Draft Screening Assessment Phthalate Substance Grouping: Sections 9 and 10 (2024)

Environment Canada
Health Canada
October 2017

Back to Sections 1-8

Appendices

• 9. Potential to cause harm to human health
  • 9.1 Exposure assessment
  • 9.2 Health effects assessment
  • 9.3 Characterization of risk to human health
  • 9.4 Uncertainties in evaluation of cumulative risk to human health
• 10. Conclusion
• References

DMP

Exposure estimates were calculated using human biomonitoring data as well as data on DMP occurrence in indoor air, soil, dust, food and cosmetics. They are presented in the SCP SOS report (Environment Canada, Health Canada 2015a) and summarized below. Since the publication of the SOS report, DMP has been analyzed in foods included in the 2013 Canadian Total Diet Study (TDS). While it was not quantified in any food composites above the method detection limit in Canada (average MDL=1.13 ng/g; Cao et al. 2015), it was reported in indoor air in homes in the United States (Tran and Kannan 2015), in agricultural soils in Canada (Khosravi and Price 2015) and in various studies internationally. However, these reported values did not change the previously presented exposure estimates derived using environmental media and food data (Environment Canada, Health Canada 2015a).

The highest exposed group according to biomonitoring data (all sources, Maternal Infant Research on Environmental Chemicals - Child Development Plus study [MIREC-CD Plus]) are male children aged 2 to 3 years, with median and 95^th percentile intakes of 0.19 and 0.66 μg/kg bw/day, respectively. For older populations (3 years and over) the highest exposed group (all sources, National Health and Nutrition Examination Survey [NHANES]) is males aged 12 to 19 years, with median and 95^thpercentile intakes of 0.042 and 0.29 μg/kg bw/day, respectively.

The subpopulation with the highest exposure to DMP from environmental media and food consisted of breastfed infants, with total daily intake of 0.019 and 0.26 μg/kg bw/dayestimated on the basis of central tendency and upper-bounding concentrations, respectively.

Estimated daily intake of DMP from use of diaper cream (infants aged 0 to 0.5 years, dermal) was 2.7 μg/kg bw/day (lower-end exposure scenario) and 8.2 μg/kg bw/day (upper-bounding exposure scenario). For adults aged 20 years and over, estimated intakes from the use of hairsprays and hair dyes were 6.6 μg/kg bw/day (lower end) and 20 μg/kg bw/day (upper bound) and 140 μg/kg bw/event (lower end) and 420 (upper bound) μg/kg bw/event, respectively^{Footnote 5}.

For more details, please refer to the SCP SOS report (Environment Canada, Health Canada 2015a).

Medium-chain phthalates

For detailed information on all 10 medium-chain phthalates in this section, please refer to the MCP SOS report (Environment Canada, Health Canada 2015b).

DIBP

Exposure estimates were calculated using biomonitoring data as well as data on DIBP occurrence in air, drinking water, dust, food, plastic items and cosmetics. They are presented in the MCP SOS report (Environment Canada, Health Canada 2015b) and summarized below. Since the publication of the SOS report, DIBP has been analyzed in foods included in the 2013 Canadian TDS. DIBP was detected in 27 of 159 composite food samples at an average MDL of 7.25 ng/g (mean of positives = 8.26 ng/g, range of 2.41 to 39.8 ng/g; Cao et al. 2015). DIBP has also been detected in indoor air in homes in the United States (Tran and Kannan 2015) and in various studies internationally. However, these reported values did not significantly change the previously presented exposure estimates derived using biomonitoring and environmental media data (Environment Canada, Health Canada 2015b).

The highest exposed group according to biomonitoring data (all sources, Canadian Health Measures Survey [CHMS]) is male children aged 6 to 11 years, with estimated mean intakes of 1.5 μg/kg bw/day and median and 95^th percentile intakes of 0.76 and 5.3 μg/kg bw/day, respectively. For older populations (12 years and over), the highest exposed group (all sources, CHMS) is 20-to-49-year-old females, with mean intakes of 0.56 μg/kg bw/day and median and 95^th percentile intakes of 0.46 and 1.4 μg/kg bw/day, respectively.

