State of the Science Report - Part 2

Phthalate Substance Grouping
Medium-Chain Phthalate Esters

Chemical Abstracts Service Registry Numbers
84-61-7; 84-64-0; 84-69-5; 523-31-9; 5334-09-8; 16883-83-3; 27215-22-1; 27987-25-3; 68515-40-2; 71888-89-6

Environment Canada
Health Canada
August 2015

Table of Contents

• Tables and Figures
• 9. Potential to Cause Harm to Human Health
  • 9.1 Exposure
    • 9.1.1 Exposure of DIBP
    • 9.1.2 Exposure of DCHP
    • 9.1.3 Exposure of DMCHP
    • 9.1.4 Exposure of CHIBP
    • 9.1.5 Exposure of BCHP
    • 9.1.6 Exposure of DBzP
    • 9.1.7 Exposure of B84P
    • 9.1.8 Exposure of DIHepP
    • 9.1.9 Exposure of BIOP
    • 9.1.10 Exposure of B79P
  • 9.2 Health effects
    • 9.2.1 Toxicokinetics of medium-chain phthalates
    • 9.2.2 Health effects of DIBP
    • 9.2.3 Health effects of DCHP
    • 9.2.4 Health effects of DMCHP
    • 9.2.5 Health effects of CHIBP
    • 9.2.6 Health effects of BCHP
    • 9.2.7 Health effects of DBzP
    • 9.2.8 Health effects of B84P
    • 9.2.9 Health effects of DIHepP
    • 9.2.10 Health effects of BIOP
    • 9.2.11 Health effects of B79P
  • 9.3 Characterization of risk to human health
    • 9.3.1 Risk characterization of DIBP
    • 9.3.2 Risk characterization of DCHP
    • 9.3.3 Risk characterization of DMCHP
    • 9.3.4 Risk characterization of DBzP
    • 9.3.5 Risk characterization of B84P
    • 9.3.6 Risk characterization of DIHepP
    • 9.3.7 Risk characterization of B79P
    • 9.3.8 Risk characterization of CHIBP, BCHP and BIOP
  • 9.4 Uncertainties in evaluation of risk to human health
• Back to the State of the Science Report - Part 1

Tables and Figures

• Table 9-1. DIBP concentrations in indoor air
• Table 9-2. DIBP concentrations in dust
• Table 9-3. Rates of DIBP migration into simulated sweat from various articles
• Table 9-4. DIBP exposure estimates from dermal contact with plastic articles for infants (0-18 months) and adults (20+ years)
• Table 9-5. Concentrations of DIBP in various types of toys and children's articles
• Table 9-6: In vivo and in vitro migration rates into saliva from children's toys and articles
• Table 9-7. Exposure estimates from mouthing toys and childcare articles
• Table 9-8. Concentrations of DIBP in cosmetics and personal care products
• Table 9-9. Estimates of exposure from cosmetics use
• Table 9-10. FUEs for DIBP primary and secondary metabolites
• Table 9-11. Metabolites used for intake calculations in CHMS and P4 analyses
• Table 9-12. CHMS biomonitoring intakes (µg/kg/day) for 3-5 year-olds (males and females)
• Table 9-13. CHMS biomonitoring intakes for males (µg/kg/day)
• Table 9-14. CHMS biomonitoring intakes for females (µg/kg/day)
• Table 9-15. P4 pregnant women and MIREC-CD Plus infants (intakes) (µg/kg/day)
• Table 9-16. Dust concentrations of DCHP
• Table 9-17. Uncorrected urinary concentrations of MCHP in various North American surveys
• Table 9-18. Rates of DEHP migration into simulated sweat from various articles
• Table 9-19. B84P exposure estimates from dermal contact with plastic articles for infants (0-18 months) and adults (20+ years)
• Table 9-20: Effects from gestational exposure to DIBP in male offspring (mg/kg bw/day)
• Table 9-21. Effects from exposure to DIBP in prepubertal/pubertal males (mg/kg bw/day)
• Table 9-22. Effects from exposure to DBP in adult males (mg/kg bw/day)
• Table 9-23. Effects from gestational exposure to DCHP in male offspring (mg/kg bw/day)
• Table 9-24. Effects from exposure to DCHP in adult males (mg/kg bw/day)
• Table 9-25. Effects from gestational exposure to MBzP in male offspring (mg/kg bw/day)
• Table 9-26. Effects from gestational exposure to BBP in male offspring (mg/kg bw/day)
• Table 9-27. Effects from exposure to BBP in prepubertal/pubertal males (mg/kg bw/day)
• Table 9-28. Reproductive effects from exposure to BBP in adult males (mg/kg bw/day)
• Table 9-29. Short-term and subchronic studies in rodents exposed to BBP
• Table 9-30. Effects from gestational exposure to DIHepP in male offspring (mg/kg bw/day)
• Table 9-31. Effects from gestational exposure to B79P, MBzP and DINP in male offspring (mg/kg bw/day)
• Table 9-32. Reproductive effects from exposure to B79P and from exposure to DINP in adult males (mg/kg bw/day)
• Table 9-33. Carcinogenicity studies in rodents exposed to DINP
• Table 9-34. Summary of critical effects levels for reproductive and/or developmental effects after oral exposure to DIBP
• Table 9-35. Summary of margins of exposure to DIBP for relevant subpopulations with highest exposure
• Table 9-36. Summary of critical effect levels for reproductive and/or developmental effects after oral exposure to DCHP
• Table 9-37. Summary of margins of exposure to DCHP for relevant subpopulations with highest exposure
• Table 9-38. Summary of margins of exposure to DMCHP for relevant subpopulations with highest exposure
• Table 9-39. Summary of critical effect levels after oral exposure to DBzP using MBzP as closest analogue
• Table 9-40. Summary of margins of exposure to DBzP for relevant subpopulations with highest exposure
• Table 9-41. Summary of critical effect levels for reproductive and/or developmental effects after oral exposure to B84P using DIBP, MBzP and BBP as closest analogues
• Table 9-42. Summary of margins of exposure to B84P for relevant subpopulations with highest exposure
• Table 9-43. Summary of critical effect levels for reproductive and/or developmental effects after oral exposure to DIHepP
• Table 9-44. Summary of margins of exposure to DIHepP for relevant subpopulations with highest exposure
• Table 9-45. Summary of margins of exposure to B79P for relevant subpopulations with highest exposure

9. Potential to Cause Harm to Human Health

9.1 Exposure

9.1.1 DIBP

Environment media and food

The predominant sources of exposure are indoor air, dust and food (see Appendix F-1). The subpopulation with the highest exposure from environmental media and food consisted of breastfed infants with total daily intakes of 1.6 and 5.9 µg/kg/day, based on central tendency and upper-bounding concentrations, respectively.

Ambient air, drinking water and soil

No Canadian data were identified for DIBP in ambient air. DIBP has been detected in ambient air internationally (see section 8.2.1; Rudel et al. 2010; Xie et al. 2007). Rudel et al. (2010) was identified as the relevant study for exposure characterization (sampling from North America), and median and maximum (3.6 ng/m³, 18 ng/m³) concentrations were used to estimate potential exposures to DIBP via ambient air.

Limited data was identified indicating the presence of DIBP in surface water. Available Canadian data was limited to wastewater concentrations downstream of industrial sites and are not relevant for estimating potential drinking water exposure. Additionally, DIBP was detected in tap water and surface water in China (Shao et al. 2013).

No Canadian data as to DIBP presence in soil was identified; however, it has been detected in agricultural, top soil and urban area soil in China at concentrations less than 4 µg/g (Liu et al. 2010; Zeng et al. 2009; Zeng et al. 2008b).

Due to limited data pertaining to the presence of DIBP in soil, potential exposures were not estimated.

Bottled water

As phthalates are plasticizers with applications in packaging, they may be present in bottled water. In Canada, Cao (2008) surveyed phthalates in bottled carbonated and non-carbonated water and detected and quantified DIBP in all 11 samples (range: 0.133–0.481 µg/L). DIBP was also detected and quantified in bottled water samples internationally (Fierens et al. 2012a; Montuori et al. 2008; Guo et al. 2012; Cao 2008; Keresztes et al. 2013; Santana et al. 2014; Shao et al. 2013; Sun et al. 2013).

In the absence of data on levels of DIBP in tap water, mean (0.225 µg/L) and maximum (0.353 µg/L) concentrations of DIBP in bottled non-carbonated water were used to estimate the general population daily intake from drinking water (Cao 2008). The highest exposed subpopulation is 0 to 0.5 year-old (formula-fed) infants, and estimates of exposure are 0.024 and 0.038 µg/kg/day based on mean and maximum concentrations, respectively. The assumption that 100% of water consumption is from bottled water is considered conservative.

Indoor air and dust

Phthalates are semi-volatile compounds and are generally present in the indoor environment, likely due to their presence in plastic products (Weschler and Nazaroff 2010; Fromme et al. 2004; Bergh et al. 2011ab; Rudel et al. 2010; Bornehag et al. 2005). While long-chain phthalates tend to partition more to settled dust and surfaces, short-chain and low-molecular-weight medium-chain phthalates may partition in greater proportions to gaseous or particle phases of indoor air (Weschler and Nazaroff 2010; Fromme et al. 2004; Bergh et al. 2011ab). DIBP is considered to be a low-molecular-weight medium-chain phthalate and has been shown to exhibit these properties in paired dust and indoor air samples taken from residences in Germany and Sweden (Fromme et al. 2004; Bergh et al. 2011a). DIBP was detected in 100% of samples in these studies (Fromme et al. 2004; Bergh et al. 2011a). Additional key studies pertaining to dust and indoor air concentrations are listed below in Tables 9-1 and 9-2.

DIBP may be used in automotive repair adhesives and as a plasticizer in the manufacturing of various vehicle components (ECHA c2007-2014b). Phthalates have been monitored in air and particulate matter in car cabins, and DBP, DEHP and DEP have been detected; DIBP, however, was not detected (Geiss et al. 2009). No submissions indicating DIBP use in the manufacturing of automotive parts or automobiles in Canada were identified (Environment Canada 2014).

As no Canadian indoor air survey was identified, median (0.130 µg/m³) and maximum (1.7 µg/m³) concentrations from a United States study (Rudel et al. 2010) were used to estimate the general population daily intake of DIBP from indoor air. While these concentrations are approximately five times lower than the levels measured in Germany (Fromme et al. 2004), Swedish studies conducted more recently show comparable levels (Bergh et al. 2011ab).

The Canadian survey (Kubwabo et al. 2013) was identified as the key study for exposure characterization and median (5.17 µg/g), and 95th percentile concentrations (16.2 µg/g) were used to estimate the Canadian general population daily intake of DIBP from dust.

Estimated intakes from indoor air and dust exposure to DIBP were 1.6 and 5.9 µg/kg/day for 0- to 6-month-old infants (highest exposed group) for central tendency and upper-bound concentrations, respectively (see Appendix F-1, Table F-1a).

Food, beverages and infant formula

Some phthalates may be present in food and beverages through their potential use in PVC tubing and gloves, food packaging films, PVC gaskets for glass jars, printing inks in food packaging and the like (Fasano et al. 2012). Consequently, they have been detected in various food packaging and processing articles and have been known to migrate into food and beverages (Alin and Hakkarainen 2011; Barros et al. 2010; Bradley et al. 2007; Gartner et al. 2009; Page and Lacroix 1992; Fierens et al. 2012; Petersen et al. 2010; Xu et al. 2010; Xue et al. 2010).

In Canada, phthalates were monitored in a targeted survey of butter and margarine, including their packaging as part of Health Canada's Total Diet Study (Page and Lacroix 1992; Page and Lacroix 1995). DIBP was not detected in any of the samples measured (Page and Lacroix 1992: limit of detection (LOD) = 1000 ppb; Page and Lacroix 1995: LOD = 50-500 ppb). Recently, a survey of phthalates was also conducted on meat, fish and cheese and their packaging films, with DIBP not being detected in any samples (Cao et al. 2014: LOD = 110 ppb).^Footnote[4]

Phthalates have been monitored in total diet surveys and duplicate diet surveys in the United Kingdom, Belgium, United States, Germany, China and Taiwan, with DIBP being detected in all surveys. Specifically, DIBP was detected in 75% of 400 food samples in Belgium (Limit of quantification (LOQ): 0.03-15 ppb), 55% of 65 food samples in the United States (LOD: 0.2 ppb), 45% of 20 total diet survey food groups in the United Kingdom (LOD: 11.1-37.0 ppb), 61% of 350 duplicate diet samples (LOD: 10 ppb) and 100% of 171 duplicate diet samples in Germany (LOD: 0.2 ppb), greater than 60% of 70 food samples in China (LOQ: 2 ppb), and a significant proportion (detection frequency not stated) of 1200 food samples in Taiwan (LOD: 25-50 ppb) (Bradley et al. 2013ab; Fierens et al. 2012a; Schecter et al. 2013; Fromme et al. 2007; Fromme et al. 2013; Guo et al. 2012; Chang et al. 2014).

Data collected for the US total diet survey (Schecter et al. 2013) was considered representative of Canadian food levels (based on vicinity and food types evaluated) and was consequently used to estimate daily intake of DIBP for the general population. Additionally, the UK total diet survey (Bradley et al. 2013b) was also used to inform data gaps.^Footnote[5]

Probabilistic dietary intakes were derived for DIBP; the methodology used and results are outlined in Appendix F-2 and Table F-1a, respectively.

The highest intake estimates, among infants and children, based on median food concentrations are for 1-3 year-olds with a dietary intake of 0.024 µg/kg/day. For adults, the highest intakes are for females 19 to 30 years of age with a dietary intake of 0.0042 µg/kg/day. For infants less than 6 months, the 90^thpercentile^Footnote[6]dietary intake was estimated to be 0.12 µg/kg/day (highest 90th percentile intake for all populations).^Footnote[7] For 1-3 year-olds, food types that drive intake estimates of DIBP are cream, crackers and ready-to-eat cereals. For adults, food types that drive intake estimates for DIBP are cream, pork and ready-to-eat cereals.

Breast milk

DIBP metabolizes to the monoester MIBP in the gut, prior to uptake, although the parent substance may also be absorbed (Koch et al. 2012). Therefore, both DIBP and MIBP may be found in breast milk. Recently, Health Canada analyzed breast milk samples in the Maternal-Infant Research on Environmental Chemicals (MIREC) survey. DIBP was observed to be detected in 27 of 305 samples (Detection frequency (DF): 9%, mean: 33 ng/g, range: less than method detection limit (MDL)-85.4 ng/g) (personal communication from Food Directorate, Health Canada, to Existing Substances Risk Assessment Bureau, Health Canada, November 2014). However, these data were not used to quantify intakes, as it is thought that a majority of DIBP will metabolize to MIBP quickly; thus, MIBP is expected to be found at greater quantities and higher detection frequency than DIBP in breast milk (Koch et al. 2012).

Calafat et al. (2004) did not detect MIBP in any samples (n = 3 pooled samples) collected in the United States, while Hogberg et al. (2008) detected MIBP in 2 out of 42 samples (0.5 and 2.1 µg/L) in Sweden. In Italy, MIBP was detected in all 62 samples at a range of 8.4 to 57.2 µg/L (Latini et al. 2009). In the most recent breast milk survey, MIBP was detected in 100% of 74 samples collected in Germany (mean: 13.8 µg/L, median: 11.8 µg/L, range: 4.4-43.8 µg/L, 95^th percentile: 27.9 µg/L) (Fromme et al. 2011).

The parent compound DIBP was not detected in 10 samples in Sweden (Hogberg et al. 2008) but was detected in 82% of 78 samples (mean: 1.5 ng/g, median: 1.2 ng/g, range: ND-5.3 ng/g) in Germany (Fromme et al. 2011).

Results from the most recent published German survey were used to estimate exposure from breast milk (Fromme et al. 2011). Since concentrations of MIBP were detected at higher concentrations than DIBP (consistent with current knowledge), MIBP concentrations were used for exposure characterization (after a molecular weight adjustment: parent MW/metabolite MW = 1.252). Mean and maximum concentrations were used to derive general population intakes which were estimated to be 1.5 and 5.4 µg/kg/day (see Appendix F, Table F-1a) for breastfed infants.

Products Used by Consumers

Manufactured items/children's articles/children's toys/textiles

DIBP may also be present in a wide variety of manufactured items including plastic sandals, balance balls, furniture, and decorative articles (Danish EPA 2011; Danish EPA 2010ab). Canadian use of DIBP was identified in toys and exercise equipment (e.g., yoga mats, balance balls) (Environment Canada 2014).

DIBP has also been reported to be present in leather articles, textiles, and apparel (ECHA c2007-2014b) and has been detected in 6 out of 10 t-shirt samples at concentrations of less than less than 0.002% (Danish EPA 2010a). Globally, other phthalates (DCHP, B79P, B84P) have been reported to be present as coatings in textiles and fabric (see Table 5-2); however no reports of this use of DIBP in Canada were identified. However, DIBP was found to be present at a low frequency rate (see Table 9-5) in children's articles, such as bibs, handbags, slippers and balls (Health Canada 2007b, 2014). Concentrations in these articles ranged from 0.003 to 61.7%.

It is expected that the general population may be exposed to DIBP from dermal contact with these articlesand, in the case of small children, potentially from mouthing articles, such as toys.

Dermal exposure

Dermal exposure to phthalates from products such as toys, balance balls and sandals has been assessed by other jurisdictions (Danish EPA 2011; Danish EPA 2010ab; NICNAS 2011). For DIBP specifically, the internal dose from dermal contact with these articles^Footnote[8] was estimated to be 0.58 to 4.92 µg/kg/day and 1.0 to 3.6 µg/kg/day for adults and children, respectively (Danish EPA 2011; Danish EPA 2010ab). The Danish EPA also estimated dermal exposure to DIBP from dermal contact with school bags, toy bags, pencil cases and erasers (range between 0.01 and 32.54 µg/kg/day, Danish EPA 2007).^Footnote[9]

A conservative exposure assessment was conducted to estimate exposure to DIBP from dermal contact with the various manufactured items mentioned above. Two representitative scenarios were developed to model exposure of infants in contact with various plastic articles (PVC, polyurethane, polyester, etc.) for 1 hour/day with 25% of their body surface area (representative of multiple diaper changes per day on a change pad) and for 4 hours/day with 50% of their body surface area (representative of holding a plastic article and being changed on a plastic change pad multiple times a day and playing on a plastic mat).

Two representitative scenarios to model exposure of adults in contact with various plastic articles were also assessed: the first for 3 hours/day with 16% of their body surface area (analogous to sitting on a couch or wearing plastic gloves), and the second for 3 hours/day with 50% of their body surface area (representative of various daily contacts with plastic articles including wearing gloves or holding a plastic steering wheel, sitting on a couch and wearing plastic clothing).

Rates of DIBP migration from plastics into a solution simulating sweat have been studied and are summarized in Table 9-3.

An average migration rate of 2.5 µg/cm²/h was derived from these studies. Note that migration rates were derived without correcting for experiment duration (assuming all plasticizer comes out in the first hour; for example, 7.9 µg/cm²/h was averaged, not 7.9 µg/cm²/16 h), as an evaluation of migration rate data shows that a majority of phthalates migrate out in the first 1 to 3 hours. Therefore, dividing the migration rate by 16 hours would lead to underestimation of exposure. This scenario assumes zero dermal lag times and does not account for plasticizer depletion, both of which are conservatisms in the scenario.

Estimates of exposure for adults and children from dermal contact with plastic articles are provided in Table 9-4.

These exposure estimates are higher than estimates of systemic exposure following dermal contact reported by Danish EPA 2011 and others and much higher than estimates of exposure from biomonitoring studies (Tables 9-12-15). Additionally, these estimates have significant uncertainty associated with them, as there is high uncertainty with respect to factors such as surface area and contact time.

Oral exposure

DIBP has also been detected in childcare articles and toys; the results are summarized in Table 9-5.

The various types of toys and articles tested included items such as dolls, figurines, building blocks and cars to modelling clays, bath toys and bibs (Health Canada 2007b; Stringer et al. 2000; Biedermann-Brem et al. 2008; and Korfali et al. 2013).

Several jurisdictions have evaluated the migration of phthalates (DINP, DEHP, DBP, etc.) from toys and childcare articles into simulated saliva (Danish EPA 2010a; RIVM 1998; RIVM 2001; NICNAS 2010; Danish EPA 2011). DIBP migration into saliva has been evaluated in one study (balance ball), while DBP^Footnote[10] has been evaluated in numerous studies^Footnote[11](Danish EPA 2010a; RIVM 2001; Niino et al. 2001, 2003).

Migration rates for DIBP and DBP have been organized according to % content and magnitude of migration, and are outlined in Table 9-6. The data indicates that DBP migration rates follow a linear relationship with % concentration and in vivo migration rates are approximately ten-fold lower. However, DBP has been shown to metabolize in saliva as concentrations of the monoester, MBP, reach 87% within 60 minutes (Niino et al. 2001, 2003). Therefore, DBP in vivo migration rates, if calculated solely on DBP appearance in saliva, may be underestimated.

Mouthing time, surface area exposed and frequency of mouthing have been evaluated and summarized in numerous publications (Babich et al. 2004; USEPA 2011; Greene 2002; Juberg et al. 2001; Xue et al. 2010). ECHA evaluated these parameters for phthalates in a recent DINP and DIDP risk assessment, and used daily mouthing durations of children's toys and articles to be 0.5 to 2 hours/day for typical and worse case scenarios, respectively. A surface area mouthed of 10 cm² was used (ECHA 2013a).

Exposure to DIBP from mouthing of toys and childcare articles was estimated as a range based on in vitro migration rates of DIBP and DBP.^Footnote[12]Exposure estimates for infants 0 to 18 months ranged from 2.47 to 251.0 µg/kg/day (see Table 9-7).

Cosmetics and personal care products

Based on notifications submitted under the Cosmetic Regulations Health Canada DIBP is not expected to be present in cosmetics in Canada (September 2014 email from the Consumer Product Safety Directorate (CPSD), Health Canada to Existing substances Risk Assessment Bureau (ESRAB), Health Canada).  DIBP has been detected in various types of cosmetics and personal care products^Footnote[13](Koniecki et al. 2011; Guo and Kannan 2013ab; Liang et al. 2013). This presence may be due to potential migration from packaging. A summary of recent studies measuring concentrations of DIBP in cosmetics and personal care products is outlined in Table 9-8.

Since the two North American studies report that detection frequencies are low (5 to 7% in Canada, 10 to 27% in USA), and a majority of the concentrations in all three studies are in the sub-ppm range, exposure from personal care products and cosmetics may not be significant.

Estimates of dermal exposure from cosmetic products associated with the highest potential exposure are presented in Table 9-9. The products presented were chosen because they are associated with leave-on application, highest frequency of use and highest DIBP concentration.

The highest estimate of dermal exposure to DIBP from cosmetics is from use of body lotion, with an estimated systemic exposure of 0.028 µg/kg/day generated using maximal concentrations from Koniecki et al. (2011). For the oral route, exposure to DIBP from presence in lipstick was estimated to be less than 1 ng/kg/day.

For baby products, Guo and Kannan et al. (2013a) surveyed shampoo, lotions and oils, baby powder, sunscreen and diaper cream, with DIBP being present in only two baby shampoo samples out of four (mean: 0.03 µg/g, max: 0.09 µg/g). Additionally, Koniecki et al. (2011) showed non-detection in 98 baby products (lotions, oil, diaper creams and shampoo). Using mean and maximum concentrations in Guo and Kannan et al. (2013b), exposure estimates were generated for infants 0 to 6 months (baby shampoo use) and were estimated to be less than 0.001 µg/kg/day.

Koniecki et al. (2011) did not generate estimates for DIBP, stating that concentrations were less than 10 µg/g, while Guo et al. 2013b generated aggregate leave-on exposure estimates of 0.0005 and 0.004 µg/kg/day using mean and max concentrations, respectively. These aggregate exposure estimates are generally consistent with the estimates presented above.

Biomonitoring

DIBP is expected to be metabolized in the body primarily to the unique monoester MIBP (Koch et al. 2012). The fractional urinary excretion (FUE) of a substance is defined as the mole ratio of the amount of metabolites excreted in urine (at 24 hours) to that of total parent compound ingested. FUEs of MIBP and the secondary metabolite 2OH-MIBP are presented below in Table 9-10.

MIBP has been monitored in the Canadian Health Measures Survey (CHMS) Cycle 2 (2009-2011) with 100% detection in all samples (Health Canada 2011, 2013). MIBP and 2OH-MIBP were also monitored by Health Canada in two cohort surveys: Plastics and Personal Care Product Use in Pregnancy survey (P4, n = 31 women, 542 individual urine spot samples; women provided multiple urine samples over two visits), and Maternal-Infant Research on Environmental Chemicals - Child Development Plus study (MIREC-CD Plus, 194 children, 2-3 years old, 1 spot sample per individual). Both these surveys reported a 100% detection of both metabolites (personal communication from Environmental Health Science and Radiation Directorate [EHSRD] to Existing Substances Risk Assessment Bureau [ESRAB], October 2013, 2014).

Finally, in the United States, the National Health and Nutrition Examination Survey (NHANES) also monitored MIBP in urine during survey years 1999-2012 and show high detection frequencies (CDC 2014).

Using the CHMS, P4 and MIREC-CD Plus datasets, reverse dosimetry intake estimates were generated. Metabolite concentrations were adjusted for urine dilution using the creatinine correction method, a commonly used method for phthalate biomonitoring assessment (Fromme 2007; Christensen et al. 2014; CHAP 2014; Frederiksen et al. 2014). Daily creatinine excretion rates for participants were estimated using the Mage equation. Biomonitoring intakes are presented in Tables 9-12 through to 9-15 below (see Appendix G for further information on the methodology).

The highest exposed group (all sources, CHMS) is 6-11 year-old male children with median and 95^th percentile intakes of 0.76 and 5.3 µg/kg/day, respectively. For older populations, the highest exposed group (all sources, CHMS) is 12-19 year-old males with median and 95^th percentile intakes of 0.53 and 1.4 µg/kg/day, respectively.

9.1.2 DCHP

Environment media and food

Ambient air, drinking water and soil

No Canadian data were identified for DCHP in ambient air, water or soil. Limited international monitoring data on DCHP presence in ambient air and surface water were identified. In ambient air, DCHP was not detected at industrial and rural sites in California (MDL: 1 ng/m³). In surface water, DCHP has been detected in the Netherlands, Germany and China (see Section 8.2 for concentrations), while in groundwater, DCHP was not detected on Belgian farmland (Fierens et al. 2012b). Monitoring data on DCHP presence in drinking water were not identified.

