22 and 57 mg a. i. /kg. At test termination (day 28), the mean number of surviving oligochaetes observed among the oligochaetes exposed to the 1.6, 3.9, 9.9, 22 and 57 mg a. i. /kg treatment levels was 20, 28, 25, 24 and 15, respectively. Statistical analysis (Bonferroni’s t-Test) determined a significant reduction in the mean number of surviving oligochaetes in the 1.6 and 57 mg a. i. /kg treatment levels compared to the control data (27 oligochaetes). However, due to the lack of dose-response in the higher treatment levels, the effect determined in the 1.6 mg a. i. /kg treatment level was not considered to be toxicant-related. Mean biomass per replicate in the 1.6, 3.9, 11, 22 and 57 mg a. i. /kg treatment levels was 37, 43, 37, 34



and 24 mg, respectively. Statistical analysis (Bonferroni’s t-Test) determined a significant difference in the 57 mg a. i. /kg treatment level compared to the control data (36 mg). Based on mean measured concentrations of Bisphenol A in the sediment, the NOEC for this exposure was determined to be 22 mg a. i. /kg. The Lowest-Observed-Effect Concentration (LOEC) for this exposure was determined to be 57 mg a. i. /kg. Since no concentration tested resulted in ≥ 50% reduction in reproduction or biomass, the EC50 value was empirically estimated to be > 57 mg a. i. /kg, the highest mean measured concentration tested.

A 28-day chronic estuarine amphipod study was performed exposing Leptocheirus plumulosus to nominal Bisphenol A concentrations in natural marine sediment of 2.6, 6.4, 16, 40 and 100 mg a. i. /kg under static renewal conditions (Picard, 2010b). The results of the analysis of sediment exposure concentrations established mean measured concentrations ranging from 75 to 81% of nominal concentrations and defined the treatment levels tested as 2.0, 5.0, 12, 32 and 78 mg a. i. /kg. At test termination (test day 28), survival observed among amphipods in the 2.0, 5.0, 12, 32 and 78 mga. i. /kg treatment levels was 82, 88, 87, 81 and 2%, respectively. Statistical analysis (Dunnett’s Test) demonstrated a significant difference in survival among amphipods exposed to the 78 mg a. i. /kg treatment level compared to the survival of the control (88%). Subsequently, data from the 78 mg a. i. /kg treatment level was excluded from further statistical analysis for determination of NOEC and LOEC due to the significant effect observed for survival. Based on mean measured sediment concentrations, the Lowest-Observed-Effect Concentration (LOEC) and No-Observed-Effect Concentration (NOEC) for amphipod survival were determined to be 78 and 32 mg a. i. /kg, respectively. The 28-day LC50 value for survival was determined to be 63 mg a. i. /kg, with 95% confidence intervals of 55 to 70 mg a. i. /kg. Growth observed among amphipods in the 2.0, 5.0, 12, and 32 a. i. /kg treatment levels averaged 1.40, 1.81, 1.40 and 1.22 mg dry weight per amphipod, respectively. Statistical analysis (Dunnett’s Test) determined no significant difference in growth among amphipods exposed to treatment levels ≤ 32 mg a. i. /kg compared to the control (1.76 mg dry weight per amphipod). The LOEC and NOEC for amphipod growth were determined to be >32 and 32 mg a. i. /kg, respectively. The 28-day EC50 value for growth was determined to be 60 mg a. i. /kg, with 95% confidence intervals of 31 to 75 mg a. i. /kg. Reproduction among amphipods exposed to the 2.0, 5.0, 12 and 32 mg a. i. /kg treatment levels averaged 7, 9, 7 and 5 offspring per amphipod, respectively. Statistical analysis (Dunnett’s Test) determined no significant difference in reproduction among amphipods exposed to treatment levels ≤ 32 mg a. i. /kg compared to the control (6 offspring per amphipod). The LOEC and NOEC for amphipod reproduction were determined to be >32 and 32 mg a. i. /kg, respectively. The 28-day EC50 value for reproduction was determined to be 38 mg a. i. /kg, with 95% confidence intervals of 11 to 52 mg a. i. /kg.

A 28-day chronic freshwater midge study was performed exposing Chironomus riparius to nominal Bisphenol A concentrations in artificial sediment of 50, 100, 200, 400 and 800 mg a. i. /kg dry weight under static renewal conditions (Picard, 2010c). The results of the analysis of sediment exposure concentrations established mean measured concentrations of 29, 54, 110, 210 and 490 mg a. i. /kg dry weight. At test termination (day 28), the mean percent emergence observed among the midges exposed to mean measured sediment concentrations of 29, 54, 110, 210 and 490 mg a. i. /kg was 83, 98, 68, 11 and 0%, respectively. Statistical analysis (Dunnett’s Test) determined a significant difference in mean percent emergence in the 110, 210 and 490 mg a. i. /kg treatment levels compared to the control organisms (93%). Mean development rate of male midges in the 29, 54, 110 and 210 mg a. i. /kg treatment levels was 0.0752, 0.0674, 0.0774 and 0.0789, respectively. Statistical analysis (Bonferroni’s t-Test) determined no significant difference in midge mean development rate in any of the treatment levels tested compared to the mean development rate of the control midges (0.0663). Based on the mean measured concentrations of applied test substance and midge emergence, the NOEC was established to be 54 mg a. i. /kg and LOEC was established to be 110 mg a. i. /kg. The 28-day EC50 based on midge emergence was determined to be 150 mg a. i. /kg dry weight, with 95% confidence intervals of 140 to 160 mg a. i. /kg dry weight. Based on the mean measured concentrations of applied test substance and development rate, the NOEC was established to be 210 mg a. i. /kg. Since no test concentration resulted in a significant reduction of development rate compared to the control, the LOEC and the 28-day EC50 were both empirically estimated to be > 210 mg a. i. /kg.

The two studies performed with measured water concentrations instead of quantified sediment concentrations are with a freshwater amphipod and a chironomid. A 42-day chronic amphipod study was performed exposing Hyalella azteca to nominal Bisphenol A concentrations in the water of 0.19, 0.38, 0.75, 1.5, and 3.0 mg/L under flow-through conditions (Springborn Smithers, 2006b, Mihaich et al., 2009). The results of the analysis of exposure solutions established mean measured time-weighted average concentrations ranging from 58-73% of nominal, defined as 0.12, 0.22, 0.49, 1.1 and 2.2 mg/L, respectively. At 42 days, the mean percent survival was 0% in the 2.2 mg/L treatment. Survival of 75, 83, 78, 72, and 77% was observed among amphipods exposed to the 0.12, 0.22, 0.49, 1.1 mg/L and controls, respectively. Survival was only statistically reduced in the 2.2 mg/L dose group. Amphipod reproduction was 12, 9.4, 13, 11 and 8.5 offspring per female in the controls, 0.12, 0.22, 0.49 and 1.1 mg/L dose groups, respectively. No significant difference was observed with respect to growth.



Based on the results of this study, the 42-day LC50 value was determined by moving average analysis to be 0.78 mg/L, with 95% confidence intervals of 0.68 to 0.90 mg/L. Based on mean measured concentrations and amphipod reproduction, the LOEC was determined to be 1.1 mg/L. The NOEC based on cumulative number of offspring per female was determined to be 0.49 mg/L.

A 96-hour midge (Chironomus tentans) acute study was performed exposing 10-day old midge to nominal Bisphenol A test concentrations in the water of 1.3, 2.2, 3.6, 6.0 and 10 mg/L under flow-through conditions (Springborn Smithers, 2005, Mihaich et al., 2009). Mean measured concentrations ranged from 84 to 110% of nominal and defined the treatment levels as 1.4, 2.1, 3.0, 5.5, and 9.4 mg/L. Following 96 hours of exposure, 100% mortality was observed among midges exposed to the 9.4 mg/L treatment level. Mortality of 10, 45, 65 and 75% was observed among midges exposed to the 1.4, 2.1, 3.0 and 5.5 mg/L treatment levels, respectively. Mortality of 0 and 5% was observed among midges exposed to the pH buffer control and the dilution water control. Based on the results of this study, the 96-hour LC50 value was determined by probit analysis to be 2.7 mg/L with 95% confidence intervals of 2.1 to 3.2 mg/L. The NOEC, based on survival, was determined to be 1.4 mg/L.