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 bw/day estimated on the basis of central tendency and upper-bounding concentrations, respectively. Daily intakes for DIBP for infants (0 to 1.5 years) from exposure from mouthing plastic toys and childcare articles were estimated at 62.8 and 251 μg/kg bw/day on the basis of lower-end and upper-bounding exposure scenarios, respectively. Estimated daily intakes for infants from dermal exposure to plastic items, based on a lower-end and upper-bounding exposure scenarios, were 30.7 and 245.3 μg/kg bw/day, respectively. For adults exposed to plastic items (females aged 20 years and over), estimated daily intakes, based on lower-end and upper-bounding exposure scenarios, were 30.8 and 96.3 μg/kg bw/day, respectively. Estimated daily intake of DIBP from use of body lotions was 0.03 μg/kg bw/day for adults aged 20 to 59 years.

DCHP

Exposure estimates were calculated using data on DCHP occurrence in dust and food. Since the publication of the MCP SOS report, DCHP has been analyzed in foods included in the 2013 Canadian TDS but was detected in only 1 of 159 composite food samples, at a concentration of 64.9 ng/g (average MDL=1.58 ng/g; Cao et al. 2015). However, this new information did not significantly change the exposure estimates presented previously (Environment Canada, Health Canada 2015b).

The subpopulation with the highest exposure from environmental media and food consisted of children (aged 6 months to 4 years), with total daily intakes of 0.0018 and 0.15 μg/kg bw/day estimated on the basis of central tendency and upper-bounding concentrations, respectively. For populations aged 12 years and over, the highest exposed group is adolescents (aged 12 to 19 years), with total daily intakes of less than 0.001 and 0.065 μg/kg bw/day estimated on the basis of central tendency and upper-bounding concentrations, respectively.

Intakes were not calculated on the basis of biomonitoring data, because the majority of samples were measured below the limit of detection (CHMS, MIREC-CD Plus, Maternal Infant Research on Environmental Chemicals study [MIREC], Plastics and Personal Care Product Use in Pregnancy survey [P4]).

DMCHP

Exposure estimates were calculated using data on DMCHP occurrence in dust. They are presented in the MCP SOS report (Environment Canada, Health Canada 2015b) and summarized below.

The subpopulation with the highest exposure from dust consisted of infants (0 to 6 months old) with total daily intake of 0.0027 and 0.054 μg/kg bw/day estimated on the basis of central tendency and upper-bounding concentrations, respectively. For populations aged 12 years and over, total daily intakes are less than 0.001 μg/kg bw/day.

DBzP

Exposure estimates were calculated using data on DBzP occurrence in dust. Since the publication of the MCP SOS report, DBzP has been analyzed in foods included in the 2013 Canadian TDS and was not quantified above the MDL (average MDL=12.7; Cao et al. 2015). Exposure estimates for DBzP are presented in the MCP SOS report (Environment Canada, Health Canada 2015b) and summarized below.

The subpopulation with the highest exposure from dust consisted of infants (0 to 6 months old), with total daily intakes of 0.016 and 0.097 μg/kg bw/day estimated on the basis of central tendency and upper-bounding concentrations, respectively. For populations aged 12 years and over, total daily intakes of less than 0.001 and 0.0011 μg/kg bw/day estimated on the basis of central tendency and upper-bounding concentrations, respectively, are observed for all age groups.

B84P

Exposure estimates were calculated using data on B84P occurrence in dust and plastic items. They are presented in the MCP SOS report (Environment Canada, Health Canada 2015b) and summarized below.

The subpopulation with the highest exposure from dust^{Footnote 6} consisted of infants (0 to 6 months old), with total daily intakes of 0.0063 and 0.047 μg/kg bw/day estimated on the basis of central tendency and upper-bounding concentrations, respectively.

Estimated daily intakes for infants from dermal exposure to B84P from plastic items, calculated on the basis of lower-end and upper-bounding exposure scenarios, were 2.7 and 21.6 μg/kg bw/day, respectively. For adults exposed to plastic items (females aged 20 years and older), lower-end and upper-bounding estimates of daily intake were 2.7 and 8.5 μg/kg bw/day, respectively.