DCHP was detected in topsoil, urban soil and agricultural soils in China and Belgium; however, concentrations (ND-0.30 µg/g) were generally lower than concentrations found in house dust (see Table 9-16) (Zeng et al. 2009; Liu et al. 2010; Fierens et al. 2012b).

Due to the absence of Canadian data on DCHP presence in ambient air, soil and water intake estimates were not generated for these sources.

Indoor air and dust

No Canadian data were identified for DCHP in indoor air.  Elsewhere, DCHP has been detected in indoor air (both volatile and particulate matter) in one survey of homes conducted in Cape Cod, USA (21% of 102 homes; arithmetic mean: 3.4; median: ND; 90^th percentile^Footnote[14]: 210; range: ND-280 ng/m³) (Rudel et al. 2003).^Footnote[15]Additionally, a survey in Norway measured DCHP in PM_2.5and PM₁₀ particulates and reported no significant differences in the presence of DCHP in both particle phases (PM_2.5: ND-20 ng/m³, mean: 4 ng/m³, PM₁₀: ND-19 ng/m³, mean: 5 ng/m³) (Rakkestad et al. 2007).

DCHP has applications as a plasticizer in the manufacturing of automobiles and automobile parts (Environment Canada 2014). For the general population, indirect exposure (e.g., off-gassing) is considered a relevant source, but no data on this exposure source has been identified, which is currently an uncertainty in the assessment.

The Canadian survey, Kubwabo et al. 2013 (median: 0.21 µg/kg, 95th percentile: 1.0 µg/kg) and Rudel et al. 2004 (arithmetic mean: 3.4, maximum: 280 ng/m³) were identified as relevant studies for exposure characterization of the general population. Estimated intakes from indoor air and dust exposure to DCHP were 0.0018 and 0.15 µg/kg/day for infants 0.5 to 4 years (highest exposed group), for central tendency and upper-bound concentrations, respectively (see Appendix F-1, Table F-2).

Food, beverages and infant formula

DCHP has been identified as a component of food packaging material (US FDA 2014). In Canada, phthalates were monitored in a targeted survey of butter and margarine, including their packaging, and as part of Health Canada's Total Diet Study (Page and Lacroix 1992; Page and Lacroix 1995). DCHP was not detected in any of the samples.^Footnote[16]

Internationally, DCHP has been monitored in three total dietary surveys conducted in the United States, United Kingdom and Belgium (Schecter et al. 2013; Fierens et al. 2012a; Bradley et al. 2013b). Schecter et al. (2013) detected DCHP in 4 out of 65 total diet survey samples,^Footnote[17] while Bradley et al. 2013b did not detect DCHP in a majority of total diet survey samples. Fierens et al. (2012a) detected DCHP in 97 out of 400 total diet survey samples, with DCHP being detected in all food groups tested. Levels detected in food in the US total diet survey (Schecter et al. 2013) were used as the most relevant data to estimate dietary exposure of the general population. Additionally, the UK total diet survey (Bradley et al. 2013b) was also used to inform data gaps.

Probabilistic dietary intake estimates were derived for DCHP, with median and 90th percentile^Footnote[18]intakes were less than0.001 µg/kg/day.

Breast milk

Recently, an analysis of breast milk samples obtained from 56 Canadian women in the P4 cohort survey showed no detection of MCHP, the monoester of DCHP (LOD: 0.009 µg/L) (personal communication from EHSRD to ESRAB, October 2013).

Products used by consumers

DCHP was not detected in any samples from an emission chamber study measuring and modelling the emission of phthalates from 101 manufactured items (e.g. shower curtains, cable/wire) sampled from the Ottawa area (NRC 2012).

With respect to DCHP reported international use in the production of plastisols used in fabrics, textiles and apparel (ECHA . c2007-2014c), due to a lack of data on the migration of DCHP or a similar phthalate, exposure estimates from this source were not generated and are currently an uncertainty in this assessment.

Finally, DCHP was detected in 1 out of 36 perfume samples at a concentration of 3 mg/kg (SCCP 2007). Dermal exposure was estimated to be less than 10 ng/kg/day.

Biomonitoring

DCHP is expected to be metabolized in the body primarily to the unique monoester MCHP (See section 9.2.1), and has been monitored in several surveys in North America. Specifically, MCHP has been monitored in CHMS Cycle 1 and 2, with 87 and 72% of samples being measured below the limit of detection (Health Canada 2011, 2013).^Footnote[19] MCHP was also monitored by Health Canada in the MIREC, MIREC-CD Plus and P4 cohort surveys (Arbuckle et al. 2014; personal communication to ESRAB from EHSRD, October 2013, 2014).

In the United States, the National Health and Nutrition Examination Survey (NHANES) monitored MCHP in urine during survey years 1999-2010. From 1999 to 2004, MCHP was detected at the 90th percentile and above. In subsequent survey years, however, MCHP was not detected at the 95th percentile (CDC 2013).

Urinary concentrations and detection frequencies of MCHP are presented in Table 9-17.

Currently, information regarding DCHP toxicokinetics in humans is limited (e.g. FUEs cannot be determined), and reverse dosimetry intake estimates could not be derived from urine concentrations measured in humans.

9.1.3 DMCHP

Environment media and food

Monitoring data were not identified for DMCHP in ambient air, indoor air, surface water or drinking water in Canada or elsewhere. Similarly, monitoring data were not identified for DMCHP in food or food packaging materials (US FDA 2014).

Although DMCHP was not reported to be in commerce above the reporting threshold in Canada (see Uses section), it was detected in dust samples collected as part of the Canadian House Dust Study (CHDS), which analysed dust from 126 homes in ten cities across Canada (Kubwabo et al. 2013). Two isomers of DMCHP were identified in selected standards (DMCHP1 and DMCHP2) and the two isomers were detected in 37 and 89% of homes, respectively (DMCHP1 levels ranged from ND to 4.1µg/g, median: ND, while DMCHP2 levels ranged from ND to 24.3 µg/g, median: 0.53 µg/g, 95th percentile: 10.7 µg/g) (Kubwabo et al. 2013).

Levels of DMCHP2 were used to characterize exposure from dust, as this substance was present in more homes and was detected at higher levels. Estimates of exposure from DMCHP presence in dust were 0.0027 µg/kg/day (median) and 0.054 µg/kg/day (95th percentile) for the 0-to-6-month age group (highest exposed age group) (see Appendix Table F-3).

Products used by consumers

Based on information collected in the section 71 survey, no import, manufacture or export was identified for DMCHP (Environment Canada 2014). DMCHP was also identified to be not in commerce as per the DSL IU initiative (Canada 2009).

Generic uses have been reported for DMCHP in other jurisdictions (see Uses), and analysis of these uses indicated potential use of DMCHP in manufactured items. However, an emission chamber study measuring and modelling the emission of phthalates in 101 manufactured items (cable/wire, shower curtains, caulking/sealant, etc.) did not detect emitted DMCHP in any of the samples (NRC 2012).

Therefore, based on the above considerations, direct exposure to DMCHP from use of products used by consumers or contact with manufactured items is not expected.

9.1.4 CHIBP

Monitoring data were not identified for CHIBP in ambient air, indoor air, surface water or drinking water in Canada or elsewhere. Similarly, monitoring data were not identified for CHIBP in food or food packaging materials (US FDA 2014).

CHIBP was an analyte in the CHDS and was not detected in any of the samples (MDL: 0.008 µg/g) (email from the EHSRD, Health Canada, to ESRAB, Health Canada, October 2013). An emission chamber study measuring and modelling the emission of phthalates from 101 manufactured items (cable/wire, shower curtains, caulking/sealant, etc.) detected CHIBP emission in six samples (NRC 2012). Specifically, CHIBP was detected in three shower curtain and three cable/wire samples, and a modelled average indoor air concentration of 2 ng/m³ was derived by NRC (NRC 2012).

Based on information collected in the section 71 survey, no import, manufacture or export was identified for CHIBP (Environment Canada 2014). Therefore, given the absence of reporting to the section 71 industry survey, non-detection in dust, negligible modelled indoor air concentrations, and the absence of information as to CHIBP presence in product databases, general population exposure to CHIBP from environmental media and products used by consumers is expected to be negligible.

9.1.5 BCHP

Monitoring data were not identified for BCHP in ambient air, indoor air, surface water or drinking water in Canada or elsewhere. Similarly, monitoring data were not identified for BCHP in food or food packaging materials (US FDA 2014).

BCHP was an analyte in the CHDS and was not detected in any of the samples (MDL: 0.008 µg/g) (email from the EHSRD, Health Canada, to ESRAB, Health Canada, October 2013). Additionally, an emission chamber study measuring and modelling the emission of phthalates from 101 manufactured items (cable/wire, shower curtains, caulking/sealant, etc.) did not detect BCHP in any of the samples (NRC 2012).

Based on information collected in the section 71 survey, no import, manufacture or export was identified for BCHP (Environment Canada 2014).

Therefore, given the absence of reporting to the section 71 industry survey, non-detection in dust and products (emission chamber study), and the absence of information as to BCHP presence in product databases, general population exposure to BCHP from environmental media and products used by consumers is expected to be negligible.

9.1.6 DBzP

Environment media and food

Monitoring data were not identified for DBzP in ambient air, indoor air, surface water or drinking water in Canada or elsewhere. DBzP may be used as an indirect additive in food packaging materials (US FDA 2014). However, monitoring data as to its presence in food were not identified. DBzP was monitored in two brands of bottled water in France and was not detected in any of the samples (Devier et al. 2013). DBzP has been identified as a candidate for monitoring as part of Health Canada's Total Diet Study (January 2014, email from the Food Directorate, Health Canada, to the Risk Management Bureau, Health Canada).

DBzP was surveyed in the CHDS; however, chromatography analysis showed that DBzP co-eluted with another phthalate (BIOP, CAS RN 27215-22-1) under experimental conditions (Kubwabo et al. 2013). These methodological issues precluded separate quantification of the compounds, and DBzP and BIOP concentrations were reported together (DF: 83%; range: less than DL-61.2 µg/g, median: 3.09 µg/g, 95th percentile: 19.1 µg/g) (Kubwabo et al. 2013). Since no BIOP manufacture, import or export was reported under the section 71 industry survey (Environment Canada 2014) and a search of databases revealed no evidence of BIOP use in products globally, chromatogram peaks in the CHDS are attributed to DBzP.

Therefore, estimates of exposure from DBzP presence in dust were derived, with the highest intakes (for infants 0 to 6 months) being 0.016 and 0.097 µg/kg/day for median (3.09 µg/g) and 95th percentile (19.1 µg/g) concentrations, respectively (see Appendix Table F-4).

Products used by consumers

DBzP was imported at quantities of less than 100 000 kg during 2008 (Canada 2009); however, it was not reported to be imported, exported or manufactured during 2012 (Environment Canada 2014).

DBzP uses, identified in the 2009 survey, were adhesives, sealants, paints and coatings, and these uses were corroborated by global uses (see Uses section). However, searches for concentrations did not identify any concentration information (HPD 2014). Finally, DBzP was measured in a study monitoring emission of phthalates in 101 manufactured items (vinyl flooring, wall coverings, caulking/sealant, shower curtains, cable/wire) and was not detected in any of the samples surveyed (NRC 2012).

Based on the above considerations, exposure to DBzP from use of products by consumers or contact with manufactured items is not expected.

9.1.7 B84P

Environment media and food

Monitoring data were not identified for B84P in ambient air, indoor air, surface water or drinking water in Canada or elsewhere. Similarly, monitoring data were not identified for B84P in food or food packaging materials (US FDA 2014).However, it is important to note that B84P has no laboratory standard, meaning that monitoring of this substance in different media is currently not feasible and unavailability of data does not indicate an absence of exposure. Additionally, a similar plasticizer with similar uses (B79P) was found to be present in 100% of dust samples in the CHDS (Kubwabo et al. 2013). It is therefore plausible that the general population may also be exposed to B84P. In the absence of an appropriate analytical method for measuring B84P in dust, and because similar quantities and uses were reported in Canada (Environment Canada 2014), B79P dust intakes were used as a surrogate for B84P dust exposure. The highest estimate exposure (for infants 0 to 6 months) being 0.0063 and 0.047 µg/kg/day for median (1.2 µg/g) and maximum (52.3 µg/g) concentrations, respectively (see Appendix Table F-5). B84P has applications in the production of automotive sealants and compounds used in the manufacturing of automobiles and vehicle parts (Environment Canada 2014). For the general population, indirect exposure (e.g., off-gassing) is considered a relevant source, but no data on this exposure source has been identified, which is currently an uncertainty in the assessment.

Products used by consumers

B84P may also be used as a plasticizer in the coating of textiles and fabrics in multiple applications (e.g., personal apparel, vehicle upholstery) (ECHA c2007-2014d). Given high import volumes (see Section 4) and its global use pattern, potential exposure to the general population to B84P, from its use as a plasticizer in manufactured items (e.g. PVC, polyurethane, polyester), was characterized.

A conservative exposure assessment was conducted to estimate exposure to B84P from dermal contact with the various manufactured items (see Table 9-19). Two scenarios were developed to model exposure of infants in contact with various plastic articles (PVC, polyurethane, polyester, etc.) for 1 hour/day with 25% of their body surface area (representative of multiple diaper changes per day on a change pad) and for 4 hours/day with 50% of their body surface area (representative of holding a plastic article and being changed on a plastic change pad multiple times a day and playing on a plastic mat).

Two scenarios to model exposure of adults in contact with various plastic articles were also assessed: the first for 3 hours/day with 16% of their body surface area (analogous to sitting on a couch or wearing plastic gloves), and the second for 3 hours/day with 50% of their body surface area (representative of various daily contacts with plastic articles including wearing gloves or holding a plastic steering wheel, sitting on a couch and wearing plastic clothing).

Migration studies have shown that various phthalates can migrate from articles (sandals, children's articles, toys, etc.) into saliva and sweat (Danish EPA 2010ab; RIVM 2001; Babich et al. 2004). Given that B84P is similar in molecular weight, log K^ow and solubility to DEHP,^Footnote[20] rates of DEHP migration into simulated sweat were used as an analogue to quantify exposure to B84P from dermal contact with plastic articles. An outline of the migration rates used for B84P is provided below (Table 9-18).

An average migration rate of 0.22 µg/cm²/h was derived from these studies. Migration rates were derived without correcting for experiment duration (assuming all plasticizer comes out in the first hour; for example, 1.7 µg/cm²/h was averaged, not 1.7 µg/cm²/16 h), as an evaluation of migration rate data shows that a majority of phthalates leach out in the first 1 to 3 hours. Dividing the migration rate by 16 hours would therefore lead to an underestimation of exposure. Note that this scenario assumes zero dermal lag times and does not account for plasticizer depletion, both of which are conservatisms.

Estimates of exposure for adults and children from dermal contact with plastic articles are provided in Table 9-19.

Conservative estimates of dermal exposure from contact with plastic articles, depending on the scenario, were 2.7 and 21.6 µg/kg/day for infants. For adults, conservative estimates of dermal exposure, depending on the scenario, were 2.7 and 8.5 µg/kg/day.

Finally, B84P was reported to be used mainly in industrial coatings applied to exterior and interior systems (Environment Canada 2014), but a do-it-yourself use is also feasible. Given that phthalates are metabolized quickly, do not bioaccumulate in the body and are known to have low acute toxicity, acute dermal exposure from incidental use of these types of products is not anticipated to contribute significantly to the overall exposure of the general population in Canada. Therefore, estimates were not generated, and the focus will be on subchronic and chronic exposure assessments (see risk characterization in Section 9.3.5 for additional information).

9.1.8 DIHepP

Environment media and food

Limited monitoring data on the presence of DIHepP in surface water were identified. Specifically, DIHepP was detected in surface water in False Creek Harbour, British Columbia, at a range of 2.91 to 153 ng/L. However, monitoring data were not identified for DIHeP in ambient air, indoor air, or drinking water in Canada or elsewhere. Similarly, monitoring data were not identified for DIHeP in food or food packaging materials (US FDA 2014).

DIHepP was surveyed in the CHDS and was detected in 96% of homes (range: ND-1223 µg/g, median: 18.90 µg/g, 95th percentile: 222.5 µg/g) (Kubwabo et al. 2013). The highest estimates of exposure (for infants 0 to 6 months) were 0.096 and 1.1 µg/kg/day for median (18.9 µg/g) and 95th percentile (222.5 µg/g) concentrations, respectively (see Appendix Table F-6).

Products used by consumers

DIHepP was imported in quantities of less than 10,000 kg in 2012 and at higher quantities (100,000 to 1,000,000 kg) in 2008 (Canada 2009; Environment Canada 2014).

DIHepP was identified to be used in flooring products (Canada 2009). Exposure from this source is predominantly expected through indirect sources (e.g., dust) and is already addressed in the environmental media and food section.

Finally, DIHepP was also reported to be used in the production of caulking and sealants (Environment Canada 2014). Caulking and sealant use was confirmed by industry MSDSs (HPD 2014); a do-it-yourself use is therefore expected. Given that phthalates are metabolized quickly, do not bioaccumulate in the body and are known to have low acute toxicity, acute dermal exposure from incidental use of these types of products is not anticipated to contribute significantly to the overall exposure of the general population in Canada. Therefore, estimates were not generated, and the focus will be on subchronic and chronic exposure assessments (see risk characterization in Section 9.3.6 for additional information).

9.1.9 BIOP

Monitoring data were not identified for BIOP in ambient air, indoor air, or drinking water in Canada or elsewhere. Similarly, monitoring data were not identified for BIOP in food or food packaging materials (US FDA 2014).

BIOP was surveyed in the CHDS, but it co-eluted with DBzP under experimental conditions (Kubwabo et al. 2013). These methodological issues precluded identification of both compounds separately, and DBzP and BIOP concentrations were reported together (DF: 83%; range: less than DL-61.2 µg/g, median: 3.09 µg/g) (Kubwabo et al. 2013).

However, DBzP was imported into Canada in quantities of less than 100 000 kg for 2008 (Canada 2009) and may also be present as an impurity in BBP, a high volume phthalate. Given that there is no reported production of BIOP in the US and Europe and no reported manufacturing or import in Canada (US EPA 2014ab; ECHA c2007-2014a; Environment Canada 2014), BIOP is not expected to be present in significant levels in Canadian home dust. Therefore, presence in dust samples will be attributed to the more likely presence of DBzP in Canadian homes.

Based on information collected in the section 71 survey, no import, manufacture or export was identified for BIOP (Environment Canada 2014). Therefore, exposure of the general population to BIOP from environmental media or products used by consumers is expected to be negligible.

9.1.10 B79P

Environment media and food

Monitoring data were not identified for B79P in ambient air, indoor air, or drinking water in Canada or elsewhere. Similarly, monitoring data were not identified for B79P in food or food packaging materials (US FDA 2014).

B79P has applications in the production of compounds used in the manufacturing of automobiles and automobile parts (Environment Canada 2014). For the general population, indirect exposure (e.g., off-gassing) is considered a relevant source, but no data on this exposure source has been identified, which is currently an uncertainty in the assessment.

B79P was surveyed in the CHDS and was detected in 95% of homes (range: ND-52.3 µg/g, median: 1.24 µg/g, 95th percentile: 9.2 µg/g) (Kubwabo et al. 2013). Estimates of exposure from B79P presence in dust for the highest exposed group (infants 0 to 6 months) were 0.0063 and 0.047 µg/kg/day based on median (1.2 µg/g) and 95th percentile (9.2 µg/g) concentrations, respectively (see Appendix Table F-7).

Products used by consumers

B79P may also be used as a plasticizer in the coating of textiles and fabrics in multiple applications (e.g., personal apparel, vehicle upholstery) (ECHA c2007-2014d). Given high import volumes (see Section 4) and its global use pattern, potential exposure to the general population to B79P, from its use as a plasticizer in manufactured items (e.g. PVC, polyurethane, polyester), was characterized.

A conservative exposure assessment was conducted to estimate exposure to B79P from dermal contact with the various manufactured items (see Table 9-19). Two scenarios were developed to model exposure of infants in contact with various plastic articles (PVC, polyurethane, polyester, etc.) for 1 hour/day with 25% of their body surface area (representative of multiple diaper changes per day on a change pad) and for 4 hours/day with 50% of their body surface area (representative of holding a plastic article and being changed on a plastic change pad multiple times a day and playing on a plastic mat).

Two scenarios to model exposure of adults in contact with various plastic articles were also assessed: the first for 3 hours/day with 16% of their body surface area (analogous to sitting on a couch or wearing plastic gloves), and the second for 3 hours/day with 50% of their body surface area (representative of various daily contacts with plastic articles including wearing gloves or holding a plastic steering wheel, sitting on a couch and wearing plastic clothing).

Given that B79P is similar in molecular weight, log K_ow and solubility to DEHP^Footnote[21], DEHP migration rates into simulated sweat were used to quantify exposure from dermal contact with selected articles.  The migration rates used, approach for characterizing dermal exposure are outlined in section 9.1.7, (Tables 9-18 and 9-19). Conservative estimates of dermal exposure from contact with plastic articles, depending on the scenario, were 2.7 and 21.6 µg/kg/day for infants 0-18 months. For adults, conservative estimates of dermal exposure, depending on the scenario, were 2.7 and 8.5 µg/kg/day.

B79P was also reported to be used in the production of sealants and coatings, for which consumer use may occur (Environment Canada 2014; ECHA c2007-2014d; SPIN 2006; 3M 2012ab). A search of industry MSDSs revealed that a majority of these products are for industrial or commercial use and are used to apply coatings to various types of surfaces (mechanical, glass, fibreglass, etc.) (3M 2013; Flexbar 2011; 3M 2012ab); however, a do-it-yourself scenario is feasible. Given that phthalates are metabolized quickly, do not bioaccumulate in the body and are known to have low acute toxicity, acute dermal exposure from incidental use of these types of products is not anticipated to contribute significantly to the overall exposure of the general population in Canada. Therefore, estimates were not generated, and the focus will be on subchronic and chronic exposure assessments (see risk characterization in Section 9.3.7 for additional information).

9.2 Health Effects

A critical effect of medium-chain phthalates consists of adverse effects on the development of the male reproductive system following exposure. Exposure to these phthalates during the critical development window of gestation have been shown to result in disturbances in the androgen-mediated development of the reproductive system in male rats, with the biological pathways leading to common effects or adverse outcomes in reproduction. The effects detected in early postnatal life include altered feminization parameters, such as decreased anogenital distance (AGD) and areolar/nipple retention (NR) in juveniles (Gray et al., 2000)^Footnote[22]. Other effects observed include reproductive tract malformations (cryptorchidism [CRY], hypospadias [HYP] and testicular pathological changes) and effects on sperm counts, motility and quality at adulthood (Gray et al. 2006). This spectrum of effects on male reproductive development has been described as the "rat phthalate syndrome" (RPS) and although primarily studied in rats, it has also been demonstrated in other species (reviewed in NAS 2008).

Conceptually, the effects associated with RPS can be divided into three subsets with different mode of action considerations. The first subset of effects is related to androgen insufficiency (decreased testicular testosterone production) in the fetal rat and is caused by altered functioning of Leydig cells. The second subset of phthalate syndrome effects has also been attributed to altered functioning of Leydig cells; however, the effects are separate from the role that testosterone plays in development. INSL3 gene expression is reduced and is attributed to a second proposed mechanism of action for cryptorchidism (McKinnell et al. 2005; Wilson et al. 2004). Finally, the third subset of phthalate syndrome effects is related to altered functioning of Sertoli cells in the fetal testes. Certain phthalates can also affect Sertoli cells in utero and may result in altered Sertoli-germ cell interactions, leading to multi-nucleated gonocytes (MNG) (Kleymenova 2005). The long-term biological significance of fetal MNGs is not clearly understood (Clewell et al. 2013). For a more detailed consideration of the current knowledge of the mode of action of phthalate-induced toxicity, please see the Category Approach Document (Health Canada 2015a).

Based on the above mentioned understanding of phthalate toxicity, the hazard assessment of each phthalate in this grouping is structured to present the evaluation of studies at three different life stages (gestational exposure [GD0-21], (pre)pubertal-pubertal [PND1-55], and adult [PND55+]), with particular focus on the male gender.^Footnote[23] The purpose of the hazard assessments is to identify the most sensitive life stage of phthalate toxicity for risk characterization if adequate information is available. As the focus of the SOS report is on presenting lines of evidence pertinent to developing a screening assessment moving forward, descriptions of effects for each life stage are structured to present a summary of effects starting from the lowest doses at which these effects were observed, from an overall database perspective in lieu of a study by study narrative. Adverse effects observed subsequent to in utero exposure to phthalates (Sections 9.2.X.1) in this grouping are presented as follows: 1) changes in hormone levels (serum or testicular); 2) feminization effects; 3) reproductive tract malformations and/or effects on fertility; and 4) other developmental effects. Each section also tabulates critical information for each of the identified studies reporting effects.

The potential reproductive developmental effects of each phthalate in female animals were also assessed in a similar manner in considering life stage and species sensitivity.

Exposure to phthalates is also associated with other systemic effects in laboratory animals. Repeated-dose studies indicated that the liver may be a target of phthalate adverse effects. Effects on other organs, such as kidneys, have also been observed. A review of studies examining these effects (i.e., repeated-dose studies, chronic/carcinogenicity studies, genotoxicity studies) is presented in the appropriate sections.

When no studies were available for a particular phthalate at a specific life-stage or exposure period, an analysis of health effects of the closest analogue as identified in the Category Approach Document (Health Canada 2015a) was conducted.

Additionally, available information on the potential effects of phthalates on humans was evaluated. The published literature was searched and human studies with an epidemiological focus were identified for further consideration. The evaluation included cross-sectional, case-control and cohort studies that encompassed 14 phthalate parent compounds and their metabolites. Given the large number of studies available in humans and the diverse outcomes identified for this substance grouping, all studies collected were scored for quality using a consistent evaluation metric^Footnote[24](Downs and Black 1988).This allowed for a reliable, objective assessment tool that captured the dimensions of study quality across various study designs. Statistically significant exposure-response associations were evaluated for each health outcome. A conclusion as to the level of evidence of association of a phthalate and each health outcome was based on the strength and consistency of the relationship as well as the quality of the epidemiology studies, as determined by the Downs and Black scores. Based on the overall score obtained from the evaluation approach, the level of evidence for an association was designated as sufficient, limited, inadequate, or evidence suggesting no association.  Studies that were rated in the lowest quartile (Quartile 1) based on the evaluation were not included in this report.  This evaluation did not consider the biological plausibility of the relationship, meaning that no causal inference was established. More detail is provided in Health Canada (2015b) available upon request.