The following information is taken into account for sediment toxicity for the derivation of PNEC:

In a 10-day Corophium volutator mortality and sublethal effect study with Bisphenol A-dosed sediment, the 10-day EC50 values based on adverse effects in acetone or directly spiked sediments were 31 and 36 mg/kg dry weight (dw), respectively. 10-day LC50 and 10-day EC50 values based on predicted interstitial water concentrations are between 1.4 and 1.5 and 1.1 and 1.3 mg/L, respectively (Whale et al, 1999). In a 28-day Lumbriculus variegatus study, the No-Observed-Effect Concentration (NOEC) for both survival and biomass was 22 mg a. i. /kg dw. The Lowest-Observed-Effect Concentration (LOEC) for both endpoints was 57 mg a. i. /kg dw. The EC50 value was empirically estimated to be > 57 mg a. i. /kg dw, the highest mean measured concentration tested (Picard, 2010a). In a 28-day chronic Leptocheirus plumulosus study, the LOEC and NOEC for amphipod reproduction were >32 and 32 mg a. i. /kg dw, respectively. The 28-day EC50 value for reproduction was 38 mg a. i. /kg dw, with 95% confidence intervals (CI) of 11 to 52 mg a. i. /kg dw (Picard, 2010b). A 28-day chronic freshwater midge study exposed Chironomus riparius to measured Bisphenol A concentrations in artificial sediment 29, 54, 110, 210 and 490 mg a. i. /kg dw. The NOEC was 54 mg a. i. /kg dw and LOEC was 110 mg a. i. /kg dw. The 28 day EC50 based on midge emergence was 150 mg a. i. /kg dw, with 95% CI of 140 to 160 mg a. i. /kg dw. The NOEC for emergence was 210 mg a. i. /kg dw. The LOEC and the 28-day EC50 were both empirically estimated to be > 210 mg a. i. /kg dw (Picard, 2010c). Two studies exposing organisms in water instead of sediment were conducted. In a 42-day chronic Hyalella azteca study, the 42-day LC50 value was 0.78 mg/L and the LOEC was 1.1 mg/L. The NOEC was 0.49 mg/L (Springborn Smithers, 2006b). In a 96-hour midge (Chironomus tentans) acute study, the 96-hour LC50 value was 2.7 mg/L while the NOEC was 1.4 mg/L (Springborn Smithers, 2005).

Value used for CSA:

EC50/LC50 for freshwater sediment: 150 mg/kg sediment dw

EC50/LC50 for marine water sediment: 31 mg/kg sediment dw

EC10/LC10 or NOEC for freshwater sediment: 22 mg/kg sediment dw

EC10/LC10 or NOEC for marine water sediment: 32 mg/kg sediment dw



PNEC sediment (freshwater): 2.2 mg/kg sediment dw


The PNEC for the freshwater sediment compartment is based on the results of the most sensitive chronic sediment study with an assessment factor of 10. There are 3 fully valid sediment studies, two with freshwater organisms and one with a marine oligochaete. In a 28-day Lumbriculus variegatus study, the No-Observed-Effect Concentration (NOEC) for both survival and biomass was determined to be 22 mg/kg dw. In a 28-day chronic freshwater midge study with Chironomus



riparius the lowest NOEC was for emergence and was established to be 210 mg/kg dw. In a 28-day chronic Leptocheirus plumulosus study, reproduction was determined to be the most sensitive endpoint with a NOEC of 32 mg a.i./kg dw. With three valid studies an assessment factor of 10 is applied to the NOEC from the Lumbriculus study of 22 mg/kg dw, resulting in a PNEC freshwater sediment of 2.2 mg/kg dw.

PNEC sediment (marine water): 0.44 mg/kg sediment dw


The PNEC for the marine sediment compartment is based on the results of the most sensitive chronic sediment study with an assessment factor of 50 to account for the greater diversity of marine species per ECHA guidance. There are 3 fully valid sediment studies, two with freshwater organisms and one with a marine oligochaete. In a 28-day Lumbriculus variegatus study, the No-Observed-Effect Concentration (NOEC) for both survival and biomass was determined to be 22 mg/kg dw. In a 28-day chronic freshwater midge study with Chironomus riparius the lowest NOEC was for emergence and was established to be 210 mg/kg dw. In a 28-day chronic Leptocheirus plumulosus study, reproduction was determined to be the most sensitive endpoint with a NOEC of 32 mg a.i./kg dw. With three valid studies an assessment factor of 50 is applied to the NOEC from the Lumbriculus study of 22 mg/kg dw, resulting in a PNEC marine water sediment of 0.44 mg/kg dw.

3.3 Terrestrial toxicity

Toxicity to soil macroorganisms except arthopods

In addition to performing a chronic study with earthworms, Johnson et al. (2005) performed a short-term, acute study. However, the acute study is considered unreliable for a number of reasons, including that the control survival did not meet the validity criteria and there were no effects in the chronic study at doses that caused mortality in the acute study. The inconsistency in results prompted a new, fully valid chronic study with the potworm Enchytraeus crypticus (Moser and Egeler, 2007a). Final concentrations of Bisphenol A in the test systems, consisting of artificial soil, were 1.0, 9.5, 17.1, 30.9, 55.6, and 100 mg test item/kg soil dw (dry weight). No effects on mortality or number of juveniles were observed at the highest dose tested of 100 mg/kg dw. This is in agreement with the findings from the earthworm chronic study (Johnson et al., 2005) where the NOEC of 100 mg/kg was also established.

The following information is taken into account for effects on soil macro-organisms except arthropods for the derivation of PNEC:

No effects on mortality or number of juveniles were observed in a 28-day potworm (Enchytraeus crypticus) reproduction study (Moser and Egeler, 2007a; Staples et al., 2010). A NOEC of 100 mg/kg dw was determined.

Value used for CSA:

Long-term EC10/LC10 or NOEC for soil macroorganisms: 100 mg/kg soil dw

Toxicity to terrestrial arthopods

No effect on survival was observed during a chronic 28-day test with the collembolan, Folsomia candida, exposed to Bisphenol A in artificial soil (Moser and Egeler, 2007b; Staples et al., 2010). The number of juveniles was reduced at the highest dose tested, 1000 mg/kg dw, compared to controls so the NOEC established in the study was 500 mg/kg dw.

The following information is taken into account for effects on soil arthropods for the derivation of PNEC:

No effects on survival were observed in a chronic reproduction study with the collembolan Folsomia candida



(Moser and Egeler, 2007b; Staples et al., 2010). A NOEC of 500 mg/kg soil was determined based on number of juveniles.

Value used for CSA:

Long-term EC10/LC10 or NOEC for terrestrial arthropods: 500 mg/kg soil dw

Toxicity to terrestrial plants

An OECD Guideline 208 study assessing emergence and growth in 6 different species of terrestrial plant exposed to Bisphenol A was performed (Hoberg, 2007; Staples et al., 2010). The plant species tested were three monocotyledons, corn (Zea mays), oats (Avena sativa) and wheat (Triticum aestivum), and three dicotyledons, cabbage (Brassica oleracea), soybean (Glycine max) and tomato (Lycopersicon esculentum). For the endpoint of percent emergence, there were no effects at the highest dose tested with oat, soybean, corn, and wheat. Tomato and cabbage were equally sensitive for the endpoint percent emergence with an LOEC of 320 mg/kg dw and a NOEC of 130 mg/kg dw. There were effects on dry shoot weight endpoint, an assessment of growth, in all species tested. The most sensitive species was tomato with a LOEC and a NOEC of 50 and 20 mg/kg dw, respectively.

The following information is taken into account for toxicity on terrestrial plants for the derivation of PNEC:

An OECD 208 study in 6 plant species was performed with Bisphenol A (Hoberg, 2007; Staples et al., 2010). The most sensitive plant was tomato with a LOEC and NOEC of 50 mg/kg dw soil and 20 mg/kg dw soil, respectively.

Value used for CSA:

Long-term EC10/LC10 or NOEC for terrestrial plants: 20 mg/kg soil dw

Toxicity to soil microorganisms

Bisphenol A is readily biodegradable. There are available soil studies, including ones with terrestrial plants, soil macroorganisms and terrestrial arthropods that adequately characterise the toxicity of Bisphenol A in the soil environment. Since Bisphenol A is only intermittently applied to soil as a possible component of sewage sludge, exposure to Bisphenol A in the soil environment would be limited and transient. Furthermore, a test towards terrestrial microorganisms should be proposed by the registrant if the chemical safety assessment indicates the need to further investigate the effects on terrestrial organisms. For the risk characterisation of the terrestrial compartment, a PNEC soil derived on initial concentrations of Bisphenol A applied to soils was used. As the risk characterisation yields a PEC/PNEC ratio smaller than 1, a test towards terrestrial microorganisms is not necessary as the risk towards the terrestrial compartment is sufficiently described based on the already available data.