9.2.1 Toxicokinetics of medium-chain phthalates

A summary of the toxicokinetics of medium-chain phthalate estersis provided in Appendix H.

9.2.2 DIBP

9.2.2.1 Reproductive and developmental effects in males

9.2.2.1.1 Early development: in utero exposure

The European Commission classified DIBP as Category 2 (causes developmental toxicity in humans) Risk phrase R61 (may cause harm to unborn child) for developmental toxicity and as Category 3 (causes concern for human fertility) Risk phrase R62 (possible risk of impaired fertility) for reproductive toxicity (ECHA 2009). Subsequent changes to the classification schemes for the hazard class within the European Union Classifying, Labelling, and Packaging (CLP) regulations (EC No 1272/2008) resulted in a change in the status of DIBP to Category 1B - reproductive toxicant (presumed human reproductive toxicant).

A literature search identified six studies examining the potential toxicity of DIBP during gestation in rats focusing on exposure during the masculinization programming window (gestational days [GD] 15-17) where any potential anti-androgenic effects would be observed. Summaries of the studies are described in Table 9-20 below. A limited study in mice using only one high dose exposure to DIBP was also identified. It should be noted, however, that most of the reproductive parameters directly pertaining to the male reproductive system as it relates to the general rat phthalate syndrome (RPS) were not measured in this species, meaning that no conclusions can be made regarding the particular potential of DIBP to induce this syndrome in mice.

In utero oral exposure to DIBP in rats causes effects in the male foetus related to RPS, which increased in severity with increasing dose. In a critical study by Saillenfait et al. (2008), pregnant Sprague-Dawley rats were administered 125, 250, 500 or 625 mg/kg bw/day of DIBP by gavage on gestation days (GD) 12-21. They observed that the more sensitive effects included decreased AGD at postnatal day 1 (PND1), nipple retention (NR) at PND12-14, and effects in sperm (oligospermia and total azoospermia) and tubular degeneration in seminiferous tubules at maturity (postnatal weeks 11 and 16) at doses starting at 250 mg/kg bw/day in the absence of maternal effects.

A more recent study presenting the potential for DIBP and other phthalates to affect foetal testosterone production (ex vivo) in pregnant SD rats showed that DIBP administration altered testicular testosterone production during gestation at doses of 200 mg/kg bw/day and higher with a calculated ED₅₀ value of 288 mg/kg bw/day (Furr et al. 2014).

At higher doses, the onset of puberty (preputial separation, PPS) was delayed along with reproductive tract malformations, such as undescended testes (cryptorchidism [CRY]), hypospadias (HYP), exposed os penis, cleft prepuce, and reduced testis, epididymis, seminal vesicles and prostate weights. Histopathological lesions were also present in testes of these males at maturity that mainly consisted of seminiferous tubule degeneration. Similar findings were also observed in earlier studies at higher doses (Table 9-20; Saillenfait et al. 2005; Borch et al. 2006; Saillenfait et al. 2006; Boberg et al. 2008).

Other effects at higher doses included embryotoxicity, reduced foetal viability and foetal visceral and skeletal malformations (Saillenfait et al. 2005, 2006; Howdeshell et al. 2008). Slight maternal toxicity became evident at dose levels starting at 500 mg/kg bw/day, with transient body weight changes becoming significant at 900 mg/kg bw/day in some studies (Saillenfait et al. 2005, 2006; Howdeshell et al. 2008), while in others, no maternal toxicity was observed at similar dose levels (Saillenfait et al. 2008; Hannas et al. 2011).

Two separate studies examined the potential of DIBP to affect steroidogenesis in the developing male foetus by measuring testicular testosterone levels. Howdeshell et al. (2008) and Hannas et al. (2011) both determined that in utero exposure to DIBP during the critical masculinization programming window caused a decrease in testicular testosterone levels at similar dose levels where effects on masculinization parameters and reproductive tract malformations were observed in other studies (300 mg/kg bw/day. See Table 8-20 and the Category Approach Document for more details on these studies (Health Canada 2015a).

In order to further investigate DIBP-induced steroidogenesis, Boberg et al. (2008) performed in vivo gene expression analysis for some of the genes known to be involved in steroidogenesis. Results showed that DIBP reduced testicular mRNA levels of SR-B1, StAR, P450c17, P450scc and INSL3 in male offspring exposed to DIBP during gestation (GD19 and GD21). Additionally, DIBP reduced testicular SF-1 mRNA levels. Hannas et al. (2011) also confirmed some gene expression changes at dose levels lower than those where testosterone levels were reduced (300 mg/kg bw/day). DIBP reduced foetal testis RNA expression levels for StAR and Cyp11a at ED₅₀ values of 191 and 171 mg/kg/day, respectively. Further analysis published in 2012 showed that DIBP reduced expression of additional relevant genes in the steroid biosynthesis pathway, such as SR-B1, 3βHSD and CYP17A1 (Hannas et al. 2012). See the Category Approach Document for more details on these studies (Health Canada 2015a).

Overall, the highest oral no-observed-adverse-effect level (NOAEL) for developmental toxicity of DIBP at the in uterolife stage was 125 mg/kg bw/day based on effects on the developing male reproductive system as seen by decreased testicular testosterone production, decreased AGD, nipple retention and effects in sperm, seminiferous tubules and decreased prostate weights (above 10%) in males at maturity at the next doses tested (lowest observed-adverse-effect level [LOAEL] of 250 mg/kg bw/day) (Saillenfait et al. 2008; Furr et al. 2014). This effect level from this study was identified as a critical effect level by other jurisdictions in recent assessments (Danish EPA 2012; US CPSC CHAP 2014; Germany 2014). No marked maternal effects were reported, with decreases in body weight gain during pregnancy occurring at 900 mg/kg bw/day (LOAEL; Howdeshell et al. 2008). As mentioned previously, the one study in mice was of limited value for examining the potential of DIBP to affect male reproductive development. No developmental studies were identified examining gestational exposure to DIBP using other species.

 9.2.2.1.2 Exposure at prepubertal/pubertal life stages

Results from repeated-dose oral exposure studies in sexually immature rats (PND1-55) have shown that administration of DIBP can causereproductive effects in male rats. Summaries of the studies are described in Table 9-21 below.

In the prepubertal rat (PND~21-39), exposure to DIBP at this life stage causes effects in sperm and in the testes. Zhu et al. (2010) exposed young Sprague-Dawley male rats to 100, 300, 500, 800 and 1000 mg/kg bw/day once on PND21 and daily for seven days (PND21-28), and observed that DIBP caused an increase in apoptotic spermatogenic cells at 500 mg/kg bw/day and above for both exposure durations. The repeated oral administration of DIBP also induced a decrease in testes weight and alterations in the distribution of vimentin filaments in Sertoli cells, which, according to the authors, correlates with sloughing of spermatogenic cells from the seminiferous epithelium. These effects were not observed in prepubertal C56BL/6N mice when tested under the same conditions, with the exception of decreased testes weight at the highest dose (1000 mg/kg bw/day) after the repeated exposure (Zhu et al. 2010).

Oishi and Hiraga (1980a) observed effects in spermatogenesis and decreased relative testes weight in Wistar rats after administration of a high dose of DIBP during PND35-42 (Table 9-21). The authors also noted a significant increase in testicular testosterone concentrations in rats administered 1212 mg/kg bw/day (P less than 0.05), but not in mice administered 2083 mg/kg bw/day (Oishi and Hiraga 1980a,b). Further, mice, but not rats, exhibited increased relative testes weight at high doses (Oishi and Hiraga 1980b), which is not consistent with the other, more recent study in mice which observed decreased testes weight (Zhu et al. 2010).

In an older study (Hodge 1954) in weanling Albino rats (species and age not provided) administered 0, 0.1, 1.0 and 5% DIBP in their diet for 16 weeks, significant decreases in body weights, and both absolute and relative testes weights were observed in the high-dose group. Slight systemic effects included increased relative liver weights at the highest dose with no histopathological effects (Hodge 1954).

Overall, the highest oral NOAEL for the reproductive toxicity of DIBP at the prepubertal/pubertal life stage was 300 mg/kg bw/day based on effects on the male reproductive system, as seen by decreased testes weight, increased number of apoptotic spermatogenic cells and alterations in the distribution of vimentin filaments in Sertoli cells at the next dose tested (500 mg/kg bw/day) (Zhu et al. 2010). The dose of 300 mg/kg bw/day from the Zhu et al. (2010) study was also identified as the NOAEL for reproductive effects by the US CPSC CHAP (2014). Studies in mice indicates that this species may not be as sensitive to reproductive effects of DIBP during this life stage (Oishi and Hiraga 1980b; Zhu et al. 2010). Mild systemic effects included increased liver weight in rats as well as increased liver and decreased kidney weights in mice at dose levels of 1212 mg/kg bw/day and above (Oishi & Hiraga 1980a,b; Hodge 1954). No studies were identified on any other species by any route of exposure at this life stage.

9.2.2.1.1 Oral exposure at the mature male adult stage

No studies examining the potential health effects of DIBP were identified in sexually mature adult male rats (PND55+) by any route of exposure. Dibutyl phthalate (DBP) (1,2-Benzenedicarboxylic acid, dibutyl ester: CAS RN 84-74-2) was identified as the "closest analogue" to DIBP based on similarity in the length and nature of the ester chains (Section 2.3.2; Health Canada 2015a). Summaries of the relevant studies conducted with DBP are described below and summarized in Table 9-22.

An examination of the effects of DBP on the reproductive system of the adult male rat showed effects on sperm count and motility starting at 500 mg/kg bw/day, with testicular pathology at similar doses. Srivistava et al. (1990) administered 250, 500 and 1000 mg/kg bw/day of DBP to adult Wistar rats by gavage for 15 days. They reported a 30% decrease in sperm count as well as evidence of disorganized seminiferous tubules, disturbed spermatogenesis, and irregular spaces devoid of sperm in rats exposed to 500 mg/kg bw/day DBP, along with alterations in the activity of enzymes related to specific events of spermatogenesis. These effects became more severe at the highest dose with a decrease of approximately 70% in sperm count and severely damaged seminiferous tubules.

A more recent 14-day study in adult male Sprague-Dawley rats reported decreased epididymal weights with evidence of epididymal tubule atrophy, hyperemia of the interstitial vasculature, and oligospermic lumina at the highest dose tested (500 mg/kg bw/day) (Zhou et al. 2011). The study authors also found significant effects in the activities of antioxidant enzymes in the epididymis.

A similar study reported decreases in serum testosterone at 500 mg/kg bw/day DBP as well as testicular pathology at higher dose ranges (750-1000 mg/kg bw/day) in Sprague-Dawley rats after the same exposure period (O'Conner et al. 2002). Mild systemic effects included increased liver weight with no corresponding histopathological indications at 500 mg/kg bw/day (O'Conner et al. 2002).

There are a significant number of studies examining the potential toxicity of DBP in adult male mice, which indicate that mice could potentially be less sensitive to the effects of DBP at this life stage compared to rats. DBP did not cause any adverse effects on fertility or in the testes at up to significantly high doses (900-3689 mg/kg bw/day and above) (Lamb et al. 1987; Morrissey et al. 1988; Marsman et al. 1995; Dobryznska et al. 2011; Hao et al. 2012). Marsman et al. (1995) compared the changes in serum testosterone in both rats and mice and reported that DBP caused a decrease in testosterone concentrations in rats at 1540 mg/kg bw/day and above after a 90-day DBP administration, but did not affect levels in mice exposed for the same duration at similar dose.

In a study by Higuchi et al. (2003), adult males rabbits (6-8 months) where administered 0 and 400 mg/kg bw/day of DBP via gavage for 12 weeks. Evidence of testicular pathology (germinal epithelial loss) and abnormalities in sperm morphology were reported in exposed animals. There were no effects on serum testosterone levels or on mating behaviour at this life stage and no effects on body weight gain. An increase in thyroid weight was reported (Higuchi et al. 2003).

Two inadequately reported studies examining the effects of DIBP in other species were described in secondary sources (cat, BASF 1961 in EC 2004; dog, Hodge 1954 in NICNAS 2008). These studies were determined to be of limited value due to the small sample size and nature of parameters measured. There was, however, some indication of decreased sperm in one male dog administered 0.1 mL/kg/day of DIBP in the diet for two months (Hodge 1954 in NICNAS 2008).

Overall, the highest NOAEL for reproductive toxicity identified for DBP was 250 mg/kg bw/day based on atrophy of epididymal tubules, hyperemia of the interstitial vascular, oligoszoospermic lumina and decreased reproductive organ weight (epididymis) as well as effects on sperm count, motility, disorganized seminiferous tubules, altered spermatogenesis and irregular spaces devoid of sperm at the next dose tested of 500 mg/kg bw/day in adult male rats in two studies (Srivastava et al. 1990; Zhou et al. 2011). The lowest LOAEL for systemic toxicity was 720 mg/kg bw/day based on increased kidney and liver weights, hepatocellular cytoplasmic alterations, and increased number of peroxisomes in male rats (Marsman et al. 1995). Limited oral studies with DIBP in cats and dogs did not provide any further information, and no studies were identified on any other species by any route of exposure at this life stage. Therefore, the NOAEL of 250 mg/kg bw/day will be used as the critical effect level for the reproductive toxicity of DIBP for this life stage.

9.2.2.2 Oral exposure in females

Eight published studies on the reproductive and developmental effects of DIBP in females were identified. Most of these studies were performed in rats, where animals were administered DIBP by the oral route during different gestational times.

The lowest oral LOAEL identified for developmental toxicity in females (500 mg/kg bw/day) is based on the same study described in the section above in rats (Saillenfait et al. 2006). Developmental toxicity occurred at doses that were not maternally toxic and included alteration of growth (statistically significant reduction of foetal body weights at 500 mg/kg bw/day and above with a NOAEL of 250 mg/kg bw/day).

The lowest LOAEL identified for reproductive toxicity in adult females (750 mg/kg bw/day) is also based on the Saillenfait et al. (2006) developmental toxicity study as well as on a similar preliminary study in which exposure occurred via gavage on GD6-20 to doses of 0, 250, 500, 750 and 1000 mg/kg bw/day (Saillenfait et al. 2005, as cited in Saillenfait et al. 2006 and ECHA 2009). The critical effects included alteration of fertility and pregnancy outcomes (statistically significant increase in post-implantation losses per litter, resorptions per litter and number of live foetuses per litter) (Saillenfait et al. 2006) and an increased number of resorptions (Saillenfait et al. 2005), occurring at doses of 750 mg/kg bw/day and higher (NOAEL of 500 mg/kg bw/day).

Overall, developmental effects in females were identified at doses of 500 mg/kg bw/day and above after oral exposure with critical endpoints related to growth alterations, alterations of reproductive development, functional deficit, lethality and mild teratogenicity. When gender effects were examined separately, effects of DIBP-induced developmental effects in males and females were observed at the same dose levels, with some studies reporting male pups as more sensitive than female pups. Reproductive effects of DIBP in females and alterations of fertility and pregnancy outcomes (embryolethality) were observed at 750 mg/kg bw/day and above, higher doses than at which reproductive effects were observed in males.

9.2.2.3 Reproductive and developmental toxicity: evidence in humans

Available epidemiological studies examining the potential relationship between observed effects and exposure to DIBP in humans were reviewed (Appendix J; Health Canada 2015b). Overall, there were no associations established for DIBP and its metabolite, mono-isobutyl phthalate (MIBP), and effects on male reproductive hormones (Joensen et al. 2012), preterm birth and gestational age (Wolff et al. 2008; Meeker et al. 2009; Ferguson et al. 2014c), birth measures (Wolff et al. 2008; Philippat et al . 2012) or any other reproductive parameters (endometriosis, gynecomastia, time to pregnancy) examined (Mieritz et al. 2012; Upson et al. 2013; Buck Louis et al. 2014) . Inadequate evidence was identified for an association for MIBP and placental gene expression (Adibi et al. 2010). There was inadequate evidence for associations between MIBP exposure and mental and psychomotor neurodevelopment, or behavioural and cognitive function ( Engel et al. 2010; Yolton et al. 2011; Whyatt et al. 2012; Téllez-Rojo et al. 2013).

More recent studies have found associations between DIBP and various endpoints, but these have not yet been assessed using the Downs and Black evaluation approach. There were inconsistent results on the association of MIBP and hormone levels (e.g. estradiol, testosterone, DHEA-S) in both genders (Ferguson et al. 2014a; Meeker and Ferguson 2014; Sathyanarayana et al. 2014; Watkins et al. 2014). There were no associations with MIBP and female puberty (Wolff et al. 2014; Watkins et al. 2014), male puberty (Ferguson et al. 2014a), gene expression in placenta (LaRocca et al. 2014), and preterm birth (Ferguson et al. 2014b). There were inconsistent associations between MIBP and neurobehavioral and cognitive functioning, and psychomotor development (Kobrosly et al. 2014; Polanska et al. 2014; Braun et al 2014).

9.2.2.4 Other systemic effects^Footnote[25]

9.2.2.4.1 Repeated-dose studies

The database for repeated-dose toxicity of DIBP is limited to a few short-term and subchronic oral studies investigating the effects of DIBP on rats and mice. Based on the available data, DIBP is of very low systemic toxicity. Summaries of the studies are described below.
In a short-term study in rat in which females were treated with 0, 50, 100, 200, 2000 mg DIBP/kg-bw/din feed for 14 days, increased changes in liver dodecanoic acid 12-hydroxylase activity and a decrease in serum triglyceride levels were observed at 100 mg/kg bw/day and higher. Increased absolute and relative liver weights, increased serum albumin levels and decreased cholesterol levels were also observed at the highest dose tested. No histopathological changes in the liver were reported (BUA 1998).

When male rats were exposed for one month to 0, 0.01, 0.1, 1, 2 or 5% DIBP in feed (equivalent to 0, 15, 142, 1417, 2975 and 8911 mg/kg bw/day according to US CPSC 2010a), reduced growth was observed in animals exposed at the highest dose (terminal body weight was approximately 62% of controls and 75% of other dose groups). A significant increase in absolute and relative liver weight and a significant increase in relative kidney weight were also observed in animals treated with 1417 and 2975 mg/kg bw/day, respectively. Histological examination did not identify any significant treatment-related lesions in both the liver and the kidneys (Hodge 1953).

In another study, two dogs were fed diets containing DIBP for two months (a male treated with 0.1 mL/kg feed and a female treated with 2.0 mL/kg feed; equivalent to 1 and 16 mg/kg bw/day, respectively, according to US CPSC 2010a). In the female, relative liver weight was increased (in comparison with the relative liver weight of the male), without histopathological lesions. However, due to the limitations of the study (small number of animals tested and lack of concurrent controls or information on historical controls), an effect level could not be derived (Hodge 1954).

In a subchronic feeding study in which male and female albino rats were exposed for 16 weeks (see Section 9.2.2.1.2 and Hodge 1954 for more details on study protocol), a decrease in body weight gain (more than 10%) was observed in males and females treated at the highest dose (males: 5960 mg/kg bw/day; females: 4861 mg/kg bw/day). Terminal body weights were significantly lower than controls among rats treated at the highest dose (decreased 43% for males and 13% for females). Increases in absolute and relative liver weight were also observed in both sexes at the highest dose tested, and absolute and relative testes weights were significantly reduced at that dose level in males. No histopathological changes were noted in the liver and kidneys (histopathology of the other organs was not performed).

In another subchronic study in which four cats were administered 1486 mg DIBP/kg-bw/d by gavage for three months, decreased body weight and food intake, diarrhoea, and emesis were observed. Survival was unaffected, and blood parameters and liver function were unchanged (BASF 1961). However, the lack of details from this study limits the interpretation of these results.

9.2.2.4.2 Carcinogenicity

DIBP has not been classified for its potential carcinogenicity by other international agencies and no chronic toxicity/carcinogenicity studies were identified for this phthalate. There was also no study available in the literature for its closest analogue DBP. However, no pre-cancer effects were noted in animals exposed to a high dose of DIBP in short-term and subchronic studies. Also, in a multigenerational study in which Sprague-Dawley rats (20/sex/group; 40/sex for controls) were given DBP at 0, 0.1, 0.5 and 1.0% in the diet (0, 52, 256 and 509 mg/kg bw/day for males and 0, 80, 385 and 794 mg/kg bw/day for females), the only systemic effect reported in F1 adults (exposed for a significant period of their lifetime) was a decrease in body weight (NTP, 1995*; Wine et al. 1997).

9.2.2.4.3 Genotoxicity

DIBP was found to be non-mutagenic in several bacterial reverse mutation assays using S. typhimurium strains TA 98, TA 100, TA 1535, TA 1537 and TA 1538, with and without metabolic activation (Simmon 1977; Zeiger 1982; Zeiger 1985; Huels AG 1988; Sato 1994). In an 8-azaguanine resistance assay, DIBP was also not mutagenic both in the presence and in the absence of metabolic activation (Seed 1982). However, DIBP induced DNA damage (single-strand breaks) in vitro in a comet assay using human cells (oropharayngeal and inferior nasal turbinate mucosal cells and lymphocytes) (Kleinsasser 2000; Kleinsasser 2001a,b).

9.2.2.4.4 Evidence of systemic toxicity in humans

Available epidemiological studies examining the potential relationship between observed effects and exposure to DIBP in humans were reviewed (Appendix J; Health Canada 2015b). Overall, there was limited evidence for an association between exposure to diisobutyl phthalate (DIBP) and its mono-isobutyl phthalate (MIBP) and diabetes (Lind et al. 2012b; James-Todd et al. 2012; Trasande et al. 2013a). There was inadequate evidence for associations between MIBP and oxidative stress (Ferguson et al. 2011; Ferguson et al. 2012),  cardiovascular function (Lind and Lind 2011; Shiue 2013; Trasande et al. 2013b; Trasande et al. 2014), and obesity (Lind et al. 2012a; Teitelbaum et al. 2012; Trasande et al 2013c; Wang et al. 2013). Inadequate evidence was identified for an inverse association with MIBP and the breast cancer risk (Lopez-Carrillo et al. 2010; Martinez-Nava et al. 2013). No associations were observed between DIBP or MIBP and asthma /allergy related symptoms (Hoppin et al. 2013; Callesen et al. 2014a), or serum levels of thyroid hormones (Meeker and Ferguson 2011).

More recent studies have found associations between DIBP and various endpoints, but these have not yet been assessed using the Downs and Black evaluation approach. Significant associations were found between MIBP and biomarkers of diabetes (Huang et al. 2014) and obesity (Christensen et al. 2014b; Buser et al. 2014). Callesen et al. (2014b) reported no associations with MIBP and asthma, atopic dermatitis or rhinoconjunctivitis. However, Bamai et al. (2014) found associations between atopic dermatitis and DIBP in floor dust, but not in multi-surface dust.  Significant association was reported between MIBP and blood pressure in women (Shiue and Hristova 2014), but no associations were reported considering both genders together (Shiue 2014a,b; Shiue and Hristova 2014), and considering men only (Shiue and Hristova 2014). MIBP was associated with oxidative stress, but not with inflammation (Ferguson et al. 2014d). No association was found between MIBP and osteoporosis (Min and Min 2014).

9.2.3 DCHP

9.2.3.1 Reproductive and developmental effects in males

9.2.3.1.1 Early development: in utero exposure

A literature search identified four studies examining the potential toxicity of DCHP during gestation in rats, all focusing on male reproductive effects during the masculinization programming window (GD15-17) where any potential anti-androgenic effects would be observed. The studies are described in Table 9-23 below. No other developmental studies were identified examining gestational exposure to DCHP in other species.

In a 2-generation toxicity study, DCHP was associated with effects in the parental (F0) and both filial generations (F1 and F2) in a dose-responsive manner. In this study, male and female rats of both F0 and F1 generations were administered 0, 240, 1200 and 6000 ppm DCHP in the diet for greater than or equal to 10 weeks of pre-mating and mating periods (see Table 9-23). Estimated dose levels for the F0 generation males were 0, 16, 80 and 402 mg/kg bw/day and 0, 18, 90 and 457 mg/kg bw/day for the F1 generation males. Developmental effects were observed in F1 and F2 male pups. There was a statistically significant decrease in AGD measured in length as well as when adjusted by body weight and retained nipples/areolae (NR) in male F2 offspring treated at 1200 ppm or higher (107 mg/kg bw/day based on amount ingested by F1 dams). In male F1 offspring, these effects were observed at the highest dose only. Foetal body weight was reduced in F1 and F2 offspring at the highest dose (Hoshino et al. 2005). Slight maternal toxicity was observed at 1200 and 6000 ppm based on reduced body weight gain and reduced food consumption and diffuse hypertrophy of hepatocytes (identified as statistically significant at P less than 0.05, but less than 10% reduction in food consumption and body weight during gestation) in F0 females (LOAEL for maternal toxicity was estimated at 104 mg/kg bw/day). As other general physical development parameters, such as eye opening, pinna unfolding, and incisor eruption, were not affected, the reproductive/developmental effects of AGD and NR in F1 male pups at the high dose level were not considered to be a secondary effect of maternal toxicity.

In the F1 generation, reproductive effects observed when F1 animals reached reproductive age included a 15 and 24% decrease in spermatid head counts in the testes and testicular atrophy in F1 males treated at the two highest doses, 90 (2 of 20)  and 457 (9 of 14) mg/kg bw/day, respectively. At the highest dose, 3 males had small and/or soft testes, with one male examined showing no sperm. Relative prostate weight was also decreased in high dose F1 males. Sex ratio, number of implantations, mating, fertility and birth indices did not differ from the control, and there were no changes in serum testosterone levels in this generation when examined at the adult stage (Hoshino et al. 2005). Reduced body weight gain and reduced food consumption were observed in F1 males at 90 mg/kg bw/day (Hoshino et al. 2005).

A developmental study conducted by Saillenfait et al. (2009) observed a dose-dependent decrease in AGD in male neonates of dams exposed to DCHP during gestation (GD6-20) at the lowest dose tested and above (250, 500 and 750 mg/kg bw/day; gavage). No effects on testicular descent (i.e., cryptorchidism) were identified. The body weights of males were decreased at the highest dose (750 mg/kg bw/day), but this dose was also associated with maternal toxicity and decreased food consumption. There was no evidence of teratogenicity or embryolethality at any of the maternal treatment doses. The LOAEL for maternal toxicity in this study was 750 mg/kg bw/day based on a significant reduction in body weight gain in dams. A LOEL for maternal toxicity of 500 mg/kg bw/day was identified based on an increase in absolute and relative maternal liver weight at the two highest doses tested with no histological findings.