Toxicity to birds

While no guideline avian reproduction studies are currently available, a number of egg injection studies exist which provide a weight of evidence, worst-case type exposure to chicks during development in ovo. The most relevant of the studies is by Halldin et al., 2001 where three different studies were performed with quail embryos or adults after in ovo exposure to 67 µg/egg (uptake study), 67 µg/egg and 200 µg/g egg (male and female behavioural and reproduction study), or 105 µg/bird (female distribution study). The most relevant studies for the assessment of reproductive effects are the male and female reproductive variable assessment studies. For the male reproductive variable assessment, 7-week old males were individually placed in metal cages. Sexual behaviour tests were performed in the 8th week after hatching. Neck grab and mount attempt were assessed upon placement with a female. Blood for testosterone analysis was collected and testis weight and gonadosomatic index were assessed. Female reproductive variables assessed were egg laying and oviduct pathology. No significant oestrogen-like effects were observed in males or females treated with Bisphenol A. There was a tendency for females exposed to the 200 µg/g egg to retain the right oviduct. Bisphenol A was readily excreted by the laying female as well as the growing embryo. In view of the data on distribution, maternal transfer, embryonic uptake and oestrogenic potency, it was concluded that the risk for adverse reproductive toxicity in avian wildlife resulting from embryonic exposure is probably low.

Sashihara et al., 2003 reported that hatchability was not significantly affected by Bisphenol A exposure to embryos after egg injection. There was no indication that Bisphenol A exposure significantly damaged chick



embryos as determined by the stage of development at death. The phenotypic ratio of male chicks in the 10 and 100 ng/µL treatment groups was slightly higher than the 50% ratio found in controls. When compared, the genotype and phenotype ratios completely matched. The results of this study indicate that low doses of Bisphenol A do not affect the hatchability or embryonic development of chickens.

Considered a supporting study to the two weight of evidence studies described above, Berg et al., 2001 assessed embryo development just prior to hatch in eggs that were treated by injection on day 3 of incubation. There was a statistically significant increase in mortality in chickens at both doses but no impact on quail. In quail embryos treated with 200 µg/g Bisphenol A, malformations of the mullerian ducts occurred in 6 of the 14 females. No impact was seen in males or in either sex in chickens. In quail embryos, Bisphenol A did not cause an increased ovotestis frequency compared to controls. In chicken embryos, the ovotestis frequency was 55% following exposure to 200 µg/g Bisphenol A, but none were found in the 67 µg/g dose group. The risk for reproductive impacts in avian wildlife, as predicted by this study, are probably low as the doses required for effects in this study were fairly high.

Although no guideline avian reproduction studies are currently available, there are a number of egg injections studies that exposed embryos of chicken and quail in ovo to various concentrations of Bisphenol A. These studies provide a weight of evidence, worst-case type of exposure which mimics maternal transfer of material to embryos. Doses are difficult to compare to a dietary intake but concentrations of Bisphenol A employed in the studies were described as low in one and high in another, spanning a range of possible exposure concentrations. Endpoints assessed included egg laying, embryonic development, phenotypic and genotypic chick sex ratios, sexual behaviour, and hatchability. Conclusions from the two weight of evidence and one supporting study were in agreement that the risk for adverse reproductive toxicity in avian wildlife resulting from embryonic exposure is probably low. Given the low potential for dietary exposure of avian populations to Bisphenol A, a full avian reproduction study with the sacrifice of many animals, is not justified. This is confirmed by the 2008 Addendum to the EU Risk Assessment that concluded that it is not considered appropriate to request a multi-generational study with birds because Bisphenol A is readily biodegradable and has a low bioaccumulation potential, the existing studies addressed several relevant end points, and the PEC/PNEC ratios are all significantly below 1.



PNEC soil: 3.7 mg/kg soil dw


The PNEC for the terrestrial compartment is based on the most sensitive NOEC from studies with terrestrial plant, potworms, and springtail. The lowest organic matter-normalised (from 1.82% organic matter to the standard 3.4%) NOEC among all the tests was for dry shoot weight of tomatoes of 37 mg/kg-dry weight. Dividing by an assessment factor of 10, a predicted no-effect concentration in soil (PNECsoil) of 3.7 mg/kg-dry weight was calculated.

PNEC oral: 13.8 mg/kg food

30 Depending on the study used, the PNECoral values calculated from a three-generation rat study and a two-generation mouse study are 33 mg/kg bw/day and 13.8 mg/kg bw/day, respectively. The two-generation study on mice, as well as the three-generation study with rats, identified a NOAEL of 50 mg/kg bw/day. The conversion factor for mice is 8.3, giving a NOEC of 415 mg/kg; applying an assessment factor of 30 gives the lower PNECoral of 13.8 mg/kg food. Although no guideline avian reproduction studies are available, there are a number of egg injections studies which provide a weight of evidence, worst-case type of exposure which mimics maternal transfer of material to embryos. Endpoints assessed included egg laying, embryonic development, phenotypic and genotypic chick sex ratios, sexual behavior, and hatchability. Conclusions from the two weight of evidence and one supporting study were in agreement that the risk for adverse reproductive toxicity in avian wildlife resulting from embryonic exposure is probably low. Given the low potential for dietary exposure of avian populations to Bisphenol A, intermittent release to the terrestrial environment, and the ready biodegradability of the compound, a full avian reproduction study with the sacrifice of many animals, is not justified. The relevant PNECoral based on robust chronic mammalian studies is 13.8 mg/kg food.



3.4 Environmental classification justification GHS classification for acute aquatic toxicity is based on acute aquatic toxicity effect levels and subdivides into categories 1 to 3. Chronic aquatic toxicity is based on chronic aquatic toxicity effect levels as well as on rapid degradability and is subdivided into categories 1 to 4. The GHS classification system was implemented by the European Union by means of the CLP regulation (Classification, Labelling and Packaging) EC No. 1272/2008 as of 16th December 2008 with the subsequent 2nd Adaptation to Technical Progress (ATP) from 10th March 2011. While establishing all four categories for chronic aquatic toxicity, the CLP regulation implemented acute aquatic toxicity category 1 only and did not adopt acute aquatic toxicity categories 2 and 3. a) Classification for acute aquatic toxicity: The most critical acute aquatic toxicity was reported by Springborn Bionomics (1985c) in a 96-hr study with the marine algae, Skeletonema costatum. The EC50 reported in the study using non-linear interpolation was 1.0 mg/L based on cell count. These original data have been analysed by the UK rapporteur in the EU Risk Assessment Update in 2008 (EC, 2008a, p. 63) using probit analysis in accordance with the OECD Guideline. The resulting EC50 for cell count was 1.1 mg/l. According to GHS criteria Bisphenol A would qualify for classification under acute aquatic toxicity category 2 (LC50/EC50 > 1 to ≤ 10 mg/l). However, acute aquatic toxicity category 2 was not implemented in the CLP regulation in the European Union and, therefore, Bisphenol A is not to be classified for hazards to the aquatic environment (acute/short-term) under CLP. b) Classification for chronic aquatic toxicity: Prior to implementation of the 2nd ATP the CLP criteria did not require a classification for chronic aquatic toxicity for Bisphenol A as Bisphenol A is readily biodegradable and not bioaccumulative. The 2nd ATP introduced the principle that classification for chronic aquatic toxicity shall be based on chronic studies if those are available and that ready biodegradability and non-bioaccumulation do not prevent a potential classification. Consequently, the classification for chronic aquatic toxicity of Bisphenol A was re-assessed based on the new criteria established by the 2nd ATP. Caunter et al. (2000) reported on a multigeneration study exposing the fathead minnow (Pimephales promelas) for 444 days, through two generations, to a dilution water control and nominal Bisphenol A concentrations of 1.0, 16, 160, 640, and 1280 µg/l. Daily observations of mortality, behaviour and appearance were made and any abnormal effects recorded for the F0, F1, and F2 generation fish. Survival, growth, reproduction, gonadal size, vitellogenin and gonadal histology were evaluated. The NOEC established based on F2 hatchability in this multigenerational exposure was 0.016 mg/l. The F2 hatchability NOEC of 0.016 mg/l from the 2-generation fathead minnow study is the most conservative endpoint from fully valid studies and was used in the derivation of chronic classification according to the CLP regulation. According to the CLP criteria introduced by means of the 2nd ATP (rapid biodegradability and NOEC > 0.01 to ≤ 0.1 mg/l) Bisphenol A is to be classified for hazards to the aquatic environment (long-term) as category chronic 2. This is by no means the result of a re-assessment of scientific data but solely a consequence of the revised classification criteria implemented by the 2nd ATP.