In a study in which pregnant rats were exposed by gavage to 0, 20, 100 and 500 mg DCHP/kg-bw/d during GD6 to PND20, effects on the reproductive system, such as prolonged preputial separation, reduced AGD, increased areola/nipples retention, hypospadia, decreased relative weights of the ventral prostate and ani/bulbocavernosus muscles, and histological changes in the testis and kidney (decreased testicular germ cells and degenerated renal proximal tubules) were observed in animals treated at the highest dose tested. The authors attributed these findings to the anti-androgenic effects of DCHP. A decrease in body weights and a slight but significant reduction in the viability index were also observed in male pups at this dose level. A NOAEL of 100 mg/kg bw/day and a LOAEL of 500 mg/kg bw/day for developmental toxicity were determined from this study. A LOEL of 100 mg/kg bw/day was identified in this study for slight maternal toxicity based on increased absolute and relative liver weights in dams at 100 and 500 mg/kg bw/day (6 and 19% compared to controls, respectively) (Yamasaki et al. 2009). No histopathological observations of the liver in dams or general physical development parameters, such as eye opening, pinna unfolding, and incisor eruption in pups were reported, limiting the interpretation of maternal health on developmental outcomes.

In another developmental study, Ahbab and Barlas (2013) administered 0, 20, 100 or 500 mg/kg bw/day DCHP to pregnant Wistar Albino rats via gavage on GD6-19 (Table 9-23). The results from this study are limited based on the dosing method, but histopathological evidence of reproductive tract malformations (in the testis, epididymis and prostate gland) and a significant increase in abnormal sperm were observed in male pups exposed to DCHP at all dose levels, although the sperm effects did not follow a dose-responsive trend. Typical dose response relationships were not observed for many metrics, and maternal health was not reported (Ahbab and Barlas 2013).

A more recent study presenting the potential for DCHP and other phthalates to alter foetal testosterone production (ex vivo) in pregnant SD rats showed that this phthalate altered testicular testosterone production during gestation at doses of 100 mg/kg bw/day and above with a calculated ED₅₀ value of 61.6 mg/kg bw/day (Furr et al. 2014).

Overall, the lowest NOAEL for developmental toxicity identified for DCHP was 21 mg/kg bw/day based on reduced AGD and increased areola mammae retention (NR) at the next highest dose of 107 mg/kg bw/day (1200 ppm) and above in a 2-generation study (Hoshino et al. 2005) conducted according to OECD guidelines. The dose level of 16-21 mg/kg bw/day from this study was also identified as the NOAEL for developmental effects by the Australian NICNAS (2008) and the US CPSC CHAP (2014). Foetal testicular testosterone production was also altered at similar dose levels (NOAEL of 33 mg/kg bw/day; Furr et al. 2014). The lowest LOAEL for maternal toxicity of DCHP was 104 mg/kg bw/day based on reduction in body weight gain in exposed F0 dams in the same study. Developmental effects of AGD and NR in male pups were not considered to be secondary to maternal toxicity.

9.2.3.1.2 Exposure at prepubertal/pubertal life stage

There was only one repeated-dose oral exposure study in sexually immature animals (PND1-55) with DCHP where most parameters related to RPS were not examined and males were only observed for reproductive tract malformations at high doses (Lake et al. 1982).

After a seven-day oral gavage exposure of DCHP to young PND30 male SD rats, testes sections from control and 1500 mg/kg/d DCHP animals had no abnormalities. However, examination of one of the five animals treated with 2500 mg/kg/d of DCHP exhibited a bilateral tubular atrophy of 30 to 40% of the germinal cells of the testes. There were no effects on testes weight in any of the high-treatment groups (1500 and 2500 mg/kg bw/day). In contrast to treatment with DCHP, administration of 1130 mg/kg bw/day with one of its metabolites, MCHP, resulted in a significant reduction in relative testes weight to 44% of control values. Morphological examination showed an almost complete bilateral atrophy of the germinal epithelium of the seminiferous tubules (Lake et al. 1982).

Overall, the lowest NOEL for the reproductive toxicity of DCHP at the prepubertal/pubertal life stage was based on a limited study where effects on the male reproductive system (bilateral tubular atrophy in 1 out of 5 animals) were observed at 2500 mg/kg bw/day. It will therefore not be used to characterize risk for this life stage based on this limitation (Lake et al. 1982). No studies were identified on any other species via any route of exposure at this life stage.

9.2.3.1.3 Oral exposure at the mature male adult stage

Using the 2-generation toxicity study described in the previous section (Hoshino et al. 2005), information was extracted to determine the effects of DCHP on the adult male (PND55+). Summaries of the studies are described in Table 9-24 below.

DCHP was predominantly associated with slight systemic toxicity in the parental (F0) animals. In parental animals treated at the highest dose, effects included reduced body weight gain, increased liver and thyroid weights, as well as increased hyaline droplets in the proximal tubules of males (Table 9-24). A LOAEL for systemic toxicity was identified based on diffuse hypertrophy of hepatocytes (identified as a "slight" effect) in F0 males (and females) and increased incidence of thyroid follicular cell hypertrophy (also identified as a "slight" effect) in F0 males at 1200 and 6000 ppm (80 and 402 mg/kg bw/day). Sex ratio, number of implantations, mating, fertility, and birth indices did not differ from the control, and there were no effects on testes or sperm parameters in F0 males.

The highest NOAEL for reproductive toxicity identified for DCHP was 402 mg/kg bw/day based on appearance of one F0 male with slight focal seminiferous tubule atrophy in a 2-generation OECD guideline study (Hoshino et al. 2005). There was evidence of systemic effects at this dose (decreased body weight gain). The lowest LOEL for systemic toxicity for males was 80 mg/kg bw/day based on slight hypertrophy of liver and thyroid in F0 males (Hoshino et al. 2005). The lowest LOAEL for systemic toxicity for F0 females was 104 mg/kg bw/day based on decreased body weight gain and increased relative liver weights with slight hypertrophy (Hoshino et al. 2005).

9.2.3.2 Oral exposure in females

Five published studies on the reproductive and developmental effects of DCHP in females were identified. These studies were performed in rats exposed to DCHP before mating only, during gestation (GD6-20, GD6-PND20) or continuous breeding via feed or gavage. One 2-generation study was identified.

The lowest LOAEL identified for developmental toxicity in females is 402 to 534 mg/kg bw/day (6000 ppm [0.6%] in the diet) derived from the 2-generation study described in Section 9.2.3.1.1 above (Hoshino et al. 2005). Developmental toxicity occurred at doses inducing maternal toxicity in the F1 generation. The critical endpoints included altered growth (statistically significant decrease in body weight and altered organ weights at 402-534 mg/kg bw/day in F1 and F2 pups with a NOAEL of 80-107 mg/kg bw/day, 1200 ppm). Maternal effects (F0) included a statistically significant decrease in body weights and food consumption. A statistically significant increase in diffuse hypertrophy of hepatocytes and in relative liver weight occurred at 104 mg/kg bw/day and above (1200 ppm; NOAEL of 21 mg/kg bw/day, 240 ppm).

The lowest LOAEL identified for reproductive toxicity in adult females is 511 mg/kg bw/day (6000 ppm [0.6%] in the diet) in F0 parents in the same study (Hoshino et al., 2005). The critical endpoint was based on altered fertility and pregnancy (statistically significant decrease in F1 body weight at 511 mg/kg bw/day, 6000 ppm with a NOAEL of 104 mg/kg bw/day, 1200 ppm). A prolongation of the estrous cycle was also reported at this dose. However, it was not considered as a direct effect of DCHP on the endocrine system, but rather on the suppression of body weight gain (observed at 104 mg/kg bw/day and above). In F1 parents, no reproductive adverse effect was identified.

Overall, the few studies on the reproductive and developmental effects of DCHP in females have indicated no evidence of teratogenicity or embryolethality. Developmental toxicity (growth alterations [organs and body weights] and lethality) and reproductive toxicity (pregnancy outcome alterations) were reported at high doses (500 mg/kg bw/day and above). When gender effects were examined separately, DCHP-induced developmental effects in males and females were observed at the same dose levels, with some studies reporting males as more sensitive than females.

9.2.3.3 Endocrine studies

In vitro studies have been conducted to examine the potential effects of DCHP on the endocrine system including steroidogenesis in mammalian systems and are related to measurements of reproductive hormones and glucocorticoid effects.

With respect to the potential mechanisms for reproductive effects, studies on estrogen receptor (ER) and androgen receptor (AR) binding by DCHP have been identified. DCHP appeared to be an inducer of MCF-7 cell proliferation in a breast cancer cell line assay for ER activation (Okubo et al. 2003), but this result was not reproducible in vivo (Hong et al. 2005). In a β-Galactosidase Activity Assay (which can measure estrogenic, anti-estrogenic, androgenic and anti-androgenic activity in Chinese Hamster Ovarian [CHO] cells), DCHP was shown to have estrogenic activity with one form of ER (ERα), but no estrogenic and possibly anti-estrogenic effects with ERβ (Takeuchi et al. 2005). DCHP was shown to be estrogenic in the Yeast Two-Hybrid Assay (Nishihara et al. 2000). Furthermore, DCHP appeared to show anti-androgenic effects on the AR. In a competitive binding assay for ER and AR, DCHP had low binding affinity for AR and ERα, but had partial inhibitory effects on binding to AR (45% inhibition at 10^-4 M DCHP) and ERα (IC₅₀ at 5.8 x 10^-8 M) (Satoh et al. 2001). DCHP was also shown to have inhibitory effects on two enzymes that are involved in testosterone production. These are 3β-hydroxysteroid dehydrogenase (HSD) and 17β-HSD3 (Yuan et al. 2012).

Other studies investigated the potential effect of DCHP on corticosteroid production. DCHP was shown in microsomal assays to inhibit 11β-HSD2 (Oshima et al. 2005; Zhao et al. 2010), which is involved in the inactivation of cortisol. This inactivation could result in mineralocorticoid excess, with systemic symptoms similar to pseudoaldosteronism. Zhao et al. (2010) speculated that the inhibition of 11β-HSD2 could also have implications on Leydig cell cortisol levels, resulting in higher tissue levels of cortisol, which could subsequently result in a decrease in testosterone production. Conversely, DCHP was shown to inhibit dibutyryl cAMP-induced cortisol secretion from H295R, an adrenal cell line that serves an in vitro model for human steroidogenic cells (Nakajin et al. 2001). DCHP has also been shown to bind to the glucocorticoid receptor (GR), but results were mixed with respect to whether this binding was associated with changes in adipocyte differentiation (Sargis et al. 2010). Liu et al. (2002) and Lu et al. (2004) evaluated the effect of DCHP on neuroendrocrine processes. DCHP suppressed Ca+2 release through the nicotinamide acetyl choline receptor (nAChR) of bovine adrenal chromaffin cells (Liu et al. 2002). This result was confirmed in a human cell line (SH-SH5Y), and the effect of DCHP on this process in SH-SH5Y was ten times greater than for estradiol (Lu et al. 2004).

There have been in vitro studies on the potential immunological effects of DCHP. One study (Ohnishi et al. 2008) investigated whether phthalate exposure could increase susceptibility to infection, but showed no adverse effect on microphages incubated with 100 μM DCHP. Estrogen-receptor mediated effects on acquired immunity were investigated by Yano et al. (2003). DCHP inhibited mouse spleen cell production of Type 1 helper T-cells (Th1) and Type 2 helper T-cells (Th2); however, this effect did not appear to be mediated by an estrogen receptor.

9.2.3.4 Reproductive and developmental toxicity: evidence in humans

Available information on the potential effects of DCHP on humans was reviewed, rated and assessed for human health risk (Appendix J; Health Canada 2015b). No associations were established for reproductive paramaters such as time to pregnancy (Buck Louis et al. 2014).

9.2.3.5 Other systemic effects^Footnote[26]

9.2.3.5.1 Repeated-dose studies

The database for repeated-dose toxicity of DCHP is limited to a few short-term and subchronic oral rat studies identified in the literature. The available health effects information for DCHP is summarized below. In a gavage study in rats, no effects were reported when animals were exposed to 200 mg DCHP/kg-bw/d, twice a week for six weeks (Bornmann 1956). In another gavage study in male rats, a LOAEL of 500 mg/kg bw/day (the lowest dose level) was identified. In this study, the animals were administered 0, 500, 1000, 1500, 2000 and 2500 mg DCHP/kg-bw/d for a shorter time period (7 days). The LOAEL was based on a dose-related increase in liver weight and induction of hepatic enzymes in treated animals. Histopathological examination (limited to the liver, kidney and testes of animals treated with 1500 and 2500 mg/kg bw/day) revealed slight hepatic centrilobular hypertrophy at those dose levels, as well as marked proliferation of the smooth endoplasmic reticulum. There was no evidence of peroxisome proliferation. The authors characterize the induction of xenobiotic metabolism observed in this study as weak, drug-type induction, different from the peroxisomal proliferation seen with DEHP (Lake et al. 1982). In a single dose, 21-day feeding study, a LOAEL of 4170 mg/kg bw/day was determined based on a number of toxic effects in treated animals, including testicular atrophy, liver enlargement, alopecia and stomach squamous cell hyperplasia. However, this study is limited since little information is available on the number of animals treated, the species or study design (Grasso et al. 1978).

A subchronic feeding study in which rats were exposed to 0, 0.05, 0.15, 0.4 or 1% (corresponding to doses of 0, 25, 75, 200 and 500 mg/kg bw/day) DCHP for 90 days yielded a LOAEL of 75 mg/kg bw/day based on increased relative liver weight in female rats. An increase in relative liver weight was observed in males only from 200 mg/kg bw/day. Those increases in liver weight were accompanied by histological changes in the liver and kidneys in both sexes at the two highest doses tested. An increase in serum alkaline phosphatase levels was observed in male rats exposed to doses of 25 mg/kg and above and in females at the highest dose. Decreased body weight gain and food consumption in males was noted at the highest dose. No mortality or clinical signs of toxicity were observed (de Ryke and Willems 1977).

In these studies, the lowest LOAEL for repeated-dose oral exposure was 75 mg/kg bw/day based on increased relative liver weight in female rats, accompanied by histological changes in the liver at exposure to a dose of 200 mg DCHP/kg-bw/day and higher.

9.2.3.5.2 Carcinogenicity

DCHP has not been classified for its potential carcinogenicity by other international agencies.

No effects were reported in rats fed DCHP at 27 mg/kg bw/day for two years or in dogs treated with 14 mg/kg bw/day for a year in their diet. No further details were available (Shibko and Blumenthal 1973). No effects were reported among rats exposed by gavage to 0.5 or 1 ml/kg of a preparation containing 25% DCHP in olive oil (approximately 100 or 200 mg DCHP/kg-bw per day) twice weekly for up to 52 weeks (Bornmann 1956). In an 18-month study in the Wistar rat, no carcinogenicity or changes in body weight were noted at low, medium or high exposure doses when compared with the control. The highest dose tested was estimated at 5 mg/kg bw/day (Lefaux 1968).

Based on available information, there is no indication that DCHP is a potential carcinogen. Lake et al. (1982) concluded that DCHP does not appear to be a peroxisome proliferative agent, as the authors did not identify an increase in the number of peroxisomes or any changes in mitochondrial structure or function.

9.2.3.5.3 Genotoxicity

As part of the National Toxicology Program's Environmental Mutagenesis Test Development Program, DCHP was tested for mutagenicity in the Ames test with and without metabolic activation (NTP 1983; Zeigler 1985). Salmonella typhimurium (S. typhimurium) strains TA1535, TA1537, TA98 and TA100 were tested either as-is or with the inclusion of the S-9 fraction of liver homogenate from Aroclor 1254 exposed Sprague-Dawley rats. The Ames test results were all negative for DCHP. The Ames test (in S. typhimurium TA 1535, TA 1537, TA 98 and TA 100), the E. coli DNA repair assay and the mesenchymal fibroblast-like cell transformation assay for the plasticizer Nuoplax 6938, which is composed of 61.2% BCHP, 21.9% DBP, 15.2% DCHP and 1.7% DMP, were also all negative (with and without metabolic activation) (Nuodex 1982a,b). No in vivo studies have been identified in the literature.

9.2.3.5.4 Evidence of systemic toxicity in humans

Available information on the potential effects of DCHP on humans was reviewed, rated and assessed for human health risk (Appendix J; Health Canada 2015b). There was inadequate evidence of an association with MCHP (metabolite of DCHP) and obesity in children and adolescents (Wang et al. 2013). No associations were established for cardiovascular function (Shiue 2013; Trasande et al. 2014).

9.2.4 DMCHP

9.2.4.1 Reproductive and developmental effects of DMCHP in males

No studies examining the potential health effects of DMCHP were identified for any species or gender. DCHP (1,2-Benzenedicarboxylic acid, dicyclohexyl ester: CAS RN 84-61-7) was identified as the "closest analogue" to DMCHP based on similarity in the nature of the ester chains (Section 2.3.2; Health Canada 2015a). Ester groups of DMCHP and DCHP both consist of cyclohexyl chains, with DMCHP having an additional methyl group on the cyclohexyl ring. Due to structural similarities, the physical-chemical properties for DMCHP and DCHP are also similar. Refer to Section 9.2.3 for summaries of the studies using DCHP for all life stages.

9.2.5 CHIBP

No studies examining the potential reproductive/developmental health effects of CHIBP were identified for any species or gender. DIBP (1,2-Benzenedicarboxylic acid, bis(2-methylpropyl) ester: CAS RN 84-69-5) and DCHP (1,2-Benzenedicarboxylic acid, dicyclohexyl ester: CAS RN 84-61-7) were identified as the "closest analogue" phthalates to CHIBP within the subcategory based on consideration of similarities in monoester metabolism as well as the length and nature of the ester chains (Section 2.3.2; Health Canada 2015a). CHIBP is expected to yield monoester metabolites identical to the monoester metabolites of DIBP and DCHP.

Based on health effects information on the analogues DIBP and DCHP, a potential health effect of concern may be associated with CHIBP. A review of the potential developmental and reproductive toxicity of the analogue(s) showed that this medium-chain phthalate could have adverse effects on the reproductive system of the developing male, in addition to systemic effects (liver, kidney).

Given the absence of reporting to the section 71 industry survey, non-detection in dust, negligible modelled indoor air concentrations, and the absence of information as to CHIBP presence in product databases, general population exposure to CHIBP from environmental media and prodcuts used by consumers is expected to be negligible. Therefore, risk to human health for this substance is not expected.

9.2.6 BCHP

No studies examining the potential reproductive/developmental health effects of BCHP were identified for any species or gender. DBP (1,2-Benzenedicarboxylic acid, dibutyl ester: CAS RN 84-74-2) and DCHP (1,2-Benzenedicarboxylic acid, dicyclohexyl ester: CAS RN 84-61-7) were identified as the "closest analogue" phthalates to BCHP within the subcategory based on consideration of similarities in monoester metabolism as well as the length and nature of the ester chains (Section 2.3.2; Health Canada 2015a).

Based on health effects information on the analogues DBP and DCHP, a potential health effect of concern may be associated with CHIBP. A review of the potential developmental and reproductive toxicity of the analogue(s) showed that this medium-chain phthalate could have adverse effects on the reproductive system of the developing male, in addition to systemic effects (liver, kidney).

Given the absence of reporting to the section 71 industry survey, non-detection in dust and products (emission chamber study), and the absence of information as to BCHP presence in product databases, general population exposure to BCHP from environmental media and products used by consumers is expected to be negligible. Therefore, risk to human health for this substance is not expected.

9.2.7 DBzP

9.2.7.1 Reproductive and developmental effects in males

9.2.7.1.1 Early development: in utero exposure

No studies examining the potential reproductive/developmental health effects of DBzP were identified for any species or gender. MBzP (1,2-Benzenedicarboxylic acid, mono[phenylmethyl] ester: CAS RN 2528-16-7) is the monoester hydrolysis product of DBzP. Ortho phthalates are generally known to be rapidly absorbed following oral exposure, and the diester is cleaved into one or more monoesters in the digestive tract. Monoesters are generally considered to be responsible for the health effects of the parent compound (Health Canada 2015a). MBzP is the monoester hydrolysis product of DBzP and is therefore suitable for inferring toxicity in oral developmental studies. Toxicological studies conducted with MBzP were examined to characterize the health effects of DBzP. Summaries of the studies are described in Table 9-25 below.

Several studies examined the potential for MBzP to induce developmental effects in rodents; however, only one study was performed during gestation in rats focusing on the male programming window (GD15-17) where any potential anti-androgenic effects would be observed (see Table 9-25 below). Pregnant Wistar rats were given MBzP via gavage at doses of 167, 250 or 375 mg/kg bw/day on GD15-17 of pregnancy, and offspring were examined on GD21 (Ema et al. 2003). Developmental effects included significant increases in the incidence of CRY, decreases in AGD and the ratio of AGD to the cubic root of body weight in male foetuses at 250 mg/kg bw/day and higher as well as significantly decreased foetal weight at 375 mg/kg bw/day (also described in Table 9-25). However, a significant and dose-dependent decrease in maternal body weight gain (22%) and food consumption (8-15%) was also noted at the lowest dose (167 mg/kg bw/day) and higher.

Saillenfait et al. (2003) evaluated the embryotoxic effects of MBzP in OF1 mice and Sprague-Dawley rats on GD8 and GD10, respectively. In mice, maternal deaths occurred, and maternal body weight gain (statistically significant at the highest dose, 1380 mg/kg bw/day) was reduced along with the corresponding developmental effects in offspring (embryolethality and teratogenicity). In rats, MBzP did not cause significant developmental effects up to doses that produced maternal mortality and/or weight loss (1380 mg/kg bw/day). Previously, Ema et al. treated rats during GD7-15 with MBzP via gavage at dose levels of 250, 313, 375, 438 and 500 mg/kg and reported skeletal malformations at doses equal to or higher than 313 mg/kg as well as increased incidences of post-implantation loss at the two highest doses (1996a). In stage-sensitivity studies by the same authors, teratogenic effects were observed in rats given MBzP (250-625 mg/kg) on GD7-9 and GD13-15. MBzP also caused a dose-related increase in the incidence of resorptions regardless of the periods of administration (Ema et al. 1996b). Maternal effects (reductions in maternal body weight with corresponding reduced food consumption) were observed at lower or equal doses compared to those at which foetal effects occurred, in both studies (see Table 9-25 for a summary).

Overall, the highest oral NOAEL for developmental toxicity of MBzP at the in utero life stage was 167 mg/kg bw/day based on increased incidence of CRY and decreased AGD in male foetuses at 250 mg/kg bw/day and higher as well as significantly decreased foetal weight at 375 mg/kg bw/day (Ema et al. 2003). The lowest dose, and above, in this study also caused slight maternal toxicity, as seen by decreased food consumption and body weight gain, with no evidence of embryolethality such as number of corpora lutea, implantations, resorptions, and dead fetuses, the incidence of post-implantation loss per litter, or the sex ratio of live fetuses  (LOAEL of 167 mg/kg bw/day). Therefore, the NOAEL of 167 mg/kg bw/day is considered the critical effect level for the developmental toxicity of DBzP for this life stage.

9.2.7.1.2 Exposure at prepubertal/pubertal life stage

There were no repeated-dose oral exposure studies in sexually immature animals (PND1-55) with DBzP via any route of exposure identified in the literature. As with the previous section, studies conducted with MBzP were reviewed to characterize the health effects of DBzP (Health Canada 2015a).

To determine the potential reproductive toxicity of phthalate diesters and monoesters on sperm parameters in young male rats, Kwack and colleagues (2009) orally exposed six 5-week-old Sprague-Dawley rats to 250 mg/kg bw/day of MBzP for four weeks. No adverse effects were found on reproductive organ weights. The authors reported that MBzP significantly lowered the sperm counts (20% decrease compared to controls) and increased sperm motility (VCL) at 250 mg/kg bw/day (Kwack et al. 2009).

Overall, the LOEL for the reproductive toxicity of MBzP at the prepubertal/pubertal life stage was 250 mg/kg bw/day based on decreased sperm counts after four weeks of exposure (Kwack et al. 2009). No studies were identified on any other species via any route of exposure for this life stage. Therefore, the LOEL of 250 mg/kg bw/day will be used as the critical effect level for the reproductive toxicity of DBzP for this life stage.

 9.2.7.1.3 Oral exposure at the mature male adult stage

No studies examining the potential reproductive toxicity of DBzP at the adult male life stage were identified. There were no studies available where MBzP was administered to adult males starting after PND55. Therefore, the effects observed in the MBzP study examining pubertal animals will be used to characterize the reproductive toxicity of DBzP for the adult life stage as administration of MBzP continued into adulthood (PND63) in these animals (Kwack et al. 2009).

9.2.7.2 Oral exposure in females

Five published studies on the reproductive and developmental effects of MBzP in females were identified. Four studies were performed in rats and one in mice, in which pregnant animals were exposed via gavage to MBzP during gestation.

A group of researchers (Ema et al.) conducted studies at different gestation times (GD7-15, GD7-9, GD13-15) in order to determine the most sensitive period. A study was also conducted in female mice and rats exposed during their period of neurulation (GD8 and GD10, respectively) in order to determine the most sensitive species.

The lowest LOAEL for developmental toxicity in females (313 mg/kg bw/day) was identified in a study in which pregnant Wistar rats were exposed by gavage to MBzP (0, 250, 313, 375, 438 and 500 mg/kg bw/day) during the whole period of organogenesis (GD7-15) (Ema et al., 1996a). The effects occurred at a dose at which maternal effects were also observed and included a statistically significant increase in skeletal malformations (primarily cleft palate, dilated renal pelvis and fusion of ribs, cervical and/or thoracic vertebral arches) at 313 mg/kg bw/day and above. A NOAEL of 250 mg/kg bw/day was identified. Maternal effects were reported as a statistically significant and dose-dependent decrease in food consumption at 250 mg/kg bw/day and above, the lowest dose tested.

The lowest LOAEL identified for the reproductive toxicity in adult females (438 mg/kg bw/day) is based on the rat study where there were observations of increased post-implantation loss along with decreased food consumption after gavage treatment of MBzP at GD7-15 (NOAEL of 375 mg/kg bw/day) (Ema et al., 1996a). One study by Zhang et al. (2011) examined the potential estrogenic effects of DBzP using an in vitro yeast estrogen screen (YES) and in vivo immature mouse uterotrophic assay (three-day exposure via oral gavage). In the uterotrophic assay, DBzP significantly inhibited the effects of E2 at the high dose (400 µg/kg-bw/d) and low dose (40 µg/kg-bw/d) (P less than 0.05), which demonstrated its strong estrogenic antagonistic ability. The authors compared these results to those using BBP. In an in vivo uterotrophic assay with BBP (2240 mg/kg bw/day and above), uterine growth was not promoted in immature females (ECJRC 2007), whereas the results obtained by Zhang et al. (2011) showed that DBzP did effect uterine growth and suggested that the (in vivo) estrogenic potency of DBzP is higher than that of BBP. In the YES assay, DBzP inhibited the agonist activity of 10^-9M E2 at 1.95×10^-6M and above. Similar results were shown for BBP in this study (the EC₅₀ value of DBzP was 8.06×10^-6 M, slightly lower than BBP at 1.17×10^-5 M) (Zhang et al. 2011).

Overall, the studies related to the reproductive and/or developmental effects of MBzP in females indicate that MBzP is teratogenic and embryolethal (313 mg/kg bw/day only at doses which also cause maternal toxicity. The gestational age at the time of exposure is critical to the teratogenic effects of MBzP. Some gender-related differences were reported (male pups more sensitive than female pups).

9.2.7.3 Reproductive and developmental toxicity: evidence in humans

No information is currently available on the potential reproductive/developmental effects of DBzP in humans.

9.2.7.4 Other systemic effects^Footnote[27]

9.2.7.4.1 Repeated-dose studies

No repeated-dose studies have been identified in the literature for DBzP. There was one repeated-dose study available for the closest analogue MBzP.

In a short-term study in which six 5-week-old Sprague-Dawley rats were orally exposed to MBzP at 250 mg/kg bw/day for four weeks, no adverse effects were found on body weight gain, food consumption or relative organ weights, or hematology measurements. Some serum parameters (glucose by ~25%, glutamate oxaloacetate transaminase by ~40%) were found to be significantly higher from controls. Leukocyte counts were also changed (data not shown) (Kwack et al. 2009).

Systemic toxicity has been reported in developmental toxicity studies in which rodents were exposed orally to this substance (Ema et al. 1996ab, 2003; Saillenfait et al. 2003). The common systemic effect reported was reduction in maternal body weight gain and, in one study, maternal death was reported at the highest dose tested (1380 mg/kg bw/day) (Saillenfait et al. 2003). The lowest LOAEL for short-term exposure was 167 mg/kg bw/day based on a dose-dependent decrease in body weight gain (22% decrease for adjusted weight gain) associated with a decrease in food consumption (8-15%) in dams in a developmental toxicity study in rats (Ema et al. 2003, as described in Section 9.2.7.1).

9.2.7.4.2 Carcinogenicity

DBzP has not been classified for its potential carcinogenicity by other international agencies, and no chronic toxicity/carcinogenicity studies were available for this phthalate. There was also no study available for the closest analogue MBzP.

9.2.7.4.3 Genotoxicity

No genotoxicity studies were identified for DBzP or its closest analogue MBzP.

9.2.7.4.4 Evidence of systemic toxicity in humans

No information has been identified on the potential effects of DBzP in humans.

9.2.8 B84P

 Reproductive and developmental effects in males

Early development: in utero exposure

No studies examining the potential reproductive/developmental health effects of B84P were identified for any species or gender. DIBP (1,2-Benzenedicarboxylic acid, bis[2-methylpropyl] ester: CAS RN 84-69-5), BBP (1,2-Benzenedicarboxylic acid, butyl phenylmethyl ester: CAS RN 85-68-7) and MBzP (1,2-Benzenedicarboxylic acid, mono[phenylmethyl] ester: CAS RN 2528-16-7) were identified as the "closest analogue" phthalates to B84P within the subcategory based on consideration of similarities in monoester metabolism (Section 2.3.2; Health Canada 2015a). The health effects of DIBP and MBzP have been characterized in sections 9.2.2.1 and 9.2.7.1 above.

The European Commission classified BBP as Category 2 (causes developmental toxicity in humans) Risk phrase R61 (may cause harm to unborn child) for developmental toxicity and as Category 3 (causes concern for human fertility) Risk phrase R62 (possible risk of impaired fertility) for reproductive toxicity (ECHA 2008). Subsequent changes to the classification schemes for the hazard class within the European Union Classifying, Labelling, and Packaging (CLP) regulations (EC No 1272/2008) resulted in a change in the status of BBP to Category 1B - reproductive toxicant (presumed human reproductive toxicant).

A literature search identified many studies examining the effects of BBP during gestation in rodents. For the purpose of characterizing effects during early male development, only studies in rats in which effects of BBP were observed at doses at and below 500 mg/kg bw/day following in utero exposure during the masculinization programming window are reported here. Summaries of the studies are described below and in Table 9-26.

Overall, adverse effects in the parameters used to measure RPS in male rat offspring after in utero exposure to BBP include decreased testicular testosterone levels, delayed preputial separation (PPS), AGD, NR, CRY, gross and testicular malformations, and effects on fertility.

The dose level at which developmental effects were first observed after gestational exposure to BBP appeared to be somewhat consistent across studies. A decrease in male rat offspring body weights, but no significant change in body weight gain, was observed at 100 mg/kg bw/day and above at birth in F1 and/or F2 pups in two separate 2-generation studies (Aso et al. 2005; Nagao et al. 2000). Decreased pup weight was also observed at higher doses in other studies (see Table 9-26) (Ema et al. 1990; Piersma et al. 1999; Tyl et al. 2004).

Effects related to feminization parameters associated with RPS have been reported starting as low as 100 mg/kg bw/day, where BBP-induced effects on AGD at birth (PND1-4) were observed in F2 offspring in rats after oral exposure in a 2-generation study (Aso et al. 2005). Other studies have also shown reduced AGD in newborn male pups at 250 mg/kg bw/day and above (Nagao et al. 2000; Hotchkiss et al. 2004; Tyl et al. 2004; Liu et al. 2005). Nipple retention (NR) in young males, when measured at one to two weeks after birth, was observed at doses as low as 500 mg/kg after short gestational exposure (GD14-18), although it was not statistically significant (Hotchkiss et al. 2004). This effect became statistically significant in a 2-generation study at higher doses when measured on PND11-13 (750 mg/kg bw/day) (Tyl et al. 2004).

Delays in PPS have been reported at doses as low as 400 mg/kg bw/day and above in F1 males in three separate 2-generation studies using BBP. It is interesting to note that a short gestational exposure (GD14-18) did not elicit delays in this endpoint at similar doses (500 mg/kg bw/day) (Hotchkiss et al. 2004).

Reproductive tract malformations, such as CRY and HYP, were reported after BBP gestational exposure at higher dose ranges. The lowest dose at which significant incidences of CRY were reported was 580 mg/kg bw/day and above after short (GD5-20) exposure to BBP (Piersma et al. 1999 in NICNAS 2008). It is of interest to note that a number of studies examining CRY in short and longer (2-generation) exposures to 500 mg/kg bw/day of BBP during gestation did not observe significantly increased occurrences of this malformation (Nagao et al. 2000; Hotchkiss et al. 2004). A separate 2-generation study by Tyl et al. (2004) noted male pups with undescended testes at 750 mg/kg bw/day in both F1 and F2 generations in the presence of maternal toxicity. Gray et al. (2000) also observed CRY at this dose level after gestational exposure (GD14-PND3). See Table 9-26.

Gross malformations of the penis (HYP) appeared to follow a somewhat similar pattern as CRY, where there were no significant incidences of HYP at 500 mg/kg and below. They did become evident at 750 mg/kg bw/day but only in F1, not F2 pups in the presence of maternal toxicity (see Table 9-26) (Nagao et al. 2000; Tyl et al. 2004). Gray et al. (2000) also observed HYP at this dose level after gestational exposure (GD14-PND3).

A decrease in postnatal relative testis weights was reported at doses of BBP as low as 270 mg/kg bw/day, the lowest dose tested, and above after gestational exposure to BBP from GD5-20 (Piersma et al. 1999 in NICNAS 2008). No histopathological effects were reported and/or measured. No changes in testes weight were observed in a separate short (GD14-18) exposure to 500 mg/kg bw/day BBP, but a decrease in relative levator ani+bulbocavernosus muscle (LABC) weight (10%) was reported (Hotchkiss et al. 2004). In multigenerational studies, relative and absolute testes weights were decreased in F1 male pups at PND22 at the highest dose tested (Nagao et al. 2000). This effect was consistent at higher doses in both F1 and F2 offspring in another study (750 mg/kg bw/day) (Tyl et al. 2004).

Histopathological effects in the testes after exposure to BBP included seminiferous tubule atrophy, Leydig cell hyperplasia, absence of gubernaculi and flaccid fluid-filled testes; these effects were observed regardless of the length of exposure. Multiple effects in the testes included aplasia/dysplasia of the epididymis, diffuse atrophy of the seminiferous tubules and Leydig cell hyperplasia at 400 mg/kg bw/day in F1 males when examined in adulthood (Aso et al. 2005). Severe seminiferous tubule degeneration and atrophy were also observed in F1 males in adulthood along with Leydig cell hyperplasia at 500 mg/kg bw/day and above in other studies (Nagao et al. 2000; Tyl et al. 2004).

The effect of gestational exposure to BBP on fertility, whether measured by sperm parameters at a young age or by reproductive success as adult males, was evident at doses higher than testicular histopathology. Aso et al. (2005) did not find any adverse effects in sperm parameters (sperm count, motility, morphology) in F1 adult males, nor were there any adverse effects related to BBP on male fertility or mating indices at doses up to and including 400 mg/kg bw/day. Similarly, Nagao et al. (2000) reported no effects on sperm parameters or the reproductive performance of F1 adult males at doses up to 500 mg/kg bw/day^Footnote[28] BBP via gavage. At higher doses, Tyl et al. (2004) reported decreases in sperm concentration and sperm motility as well as decreased mating and fertility indices at 750 mg/kg bw/day as adults. Pregnancy indices were also reduced by 15% at this dose as well. It should be noted that these males also exhibited decreased body weights and decreased relative adrenal, brain and pancreas weights at this dose level.

Alterations of steroidogenesis have been reported with exposure to BBP as decreases in serum and testicular testosterone levels when measured post birth. Only two studies examined serum testosterone levels; one found a statistically significant (44%) decrease in levels in F1 adult males at 500 mg/kg bw/day (highest dose tested) (Nagao et al. 2000), while another reported a similar, non-statistically significant decrease at lower doses of 400 mg/kg bw/day in F0 adult males (Aso et al. 2005). Testicular testosterone levels were measured in three short-term gestational studies (GD18) immediately after cessation of in utero exposure. Howdeshell et al. (2008) reported a dose-dependent decrease in testosterone levels (from 27 up to 91% at 900 mg/kg bw/day) in foetal pup testes at 300 mg/kg bw/day and above. It should be noted that maternal toxicity was evident through decreased body weight gain during gestation at this dose and above as well. Both testicular testosterone production and testosterone concentrations were statistically significantly reduced in GD18 male foetuses at 500 mg/kg bw/day with no apparent maternal toxicity (Hotchkiss et al. 2004). A more recent study presenting the potential for BBP and other phthalates to alter foetal testosterone production (ex vivo) in pregnant SD rats showed that this phthalate disturbed testicular testosterone production during gestation at doses as low as 100 mg/kg bw/day and above with a calculated ED₅₀value of 172.4 mg/kg bw/day (Furr et al. 2014).

Maternal toxicity was examined in relation to the effects in offspring at similar or lower doses. Overall, maternal toxicity after exposure to BBP became evident in the form of decreases in body weight gain, increased kidney and liver weights, and changes in reproductive organ weights (ovaries and uterus), although not consistently. In multi-generational studies, decreased body weight gain was observed at 500 mg/kg bw/day in F1 (data not shown) (Nagao et al. 2000) and 750 mg/kg bw/day (Tyl et al. 2004) in both F0 and F1 females. No changes in body weight gain were observed at lower doses (Aso et al. 2005; Nagao et al. 2000). In shorter term studies (GD5-20), maternal body weight gain was significantly reduced at 300 mg/kg bw/day (18%) and above (Ema et al. 1990; Piersma et al. 1999; Howdeshell et al. 2008).

Liver weight changes (relative and absolute) occurred at doses as low as 200 mg/kg bw/day BBP, without an increase in magnitude or histopathological lesions in F0 but not F1 females (Aso et al. 2005). Tyl et al. (2004) observed significant (16%) increases in relative and absolute liver weights in F0 females at necropsy, but not in F1 females at the highest dose tested (750 mg/kg bw/day), with histopathological lesions in both generations at this dose.

Kidney weight changes (relative and absolute) occurred at doses as low as 400 mg/kg bw/day BBP, without an increase in magnitude or histopathological lesions in F0 but not F1 females (Aso et al. 2005). Similar results were reported by Tyl et al. (2004), where a small increase (less than 10%) in kidney weights was reported in both F0 and F1 females up to as high as 750 mg/kg bw/day, without histopathological outcome.

Nagao et al. (2000) observed a significant decrease (12%) in ovary weight in P0 dams at necropsy at 500 mg/kg bw/day. This was not observed in F1 females at the same dose levels of BBP. In another multi-generational study, relative uterine weights were decreased at 200 mg/kg bw/day. However, according to the authors, this is most likely not due to BBP treatment, as it is not dose-dependent and did not occur in F1 females (Aso et al. 2005). This effect appeared more consistent across generations and increased in severity at higher doses (Tyl et al. 2004). See Table 9-26.

A search of the available literature revealed only two studies examining the effects of gestational exposure of BBP in mice, tested prior to the masculinization programming window. Neither examined the parameters used to measure those related to RPS (see Table 9-26).

Overall, the highest NOAEL for developmental toxicity identified for BBP was 50 mg/kg bw/day based on pup body weights (both male and female) at 100 mg/kg bw/day and decreased AGD at birth in males at 100-250 mg/kg bw/day and above (Aso et al. 2005; Nagao et al. 2000; Tyl et al. 2004). Foetal testicular testosterone was also reduced at this dose level and above (Furr et al. 2014). The lowest LOAEL for maternal toxicity of BBP was 300 mg/kg bw/day based on significantly reduced maternal body weight gain (Howdeshell et al. 2008). The available information indicates that BBP causes developmental effects in male pups at lower doses than the other two analogues, DIBP and MBzP. Refer to sections 9.2.2.1 and 9.2.7.1 above for summaries of the studies conducted with DIBP and DBzP (MBzP), respectively.

Therefore, the critical effect level for developmental toxicity of B84P for this life stage, based on effects observed after exposure to DBP and DIBP, is 100-250 mg/kg bw/day.

9.2.8.1.2 Exposure to B84P at prepubertal/pubertal life stage

There were no repeated-dose oral exposure studies in sexually immature animals (PND1-55) with B84P via any route of exposure. As with the previous section, DIBP and MBzP were identified as the most appropriate candidates for read-across. Refer to sections 9.2.2.1 and 9.2.7.1 above for summaries of the studies using DIBP and DBzP (MBzP), respectively. Summaries of the studies are described in Table 9-27 below.

To determine the potential reproductive toxicity of phthalate diesters on sperm parameters in young male rats, Kwack and colleagues (2009) orally exposed six 5-week-old Sprague-Dawley rats to 500 mg/kg bw/day of BBP for four weeks. Adverse effects included decreased body weight gain as well as increased relative liver weights. The authors reported that BBP significantly lowered the sperm counts (31% decrease compared to controls) and decreased sperm motility at 500 mg/kg bw/day (Kwack et al. 2009). Effects on sperm parameters could potentially be secondary to systemic toxicity.

No changes in reproductive organ weights were observed after ten-day oral BBP exposure up to and including 500 mg/kg bw/day in young castrated male rats in a Hershberger assay (Lee and Koo 2007). The authors did note a small but statistically significant decrease in serum testosterone levels and a slight increase in serum luteinizing hormone levels at 100 mg/kg bw/day and above.

Overall, the lowest LOEL for reproductive toxicity identified for BBP at the prepubertal/pubertal life stage was 500 mg/kg bw/day based on decreased sperm count and sperm motility in young male rats, effects that could potentially have been secondary to systemic effects (Kwack et al. 2009). As described in Section 9.2.7.1, the lowest LOEL for reproductive toxicity identified for MBzP at the prepubertal/pubertal life stage was 250 mg/kg bw/day based on decreased sperm counts after four weeks of exposure (Kwack et al. 2009). Therefore, the lowest LOEL of 250-500 mg/kg bw/day will be used as the critical effect level range for the reproductive toxicity of B84P for this life stage based on the effects observed after exposure to MBzP and BBP, respectively.

9.2.8.1.3 Oral exposure at the mature male adult stage

As with the previous life stages, no studies examining the potential reproductive toxicity of B84P at the adult male life stage (PND55+) were identified. Studies conducted with BBP were reviewed to characterize the health effects of B84P for this life stage (Health Canada 2015a). Summaries of the studies are described in Table 9-28 below. As mentioned in previous sections, no studies examining the potential reproductive toxicity of DIBP and MBzP at the adult male life stage were identified.

Overall, the reproductive effects of BBP in adult male rats have included reduced mating and fertility, decreased testes weights, histopathological effects in the testes as well as decreases in serum testosterone levels. See Table 9-28 below for a summary of the effects in adult male rodents after oral exposure to BBP. In an NTP ten-week modified mating study, male F344 rats were exposed to BBP in the diet at levels of 0, 300, 2800 or 25 000 ppm (0, 20, 200 or 2200 mg/kg bw/day) for ten weeks with a corresponding two-day recovery period (NTP 1997c). The rats were then mated with untreated females and necropsied with a full histological examination of the control and high-dose group only. However, the testis and epididymis, seminal vesicle and prostate were examined in all groups. Males in the high-dose group (2200 mg/kg bw/day) exhibited reduced absolute and relative testis and prostate weights, along with marked degeneration in the testis and epididymis. Epididymal sperm concentration was 87, 70, and 0.1% of the control in the 20, 200 and 2200 mg/kg bw/day groups, respectively. Reproductive success (pregnancies) and sperm motility and morphology were comparable between the controls and the low- and mid-dose groups, but these parameters were not measured in the high-dose group due to the absence of sperm; no females were pregnant after mating with the males. The significant reduction in sperm count observed in the 200 mg/kg bw/day group was not considered adverse by the NTP Expert Panel (2002), as sperm counts might have been affected by the shorter recovery period from the time between mating to necropsy in this group compared to the other dose groups. Further, a European Union risk assessment on BBP (2007) performed a covariate analysis of variance taking into account days of recovery and concluded that the decrease in spermatozoa concentration was not statistically significant at 200 mg/kg bw/day (at the 5% level; p = 0.07) (ECJRC 2007), although the response was still dose-dependent.

A more recent 2-generation study in Crj:CD Sprague-Dawley IGS rats administered 0, 100, 200 and 400 mg/kg bw/day of BBP by gavage starting at five weeks of age (F0) and three weeks of age (F1) for ten weeks prior to mating through weaning (Aso et al. 2005). Effects in F0 males included reduced absolute epididymal weight, hyperplasia of the Leydig cells in the testes and decreased spermatozoa in the lumina of the epididymis at the 400 mg/kg bw/day dose level.

Overall, the highest NOAEL for reproductive toxicity identified for BBP was 200 mg/kg bw/day based on histopathological effects in testes of adult F0 males, which included reduced absolute epididymal weight, hyperplasia of the Leydig cells in the testes and decreased spermatozoa in the lumina of the epididymis at 400 mg/kg bw/day (Aso et al. 2005). A NOAEL of 200 mg/kg bw/day was also determined from an earlier study in F344 rats, where a high rate of infertility (decreased numbers of pregnancies), marked histopathology in the testes and epididymis, and lower sperm counts were observed at 2200 mg/kg bw/day (NTP 1997b). These endpoints were selected, as they represent reproductive effects in adult male animals that were exposed specifically during this life stage. Systemic effects appeared to be mostly limited to increased kidney and liver weights, with the lowest LOAEL at 313 mg/kg bw/day based on a significant increase in relative liver and kidney weights, accompanied by histological changes in the liver at higher doses in male F344 rats administered BBP for 14 days (Agarwal et al. 1985). No studies were identified for B84P on any other species via any other route of exposure (dermal, inhalation) at this life stage. Therefore, the NOAEL of 200 mg/kg bw/day will be used as the critical effect level for the reproductive toxicity of B84P for this life stage.

9.2.8.2 Oral exposure in females

The potential reproductive/developmental effects of DIBP and DBzP (MBzP) in females were summarized in sections 9.2.2.2 and 9.2.7.2 above, respectively.

As previously noted, there was many studies identified examining the reproductive and developmental effects of BBP. Over 20 studies were performed in rats and 3 studies in mice. The route of exposure was oral in all instances, either gavage or feed. The lowest LOAEL identified for developmental toxicity of BBP in females was 100 mg/kg bw/day and was determined from the previously described 2-generation studies in Section 9.2.8.1 (Aso et al., 2005; Nagao et al., 2000). In the first study (Aso et al., 2005), effects included altered reproductive development (statistically significant increase of AGD in F1 and decreased pup weight on PND0 in F2 female offspring at 100 mg/kg bw/day and above, the lowest dose tested). Maternal effects were reported as a statistically significant increase in absolute and relative liver and kidney weights in F0 females at 400 mg/kg bw/day (NOAEL of 200 mg/kg bw/day).

In the second study (Nagao et al., 2000), altered growth and functional deficits (thyroid) included a statistically significant decrease in mean body weights (on PND0) and of serum T3 (at weaning) at 100 mg/kg bw/day BBP and above in F1 offspring only (NOAEL of 20 mg/kg bw/day). Maternal effects included a statistically significant decrease in relative ovary weights in F0 females and decreased body weights in F1 at 500 mg/kg bw/day (NOAEL of 100 mg/kg bw/day).
The lowest LOAELs identified for the reproductive toxicity of BBP in adult females ranged from 500 to 590-2330 mg/kg bw/day, with NOAEL ranging from 100 to 1100 mg/kg bw/day. The critical effects referred principally to pregnancy outcomes, but also to alterations in reproductive organ weights and hormone levels (serum prolactin).

In summary, studies that examined the reproductive/developmental effects of BBP in females were principally performed at high doses of exposure. Their results suggest that BBP is a reproductive and developmental toxicant at 100 mg/kg bw/day and above, at the same dose level as that for which effects were observed in male offspring. The critical endpoints were related to growth alteration, lethality, altered reproductive organ weights, delay of puberty and teratogenicity (variations and skeletal or visceral malformations). The critical reproductive endpoints were related to pregnancy outcomes, alterations of reproductive organ weights, hormone levels (progesterone and prolactin) and reproductive-related organ visual examination and histopathology (mammary gland). Gender-related differences were reported (males are apparently more sensitive than females).

9.2.8.3 Endocrine studies

One targeted study by Clewell et al. (2010) examined the effects of MBzP on progesterone and testosterone synthesis in the immortalized mouse Leydig cell tumour (MA-10) assay system. MBzP was a weak inhibitor of testosterone synthesis because, according to the authors, this inhibition was statistically significant at concentrations of 3, 30 and 100 µM, and none of the treatment concentrations caused more than a 35% decrease in testosterone. A more recent study by the same group using a rat Leydig cell line (R2C) also showed that this monoester slightly reduced testosterone production at concentrations equal to and greater than 30 µM (Balbuena et al. 2013).

In the same study described in Section 9.2.8.1 by Saillenfait et al. (2003), mouse and rat whole embryos were cultured at comparable developmental stages and exposed to 0 to 5 mM MBzP for 48 hours. It appeared as though the mouse embryo was not intrinsically more sensitive to MBzP than the rat embryo. Similar to in vivoobservations, the central nervous system was a target of MBzP, as indicated by the occurrence of open neural tubes (Saillenfait et al. 2003).

9.2.8.4 Reproductive and developmental toxicity: evidence in humans

No information is available on the potential reproductive/developmental effects of B84P in humans.

9.2.8.5 Other systemic effects^Footnote[29]

9.2.8.5.1  Repeated-dose studies

No short-term and subchronic studies have been identified in the literature for B84P. Studies conducted on its closest analogues DIBP, MBzP and BBP (as described in sections 9.2.2.4 and 9.2.7.4, respectively) were used to characterize the health effects of B84P.
Common systemic effects of these analogues are organ changes (organ weight changes, histopathological changes) and decreases in body weight gain in exposed dams.

The lowest LOAEL for short-term exposure was 167 mg/kg bw/day based on dose-dependent decreases in body weight gain (22% decrease for adjusted weight gain) associated with a decrease in food consumption (8-15%) in dams in a developmental toxicity study in rats after exposure to MBzP (Ema et al. 2003, as described in Section 9.2.8.1).

Short-term and subchronic oral studies looking at the effects of DIBP on rodents have also been identified in the literature. Refer to Section 9.2.2.4 for summaries of the studies for this analogue.

Short-term and subchronic studies were also identified for BBP. Most of the repeated-dose studies for this phthalate have been conducted in rats. Only one study in mice and one study in dogs were reported. Also, most studies have used the oral route to study the potential effects of BBP exposure. However, a few inhalation studies and one dermal study were identified. The main effects reported are decreases in body weights and increases in organ weights. The available studies are summarized below. Critical effect levels identified from these studies are presented in Table 9-29.

In a short-term study in which male F344 rats were exposed to 0, 0.625, 1.25, 2.5 or 5.0% BBP (0, 313, 625, 1250 or 2500 mg/kg bw/day) through diet for 14 days, a LOAEL of 313 mg/kg bw/day was identified based on increased LH levels and an increase in relative liver and kidney weights, accompanied by histological changes in the liver (mild multifocal hepatitis) at the highest dose tested. At the two highest doses, a relative decrease in testes, seminal vesicle and thymus weight were observed, and dose-related histopathological changes in seminal vesicles, testes and prostate were noted, as were a decrease in bone marrow cellularity and an increase in FSH. A relative decrease in epididymal weight and cortical lymphocytololysis in the thymus was also observed at the highest dose tested, as well as an increase in testosterone levels (Agarwal et al. 1985). In a 14-day gavage study in male Sprague-Dawley rats, historical changes in the testes were observed in animals administered 480 mg BBP/kg bw/day. Testicular atrophy, a decrease in body weight, liver enlargement and ultrastructural changes with increased peroxisome numbers in the liver were observed at 1600 mg/kg bw/day, the highest dose tested (Lake et al. 1978). In another two-week gavage study in which Wistar rats were given similar doses of BBP (0, 480 or 1600 mg/kg bw/day), no effects were reported at 480 mg/kg bw/day, and decreased body weight and testicular atrophy were observed at the highest dose tested. Microscopic changes in the liver were not determined (Hammond et al. 1987).