4. HUMAN HEALTH HAZARD ASSESSMENT

The human health hazard assessment is based on the initial EU Risk Assessment Report 2003 on Bisphenol A and the recent EU Risk Assessment Update in 2008. For each endpoint the conclusion of the EU Risk Assessments in 2003 and 2008 are cited and, if available, additional reliable and relevant information not evaluated in one of the Risk Assessments is indicated.

Due to the large data set on Bisphenol A, criteria were established to define key studies, supporting studies and additional studies to cover all available data for the human health hazard assessment. Key studies are defined as comprehensive studies conducted according to scientifically accepted methods and performed according to or exceeding validated guidelines (e. g. OECD testing guidelines). Supporting studies are defined as comprehensive studies conducted to scientifically accepted methods and performed similar to validated guidelines with only very minor deviations.

All additional studies are included in the IUCLID database without further specification of the purpose flag in chapters 7.10 and 7.12. Since this dataset is based on the IUCLID4 database prepared for and evaluated by the EU Risk Assessment in 2003, additional studies entered into the database before the EU Risk Assessment in 2003 are transferred to IUCLID5 without further modification. Additional studies evaluated during the EU Risk Assesment Update in 2008 or published more recently are added to the database as “additional information” and a short description of the method used and the key results is given.

DNEL are calculated according to the EU Annex XV Report submitted by United Kingdom, dated 30 November 2008.

4.1 Toxicokinetics, metabolism and distribution

The 2003 EU Risk Assessment Report concluded:

"Animal data indicates that absorption of BPA from the gastrointestinal tract is rapid and extensive following oral administration, although it is not possible to reliably quantify the extent of absorption. Following dermal exposure, available data suggest limited absorption of about 10% of the applied dose. BPA is removed rapidly from the blood and the data indicate extensive first pass metabolism following absorption from the gastrointestinal tract. In view of this first pass metabolism, the bioavailability of unconjugated BPA is probably limited following oral exposure to no more than 10 to 20% of the administered dose. The major metabolic pathway in rats involves glucuronide conjugation, with approximately 10% and 20% of the administered dose recovered in urine as the glucuronide metabolite in males and females, respectively. The major route of excretion is via the faeces with the urinary route being of secondary importance. Over seven days post dosing, approximately 80% and 70% of the administered dose was eliminated in the faeces in male and female rats, respectively. The first pass metabolism and extensive and rapid elimination of BPA suggest that the potential for transfer to the foetus and bioaccumulation may be limited. There are no data on the toxicokinetics of BPA following inhalation exposure."

The 2008 updated EU Risk Assessment Report concluded:

"New information on the toxicokinetics of BPA in humans and in pregnant and non-pregnant rodents of different ages provides an important contribution to the knowledge of kinetic properties of BPA. Human studies have demonstrated that at comparable exposure levels the blood concentrations of free BPA in humans are much lower than those in rodents. In rats, mice, monkeys, and humans, the available evidence suggests that following oral administration, BPA is rapidly and extensively absorbed from the gastrointestinal tract. For the purposes of risk characterisation, absorption via the oral and inhalation routes will be assumed to be 100%; dermal absorption will be taken to be 10%. A number of studies in rats suggest that BPA metabolites and free BPA have a limited distribution to the embryo/foetal or placental compartments following oral administration. Maternal and embryo/foetal exposure to free BPA did occur, but systemic levels were found to be low due to extensive first-pass metabolism. There are differences between humans and rodents in the distribution of BPA. After oral administration, BPA is rapidly metabolised in the gut wall and the liver to BPA glucuronide. In humans, the glucuronide is released from the liver into the systemic circulation and cleared by urinary excretion. In contrast, BPA glucuronide is eliminated in bile in rodents and undergoes enterohepatic recirculation after cleavage to BPA and glucuronic acid by glucuronidase in the intestinal tract."



There is no reliable and significant new information on the toxicokinetics of Bisphenol A.

The following information is taken into account for any hazard / risk assessment:

After oral administration, Bisphenol A is rapidly metabolised by intestinal tissue and the liver to Bisphenol A glucuronide. In humans, the glucuronide is released from the liver into the systemic circulation and cleared rapidly by urinary excretion. In contrast, Bisphenol A glucuronide is primarily eliminated in bile in rodents with some partial urinary excretion. The Bisphenol A glucuronide excreted via the bile undergoes enterohepatic recirculation after cleavage to Bisphenol A and glucuronic acid by glucuronidase in the intestinal tract.

4.2 Acute toxicity


"No useful information is available on the effects of single exposure to Bisphenol A in humans. Oral LD50 values beyond 2,000 mg/kg are indicated in the rat and mouse, and dermal LD50 values above 2,000 mg/kg are indicated in the rabbit. Few details exist of the toxic signs observed or of target organs. For inhalation, a 6-hour exposure to 170 mg/m3 (the highest attainable concentration) produced no death in rats; slight and transient slight nasal tract epithelial damage was observed. These data indicate that Bisphenol A is of low acute toxicity by all routes of exposure relevant to human health."

There is no reliable and significant new information on the acute toxicity of Bisphenol A in the updated 2008 EU Risk Assessment Report or elsewhere.

Value used for CSA:

LD50 (oral): 2000 mg/kg bw

LD50 (dermal): 3000 mg/kg bw

Justification for classification or non classification

Bisphenol A is included in Annex 1 of Regulation 67/548/EEC and Annex VI of Regulation (EC) No 1272/2008.

No classification regarding acute toxicity is required.

4.3 Irritation / Corrosion


"Limited human anecdotal information of uncertain reliability is available from written industry correspondence suggesting that workers handling BPA have in the past experienced skin, eye and respiratory tract irritation. It cannot be determined whether the reported skin reactions were related to skin sensitisation or irritation. However, a recent well conducted animal study clearly shows that BPA is not a skin irritant. A recent well conducted animal study shows that BPA is an eye irritant; effects persisted until the end of the study (day 28 postinstillation) in 1 of 3 rabbits. Overall, taking into account the animal and human evidence, BPA has the potential to cause serious damage to the eyes. Slight and transient nasal tract epithelial damage was observed in rats exposed to BPA dust at 170 mg/m3 for 6 hours. Slight local inflammatory effects in the upper respiratory tract were observed in rats exposed to 50 mg/m3 and 150 mg/m3 of BPA in 2 and 13 week repeat inhalation studies, but were not observed at 10 mg/m3 in the same studies. Taken together with anecdotal human evidence, these data suggest BPA has a limited respiratory irritation potential."

There is no reliable and significant new information on irritation/corrosion of Bisphenol A in the updated 2008 EU Risk Assessment Report or elsewhere.



Value used for CSA:

Skin irritation / corrosion: not irritating

Eye irritation: irritating

Respiratory irritation: irritating


Bisphenol A is classified Xi; R37-41 according to Annex 1 of Regulation 67/548/EEC and Eye Dam. 1 and STOT SE 3 according to Annex VI of Regulation (EC) No 1272/2008.

4.4 Sensitisation

Skin sensitisation


"Based on the findings from the most robust animal study, BPA may possess a skin sensitisation potential, albeit a limited one, and that the available data suggest that BPA is considered capable of producing skin sensitisation responses in humans."


"Overall the new information does not confirm the previously reported evidence of a skin sensitisation potential of BPA. While the data do not exclude a skin sensitising activity of BPA at high concentrations (>30%), there is no evidence that this is a concern for workers in current BPA manufacturing plants (such workers are believed to represent the group most likely to be exposed to BPA in dust)."

There is no reliable and significant new information on the skin sensitisation of Bisphenol A that was not discussed in the 2003 or 2008 EU Risk Assessment Reports.

Value used for CSA: not sensitising

Respiratory sensitisation

The 2003 EU Risk Assessment Report concluded that there are no data from which to evaluate the potential of Bisphenol A to be a respiratory sensitiser.

The 2008 updated EU Risk Assessment Report came to the same conclusion but added: "However, based on the lack of reports of cases of respiratory sensitisation, there are no grounds for concern for this endpoint."

There is no reliable and significant new information on the respiratory sensitisation of Bisphenol A that was not discussed in the 2003 or 2008 EU Risk Assessment Reports.


Bisphenol A is classified R43 according to Annex 1 of Regulation 67/548/EEC and Skin Sens. 1 according to Annex VI of Regulation (EC) No 1272/2008.



4.5 Repeated dose toxicity


"No useful information on the effects of repeated exposure to BPA in humans is available, but experimental studies in rats, mice, and dogs are available. In rat inhalation studies, the principal effect of repeated exposure was the same as observed following a single exposure: slight upper respiratory tract epithelium inflammation, with a NOAEC of 10 mg/m3 and a LOAEC of 50 mg/m3. Dietary studies in rats have reported reductions in reproductive organ weights and testicular toxicity at 235 mg/kg and a NOAEL of 74 mg/kg was established in a two-year study based on marginal effects on body weight gain at the next dose level of 148 mg/kg. In mice, the LOAELs of 120 mg/kg in males for multinuclear giant hepatocytes and 650 mg/kg in females for a reduction in body weight gain of unknown magnitude were identifed in a two-year study. There are no animal data available for repeated dermal exposure."


"Oral studies in rats and mice have shown that the repeated dose toxicity of BPA involve[s] effects on bodyweight gain, liver and kidney. A NOAEL of 50 mg/kg/day has been identified in a recent 2-generation study in mice for these effects. This NOAEL rather than the original NOAEL of 120 mg/kg/day for liver effects from the published report is taken forward to the risk characterisation."

There is no reliable and significant new information on repeated dose toxicity of Bisphenol A.

Value used for CSA (route: oral):

NOAEL: 50 mg/kg bw/day

Value used for CSA (route: inhalation):

NOAEC: 10 mg/m³


Bisphenol A is included in Annex 1 of Regulation 67/548/EEC and Annex VI of Regulation (EC) No 1272/2008. No classification regarding repeated dose toxicity is required.

4.6 Genetic toxicity


"No human data regarding mutagenicity are available. However, Bisphenol A appears to have demonstrated aneugenic potential in vitro, positive results being observed without metabolic activation in a micronucleus test in Chinese hamster V79 cells and in a non-conventional aneuploidy assay in cultured Syrian hamster embryo cells. Additionally, in cell-free and cellular systems there is information that shows Bisphenol A disrupts microtubule formation. Bisphenol A has been shown to produce adduct spots in a post-labelling assay with isolated DNA and a peroxidase activation system, but it does not appear to produce either gene mutations or structural chromosome aberrations in bacteria, fungi or mammalian cells in vitro. However, some deficiencies in the conduct of these studies have been noted and the negative results cannot be taken as entirely conclusive. Bisphenol A does not appear to be anuegenic in vivo, since a recently conducted, standard mouse bone marrow micronucleus test has given a negative result. Bisphenol A was negative in a briefly reported dominant lethal study in rats but, given the limited details provided, this is not regarded as an adequate negative result. The only other data in somatic cells in vivo are from a 32P-postlabelling assay, which showed that Bisphenol A is capable of producing DNA adduct spots in rat liver following oral administration. These adduct spots were not characterised fully.

Considering all of the available genotoxicity data, and the absence of significant tumour findings in animal carcinogenicity studies (see below), it does not appear that Bisphenol A has significant mutagenic potential in vivo. Any aneugenic potential of Bisphenol A seems to be limited to in vitro test systems and is not of concern. The relevance of the finding that Bisphenol A can produce rat hepatic DNA adduct spots in a postlabelling assay is not entirely clear. However, given the absence of positive results for gene mutation and clastogenicity



in cultured mammalian cell tests, it seems unlikely that these are of concern for human health."


"New data from a study indicating effects of BPA on meiosis in female mice cannot be taken as conclusive evidence of an effect of BPA on germ cell meiosis because of the several methodological weaknesses and flaws identified in the study, the reporting inadequacies, and the known mutagenicity and toxicity profile of BPA. In addition, these findings have not been confirmed in more recent publications. Thus, the original conclusion that BPA has no significant mutagenic potential in vivo is still valid."

Additional recent information concerning the observations discussed in the 2008 updated EU Risk Assessment Report:

Two in vivo studies (Hunt et al. (2003) Curr Biol 13, 546-53) and Susiarjo et al. (2007) PLOS Genetics 3(1), 1-8) evaluated during the 2008 EU Risk Assessment Report reported that short-term oral exposure to low doses of Bisphenol A (≥ 0.020 mg/kg bw/day) in peripubertal or pregnant mice can interfere with meiotic divisions in development of female germ cells (“egg” or “oocyte”). An increase in hyperploid (aneuploid) metaphase II oocytes was observed following treatment with 0.020 mg/kg bw/day. There was not a significant increase in aneuploid embryos.

Two subsequent in vivo studies (Pacchierotti et al. (2008), Eichenlaub-Ritter et al. (2008)) attempted to replicate these findings. Consistent with the previous findings, they detected no significant effects of Bisphenol A exposure on the frequency of aneuploidy in “zygotes” (fertilised oocytes) produced from female mice treated before puberty or as adults with a similar range of doses. In addition, Eichenlaub-Ritter et al. (2008) found no effects of Bisphenol A exposure on aneuploid oocytes and Pacchierotti et al. (2008) found no increase in aneuploid or diploid sperm following exposure of male mice to Bisphenol A. The authors concluded that the aneuploidy predicted by the Hunt group could not be confirmed.

In addition, in a recent study published by the Hunt group, Muhlhauser et al. (2009), the authors could not replicate their initial findings on “congression failure” but report effects on chromosome alignment and/or spindle formation. The authors state “After publishing our findings [Hunt et al., 2003], we initiated studies to assess the effect of long term BPA exposure on the growing follicle. To our surprise, levels of BPA that were sufficient to elicit an effect on meiotic chromosome dynamics during the previous two years of study suddenly produced little or no effect. In an analysis of possible changes in experimental protocol, the only change identified was the lot of animal feed. ” The authors report frequencies of abnormal oocytes in the absence and presence of BPA in two different diets (casein based and soy based). The reported frequencies of abnormal oocytes of the BPA/casein group are lower than the background value reported in the soy-based diet.

Overall, the initial observations reported by the Hunt laboratory were not reproduced in the same laboratory or in other independent laboratories. Therefore, the conclusion from the 2003 EU Risk Assessment Report and the 2008 EU Risk Assessment Report Update is still valid; Bisphenol A has no significant mutagenic potential in vivo.

Value used for CSA: Genetic toxicity: negative


Bisphenol A is included in Annex 1 of Regulation 67/548/EEC and Annex VI of Regulation (EC) No 1272/2008. No classification regarding genetic toxicity is required.

4.7 Carcinogenicity


"There are no human data contributing to the assessment of whether or not BPA is carcinogenic, but a dietary carcinogenicity study in rats and mice concluded that BPA was not carcinogenic in either species because the tumour findings were not considered toxicologically significant. No inhalation or dermal carcinogenicity studies were available, although in repeat exposure inhalation toxicity studies, BPA did not exhibit properties that raised concern for potential carcinogenicity. Taking into account all the animal data available, it was concluded that the animal evidence suggests that BPA does not have carcinogenic potential."




"The new information on the potential carcinogenic and/or promoting effects of BPA in prenatal and neonatal rat models supports the original conclusion that BPA does not possess any significant carcinogenic potential. This is based on one new study in which the full carcinogenic potential of BPA on the mammary gland was examined in a prenatal model. This study claimed that BPA induced preneoplastic and neoplastic lesions of the mammary gland, but its validity was hampered by serious methodological limitations and its findings are inconsistent with the absence of preneoplastic lesions of the mammary gland in the offspring from several standard multi-generation studies in rats and mice. Other new studies suggest that prenatal or neonatal exposure to BPA does not exert promoting activity on the carcinogenesis induced by established carcinogens/initiators in specific organs."

There is no reliable and significant new information on the carcinogenicity of Bisphenol A.


Bisphenol A is included in Annex 1 of Regulation 67/548/EEC and Annex VI of Regulation (EC) No 1272/2008. No classification regarding carcinogenicity is required.