In a four-week range finding study, Sprague-Dawley rats of each sex were exposed to 0, 500, 1000, 1500, 2000, 3000 and 4000 mg/kg bw/day of BBP in feed. A NOAEL of 1000 mg/kg bw/day and a LOAEL of 1500 mg/kg bw/day were identified based on dose-related decrease in body weight gain in both sexes (more pronounced in males), increased mortality in males and a dose-related increase in histopathological changes in the testes. Stiffness while walking was noted in exposed animals from 2000 mg/kg bw/day, as was bleeding around the nares in animals exposed to the highest dose. Rats exposed to the highest dose that died during the study exhibited testicular atrophy, dehydration and blue discoloration and/or inflammation of the extremities and had gross and microscopic evidence of widespread haemorrhaging in body tissues (Hammond et al. 1987). When young male Cpb:WU rats were exposed for the same duration (28 days) via gavage at doses up to 2100 mg/kg bw/day, the NOEL and the LOEL for systemic toxicity were 580 and 750 mg/kg bw/day, respectively, based on statistically significantly increased relative liver weight and liver palmitoyl CoA (PCoA), an index of peroxisome proliferation, at 750 mg/kg bw/day and above. A dose-dependent increase trend in relative kidney weight and a trend toward a decrease in thymus and thyroid weight were reported from 750 mg/kg bw/day, but none of these changes were statistically significant. A statistically significant decrease in testosterone levels was reported from 450 mg/kg bw/day, severe atrophy of the testes was noted from 970 mg/kg bw/day and a significant decrease in relative testes weight and a significant increase in FSH were reported from 1250 mg/kg bw/day (Piersma et al. 1999).

In a six-week range finding investigation studying the neurotoxicity of BBP after oral administration in feed, there were no adverse histopathological effects on the nervous system of rats exposed to up to 3000 mg/kg bw/day, although reversible clinical signs were observed (Robinson 1991).

In a subchronic diet study, Wistar rats of each sex were exposed to 0, 151, 381 and 960 mg BBP/kg-bw/d in feed for three months. A reduction in body weight gain was reported in low-, mid- and high-dose groups. However, only the reduction at the highest dose was considered compound-related since food consumption was decreased in the low- and mid-dose groups but not in the highest-dose group. Slight anaemia at the highest dose and decreased urinary pH at mid and high doses were reported in males. A significant increase in relative liver weights was observed at all dose levels in females (small increases at the low and mid doses) and at the highest dose in males. A significant increase in relative kidney weight was noted in both sexes at the mid and high doses. While the relative cecum weight was unchanged in males, a dose-related increase was observed at all dose levels in females.

Gross pathological lesions were limited to increased incidence of red spots on the liver at 381 mg/kg bw/day and higher in males. Histopathological changes were reported in the pancreas of males exposed at mid and high doses, and included islet enlargement with cell vacualoation and peri-islet congestion. Small areas of cellular necrosis were also observed in the liver of males exposed at the highest dose. No histopathological changes were reported in females. In this study, a NOAEL of 151 mg/kg bw/day and a LOAEL of 381 mg/kg bw/day was identified for males based on histopathological changes in the pancreas, gross pathological alterations in the liver and a significant increase in relative kidney weight. For females, a LOEL of 151 mg/kg bw/day was identified based on marginal increases in relative liver and cecum weights in the absence of gross and histopathological changes (Hammond et al. 1987).

In another three-month study reported in Hammond et al. 1987, Sprague-Dawley rats of both sexes were administered 0, 188, 375, 750, 1125 and 1500 mg BBP/kg-bw/d for three months. A NOEL of 375 mg/kg bw/day and a LOEL of 750 mg/kg bw/day were identified based on a significant increase in relative liver and kidney weights in females. In males, no increase in kidney weight was noted but a significant increase in relative liver weight was reported at 1125 mg/kg bw/day. No compound-related lesions were observed in this strain of rats upon histopathological examination in the organ tissues (liver, testes, and pancreas) (Hammond et al. 1987).

In a subchronic study in dogs (three per sex per group) given BBP through diet for three months, no adverse effects were reported at doses up to 50,000 ppm (equivalent to 1850 mg/kg bw/day in males and 1973 mg/kg bw/day in females). The only effect reported was a decrease in body weight gain in the highest-dose group in males and in the two highest-dose groups in females, but these decreases were associated with a decrease in food consumption (Hammond et al. 1987).

When mice were exposed for 90 days to 0, 240, 464, 946, 1875 and 3750 mg BBP/kg-bw/d through diet, no adverse effects were observed again at any BBP doses studied. While a decrease in body weight gain was observed at all doses tested in male mice and at 1875 mg/kg bw/day and higher in females, food consumption was not reported. Due to this, the author identified a LOEL for male mice at 240 mg/kg bw/day and a LOEL for female mice at 1875 mg/kg bw/day (NOEL at 946 mg/kg bw/day) based on decreased body weight gain (NTP 1982b).

In a NTP study in which F344 male rats were exposed for 26 weeks to 0, 300, 900, 2800, 8300 and 25 000 ppm DIDP (approximately 0, 30, 60, 180, 550 and 1650 mg/kg bw/day) through diet, a NOAEL of 180 mg/kg bw/day and a LOAEL of 550 mg/kg bw/day were identified based on increase in the mean cell hemoglobin found on days 60-180, which may be associated with macrocytic anemia found at the next dosing  level (1650 mg/kg bw/day) on days 30-180, and an increase in relative liver and kidney weights. At the highest dose, a decrease in total body weight was observed, presumably due to a reduction in food consumption. This reduction in actual food consumption made it difficult for the dose to be calculated; the 1650 mg/kg bw/day dose is estimated from the intake levels of the lower doses. Since the dose is based on the amount of food intake, the results seen may be due to a lower dose than what was calculated. Testicular effects (hypospermia, atrophy) were also reported at that dose level (NTP 1997b).

In a short-term inhalation study in which Sprague-Dawley rats were exposed to 0, 360, 1000 and 2100 mg/m³ BBP given through aerosol/vapour for six hours per day, five days per week for four weeks, toxicological effects such as a statistically significant decrease in body weight gain (33% for males and 13% for females), death (3/20 males and 4/20 females) and atrophy of the spleen and reproductive organs (in males only) were observed in animals in the highest-dose groups. The NOAEC was 1000 mg/m³ and the LOAEC was 2100 mg/m³ based on decreased body weight gain and atrophy of the spleen and testes (Monsanto 1981). In another similar four-week study, BBP (0, 49, 144 and 526 mg/m³) was also given as an aerosol/vapour to Sprague-Dawley rats for six hours per day, five days per week. A reduction in body weight gain was noted in both sexes exposed at the highest dose. No changes in clinical parameters, organ weights or microscopic abnormalities were observed. In this study, the NOAEC was 144 mg/m³ and the LOAEC was 526 mg/m³ based on decreased body weight gain (Hammond et al. 1987).

In a subchronic inhalation study, a group of 25 male and female Sprague-Dawley rats was exposed to concentrations of 0, 51, 218 or 789 mg/m³, six hours per day, five days per week for 13 weeks. Significant increases in absolute and/or relative liver and kidney weights were observed in both sexes. In males, a marked decrease in serum glucose was observed at 789 mg/m³. No such increase was noted in females. No compound-related macroscopic or microscopic lesions were detected in any tissues. The NOEC was 218 mg/m³ and the LOEC was 789 mg/m³ based on an increase in liver and kidney weights (both sexes) and an increase in serum glucose in males (Monsanto 1982; Hammond et al. 1987).

Finally, in a dermal study, only local irritation was reported after repeated skin applications of BBP at doses of 1, 5, 10 and 100 mg/kg-bw for five months. However, this study was poorly reported and no information regarding the species used was provided (Statstek 1974).

Overall, the lowest LOAEL for short-term oral exposure was 167 mg/kg bw/day based on a dose-dependent decrease in body weight gain and a decrease in food consumption in dams in a developmental toxicity study in rats exposed to MBzP (Ema et al. 2003). For DIBP, the lowest LOAEL for short-term oral exposure was at 900 mg/kg bw/day based on decreased body weight gain in dams in a developmental toxicity study in rats (Howdeshell et al. 2008). For BBP, the lowest LOAEL for short-term oral exposure was 313 mg/kg bw/day based on an increase in relative liver and kidney weights, accompanied by histological changes in the liver and increased LH levels in male rats (Agarwal et al. 1985).

The lowest LOAEL for subchronic oral exposure for DIBP was 4861-5960 mg/kg bw/day based on a decrease in body weight gain in male and female rats in a 16-week study (Hodge 1954). For BBP, the lowest oral LOAEL was 381 mg/kg bw/day (NOAEL of 151 mg/kg bw/day) based on histopathological changes in the pancreas, gross pathological alterations in the liver and a significant increase in relative kidney weight in male Wistar rats exposed for three months (Hammond et al. 1987). However, it is important to note that those histopathological effects were not observed in other strains of rats exposed orally to BBP for a similar or longer-term duration (Sprague-Dawley and F344, respectively). Also, no adverse effects were observed in mice and dogs exposed to high doses of BBP for three months.

Three inhalation studies were available for BBP. Among these studies, the lowest LOAEC for short-term exposure was 526 mg/m³ (NOAEC was 144 mg/m³) based on decreased body weight gain in rats (Hammond et al. 1987). In a subchronic inhalation study, the NOEC was 218 mg/m³ and the LOEC was 789 mg/m³ based on an increase in liver and kidney weights (both sexes) and an increase in serum glucose in male rats (Monsanto 1982; Hammond et al. 1987).

Finally, in a dermal study identified for BBP, a LOEL of 100 mg/kg bw/day was identified based on local irritation. However, this study was poorly reported (Statstek 1974).

9.2.8.5.2 Carcinogenicity

B84P has not been classified for its potential carcinogenicity by other international agencies.

No chronic toxicity/carcinogenicity studies were available for this phthalate or for its closest analogues MBzP or DIBP.

Chronic toxicity and carcinogenicity data have been identified in the literature for another analogue of B84P, BBP. The available data for BBP have been reviewed previously in a Priority Substances List (PSL) Assessment Report published by Environment Canada and Health Canada in 2000. Complete information from this report is available in Appendix I: Supporting information on the chronic toxicity and carcinogenicity of butylbenzylphthalate (BBP).

No new chronic studies on the carcinogenicity of BBP have been published since the PSL Assessment Report. This review has shown that no liver tumours were found to be associated with BBP oral exposure. An increase in mononuclear cell leukemias observed in female rats in a 1982 study was not confirmed in a 1997 repeat study (NTP 1982, 1997a). BBP induced an increase in pancreatic tumours (pancreatic acinar cell adenoma and combined adenoma and carcinoma) primarily in male rats, the full expression of which was prevented in a dietary restriction protocol (NTP 1997a). Also, a marginal increase in bladder tumours was observed in female rats, which was delayed upon dietary restriction (NTP 1997a). There was no evidence of carcinogenicity in mice (NTP 1982). Since the weight of evidence of genotoxicity was negative, it was suggested that BBP can be considered, at most, possibly carcinogenic to humans, likely inducing tumours through a non-genotoxic (albeit unknown) mechanism (Environment Canada and Health Canada, 2000). Non-cancer effects were observed in exposed rats (both sexes) but not in exposed mice. The lowest non-neoplastic LOAEL was 300 mg/kg bw/day based on a significant increase in incidence of nephropathy noted in all groups of exposed females (NTP 1997a; Environment Canada and Health Canada 2000).

Since the publication of the PSL Assessment Report (Environment Canada and Health Canada 2000), data on the carcinogenicity of BBP have also been reviewed by the International Agency for Research on Cancer (IARC 1999), the European Chemicals Bureau (ECB 2007), and the Office of Environmental Health Hazard Assessment (OEHHA) of the California Environmental Protection Agency (EPA). IARC has classified BBP as Group 3 "Not classifiable as to its carcinogenicity to humans" based on inadequate evidence of the carcinogenicity of BBP in humans and limited evidence in experimental animals (IARC 1999). ECB published a risk assessment report on BBP (ECB 2007) and concluded that BBP is not mutagenic. ECB suggested that BBP may be a borderline case between being not classifiable with respect to its carcinogenicity and being classified as a Category 3 carcinogen. In the end, no classification was proposed by ECB (ECB 2007). More recently, the OEHHA of the California EPA developed a document on evidence of the carcinogenicity of BBP. Members of the Carcinogen Identification Committee (CIC) concluded that BBP has not been clearly shown to cause cancer and should not be listed under Proposition 65 as a carcinogen (OEHHA 2013a).

Information on the mode of action and human relevance of the different types of tumours observed in BBP-treated animals is available in Health Canada (2015c).

9.2.8.5.3 Genotoxicity

B84P was not mutagenic in an OECD Guideline 471 study at doses less than 10 µg/plate (0.01, 0.04, 0.2, 1.0, 3.0 and 10.0 µl/plate) in Salmonella strains TA98, TA100, TA1535, TA1537 and TA1538, with or without metabolic activation. No microbial toxicity was observed in any of the five strains at 10 µg/plate (with or without metabolic activation), although levels of 3 µg/plate and higher exceeded the solubility of the test material (Monsanto Research Corporation 1982 cited in US EPA 2006, 2010).

9.2.8.5.4 Evidence of systemic toxicity in humans

No information is currently available on the potential effects of B84P in humans.

9.2.9 DIHepP

 9.2.9.1 Reproductive and developmental effects in males

9.2.9.1.1 Early development: in utero exposure

A literature search identified three oral studies examining the effect of DIHepP when administered during gestation in pregnant rats and the potential toxicity of DIHepP during gestation in rats, all focusing on male reproductive effects during the foetal masculinization programming window (GD15-17). Summaries of the studies are described in Table 9-30 below. No other developmental studies were identified examining gestational exposure to DIHepP via other routes of exposure or using other species.

In a 2-generation reproductive toxicity study, DIHepP appears to cause a multitude of effects in the male foetus related to RPS. A critical study by McKee et al. (2006) administered 0, 1000, 4500 or 8000 ppm of DIHepP in diet to Sprague-Dawley rats 70 days prior to mating, through the mating period, and during gestation and lactation (or until the termination period for males) (approximately 64-168, 304-750 and 532-1360 mg/kg bw/day for F0 and F1). Developmental effects in the F1 generation, observed primarily at 532-1289 mg/kg bw/day, included a significant reduction in AGD, a significant increase in incidence of NR and testicular abnormalities (HYP and CRY), a significant reduction in weights of testes and male accessory reproductive organs, a significant decrease in testicular sperm counts and daily sperm production, significant delays in PPS and a significant decrease in fertility. Reduction in AGD was also observed in the F2 generation at 309-750 mg/kg bw/day. Additional effects observed in the study included reductions in body weights, along with increased liver and kidney weights (mid dose and above) in both generations and increased pituitary weights in F1 males at the highest dose. The NOAEL for parental systemic toxicity in the F0 and F1 generations was 50-168 mg/kg bw/day (male and female dose level ranges) based on liver and kidney effects (see Section 9.2.9.5). The LOAEL for reproductive/developmental toxicity was 309-750 mg/kg bw/day, based on a significant reduction in AGD in male F2 pups exposed at the mid dose and above (Wil Research Laboratories Inc., 2003; McKee et al. 2006).

An oral developmental toxicity study was also performed by McKee et al. (2006), where pregnant rats were administered DIHepP by gavage at doses of 0, 100, 300 or 750 mg/kg bw/day during GD6-20. Offspring were examined on GD21. A significant decrease in maternal body weight gain resulting from a lower mean body weight was noted at the highest dose (750 mg/kg bw/day) and was primarily due to uterine content. There were also statistically significant dose-related increases in mean absolute and relative maternal liver weights in the 300 and 750 mg/kg bw/day dams compared with controls. Developmental effects included a significant reduction in the number of viable foetuses/dam and significant increases in post-implantation loss and resorptions/dam, a reduction in mean foetal weights and a significant increase in external, visceral and skeletal malformations and variations in foetuses at 750 mg/kg bw/day. The principal external malformations in the high-dose group included stunting and anophthalmia; the visceral observations included ectopic testes, ectopic ovaries, and elongated and malformations of the subclavian and innominate arteries. The skeletal variations and malformations included both rib and vertebral anomalies. The NOAEL for maternal toxicity in this study was 750 mg/kg bw/day. While a significant increase in relative and absolute liver weight was observed in dams treated at the mid and high doses in comparison with controls, the increase, consistent with the occurrence of peroxisomal proliferation, is not considered an adverse effect (LOEL = 300 mg/kg bw/day) (McKee et al. 2006). In this study, the developmental NOAEL was established at 300 mg/kg bw/day, with a LOAEL of 750 mg/kg bw/day based on increased resorptions, viability and malformations.

A more recent study presenting the potential for DIHepP and other phthalates to perturb foetal testosterone production (ex vivo) in pregnant SD rats showed that this phthalate disturbed testicular testosterone production during gestation at 750 mg/kg bw/day (only dose tested) with a calculated ED₅₀ value of 361.6 mg/kg bw/day (Furr et al. 2014). Further details were not provided.

Overall, the highest oral NOAEL for developmental toxicity of DIHepP at the in utero life stage was 64-168 mg/kg bw/day based on effects on the developing male reproductive system, as seen by decreased AGD at the next highest dose (309-750 mg/kg bw/day; McKee et al. 2006). The same effect level from this study was set by NICNAS (2008). No marked maternal toxicity that would affect the reproductive development of offspring was reported, as effects included increases in liver and kidney weight occurring at 304-750 mg/kg bw/day (LOAEL; McKee et al. 2006).

9.2.9.1.2 Exposure at prepubertal/pubertal life stages

No reproductive toxicity studies with exposure to DIHepP for this life stage have been found in the literature. Therefore, results from the 2-generation study described in Section 9.2.9.1.1 above were used, particularly those from the F1 males that were exposed in utero, through lactation to DIHepP and until PND54.

In the 2-generation study, significant decreases (with or without analysis relative to the cube root of pup body weight; only measured in high-dose group) in AGD, an increase in retained thoracic nipples, external genitalia disorder such as hypospadias (7/30), lacking testes (2/30) and undescended testes (2/30) were observed in F1 males at 8000 ppm (419-764 mg/kg bw/day). Increased balanopreputial separation was also statistically delayed compared to the control (46.1 vs. 50.3 days) in the high-dose group. Observation of the F1 generation at the adult stage also showed effects on reproductive organs (severe degeneration of seminiferous tubules) and decreased fertility at high doses (McKee et al. 2006).

Overall, the NOAEL for reproductive toxicity identified for DIHepP for this life stage was 227-416 mg/kg bw/day. However, it is difficult to determine whether the effects observed were due to in utero exposure of DIHepP or during postnatal and lactation exposure of F1 males. The same effect level from this study was set by NICNAS (2008).

9.2.9.1.3 Oral exposure at the mature adult stage

No reproductive toxicity studies with exposure to DIHepP for this life stage were identified in the literature. Effects from exposure of adult F0 males to DIHepP are reported in this section.

Reproductive effects in this 2-generation study, described above, showed that there were no significant differences in body weights in either males or females during the study (McKee et al. 2006). There were no significant differences in mating success or in gestational period length. Weights of the reproductive organs were not significantly different from the concurrent control values in the F0 generation and there were no histological changes suggestive of treatment-related effects in any of these organs. There were no differences in sperm parameters among the males. No effect on reproduction was noted in F0 males. A NOAEL of 404-623mg/kg bw/day was therefore established for this life stage.

9.2.9.2 Oral exposure in females

Three studies on the reproductive and developmental effects of DIHepP in females were identified. They include a reproductive (2-generation) and developmental toxicity study in rats administered DIHepP through continuous breeding or during gestation (GD6-20), respectively (McKee et al. 2006).

The lowest LOAEL identified for developmental toxicity in females, obtained from the 2-generation study described in previous sections, was 404-1289 mg/kg bw/day for F1 offspring and 404-1360 mg/kg bw/day for F2 offspring (8000 ppm in diet) based on reduced body weight on PND4-21 in F1 and PND14-21 in F2 offspring (McKee et al., 2006). A significant reduction in the female mating and fertility index, and a statistically significant reduction of ovarian weight at 404-1360 mg/kg bw/day (8000 ppm) were also observed in F1 parents (NOAEL of 222-750 mg/kg bw/day). No reproductive adverse effects were observed in F0 parents.

Overall, the few available studies available indicate that DIHepP induced reproductive (alterations of reproductive performance and pregnancy outcomes) and developmental (alterations of growth, functional deficit, lethality and teratogenicity) adverse effects at high doses (404-1360 mg/kg bw/day and above). Alterations of reproductive performance were observed in F1 parents only (after in utero and subsequent exposures).

9.2.9.3 Endocrine studies

DIHepP was inactive in in vitro screening tests for competitive binding and gene expression using the estrogen receptor. The test concentrations used were up to 2000 mg/kg. McKee et al. (2004) reported that the monoester corresponding to DIHepP (MHepP) was inactive in in vitro assays to assess androgen receptor activity.

DIHepP did not exhibit any estrogenic activity when tested in most in vitro and in vivo assays, with only an isomeric mixture demonstrating weak estrogenic activities in a human oestrogen receptor (ER) α (but not β) reporter gene assay (Zacharewski et al. 1998; McKee et al. 2004; Nishihara et al. 2000; Takeuchi et al. 2005; and Toda et al. 2004).

9.2.9.4 Reproductive and developmental toxicity: evidence in humans

No information is currently available on the potential reproductive/developmental effects of DIHepP in humans.

9.2.9.5 Other systemic effects^Footnote[30]

9.2.9.5.1 Repeated-dose studies

Although no long-term studies have been identified, the toxicity of DIHepP has been investigated in a shorter term study in which male rats and mice were exposed to DIHepP through diet (0, 50 or 600 mg/kg bw/day in rats; 0, 65 or 780 mg/kg bw/day in mice) for two or four weeks (Smith et al. 2000). However, only the livers were examined. In both species, effects indicative of peroxisome proliferation were observed. In rats, elevated relative liver weight and increased periportal DNA synthesis in liver were observed after two or four weeks of treatment at 50 mg/kg bw/day and higher. In mice, increased periportal DNA synthesis in liver was observed after two weeks of treatment at 65 mg/kg bw/day and higher. In both rats and mice, increased peroxisomal beta-oxidation (PBOX) in liver was also observed at two and four weeks at the highest dose. The LOEL for repeated-dose oral exposure was 50 mg/kg bw/day, based on elevated relative liver weight and increased periportal DNA synthesis in the liver of male rats.

Results of the 2-generation reproductive toxicity study described above can also be used to evaluate this endpoint (McKee et al. 2006; Section 9.2.9.1.1). No treatment-related changes were observed in body weights, clinical observations or food consumption. Dose-related increases in liver and kidney weights were seen in both sexes of parental F0 rats treated with doses of 222-716 mg/kg bw/day. Histopathological findings observed at mid and high doses in the liver, kidney and pituitary gland included centrilobular hepatocellular hypertrophy and vacuolation, dilated renal pelvis/hydronephrosis and hypertrophy of the pars distalis of the pituitary gland. The NOAEL for systemic toxicity in F0 animals was determined to be approximately 50-168 mg/kg bw/day, with a LOAEL of 222-716 mg/kg bw/day based on liver and kidney effects (McKee et al. 2006).

9.2.9.5.2 Carcinogenicity

DIHepP has not been classified for its potential carcinogenicity by other international agencies, and no chronic toxicity/carcinogenicity studies were identified for this phthalate. In the multigenerational study described above, when SD rats (30/sex/group) were given DIHepP at up to 8000 ppm in diet (419-764 mg/kg bw/day for males and 833-1360 mg/kg bw/day for females), the systemic effects reported in F1 adults (exposed for a significant period of their lifetime) were increases in liver and kidney weights associated with centrilobular hypertrophy in males and females in mid-dose animals (227-750 mg/kg bw/day; McKee et al. 2006). Hepatocellular vacuolation was also noted in male at high doses. Based on this, the carcinogenic potential of DIHepP is likely limited. However, in a repeated-dose where rats and mice were exposed to DIHepP for two or four weeks (Smith et al. 2000; described above), effects indicative of peroxisome proliferation were observed, which could potentially result in the increased incidence of hepatocellular tumours. However, this has been mostly observed at high doses.

9.2.9.5.3 Genotoxicity

In in vitro assays, DIHepP was not mutagenic in a bacterial mutation assay using Salmonella typhimurium, with and without activation (Exxon Biomedical Sciences, Inc. 1995). Similarly, a negative response was observed in an assay for chromosomal aberration in Chinese hamster ovary cells in the presence and absence of metabolic activation (Hazleton Laboratories America, Inc. 1991). In vivo genotoxicity studies for DIHepP have not been identified in the literature.

9.2.9.5.4 Evidence of systemic toxicity in humans

No information is currently available on the potential effects of DIHepP in humans.

9.2.10 BIOP

No studies examining the potential reproductive/developmental health effects of BIOP were identified for any species or gender. DIOP (1,2-Benzenedicarboxylic acid, diisooctyl esters, C7-rich: CAS RN 27554-26-3) and MBz (1,2-Benzenedicarboxylic acid, mono[phenylmethyl] ester: CAS RN 2528-16-7) were identified as the "closest analogue" phthalates to BIOP within the subcategory based on consideration of similarities in monoester metabolism as well as the length and nature of the ester chains (Section 2.3.2; Health Canada 2015a).

Based on health effects information on the analogues MBzP and DIOP, a potential health effect of concern may be associated with BIOP. A review of the potential developmental and reproductive toxicity of the analogue(s) showed that this medium-chain phthalate could have adverse effects on the reproductive system of the developing male, in addition to systemic effects (liver, kidney).

Given the absence of reporting to the section 71 industry survey and the absence of information as to BIOP presence in product databases, general population exposure to BIOP from environmental media and products used by consumers is expected to be negligible. Therefore, risk to human health for this substance is not expected.

9.2.11 B79P

9.2.11.1 Reproductive and developmental effects in males

9.2.11.1.1 Early development: in utero exposure

Only one oral rat study was found focusing on the effects of B79P during development. This study examined the effects of B79P when administered during gestation in pregnant rats during the foetal masculinization programming window (GD15-17). Summaries of the studies are described in Table 9-31 below.