4.8 Toxicity to reproduction

Effects on fertility


"No human data on reproductive toxicity of BPA are available. BPA has been shown to have endocrine modulating activity in a number of screening assays, with a potency that generally ranged from 3 to 5 orders of magnitude less than that of oestradiol. The effects of BPA on fertility and reproductive performance have been investigated in two-generation and multi-generation studies in the rat and a continuous breeding study in mice. Effects were seen in both species at approximately the same dose level and it is considered that the NOAEL of 50 mg/kg/day identified in the rat multi-generation study is also likely to produce no adverse effects in mice for which there is only a LOAEL of 300 mg/kg/day for a small decrease in epididymal weight in F1 males. The NOAEL of 50 mg/kg/day from the multi-generation study will be used for risk characterisation purposes, in relation to effects on fertility."


"A new two-generation study in mice by Tyl et al. (published in 2008) provides a comprehensive and definitive investigation on the effects of BPA on reproduction at exposure levels spanning the low (ug/kg/day) to high (mg/kg/day) ranges. This study showed that BPA causes adverse effects on pregnancy and the offspring at 600 mg/kg/day, an exposure level that also caused mild parental toxicity. Fertility was not affected by BPA exposure. A NOAEL for reproductive toxicity of 50 mg/kd/day was identified and should be used in the risk assessment."

There is no reliable and significant new information on the reproductive or developmental toxicity of Bisphenol A.

Value used for CSA (route: oral): NOAEL: 50 mg/kg bw/day

Developmental toxicity

The 2003 EU Risk Assessment Report concluded that in standard developmental studies in rodents, there is no convincing evidence that Bisphenol A is a developmental toxicant. Available and apparently conflicting data from studies conducted using low doses raise uncertainties that the Competent Authorities required to be resolved through further testing. A provisional NOAEL of 50 mg/kg/day for developmental effects, derived from a rat multi-generation study, should be used in the risk characterisation in the interim while awaiting the outcome of further testing.



The 2008 updated EU Risk Assessment Report concluded that no conclusions could be drawn from new developmental toxicity studies or from a human study investigating recurrent miscarriage and Bisphenol A exposure, so these studies do not influence the conclusions of the original risk assessment report.

There is no reliable and significant new information on the reproductive or developmental toxicity of Bisphenol A that was not discussed in the 2003 or 2008 EU Risk Assessment Reports.

Value used for CSA (route: oral): NOAEL: 50 mg/kg bw/day


Bisphenol A is classified Repr. Cat.3; R62 according to Annex 1 of Regulation 67/548/EEC and Repr. 2 according to Annex VI of Regulation (EC) No 1272/2008.

4.9 Specific investigations

Neurotoxicity

A robust study in rats conducted under OECD and US EPA guidelines (Stump, 2009) reported no evidence of developmental neurotoxicity at any of the tested concentrations of Bisphenol A. Bisphenol A was continuously administered to rat dams during gestation and lactation via test diets containing 0, 0.15, 1.5, 75, 750, and 2250 ppm Bisphenol A. Reduced body weight and body weight gain was observed in dams and F1 males and females at 750 and 2250 ppm Bisphenol A. There were no effects observed at any exposure level in the neurobehavioural and neuropathological evaluations of F1 animals. A systemic NOAEL of 75 ppm (corresponding to 5.85 mg/kg/day during gestation and 13.1 mg/kg/day during lactation) for reduction in body weight and weight gain of dams and offspring was determined. A NOAEL for developmental neurotoxicity of 2250 ppm (corresponding to 164 mg/kg/day during gestation and 410 mg/kg/day during lactation) in offspring was determined, as there was no evidence of developmental neurotoxicity at any tested dose, including the highest dose of 2250 ppm Bisphenol A.


Bisphenol A is included in Annex 1 of Regulation 67/548/EECand Annex VI of Regulation (EC) No 1272/2008. No classification regarding Neurotoxicity is required.

Immunotoxicity

There are no studies conducted with validated methods to assess whether there is an association between Bisphenol A exposure and immunotoxicity.


Bisphenol A is included in Annex 1 of Regulation 67/548/EEC and Annex VI of Regulation (EC) No 1272/2008.

Specific investigations: other studies

Ashby and Odum (2004) conducted immature rat uterotrophic assays according to OECD protocols (Kanno et al., 2003a; Kanno et al., 2003b) and examined uterine growth and gene expression in rats administered Bisphenol A orally over the dose range of 2 µg/kg/day to 800 mg/kg/day. Expression of three oestrogen-responsive uterine genes and two control genes was determined using real-time PCR. Increases in gene expression were observed over the uterotrophic dose range (approximately 200 to 800 mg/kg/day Bisphenol A) but not over the dose range of 2 to 20 mg/kg/day, in which there was no uterotrophic response and no increase in gene expression. Thus, the no effect level for uterotrophic activity of Bisphenol A coincided with the no transcriptional effect level for uterine gene expression.

4.10 Derivation of DNEL(s) / DMEL(s)



Available dose-descriptor(s) per endpoint for the submission substance as a result of its hazard assessment

Endpoint Dose descriptor Qualitative assessment Remarks on study

Acute toxicity oral

LD50: 2000 mg/kg bw

Acute toxicity dermal

LD50: 3000 mg/kg bw

Irritation / Corrosivity

skin

not irritating


eye

irritating


respiratory tract

irritating

Sensitisation skin

not sensitising

Repeated dose toxicity: sub-acute / sub-chronic / chronic

oral


Repeated dose toxicity: sub-acute / sub-chronic / chronic

inhalation

NOAEC: 10 mg/m³

Mutagenicity in vitro / in vivo

Genetic toxicity: negative

Reproductive toxicity: fertility impairment

oral


Reproductive toxicity: developmental toxicity

oral




DN(M)ELs for workers

Exposure pattern

Route Descriptor DNEL / DMEL (Corrected) Dose descriptor *) Most sensitive endpoint Justification

Acute - systemic effects

Dermal DNEL (Derived No Effect Level)

1.4 mg/kg bw/day NOAEL: 49.0 mg/kg bw/day (based on AF of 35) repeated dose toxicity


Inhalation DNEL (Derived No Effect Level)

10 mg/m³ NOAEC: repeated dose toxicity

Acute - local effects

Dermal



10 mg/m³ NOAEC: irritation (respiratory tract)

Long-term - systemic effects






Long-term - local effects

Dermal




*) The (corrected) dose descriptor starting points have been automatically calculated by multiplying the values of the fields "D(N)MEL" and "Assessment factor" provided in the Endpoint summary of IUCLID section 7. Toxicological information. It reflects the value after any corrections, e.g. route-to-route extrapolation. See column "Justification" for the rationale behind such modifications and the use of assessment factors.



Discussion DNELs for workers

The human health hazard assessment is based on the initial EU Risk Assessment Report 2003 on Bisphenol A and the recent EU Risk Assessment Update in 2008. For each endpoint the conclusion of the EU Risk Assessments in 2003 and 2008 are cited and, if available, additional relevant information not evaluated in one of the Risk Assessments is indicated.

Due to the large data set on Bisphenol A, criteria were established to define Key-Studies, Supporting-Studies and additional studies to cover all available data for the human health hazard assessment. Key studies are defined as comprehensive studies conducted according to scientifically accepted methods and performed according to or exceeding validated guidelines (e. g. OECD testing guidelines). Supporting studies are defined as comprehensive studies conducted to scientifically accepted methods and performed similar to validated guidelines with only very minor deviations.

All additional studies are included in the IUCLID database without further specification of the purpose flag. Since this dataset is based on the IUCLID4 database prepared for and evaluated by the EU Risk Assessment in 2003, additional studies entered into the database before the EU Risk Assessment in 2003 are transferred to IUCLID5 without further modification. Additional studies evaluated during the EU Risk Assesment Update in 2008 or published more recently are added to the database as “additional information” and a short description of the method used and the key results is given.

DNEL are calculated according to the EU Annex XV Report submitted by United Kingdom, dated 30 November 2008. In the report it is stated that it is not a proposal for a restriction although the format is the same.

The human health endpoints for which concerns have been identified are:

respiratory tract irritation

skin sensitisation

repeat dose toxicity

reproductive toxicity

Inhalation toxicity

Concerning inhalation toxicity the EU RA concluded that the IOELV should be used as DNEL based on a scientific evaluation of the SCOEL documentation. The EU RA indicates: “In May 2004, the European Commission’s Scientific Committee on Occupational Exposure Limits (SCOEL) published a recommendation for an IOELV for Bisphenol A of 10 mg/m3 (8-hr TWA (time-weighted average)) and this recommendation has been included in the draft 3rd IOELV Directive which is due to be adopted early in 2009. In making its recommendation, SCOEL took account of evidence for site of contact inflammation in the respiratory tract of rats inhaling Bisphenol A and evidence for systemic toxicity in oral dosing studies (SCOEL, 2004). Although new information has become available since the SCOEL recommendation was finalised it does not change the no-observed adverse effect level (NOAEL) on which the IOELV proposal is based. In accordance with the CSA guidance theproposes to adopt 10 mg/m3as the long-term inhalation worker-DNEL. ”

Since mild nasal olfactory epthelium inflamation was seen at 50 mg/m3 in rats the short term DNEL should be equal to the long term DNEL.

Skin sensitisation

With regards to skin sensitisation, a DNEL/DMEL cannot be calculated from the available information. It is possible to make a qualitative judgement about the skin sensitising potency of Bisphenol A. The REACH technical guidance document indicates that skin sensitisers for which the EC3 value is greater than 2% in a LLNA should be regarded as moderate sensitisers. In the case of Bisphenol A, no sensitising activity was observed with a concentration of 30% and therefore, if an EC3 value were to be obtained it would be greater than 30%. On this basis, it can be concluded that Bisphenol A should be regarded as a moderate skin sensitiser.

Worker-DNEL long-term dermal route

There are no studies available to characterise the dose response relationship for systemic effects. Therefore, it will be necessary to obtain the worker DNEL long-term dermal by extrapolation from oral studies. A NOAEL of



50 mg/kg/day was identified in oral studies in rodents for the effects of Bisphenol A on reproduction, bodyweight, kidney and liver. The extensive first-path mechanism has to be taken into account for effects on reproduction, bodyweight and kidney. It is not necessary to take account of first pass metabolism when conducting the oral to dermal extrapolation for the effects of Bisphenol A in the liver, since this organ potentially receives all of the orally dosed material.

Dermal DNEL derived for repeated dose effects on the liver

Modification of starting point:

A NOAEL of 50 mg/kg/day has been identified from a 2-generation study in the mouse for the liver effects. Dermal absorption in both humans and animals is assumed to be 10%; whereas, the absorption following oral administration is estimated at 100%. Differences in absorption between routes have to be taken into account in order to conduct a route-to-route extrapolation; therefore, the oral NOAEL of 50 mg/kg/day is multiplied by 10 to give the equivalent dermal NOAEL. As indicated above it is not necessary to take account of first pass effects when conducting the route-to-route extrapolation for the liver. The corrected starting point defined by the EU RA is therefore 500 mg/kg/day.

Assessment factors

Interspecies differences 17.5 (2.5x7):

EU RA: “The dose descriptor is obtained from an oral study in the mouse. To use a value extrapolated from a mouse oral study to assess dermal exposure in humans it is necessary to apply an allometric scaling factor of 7 to take account of differences in basal metabolic rates between mice and human. For effects on the liver, the rodent-humans differences in the systemic availability of free unconjugated Bisphenol A are unimportant. On this basis a default factor of 2.5 to account for other species differences will be applied giving an overall AF of 17.5.”

Intraspecies differences 5:

EU RA: “There are no data to quantify variability in susceptibility to the effects of long-term exposure to Bisphenol A in the human population. The default factor of 5 for workers will therefore be used to take account of intraspecies differences. ”

Differences in duration of exposure 1:

EU RA: “Although the experimental conditions involved subchronic rather than chronic exposure, the evidence suggests that the severity of the effects does not increase when duration of exposure increases from 90 days to 2 years. A NOAEL of 50 mg/kg bw/day was identified in parental generations in subchronic reproductive toxicity studies in the mouse. The LOAEL was 600 mg/kg bw/day. In a chronic study with mice, some liver effects were observed at a dose level of 120 mg/kg bw/day but without an increase in severity at 650 mg/kg bw/day. Therefore, it is judged that an additional factor to extrapolate the subchronic NOAEL to chronic exposure is not necessary.”

Dose response and endpoint specific/severity issues 1:

EU RA: “There is more than one order of magnitude between the LOAEL (600 mg/kg/day) and the NOAEL (50 mg/kg/day) and only minor effects were observed at the LOAEL. It is, therefore, not necessary to apply an additional factor.”

Quality of database 1:

EU RA: “Several robust repeated exposure studies and a series of good quality 2- and multi-generation reproductive toxicity studies are available from which to characterise the effects of Bisphenol A in the liver. The quality of the database is therefore not considered to contribute uncertainty to this assessment and hence it is not necessary to apply an additional factor.”

Overall assessment factor: 87.5

Endpoint specific DNEL: 500/87.5 = 6 mg/kg/day



Dermal DNEL derived for repeated dose effects on bodyweight and kidney

Starting point

A NOAEL of 50 mg/kg/day has been identified in a 2-generation study in the mouse for effects on bodyweight and kidney. Due to the first pass metabolism there is a need to adjust the NOAEL by a factor of 10 to give an internal oral NAEL (no adverse effect level) of 5 mg/kg/day. Since dermal absorption is 10% of the applied dose in both humans and animals the internal NAEL of 5 mg/kg/day has to be multiplied by 10 to give the equivalent dermal NOAEL. The corrected starting point is therefore 50 mg/kg/day.

Assessment factors

Interspecies differences 7:

EU RA: “The dose descriptor is obtained from an oral study in the mouse. To use a value extrapolated from a mouse oral study to assess dermal exposure in humans it is necessary to apply an allometric scaling factor of 7 to take account of differences in basal metabolic rates between mice and humans. In view of the evidence that much lower levels of free Bisphenol A reach the systemic circulation in humans compared to the levels observed in rats and mice indicating that humans are likely to be less sensitive to the effects of Bisphenol A it is not considered necessary to apply an additional factor to take account of other species differences.”


Default factor for workers; see above


See above


See above


See above

Overall assessment factor: 35

Endpoint specific DNEL: 50/35 = 1.4 mg/kg/day

Dermal DNEL derived for reproductive toxicity

Starting point

As for effects on bodyweight and kidney the overall starting point is considered to be 50 mg/kg/day.

Assessment factors

Interspecies differences 4:

EU RA: “The same NOAEL of 50 mg/kg/day has been identified from mouse and rat data; however, since the available information (see section 4.1.3.1 of Risk Assessment Report) shows that the rat is a better model for humans than the mouse the allometric scaling factor for the rat (4) is considered more appropriate than that for the mouse. In view of the evidence that much lower levels of free Bisphenol A reach the systemic circulation in humans compared to the levels observed in rats and mice indicating that humans are likely to be less sensitive to the effects of Bisphenol A it is not considered necessary to apply an additional factor to take account of other species differences.”


Default factor for workers; see above




See above


See above


See above


Endpoint specific DNEL: 50/20 = 2.5 mg/kg/day

Overall conclusion on worker-DNEL long-term dermal route:

The worker DNEL long-term dermal route for systemic effects is 1.4 mg/kg/day based on effects on bodyweight and kidney at higher doses.

The short term DNEL should be equal to the long term DNEL.

Bisphenol A is considered to be a moderate skin sensitiser.



DN(M)ELs for the general population

Exposure pattern

Route Descriptor DNEL / DMEL (Corrected) Dose descriptor *) Most sensitive endpoint Justification








Oral DNEL (Derived No Effect Level)

0.05 mg/kg bw/day NOAEL: repeated dose toxicity


Dermal









0.25 mg/m³ NOAEC: repeated dose toxicity


Oral DNEL (Derived No Effect Level)

0.05 mg/kg bw/day NOAEL: repeated dose toxicity


Dermal




*) The (corrected) dose descriptor starting points have been automatically calculated by multiplying the values of the fields "D(N)MEL" and "Assessment factor" provided in the Endpoint summary of IUCLID section 7. Toxicological information. It reflects the value after any corrections, e.g. route-to-route extrapolation. See column "Justification" for the rationale behind such modifications and the use of assessment factors.


2011-10-20 Additional information 34

Discussion DNELs for the general population

Oral DNEL:

The Scientific Panel on Food Additives, Flavourings, Processing Aids and Material in Contact with Food (EFSA) evaluated Bisphenol A in 2006. The panel “concluded, in view of the well described species differences in toxicokinetics, showing a low level of free BPA in humans compared with rats, that a default uncertainty factor of 100 applied to the overall NOAEL from the rodent studies can be considered as conservative. The Panel therefore established a full TDI of 0.05 mg BPA/kg bw, derived by applying a 100-fold uncertainty factor to the overall NOAEL of 5 mg/kg bw/day.” This evaluation was reconfirmed by EFSA in 2009. The TDI is taken as a DNEL long-term oral.