In the extended 1-generation reproductive and developmental toxicity study, parental female SD rats were administered 0, 750, 3750 or 7500 ppm of B79P in diet from GD6, through lactation to PND21. The dose for maternal rats was estimated to be 0, 50, 250 or 500 mg/kg bw/day based on food consumption. In male pups, a statistically significant reduction in AGD on PND21 was observed at 250 mg/kg bw/day and above. It was noted that a statistically significant reduction in AGD was seen in all treatment groups in both males and females at birth. A dose-related and statistically significant increase in the percentage of male pups with a defect of the penis (epispadias) was observed on PND21 at mid dose and above (0, 1.5, 14 and 21% at 0, 50, 250 or 500 mg/kg bw/day, respectively). Also, the percentage of male pups with one or more retained areolae at PND11-13 was statistically significantly increased at 500 mg/kg bw/day only (27% compared with 2.8% in the controls). No differences in AGD, epispadias or areolae were observed in F1 males of any of the dose groups assessed at PND75, suggesting transient effects. Treatment-related histopathological lesions of the left testis (dilated seminiferous tubule lumina) were observed in the 500 mg/kg bw/day group at PND75 (no histopathology performed on other groups). A slight increase in the incidence of undescended testes (cryptorchidism, CRY) was also observed at PND21, but not at PND75, in F1 males from all treated groups. During the lactational period only, F1 males and females showed reduced body weight gain in all treatment groups.

Maternal toxicity for the study was noted at 250 mg/kg bw/day and above based on treatment-related reductions in body weight changes seen on GD6-9 only (with no effects on final body weights at PND21) and increased organ weights (liver and kidneys) reported at PND21. Maternal reproductive and lactational parameters (including fertility and gestational indices, and number of live births) indicated no treatment-related effects at any dose level. The NOAEL for F0 maternal systemic toxicity was considered to be 50 mg/kg bw/day based on decrease body weight gain during gestation and liver and kidney effects at 250 mg/kg bw/day. For F1 male developmental toxicity, 250 mg/kg bw/day was considered the LOAEL based on a significant reduction in AGD and increased epispadias, particularly in the mid-dose group and above (REACH dossier; ECHA 2013b).

Since only one study was identified for B79P, MBzP (1,2-Benzenedicarboxylic acid, mono[phenylmethyl] ester: CAS RN 2528-16-7) and DINP (1,2-Benzenedicarboxylic acid, diisononyl ester: CAS RN 68515-48-0) were identified as the "closest analogue" phthalates to B79P within the subcategory based on consideration of similarities in monoester metabolism as well as the length and nature of the ester chains (Section 2.3.2; Health Canada 2015a).

Refer to Section 9.2.2.1.1 of the assessment for DINP (Environment Canada and Health Canada 2015b) for a review of the potential reproductive/developmental effects of DINP, and Section 9.2.7.1 of this assessment for MBzP (CAS RN 2528-16-7; included in health effects of DBzP) for all life stages.

Overall, the highest NOAEL for developmental toxicity of B79P at the in utero life stage was 50 mg/kg bw/day based on decreased AGD in males and an increased incidence of epispadias at mid-dose and above at the next dose tested (250 mg/kg bw/day) (REACH dossier; ECHA 2013b). Evidence of retained nipples was also reported, but at higher doses. The NOAEL for maternal systemic toxicity was considered to be 50 mg/kg bw/day based on decreased body weight gain during gestation and liver and kidney effects at 250 mg/kg bw/day.

For the analogues MBzP and DINP, the lowest oral LOAEL for developmental toxicity of DINP at the in utero life stage was 159-395 mg/kg bw/day based on decreased pup weight after birth in two diet studies (Watermann et al. 2000; Clewell et al. 2013). Other effects at that dose included significantly reduced foetal testicular testosterone levels and evidence of testicular pathology (MNGs) (Clewell 2011 in ECHA 2013a; Clewell et al. 2013). For MBzP, the lowest oral LOAEL for developmental toxicity at the in utero life stage was also 250 mg/kg bw/day based on an increased incidence of cryptorchidism and decreased AGD in male foetuses (Ema et al. 2003).

Therefore, the critical effect level of 250 mg/kg bw/day will be used to characterize the risk of developmental toxicity of B84P for this life stage.

9.2.11.1.2 Exposure at prepubertal/pubertal life stage

There were no repeated-dose oral exposure studies in sexually immature animals (PND1-55) with B79P via any route of exposure. As with the previous life stage, MBzP and DINP were identified as the most appropriate candidates. Refer to Section 9.2.2.1.2 of the assessment for DINP (Environment Canada and Health Canada 2015b) for a review of the potential reproductive/developmental effects of DINP, and Section 9.2.7.2 of this assessment for MBzP.

Overall, the LOEL for the reproductive toxicity of MBzP and DINP at the prepubertal/pubertal life stage was 250 and 500 mg/kg bw/day based on decreased sperm counts and motility (for DINP only) after four weeks of exposure, respectively (Kwack et al. 2009). Therefore, the LOEL of 250 mg/kg bw/day will be used as the critical effect level for the reproductive toxicity of B79P for this life stage.

9.2.11.1.3 Oral exposure at the mature male adult stage

A search of the public literature identified only one study examining the potential reproductive toxicity of B79P at the adult male life stage. In a three-week feeding study in which male Sprague-Dawley rats were given two different formulations of B79P (one EU and one US) at around 60, 600 and 1200 mg/kg bw/day, minimal testicular degeneration was observed at 1200 mg/kg bw/day in several rats given the different B79P versions (ECHA 2013b). Neither compound affected absolute weights of the epididymis, testis or brain, whereas both test materials produced statistically significant increases in relative liver weight at 600 and 1200 mg/kg bw/day. The test materials produced a statistically significant and dose-related reduction in body weight gain from 600 mg/kg bw/day, particularly during the first week. The LOAEL for this study was 600 mg/kg bw/day based on reduced body weight gain and increases in liver weight and acyl-CoA oxidase activity. The EU version of the test material had a slightly greater effect, but the difference between the two versions was minimal.

No studies were available where MBzP was administered to adult males starting after PND55. Toxicological studies conducted with DINP were examined to characterize the health effects of B79P. Refer to Section 9.2.2.1.3 of the screening assessment for DINP for a review of the studies available (Environment Canada and Health Canada 2015b).

Overall, the highest NOEL for reproductive toxicity identified for DINP was 276 mg/kg bw/day based on reduced relative and absolute reproductive organ weights at the next dose (LOEL = 742 mg/kg bw/day) (Moore 1998b) in adult male mice. This endpoint was selected in other international assessments (NICNAS 2008; EURAR 2003; ECHA 2013a). In a two-year study, testicular interstitial cell hyperplasia was also observed at the highest dose of 553 mg/kg bw/day (Babich 1998). However, the DINP CAS number was not provided in the study and was reported as 71549-78-5 (Babich, 1998). Therefore, the NOEL of 276 mg/kg bw/day will be used as the critical effect level for the reproductive toxicity of B79P for this life stage.

9.2.11.2 Oral exposure in females

No studies addressing female reproductive and developmental effects of B79P were identified. Data indicate that DINP is a developmental and reproductive toxicant at higher doses (600 mg/kg bw/day and above) in females more so than in males.

9.2.11.3 Reproductive and developmental toxicity: evidence in humans

No information is currently available on the potential reproductive/developmental effects of B79P in humans.

9.2.11.4 Other systemic effects^Footnote[31]

9.2.11.4.1 Repeated-dose studies

One repeated-dose study has been identified in the literature for B79P.

In a three-week feeding study in which male Sprague-Dawley rats were given two different formulations of B79P (one EU and one US) at around 60, 600 and 1200 mg/kg bw/day, effects observed from 600 mg/kg bw/day included reduced body weight gain and increases in relative liver weight and acyl-CoA oxidase activity (an indication of peroxisome proliferation) (ECHA 2013b). Neither compound affected absolute weights of the epididymis, testis or brain. There were apparently no gross necropsy findings at sacrifice. An initial decrease in food consumption was observed during the first two weeks at the highest dose (and returned to normal thereafter). In conclusion, no statistically significant adverse effects were seen in male rats ingesting 60 mg/kg bw/day (considered the NOAEL for this study) for three weeks. The LOAEL was 600 mg/kg bw/day based on reduced body weight gain and increases in liver weight and acyl-CoA oxidase activity.

Studies conducted with DINP were also examined to characterize the health effects of B79P. Refer to Section 9.2.1.2 for a complete summary of the available repeated-dose studies for DINP (Environment Canada and Health Canada 2015b).

The lowest LOAEL for short-term oral exposure identified for DINP was 200 mg/kg bw/day, based on histopathological changes in the liver and kidney of mice exposed for 14 days (Ma et al. 2014).. No systemic effects were noted in rats exposed to DINP in one dermal study (six-week duration) (Hazleton 1969). The lowest LOAEL for subchronic oral exposure was 60 mg/kg bw/day, based on an increased frequency of kidney lesions in all exposed males in a 13-week study in rats (Hazleton 1981a). In dogs, the NOAEL for subchronic exposure was 37 and 160 mg/kg bw/day based on increases in liver and/or kidney weights, accompanied by histopathological changes in males and females at 160 and 2000 mg/kg bw/day, respectively, in a 13-week study (Hazleton Laboratories 1971b).

The NOAEL of 500 mg/kg bw/day identified from a short-term and a subchronic study in monkeys indicates that monkeys and probably humans are less sensitive than rodents and dogs to liver effects, which is consistent with the hypothesis that species differences in the activation of PPARα or its signaling network by peroxisome proliferation may exist.

9.2.11.4.2 Carcinogenicity

B79P has not been classified for its potential carcinogenicity by other international agencies and no chronic toxicity/carcinogenicity studies were available for this phthalate. There was also no study available for the closest analogue MBzP. The OEHHA has recently reviewed evidence of the potential carcinogenicity of the analogue DINP and it has been concluded that DINP has been clearly shown, through scientifically valid testing according to generally accepted principles, to cause cancer and should be listed under Proposition 65 (OEHHA 2013b). Accordingly, DINP has been listed at the end of 2013 (OEHHA 2013c). DINP has not been classified for its potential carcinogenicity by other international agencies.

Refer to Section 9.2.2.1 for a summary of the available studies for the closest analogue DINP (Environment Canada and Health Canada 2015b).

Chronic studies conducted with DINP were examined to characterize the health effects of B79P. The most relevant studies are presented in Table 9-33.

Overall, the lowest oral doses associated with a significant increase in the incidence of tumours are 331-336 mg/kg bw/day based on a significant increase in hepatocellular tumours in female rats and mice, respectively (Bio/dynamics 1986; Moore et al. 1998b).

The lowest oral dose associated with chronic non-cancer effects was 27 mg/kg bw/day based on histologic changes in the liver of male rats exposed to DINP in a two-year carcinogenicity study (Bio/dynamics 1986); however, the incidence of those changes was not dose-related.

DINP CAS number was not provided in the Bio/dynamics study; however, it was reported as DINP-A (71549-78-5) in a risk assessment report of the US Consumer Product Safety Commission (2010). DINP-A has an isomeric composition similar to DINP-2 (28553-12-0). In comparison, the LOAEL in the Lington study was 152-184 mg/kg bw/day based also on liver effects in both male and female rats exposed to DINP-1. According to US CPSC (2010b), the difference in the toxic potency between the Bio/dynamics and Lington studies may be due to differences in dose selection, differences in toxicity between the two forms of DINP, and/or the use of a different rat strain. Since effects in the Biodynamic study (1986) were not found to be dose-related, the LOAEL of 152-184 mg/kg bw/day (NOAEL of 15-18 mg/kg bw/day) from the Lington et al. (1997) study is considered more relevant.

 9.2.11.4.3 Genotoxicity

In vitro, B79P was not mutagenic in a bacterial assay with S. typhimurium strains TA98, TA 100, TA 1535, TA 1537 and TA 15538, with or without metabolic activation (Monsanto 1982).

In the available genotoxicity studies for the closest analogue DINP, negative results were observed in in vitro and in vivo assays. Refer to Section 9.2.1.1 for a summary of the available studies for DINP (Environment Canada and Health Canada 2015b).

 9.2.11.4.4 Evidence of systemic toxicity in humans

No information is currently available on the potential effects of B79P in humans.

9.3 Characterization of risk to human health

The health effects data for medium-chain phthalates shows that there is evidence of developmental, reproductive and systemic effects. Of these, the critical effects for risk characterization are developmental effects on males, as the information available at this time is strongest for effects on the development of the reproductive system, such as alterations of feminization parameters, reproductive tract malformations and effects on fertility. Below are the various aspects taken into consideration for the characterization of risk to human health.

Relevant sources and durations of exposure

Sources of exposure for medium-chain phthalates are predominantly from indoor air, dust, food and breast milk. Due to the identified presence of some of these substances in manufactured items that may come into contact with skin, two scenarios were conducted to evaluate dermal exposure of these substances from dermal contact (adults and infants). Finally, since DIBP may also be present in children's toys and articles, oral exposure from mouthing these products was also evaluated.

With respect to the use of adhesives, sealants and coatings which contain medium-chain phthalates, exposure would not be considered to be of concern for human health based on the following:
Dermal absorption of medium-chain phthalates in rats is low (2-10%), and evidence shows that human skin is less permeable than rat skin to phthalate diesters. Also, retention in skin is 3 to 6 fold higher in rat compared to human (Mint and Hotchkiss 1993; Mint et al.1994). Distribution in tissues of rats is generally low, showing no accumulation, and excretion is rapid, within hours to days.

Exposure from use of these products would be of very short duration (acute) via the dermal route.

Phthalates in general are not considered acute toxicants, with LD₅₀ levels from dermal exposure being at minimum 2 to 5 fold higher than oral values (Draize et al. 1948; Eastman Kodak 1978; David et al. 2001; Monsanto Company 1970, cited in US EPA 2006, 2010).

Acute dermal toxicokinetic information indicates that reproductive organs are not a target organ, and that presence and residence time in other tissues (adipose and muscle) is extremely low after 7 days (0.02 to 0.3% of applied dose) (Elsisi et al. 1989).

This is consistent with the assessments of other jurisdictions who have focused their assessment on repeated exposures (ECHA 2013a; US CPSC CHAP 2014).

Adversity of apical endpoints

Effects seen after in utero exposure to certain phthalates are similar to effects observed with other chemicals, such as vinclozolin, linuron, flutamide and finasteride, which cause decreased AGD at birth and retained nipples (NR), but also otherreproductive tract malformations, including, but not limited to, hypospadias (HYP) and undescended testes (CRY) in male rats (Gray et al. 1999; Mylchreest et al. 1999, 2000; McIntyre et al. 2000, 2001, 2002; Barlow et al. 2002; Bowmann et al. 2003).

Health Canada considers that both significantly reduced AGD and NR in male rats after exposure to certain phthalates during the critical developmental window in gestation are established indicators of androgen deficiency during prenatal development, which can have severe and irreversible effects on the male reproductive system and may, in turn, interfere with fertility based on the following:

1. Reduced AGD and NR have been demonstrated to persist into adulthood (McIntyre et al. 2001; Bowmann et al. 2003; Hotchkiss et al. 2004) and were found to be predictors of compromised reproductive capacity in adulthood (Hotchkiss et al. 2004). Decreased AGD was also a sensitive predictor of lesions in the male reproductive tract (Barlow et al. 2004);
2. The measurement of AGD is part of the OECD Guidance for mammalian reproductive toxicity testing and assessment and, further, the measurement of NR was also recommended, both being based on the same premise as (1) above (OECD 2008);
3. Several international regulatory agencies and scientific bodies have used AGD and NR as points of departure in characterizing the potential risk of phthalates on the developing male reproductive system (ECJRC 2007; NICNAS 2008; Danish EPA 2012; ECHA 2013a; Germany 2014; US CPSC CHAP 2014).

For these reasons, AGD and NR, along with other adverse effects related to RPS, will be considered in the characterization of risk for phthalates.

Human relevance of reproductive/developmental effects

The specific MOA of phthalate-induced effects on the male reproductive system has not been fully elucidated and the proposed mechanism(s) of action is reviewed elsewhere (section 9.2; NAS 2008). The effects seen in the developing reproductive tract of male rats show excellent concordance with the endpoints of concern in human males, namely infertility, decreased sperm count, cryptorchidism, reproductive tract malformations, hypospadias and testicular tumours (germ-cell-derived in humans and Leydig-cell-derived in rats), which have been postulated to comprise the human testicular dysgenesis syndrome (TDS) (Health Canada 2015a; NAS 2008). It is noted, however, that there are no consistent human data directly linking the hypothesized syndrome with exposure to phthalates (NAS 2008). Regardless, several attempts have been made to determine whether these effects are observable in human tissues.

Overall, limited data on the effects on human foetal testes suggest effects such as a reduction in the number of germ cells and an increase in MNGs, with no consistent effects on testosterone biosynthesis (Hallmark et al. 2007; Lambrot et al. 2009; Yuan et al. 2012; Desdoits-Lethimonier et al. 2012). Recent reviews identified a number of limitations to the interpretation of recent xenograft studies (Mitchell et al. 2012; Heger et al. 2012; Spade et al. 2014), such as, but not limited to: 1) the substantial individual variability associated with the use of human biological material and the use of pooled testes at different ages; 2) the methods of material collection of human foetal testes, which are highly variable compared to animal models; 3) the potentially short exposure period being insufficient; 4) other sources of reproductive hormones not being considered; and 5) the potential difference in metabolizing capabilities of the animal hosts and humans (Albert et al. 2014; Habert et al. 2014).

This method was recently used with prepubertal primate testes xenographted into mice, showing that exposure to certain phthalates caused perturbations of steroidogenic gene expression, impaired tubule formation and germ cell differentiation as well as decreased spermatogonial numbers after subchronic exposure (Rodriguez-Sosa et al. 2014).

Habert et al. (2014) also cautioned that in vitrointerspecies comparisons need to be carried out carefully by selecting appropriately comparable stages of gestation, using identical methods that can measure gametogenesis and steroidogenesis across the different species, and using non-contaminated explants of very similar size.

Similarly to the position presented by ECHA (2013a) in their assessment of DINP and DIDP, it is acknowledged that there are differences between human and rat steroidogenesis, but the processes involved in male reproductive development are similar. The critical enzymes involved in steroidogenesis are identical in rats and humans, and all mammals are believed to have parallel activation mechanisms for androgen-dependent processes. As ECHA (2013) stated, it is possible that a sufficient exposure may cause anti-androgenic effects in human foetuses similar to those observed in animals. Habert et al. (2014), in a report presented at the 7th Copenhagen Workshop on Endocrine Disrupters, supported by the Danish EPA and the Society for Reproduction and Fertility, also stated that the rat model is relevant and important to human health risk assessment when choosing a common effect in both species. This position was also supported by the US CPSC CHAP in their cumulative risk assessment (US CPSC CHAP 2014).

With respect to indicators of androgen deficiency in human males regardless of the cause, reduced AGD and penile size were reported in human boys with cryptorchidism and hypospadias (Thankamony et al. 2014). Inverse associations with male AGD and environmental chemicals (phthalates and BPA) have been reported in boys from US, China and Japan, although consistency in methods and reproducibility of this endpoint has been a challenge (Swan et al. 2006; Miao et al. 2011; Suzuki et al. 2012). Although direct evidence of prenatal anti-androgen exposure and reduced adult reproductive capacity in humans is currently lacking, associations between AGD, decreased penile size, decreased semen quality, infertility and decreased serum levels of testosterone in adult men have been reported (Mendiola et al. 2011; Eisenberg et al. 2011, 2012a, 2012b; Bornehag et al. 2014; Bustamante-Montes et al. 2013). A recent review by Juul et al. (2014) stated that, like in animal studies, AGD measurements are a useful continuous gauge of androgen exposure in utero in humans.

Consideration of human relevance of other systemic effects

It is well documented that phthalates can induce peroxisome proliferation in the liver as well as increased liver weight in rats and mice. In some cases, liver cancer was also observed following longer-term oral administration of high doses of phthalates. It is well established that the peroxisome proliferator-activated receptor (PPAR) α plays a role in peroxisome proliferation-induced liver effects (Corton and Lapinskas 2005). However, the relevance of the hepatotoxic effects of phthalates observed in rodents is difficult to establish due to the species-specific differences in peroxisomal proliferation response (rodents being significantly more sensitive than humans to PPARα-mediated induction of peroxisome proliferation) (ECB 2008, NICNAS 2010, US CPSC 2010c). Several recent studies have suggested that the mechanisms of liver toxicity of peroxisome proliferators have not been entirely elucidated and that multiple pathways may exist, some of those likely PPARα-independent (Ito et al. 2007, Yang et al. 2007, Eveillard et al. 2009, Ren et al. 2010, IARC 2012). Based on this, liver effects cannot be precluded as an effect potentially relevant to humans and should be included in the characterization of the health effects of phthalates. More detailed information on the mode of action of liver carcinogenicity in rodents with peroxisome proliferators is available in Health Canada (2015c).

9.3.1 DIBP

Based principally on consideration of the weight-of-evidence-based classification of DIBP by the European Commission as Category 1B - reproductive toxicant (EC No 1272/2008) and consideration of  the available information, critical effects associated with exposure to DIBP are developmental effects on the male reproductive system. Adverse effects in the parameters used to measure RPS after in utero exposure to DIBP included decreased testicular testosterone levels, decreased AGD, nipple retention, delayed preputial separation, reproductive tract malformations (cryptorchidism, hypospadias, exposed ospenis, cleft prepuce), testicular pathology as well as potential effects on fertility through abnormalities in sperm. DIBP was also shown to have inhibitory effects on the expression of genes that are involved in testosterone production in vivo.

Based on the available information at this time, it appears as though the foetal male rat is the life stage most sensitive to the effects of exposure to DIBP. No conclusions can be made on whether the foetal mouse is less or more sensitive than the rat, as no studies examining the parameters used to measure RPS in this species using DIBP were available for this life stage. However, some evidence in the prepubertal/pubertal life stage and in the adult stage using DBP as an analogue indicated that mice were not as sensitive to the reproductive effects of DIBP later on in life. Rabbits appear to be as sensitive to the adverse reproductive effects of DBP as rats at similar doses. No studies were available identifying the potential reproductive/developmental toxicity of DIBP via any other route of exposure. See Table 9-34 for a summary of the critical effects of DIBP that will be used for risk characterization.

Developmental effects in female rats were identified at equal or higher doses than males after oral exposure with critical endpoints related to growth alterations, alterations of reproductive development, functional deficit, lethality and mild teratogenicity. Reproductive effects of DIBP in females (alterations of fertility and pregnancy outcomes) were observed at 750 mg/kg bw/day and above.

The potential sources of exposure to DIBP for the general population from environmental media and food is expected to be from food (oral ingestion), breast milk, indoor air (inhalation) and house dust (oral ingestion), with the major drivers of exposure being breast milk and indoor air. With respect to products used by consumers, potential sources of exposure may occur from oral mouthing of plastic toys and articles (infants 0 to 18 months), dermal contact with plastic articles (e.g., manufactured items such as exercise equipment and floor coverings) and DIBP presence in cosmetics. Finally, urinary metabolite concentrations of DIBP (MIBP and 2OH MIBP) were evaluated, and reverse dosimetry was used to calculate DIBP intakes. These intakes were derived from internal urinary concentrations and therefore represent exposure from all routes and sources at a given time.

Upper-bounding intakes, sources and respective margins of exposure for the relevant age-specific populations (when points of departure related to the MOA of antiandrogenicity are used) are presented in Table 9-35.

The above MOEs are considered adequate to account for uncertainties in the exposure and health effects databases, and protective of the potential reproductive effects of phthalate toxicity not only in males at older life stages but also in females, in addition to effects in other organ systems (systemic toxicity).

Based on the information available, there is evidence that DIBP has effects on the developing male reproductive system, indicative of RPS, and may have a common mode of action with other phthalates in the grouping. Although the above MOEs are considered adequate on an individual basis, this does not address the potential risk of concurrent exposure to DIBP and other phthalates exhibiting a similar mode of action.

9.3.2 DCHP

Based principally on the weight of evidence from the available information, critical effects associated with exposure to DCHP are developmental effects on the male reproductive system following exposure in utero as well as systemic effects on the liver and kidney after subchronic exposure.

Adverse effects in the parameters used to measure RPS after in utero exposure to DCHP included decreased foetal testosterone production, decreased AGD, retention of areole mammae and delayed preputial separation, along with testicular pathology. DCHP was also shown to have inhibitory effects on two enzymes that are involved in testosterone production in vitro, although this effect was not confirmed with confidence in in vivostudies.

Based on the available information from the 2-generation OECD Guideline study (Hoshino et al. 2005), it appears as though the foetal male rat is the most sensitive to adverse effects following exposure to DCHP compared to rats at other life stages; effects in F1 adult males, although occurring at the same dose levels of 1200 ppm, were less severe in nature and did not affect the overall reproductive success of the males. Further, there were no adverse reproductive effects when DCHP was administered to adult F0 males for 14 weeks (10 weeks prior to mating until 26 days after confirmed copulation). No conclusions can be made on whether the mouse is less or more sensitive than the rat as no studies examining the parameters used to measure RPS in this species using DCHP were available. Further, no studies were available identifying the potential reproductive/developmental toxicity of DCHP via any other route of exposure. See Table 9-36 for a summary of critical effects of DCHP that will be used for risk characterization.

Developmental effects in female rats were identified at higher doses than males after oral exposure, with critical endpoints related to growth alterations (organs and body weights) and lethality. Reproductive effects of DCHP in adult females (pregnancy outcome alterations) were reported at high doses.

The database for repeated-dose toxicity of DCHP suggests that liver and kidneys are also the main target organs for this phthalate. No effects were reported following chronic exposure of dogs and rats. Consequently, the carcinogenic potential of DCHP is considered limited. The lowest LOAEL identified from repeat-dose studies was 75 mg/kg bw/day (NOAEL of 25 mg/kg bw/day) based on increases in liver weight (females), accompanied by histological changes in the liver and kidneys in both sexes at the two highest doses (200 and 500 mg/kg bw/day) in a subchronic feeding study in rats (de Ryke and Willems 1977).

The principal source of exposure to DCHP for the general population is expected to be from exposure to house dust and indoor air and from use of products used by consumers, such as sealants and adhesives. Although DCHP is present in food, intakes were calculated and exposure was estimated to be negligible.