Inhalation DNEL:

The EU RA concluded that there is apparently no need for further risk reduction measurements; conclusion (ii). Therefore, no DNELs were derived in the EU Annex XV report submitted byfor the general population.

Based on the long-term inhalation worker-DNEL of 10 mg/m3 derived from an IOELV a short term inhalation DNEL for the general population can be estimated taking into account a factor of 2 for the differences in the default intraspecies factor for worker (5) and for the general population (10). 5 mg/m3 is proposed as the inhalation short term DNEL for the general population

The long term inhalation DNEL for systemic effects is derived based on the EU Annex XV Report calculations for Worker-DNEL long-term dermal route by taking into account differences in absorption between the dermal and the inhalation route and the default intraspecies factor for the general population instead of the default factor for worker. For a detailed discussion on the individual assessment factors see chapter Worker-DNEL long-term.

A NOAEL of 50 mg/kg/day was identified in oral studies in rodents (rats or mice) for the effects of Bisphenol A on reproduction, bodyweight, kidney and liver. The extensive first-path mechanism has to be taken into account for effects on reproduction, bodyweight and kidney. It is not necessary to take account of first pass metabolism when conducting the oral to dermal or inhalation extrapolation for the effects of Bisphenol A in the liver, since this organ potentially receives all of the orally dosed material.

Inhalation DNEL derived for repeated dose effects on the liver


A NOAEL of 50 mg/kg/day has been identified in the EU Risk Assessment Report from a 2-generation study in the mouse for the liver effects. Absorption by inhalation and by oral ingestion is assumed to be 100%. As indicated above it is not necessary to take account of first pass effects when conducting the route-to-route extrapolation for the liver. For further calculation a body weight of 70 kg and an inhalation volume of 20m3/day is taken. Taking these parameters into account the corrected starting point is 175 mg/m3/day.

Assessment factors

Interspecies differences 17.5 (2.5x7): see EU RAR or discussion on worker

Intraspecies differences 10: default factor according to REACH TGD

Differences in duration of exposure 1: see EU RAR or discussion on worker

Dose response and endpoint specific/severity issues 1: see EU RAR or discussion on worker

Quality of database 1: see EU RAR or discussion on worker


Endpoint specific DNEL: 175/175 = 1 mg/m3/day


2011-10-20 Additional information 35

Inhalation DNEL derived for repeated dose effects on bodyweight and kidney


A NOAEL of 50 mg/kg/day has been identified in a 2-generation study in the mouse for effects on bodyweight and kidney. Due to the first pass metabolism there is a need to adjust the NOAEL by a factor of 10 to give an internal oral NAEL (no adverse effect level) of 5 mg/kg/day (EU RAR or discussion on worker). Absorption by inhalation and by oral ingestion is assumed to be 100%. The corrected starting point is therefore 17.5 mg/m3/day.

Assessment factors

Interspecies differences 7: see EU RAR or discussion on worker



Dose response and endpoint specific/severity issues 1: EU RAR or discussion on worker



Endpoint specific DNEL: 17.5/70 = 0.25 mg/m3/day

Inhalation DNEL derived for reproductive toxicity


A NOAEL of 50 mg/kg/day has been identified in the EU Risk Assessment Report from a 3-generation study in the rat for reproductive toxicity. Due to the first pass metabolism there is a need to adjust the NOAEL by a factor of 10 to give an internal oral NAEL (no adverse effect level) of 5 mg/kg/day (EU Risk Assessment Report or discussion on worker). Absorption by inhalation and by oral ingestion is assumed to be 100%. The corrected starting point is therefore 17.5 mg/m3/day.

Assessment factors

Interspecies differences 4: see EU RAR or discussion on worker



Dose response and endpoint specific/severity issues 1: EU RAR or discussion on worker



Endpoint specific DNEL: 17.5/40 = 0.44 mg/m3/day

Overall conclusion:

The inhalation-DNEL long-term for the general population is 0.25 mg/m3/day.

Dermal DNEL: Based on the long term dermal DNEL for workers and taking into account the different default factors for intraspecies variability for worker (5) and for the general population (10) a long term dermal DNEL for the general population would be 0.7 mg/kg.


2011-10-20 Additional Information 36

5. PBT and vPvB ASSESSMENT

In order to be persistent, bioaccumulative and toxic (PBT) a substance must exceed the “P” and “B” and “T” criteria. In order to be very persistent and very bioaccumulative (vPvB) a substance must exceed the “vP” and “vB” criteria, while toxicity is not part of the vPvB assessment. In ECHA (2008) “Guidance on information requirements and chemical safety assessment, Part C: PBT Assessment”, screening criteria are given in Table C.1-2 and more definitive criteria are given in Table C.1-1.

Persistence assessment

The conclusions of the persistence assessment of Bisphenol A are summarised in the following table:

Persistence assessment Persistence evaluation Persistence decision

Ready biodegradation Bisphenol A is readily biodegradable meeting the 10-day window

Not “P” and not “vP”

Microbial degradation Bisphenol A was found to be rapidly biodegraded by microbial consortia found in many natural waters and sediments, with lag times and half-lives on the order of magnitude of a few days


Degradation in soil Bisphenol A dissipates in soil in less than 3 days, with a major route being conversion to non-extractable bound residues


Degradation in air Bisphenol A has an estimated half-life in air of 0.13 days


Persistence decision:

Bisphenol A does not meet the “Persistent” or the “very Persistent” criteria. Bisphenol A is not “P” and not “vP”.

Bioaccumulation assessment

The conclusions of the bioaccumulation assessment of Bisphenol A are summarised in the following table:

Bioaccumulation assessment Bioaccumulation evaluation Bioaccumulation decision

Potential for Bioaccumulation Bisphenol A has low potential for bioaccumulation based on laboratory bioconcentration factors that are less than or equal to 73 L/kg in fish. The measured bioconcentration factors are less than the bioaccumulation criteria of 2000 (bioaccumulative) or 5000 (very bioaccumulative).

Not “B” and not “vB”

Partition coefficient (log Pow) The log Pow value of bisphenol A is 3.4 Not “B” and not “vB”

Bioaccumulation decision:

Bisphenol A does not meet the "Bioaccumulative" or the "very Bioaccumulative" criteria. Bisphenol A is not “B” and not “vB”.



Toxicity assessment The conclusions of the toxicity assessment of Bisphenol A are summarised in the following table:

Toxicity assessment Lowest LC50/EC50 or NOEC Toxicity decision

Acute freshwater fish LC50

Acute marine water fish LC50

4.6 mg/L

9.4 mg/L

-

Chronic freshwater fish NOEC

Chronic marine water fish NOEC

0.016 mg/L

0.066 mg/L

Not “T”

Acute freshwater invertebrate LC50

Acute marine water invertebrate LC50

10.2 mg/L

1.1 mg/L

-

Chronic freshwater invertebrate NOEC 0.025 mg/L Not “T”

Acute freshwater algae EC50

Acute marine water algae EC50

Chronic freshwater algae NOEC

Chronic marine water algae NOEC

2.73 mg/L

1.1 mg/L

1.36 mg/L

0.4 mg/L

Not “T”

Acute freshwater sediment organisms LC50

Acute marine water sediment organisms LC50

Chronic freshwater sediment organisms NOEC

Chronic marine water sediment organisms NOEC

150 mg/kg dw

31 mg/kg dw

22 mg/kg dw

32 mg/kg dw

Not “T”

Other organisms NOEC 0.5 mg/L Not “T”

Soil organisms NOEC 100 mg/kg dw Not “T”

Terrestrial plants NOEC 20 mg/kg dw Not “T”

None of the acute LC/EC50 or chronic NOEC are below 0.01 mg/L. However, Bisphenol A is classified Repr. 2 according to Regulation EC No 1272/2008. Toxicity decision:

Because of the classification of Bisphenol A as Repr. 2 according to Regulation EC No 1272/2008, Bisphenol A meets the toxicity criteria. Summary and overall conclusions on PBT or vPvB properties Bisphenol A does not meet any of the P or B criteria or any of the vP or vB criteria, but does meet one of the T criteria. In conclusion, Bisphenol A is not a PBT or a vPvB substance.



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