Comparisons of upper-bounding estimates for oral exposure to DCHP from dust and indoor air for the most exposed age groups, with the appropriate critical effect levels, result in MOEs ranging from 166 667 to over 1 million, which are considered adequate to address uncertainties in the exposure and health effects databases. Further, these MOEs are considered protective of potential reproductive effects of phthalate toxicity not only in males at older life stages but also in females, in addition to effects in other organ systems (systemic toxicity).

Upper-bounding intakes, sources and respective margins of exposure for the relevant age-specific populations (when points of departure related to the MOA of antiandrogenicity are used) are presented in Table 9-37 below.

Based on the information available, there is evidence that DCHP has effects on the developing male reproductive system, indicative of RPS, and may have a common mode of action with other phthalates in the grouping. Although the above MOEs are currently considered adequate on an individual basis, this does not address the potential risk of concurrent exposure to DCHP and other phthalates exhibiting a similar mode of action.

9.3.3 DMCHP

Based principally on the weight of evidence from the available information on its analogue, DCHP, the critical effects associated with exposure to DMCHP are developmental effects on the male reproductive system following exposure in utero as well as systemic effects on the liver and kidney after subchronic exposure. See above section and Table 9-36 for a summary of the critical health effects used for this phthalate.

A potential source of exposure for DMCHP for the general public is house dust; however, no other information as to monitoring of DMCHP in other media was identified.

A comparison of upper-bounding estimates for oral exposure to DMCHP from ingestion of dust for all age groups with the appropriate critical effect levels identified from studies conducted with DCHP (see Section 9.3.2) results in MOEs ranging from 462 963 to over 1 million. These MOEs are considered adequate to address uncertainties in the exposure and health effects databases, and protective of the potential reproductive effects of phthalate toxicity not only in males at older life stages but also in females, in addition to effects in other organ systems (systemic toxicity).

Upper-bounding intakes, sources and respective margins of exposure for the relevant age-specific populations (when points of departure related to the MOA of antiandrogenicity are used) are presented in Table 9-38 below.

Based on the information available for DCHP, there is evidence that DMCHP has potential effects on the developing male reproductive system, indicative of RPS, and may have a common mode of action with other phthalates in the grouping. Although the MOEs are currently considered adequate on an individual basis, this does not address the potential risk of concurrent exposure to DMCHP and other phthalates exhibiting a similar mode of action.

9.3.4 DBzP

Based on an examination of the health effects database for the closest analogue to DBzP, MBzP (see Category Approach Document, Health Canada 2015a), the critical effects for the characterization of risk associated with exposure to DBzP are considered to be developmental effects on the male reproductive system following exposure in utero and systemic effects, such as decreased body weight gain and food consumption.

Adverse effects in the parameters used to measure RPS after in utero exposure to MBzP included decreased AGD and reproductive tract malformations (CRY). Based on the criteria used in selecting the appropriate analogue, it is considered appropriate to use the information on MBzP to characterize the toxicological profile of DBzP with respect to these reproductive/developmental effects.

Based on the limited available information, there appears to be no difference in sensitivity to the developmental effects of exposure to MBzP at different life stages. No conclusions can be made on whether the foetal mouse is less or more sensitive than the rat, as no in vivo studies examining the parameters used to measure RPS in this species using MBzP or DBzP were available for this life stage; however, there was some indication that MBzP was more toxic to pregnant mice and their offspring, but not rats at similar doses. No studies were available identifying the potential reproductive/developmental toxicity of DBzP via any other route of exposure. See Table 9-39 for a summary of the critical effects of MBzP that will be used for risk characterization.

Based on the information extrapolated from studies using MBzP, developmental effects in female rats were identified at doses equal to those of males, as MBzP is teratogenic and embryolethal at maternally toxic doses. One rat study (Ema et al. 2003) reported a higher sensitivity in male offspring to the developmental toxicity of MBzP.

With regards to systemic effects, the lowest LOAEL for short-term exposure was 167 mg/kg bw/day based on a dose-dependent decrease in body weight gain (22% decrease for adjusted weight gain) associated with a decrease in food consumption (8-15%) in dams in a developmental toxicity study in rats (Ema et al. 2003).

A potential source of exposure to DBzP for the general public is house dust; however, no other information as to DBzP monitoring in other media was identified. DBzP was identified as a potential food contact phthalate in the US; however, no monitoring as to its presence in food was identified (DBzP has been identified as a candidate for monitoring as part of future Health Canada total diet surveys). While association of potential DBzP use in products used by consumers was observed, no submissions as to its use, manufacturing and import in Canada were identified. Therefore, no direct consumer product exposure is expected.

A comparison of upper-bounding estimates for oral exposure to DBzP from ingestion of dust for all age groups with the appropriate critical effect levels results in MOEs over 1 million, considered adequate to address uncertainties in the exposure and health effects databases for DBzP on an individual substance basis. Further, these MOEs are considered protective of the potential reproductive effects of phthalate toxicity not only in males at older life stages but also in females, in addition to effects in other organ systems (systemic toxicity).

Upper-bounding intakes, sources and respective margins of exposure for the relevant age-specific populations (when points of departure related to the MOA of antiandrogenicity are used) are presented in Table 9-40 below.

Based on the information available, there is evidence that DBzP has effects on the developing male reproductive system, indicative of RPS, and may have a common mode of action with other phthalates in the grouping. Although the above MOEs are currently considered adequate on an individual basis, this does not address the potential risk of concurrent exposure to DBzP and other phthalates exhibiting a similar mode of action.

9.3.5 B84P

Based on an examination of the health effects database for BBP, MBzP and DIBP as analogues, the critical effects for characterization of the risk associated with exposure to B84P are considered to be carcinogenicity and developmental effects on the male reproductive system.

No chronic/carcinogenicity studies were available for B84P. Its close analogue BBP has been classified by IARC as Group 3 “Not classifiable as to its carcinogenicity to humans” (IARC 1999). Also, the California EPA recently concluded that BBP should not be listed under Proposition 65 as a carcinogen (OEHHA 2013a). Mononuclear cell leukemia was reported in female Fischer rats exposed to BBP in a 1982 study but not in a 1997 repeat study (NTP 1982, 1997a). It has been proposed that this type of lesion is specific to aging rats of this strain and is likely to be of no relevance to humans. BBP also induced an increase in pancreatic tumours primarily in male rats, the full expression of which was prevented in a dietary restriction protocol (NTP 1997a). There was no evidence of carcinogenicity in mice (NTP 1982). Pancreatic acinar cell carcinoma is rare in F344 male rats, having never been observed in untreated male F344 rats in NTP studies (NTP 1997a). Klaunig et al. has proposed a PPARα-dependent mode of action for the induction of pancreatic acinar cell tumours by BBP (Klaunig et al. 2003). However, there are data gaps in this proposed MOA and there are no existing data suggesting that PPARα is actually involved in the induction of pancreatic acinar cell tumours by BBP or in BBP tumorigenesis in general (Klaunig et al. 2003; OEHHA 2013a). Based on the little information available on the potential MOA involved in the increase in incidence of pancreatic tumours in rats and its potential occurrence in humans, it is considered that this type of tumour is of unclear relevance to humans.

In a Priority Substances List (PSL) report published by Environment Canada and Health Canada in 2000, it was suggested that BBP could be considered, at most, possibly carcinogenic to humans, likely inducing pancreatic tumours through a non-genotoxic (albeit unknown) mechanism (Environment Canada and Health Canada, 2000). Considering this and the available information for B84P regarding genotoxicity, which indicates that this phthalate is not likely to be genotoxic, a threshold approach is used to characterize risk to human health from exposure to B84P. An examination of chronic studies conducted on BBP indicates that potential pancreatic tumours would occur at doses higher than those at which developmental effects have been observed.

With respect to non-cancer effects, the lowest LOAEL for subchronic oral exposure was 381 mg/kg bw/day (NOAEL of 151 mg/kg bw/day) based on histopathological changes in the pancreas, gross pathological alterations in the liver and a significant increase in relative kidney weight in male Wistar rats exposed to BBP for three months (Hammond et al. 1987). No adverse effects were observed in mice and dogs exposed to high doses of BBP for three months (NTP 1982b; Hammond et al. 1987). The lowest LOAEL for chronic oral exposure was 300 mg/kg bw/day based on a significant increase in the incidence of nephropathy noted in all groups of exposed females in a two-year study in rats (NTP 1997a).

With regard to developmental effects, adverse effects in the parameters used to measure RPS after in utero exposure included decreased testicular testosterone levels, decreased AGD, NR, delayed PPS, reproductive tract malformations (CRY, HYP, exposed os penis, cleft prepuce), testicular pathology and potential effects on fertility through abnormalities in sperm. DIBP was also shown to have inhibitory effects on the expression of genes that are involved in testosterone production in vivo. MBzP also induced increased incidences of CRY and decreased AGD in the foetal rat at similar dose levels. Based on the criteria used in selecting the appropriate analogues, it is considered appropriate to use the information on BBP to characterize the toxicological profile of B84P with respect to these reproductive/developmental effects, as it appears to be the most potent of the three analogues. BBP has been classified by the European Commission as Category 1B - reproductive toxicant (EC No 1272/2008).

Using BBP as an analogue for the adult life stage indicates that the adult appears to be less sensitive compared to younger animals, although no clear conclusions can be made with any confidence. Further, no conclusion can be made on whether the foetal mouse is less or more sensitive than the rat, as no in vivo studies examining the parameters used to measure RPS in this species using the three analogues were available for this life stage; however, there was some indication that MBzP was more toxic to pregnant mice and their offspring, but not rats at similar doses. No studies were available identifying the potential reproductive/developmental toxicity of B84P via any other route of exposure. See Table 9-41 for a summary of the critical effects of BBP that will be used for risk characterization of B84P.

Based on the information extrapolated from studies using MBzP, DIBP and BBP, developmental effects in female rats were identified at doses equal to or higher than those for male offspring and were based on growth alteration, lethality, altered reproductive organ weights, delay of puberty and teratogenicity (variations and skeletal and/or visceral malformations) after gestational exposure as well as pregnancy outcomes, alterations of reproductive organ weights, hormone levels (progesterone and prolactin) and reproductive-related organ visual examination and histopathology in adulthood.

Potential exposure to B84P for the general population is expected to be from oral dust ingestion. However, dust intakes were calculated using B79P dust concentrations for B84P, as no laboratory standard is available for B84P. Uncertainty therefore exists with respect to estimating exposure from this source. Additionally, since B84P is a medium-volume chemical (greater than 100 000 kg, section 71 reporting) and has been notified to be used in textile applications in other jurisdictions, dermal exposure from handling plastic articles was assessed for infants (0 to 18 months) and adults (20+ years). The exposure estimates and respective margins of exposure are outlined in Table 9-42.

Comparisons of upper-bounding estimates for dermal exposure to B84P from contact with plastic articles (textiles, upholstery, etc.) for all age groups, with the appropriate critical effect levels, result in MOEs ranging from 2352 to 6991, which are considered adequate to address uncertainties in the exposure and health effects databases for B84P. Comparisons of upper-bounding estimates for oral exposure to B84P from dust ingestion for children 0 to 6 months of age result in MOEs over 1 million, which are also considered adequate to address uncertainties in the exposure and health effects databases for B84P. Further, these MOEs are considered protective of the potential reproductive effects of phthalate toxicity not only in males at older life stages but also in females, in addition to effects in other organ systems (systemic toxicity).

Upper-bounding intakes, sources and respective margins of exposure for the relevant age-specific populations (when points of departure related to the MOA of antiandrogenicity are used) are presented in Table 9-42 below.

Based on the information available on DIBP, BBP and MBzP, there is evidence that B84P has potential effects on the developing male reproductive system, indicative of RPS, and may have a common mode of action with other phthalates in the grouping. Although the MOEs are currently considered adequate on an individual basis, this does not address the potential risk of concurrent exposure to B84P and other phthalates exhibiting a similar mode of action.

9.3.6 DIHepP

Based principally on the weight of evidence from the available information, the critical effects associated with exposure to DIHepP are developmental effects on the male reproductive system following exposure in utero and liver and kidney effects following exposure at the adult stage. Adverse effects in the parameters used to measure RPS after in utero exposure to DIHepP included decreased foetal testicular testosterone production, decreased AGD, NR, delayed PPS, testicular pathology and potential effects on fertility through abnormalities in sperm.

Based on the available information at this time, it appears as though the foetal male rat is the life stage most sensitive to the developmental effects of exposure to DIHepP, although it should be noted that there is a lack of studies examining the effects of this substance in prepubertal/pubertal males. No conclusions can be made on whether the mouse is less or more sensitive than the rat, as no studies examining the parameters used to measure RPS in this species using DIHepP were available. Further, no studies were available identifying the potential reproductive/developmental toxicity of DIHepP via any other route of exposure. See Table 9-43 for a summary of critical effects of DIHepP that will be used for risk characterization.

Based on the limited information from studies using DIHepP, developmental effects in female rats were identified at doses higher than those for male offspring and were based on alterations of growth, functional deficit, lethality and teratogenicity after gestational exposure as well as alterations of reproductive performance and pregnancy outcomes in adulthood.

No carcinogenicity studies were available for DIHepP. However, based on the results of the multigenerational study, the carcinogenic potential of DIHepP is likely limited. Effects indicative of peroxisome proliferation were observed in a repeated-dose study conducted in rats and mice. This could potentially result in the increased incidence of hepatocellular tumours. However, this has been mostly observed at high doses. The mechanisms of liver carcinogenicity in rodents with peroxisome proliferators have not been fully elucidated. Consequently, relevance in humans remains unclear and cannot be ruled out.

Consideration of the limited available information on genotoxicity indicates that DIHepP is not likely to be genotoxic.

The lowest LOAEL for subchronic and chronic oral exposure was 222-716mg/kg bw/day (NOAEL 50-162 mg/kg bw/day) and 227-750 mg/kg bw/day (NOAEL 50-168 mg/kg bw/day), respectively, based on liver and kidney effects in the 2-generation study (McKee et al. 2006).

Potential exposure to DIHepP for the general population is expected to be from oral dust ingestion. DIHepP use was identified in adhesives and sealants; however, no products used by consumers that would lead to direct subchronic and chronic exposures were identified.

Comparisons of upper-bounding estimates for oral exposure to DIHepP from ingestion of dust for children 0 to 0.5 year of age and adolescents 12 to 19 years of age and above, with the appropriate critical effect levels, result in MOEs ranging from 45 455 to over 1 million, which are considered adequate to address uncertainties in the exposure and health effects databases for DIHepP on an individual substance basis. Further, these MOEs are considered protective of the potential reproductive effects of phthalate toxicity not only in males at older life stages but also in females, in addition to effects in other organ systems (systemic toxicity).

Upper-bounding intakes, sources and respective margins of exposure for the relevant age-specific populations (when points of departure related to the MOA of antiandrogenicity are used) are presented in Table 9-44 below.

Based on the information available, there is evidence that DIHepP has effects on the developing male reproductive system, indicative of RPS, and may have a common mode of action with other phthalates in the grouping. Although the MOEs are currently considered adequate on an individual basis, this does not address the potential risk of concurrent exposure to DIHepP and other phthalates exhibiting a similar mode of action.

9.3.7 B79P

Based principally on the weight of evidence from the available information, a critical effect associated with oral exposure to B79P is carcinogenicity.

B79P and the analogue MBzP have not been classified for their potential carcinogenicity by other international agencies, and no chronic/carcinogenicity study was identified for either phthalate. The OEHHA has recently reviewed the evidence of the potential carcinogenicity of the analogue DINP and it has been concluded that DINP has been clearly shown, through scientifically valid testing according to generally accepted principles, to cause cancer and should be listed under Proposition 65 as a carcinogen (OEHHA 2013b). Accordingly, DINP has been listed at the end of 2013 (OEHHA 2013c). DINP has not been classified for its potential carcinogenicity by other international agencies.

DINP has been tested in a number of carcinogenicity studies in rats and mice. Statistically significant increases in many tumour types were observed in animals exposed to high doses of this phthalate, such as an increase in hepatocellular tumours in rats and mice, mononuclear cell leukemia of the spleen in Fischer rats and renal tubular cell carcinomas in rats. Overall, mechanisms of carcinogenicity of DINP in rodents have not been fully elucidated, and formation of tumours following chronic exposure to DINP is of unclear or questionable relevance to humans. Nevertheless, the potential for DINP, and therefore B79P, to be carcinogenic cannot be ruled out.

Consideration of the available information on genotoxicity for B79P and DINP indicates that B79P is not likely to be genotoxic.

For non-cancer effects, the lowest oral LOAEL identified was 152-184 mg/kg bw/day (NOAEL of 15-18 mg/kg bw/day) based on liver effects in both male and female rats exposed to DINP (Lington et al. 1997).

B79P is also associated with developmental effects on the male reproductive system following exposure in utero,but at doses higher than those for which B79P is presumed to induce effects on liver and kidneys, based on a review of the health effects of its analogue DINP. Refer to Section 9.2.2.1 of the assessment for DINP (Environment Canada and Health Canada 2015b) for a review of the potential reproductive/developmental effects of DINP for all life-stages. Developmental effects of B79P were reported to include decreased AGD, retained nipples and reproductive tract malformations (cryptorchidism and epispadias). Information from analogues have also reported reproductive effects at later life stages, such as decreased sperm counts and motility as well as reproductive organ weights in adulthood.

Potential exposure to B79P for the general population is expected to be from oral dust ingestion. Additionally, since B79P is a medium-volume chemical (greater than 100 000 kg, section 71 reporting) and has been notified to be used in textile applications in other jurisdictions dermal exposure from handling plastic articles was assessed for infants (0 to 18 months). The exposure estimates and respective margins of exposure are outlined in Table 8-49.

Comparisons of upper-bounding estimates for dermal exposure to B79P from contact with plastic articles (textiles, upholstery, etc.) for infants 0 to 18 months of age, with a NOAEL of 15-18 mg/kg bw/day for DINP based on liver effects in both male and female rats (Lington et al. 1997), result in an MOE of 694. Comparisons of upper-bounding estimates for oral exposure to B79P from dust ingestion for children 0 to 0.5 year of age and adolescents 12 to 19 years of age, with the same endpoint, result in MOEs ranging from 319 149 to over 1 million. In both cases, MOEs are considered adequate to address uncertainties in the exposure and health effects databases for B79P. See Table 9-45 below.

Based on the information available, there is evidence that B79P has effects on the developing male reproductive system, indicative of RPS, and may have a common mode of action with other phthalates in the grouping. Although the MOEs are currently considered adequate on an individual basis, this does not address the potential risk of concurrent exposure to B79P and other phthalates exhibiting a similar mode of action.

9.3.8 CHIBP, BCHP and BIOP

An examination of the potential developmental and reproductive toxicity of CHIBP, BCHP and BIOP using appropriate analogues for read-across revealed that these medium-chain phthalates have the potential to have significant effects on the developing male, in addition to systemic effects (liver, kidney).

Based on the information available, it can be concluded that CHIBP, BCHP and BIOP meet the criteria for inclusion in the evaluation of the potential cumulative risk of phthalates on the developing male reproductive system based on evidence of the effects of their analogues; however, as there is no exposure at this time, they will not be included in risk characterization in a cumulative context.

Results from a section 71 industry survey for 2012 suggest that CHIBP, BCHP and BIOP are not currently in use above the specified reporting threshold, and the likelihood of exposure to the general population in Canada is considered to be negligable. Consequently, the risk to human health for these substances is not expected.

9.4 Uncertainties in evaluation of risk to human health

Empirical health effects data for medium-chain phthalates range from robust to very limited and create uncertainty in the evaluation of risk to humans. There is some uncertainty associated with the use of analogues to characterize the human health effects of phthalates with limited or no available toxicological information. This lack of available toxicological information applies to DMCHP, CHIBP, BCHP, DBzP, B84P, BIOP and B79P.

There are no studies by any route of administration on developmental neurotoxicity for any of these phthalates. In the case of DIBP, DMCHP, CHIBP, BCHP, DBzP, B84P, BIOP and B79P, there are no 2-generation studies. The majority of the reproductive and developmental toxicity data for these medium chain phthalates is generally limited to one species (rat) and to males. There is some uncertainty associated with not only the potential biological significance of effects, but also the sensitivity of effects after exposure to this substance group in both female and male humans.

There is also limited or no information on repeated-dose effects via the inhalation and dermal routes of exposure for the majority of phthalates in this grouping.

There is uncertainty regarding the potential carcinogenicity of some medium-chain phthalates due to the lack of long-term studies (DIBP, DMCHP, CHIBP, BCHP, DBzP, DIHepP, B84P, BIOP and B79P). However, there is available information from carcinogenicity studies for certain analogues (BBP and DINP) to address this endpoint for B84P and B79P, respectively.

There is uncertainty associated with the mode of induction of tumours for BBP and DINP. Postulated mechanisms have been identified for some tumour-types, but the mechanisms have not been fully elucidated.

Studies used for risk characterization for medium-chain phthalates ranged from high-quality OECD Guideline studies to those with limited information. This uncertainty was addressed in the selection of precautionary target MOEs, where required.

Although a rigorous evaluation approach was conducted with the available human epidemiological data, uncertainty still exists as to the relevance of these studies implicating the potential hazard that certain phthalates pose to humans.Thoroughly conducted epidemiologic studies showing robust and consistent associations between an exposure factor and an outcome may provide strong implication for causal inference. However, observational studies in diverse populations pose challenges in both the measure of exposure and the measure of the outcome, and inherently have biases and confounding factors (Lucas and McMichael 2005). The majority of epidemiological studies examined were cross-sectional in which a temporal sequence whereby exposure precedes the outcome cannot be established. In addition, several outcomes associated with phthalate exposure in human epidemiological studies have long latencies (such as cancer, diabetes, obesity, cardiovascular disease) and multifactorial etiologies (geographical location, socioeconomic status, diet, lifestyle factors, genetic propensity, nonchemical stressors) and are chronic in nature, whereas phthalates have short biological half-lives and their measurement therefore reflects a snapshot of recent exposure. Moreover, biomonitoring data showed that exposure to certain phthalates is ubiquitous and therefore cannot be dichotomized as present or absent but is instead a continuous variable, often with a limited range.

While it has been argued that even in the absence of consistent methods, a robust association should yield consistent findings (La Kind et al., 2012), poor reproducibility continues to feature prominently in epidemiological studies involving phthalates. Adding to the lack of clarity is the fact that humans are simultaneously exposed to multiple phthalates from multiple sources via multiple routes, as well as other environmental agents that may share coinciding effect domains, including bisphenol A, certain metals and organochlorine compounds, such as PCBs, dioxins and various persistent organic pesticides. In its final report in 2014, the US Chronic Hazard Advisory Panel (CHAP) on Phthalates concluded that although there is a growing body of studies reporting associations between phthalate exposure and human health, and many of the reported health effects are consistent with testicular dysgenesis syndrome in humans, there are acknowledged limitations of these studies similar to those described above. These were therefore not used in risk characterization (US CPSC CHAP 2014). Another recent review also found that epidemiological evidence for associations with reproductive and developmental effects from phthalates is minimal to weak in most cases (Kay et al. 2014).

No monitoring data on the presence of BCHP, CHIBP and BIOP in environmental media and food were identified for Canada or elsewhere. Based on information submitted to Environment Canada and known international use patterns, exposure to the general population is not expected.

There are uncertainties associated with estimating intakes of phthalates from environmental media due to minimal monitoring data available for these phthalates in air, drinking water and soil. Confidence is moderate to high that derived intake estimates from household dust are representative of the potential exposure of the general Canadian population, since the exposure estimates are based on a Canadian house dust monitoring study. However, B84P was not monitored in dust, as no standard was available. Moreover, despite the presence of BIOP and DBzP peaks in dust analysis chromatograms, BIOP was assumed not to be present in dust based on a zero kilogram production volume in Canada. Therefore, uncertainty exists in the quantification of dust exposure for these two substances. However, despite these collective uncertainties, there is confidence that assumptions made in estimating exposure are conservative enough to account for these uncertainties (use of higher metric concentrations).

For phthalate presence in food, there is uncertainty in the literature regarding the presentation of LODs and LOQs, as a portion of publications present the LOD of the instrument rather than incorporating the background level of phthalate contamination.

For quantification of food exposure, from DIBP and DCHP presence in food, US and UK surveys were used for analysis. Uncertainty therefore exists, as these intakes are extrapolated for the general Canadian population. There is uncertainty associated with potential intake from food containing DBzP since this substance is present on international databases indicating potential food contact exposure. However, no monitoring data as to its presence in food was identified.

There is also uncertainty associated with exposure estimates calculated from DIBP monoester (MIBP) presence in breast milk. This is related to the quantification of exposure (conversion of metabolite exposure to parent phthalate exposure) and the evaluation of margins of exposure between exposure intakes derived from metabolite exposure (infants ingesting breast milk containing DIBP), and toxicology studies evaluating effects of parent phthalate exposure.

For DIBP, B84P and B79P, there is uncertainty as to the estimation of dermal exposure from contact with manufactured items containing these phthalates, based on limited substance-specific information with regard to the presence and migration over time of phthalates from these products. Therefore, it is unclear as to how much phthalate is available for transfer to skin from contact. Additional uncertainty is associated with the parameters used (e.g., dermal absorption and migration rate) in estimating exposure from manufactured items; however, there is confidence that the assumptions used were conservative.

A number of assumptions have been made to derive intake estimates from biomonitoring data that represent a source of uncertainty, that is to say, the assumption that spot urine samples are representative of steady state daily concentrations and assumptions around the use of creatinine-corrected concentrations. However, there is confidence that the assumptions used in deriving estimates of intakes are appropriate and conservative. Also, confidence in the biomonitoring database for DIBP is high, as it represents a substantially large number of data points collected recently in Canadian individuals spanning a wide age spectrum, including subpopulations such as pregnant woman.

Due to the lack of or limited health effects data for all relevant routes and durations of exposure, route-to-route extrapolation was required, and/or use of effect levels from studies with a longer or shorter duration of exposure than the exposure scenarios was applied. Conservatism in the derivation of exposure estimates is recognized by the application of conservative dermal absorption values.

Uncertainty is recognized in the potential oral bioavailability of medium-chain phthalates, in particular the estimated internal dose at which effects were observed in animal studies after administration. Information exists that absorption of these phthalates is highly variable (30-95%) and is influenced by rates of metabolism and excretion of an organism and by different routes at any given time of measurement. These limitations do not allow for accurate adjustments in risk characterization for each phthalate; however, estimated MOEs are considered adequate to account for this uncertainty.

Date modified:

2016-05-24