DIHepP

Exposure estimates were calculated using data on DIHeP occurrence in dust. They are presented in the MCP SOS report (Environment Canada, Health Canada 2015b) and summarized below.

The subpopulation with the highest exposure from dust consisted of infants (0 to 6 months old), with total daily intakes of 0.096 and 1.1 μg/kg bw/day estimated on the basis of central tendency and upper-bounding concentrations, respectively (for more details refer to Environment Canada, Health Canada 2015b). For populations aged 12 years and over, the highest exposed group is adolescents (aged 12 to 19 years), with total daily intakes of 0.0011 and 0.013 μg/kg bw/day estimated on the basis of central tendency and upper-bounding concentrations, respectively.

B79P

Exposure estimates were calculated using data on B79P occurrence in dust and plastic items. They are presented in the MCP SOS report (Environment Canada, Health Canada 2015b) and summarized below.

The subpopulation with the highest exposure from dust consisted of infants (0 to 6 months old), with total daily intakes of 0.0063 and 0.047 μg/kg bw/day estimated on the basis of central tendency and upper-bounding concentrations, respectively. For populations aged 12 years and over, total daily intakes are less than 0.001 μg/kg bw/day.

Estimated daily intakes for infants from dermal exposure to B79P from plastic items, as calculated on the basis of lower-end and upper-bounding exposure scenarios, were 2.7 and 21.6 μg/kg bw/day, respectively. For adults exposed to plastic items (females aged 20 years and over), estimated daily intakes, based on lower-end and upper-bounding exposure scenarios, were 2.7 and 8.5 μg/kg bw/day respectively.

DINP

Exposure estimates were calculated using biomonitoring data as well as data on DINP occurrence in dust, food, and plastic items. They are presented in the DINP SOS report (Environment Canada, Health Canada 2015c) and summarized below. Recently, DINP has been reported internationally in soils (Tran et al. 2015), indoor air (Blanchard et al. 2014, Takeuchi et al. 2014), settled dust (Blanchard et al. 2014; Luongo and Östman 2015) and tap water (Yang et al. 2014). However, these reported values did not change the exposure estimates presented previously.

The highest exposed group according to biomonitoring data (all sources, NHANES) is male children aged 6 to 11 years, with mean intakes of 4.6 μg/kg bw/day, and median and 95^thpercentile intakes of 4.2 and 25 μg/kg bw/day, respectively. For older populations (aged 12 years and over), the highest exposed group (all sources, NHANES) is 12-to-19-year-old males, with mean intakes of 3.0 μg/kg bw/day, and median and 95^thpercentile intakes of 2.6 and 33 μg/kg bw/day, respectively. For adults aged 20 years and over, daily intakes are 2.8 μg/kg bw/day (mean), 2.4 μg/kg bw/day (median), and 24 μg/kg bw/day (95^th percentile) for males and 2.3 μg/kg bw/day (mean), 1.9 μg/kg bw/day (median), and 23 μg/kg bw/day (95^th percentile) for females.

The subpopulation with the highest exposure from dust and food consisted of children (aged 6 months to 4 years), with total daily intakes of 1.8 and 19.7 μg/kg bw/day estimated on the basis of central tendency and upper-bounding concentrations, respectively. For populations aged 12 years and over, the highest exposed group is adolescents (aged 12 to 19 years), with total daily intakes of 1.0 and 11.4 μg/kg bw/day estimated on the basis of central tendency and upper-bounding concentrations, respectively.

Daily intakes for DINP for infants (aged 0 to 1.5 years) from mouthing plastic toys and childcare articles, were estimated on the basis of lower-end and upper-bounding exposure scenarios, to be 30 and 120 μg/kg bw/day, respectively. Estimated daily intakes for infants from dermal exposure to plastic items, based on lower-end and upper-bounding exposure scenarios, were 1.1 and 8.6 μg/kg bw/day, respectively. For females aged 20 years and over, estimated daily intakes from dermal exposure to plastic items, based on lower-end and upper-bounding exposure scenarios, were 1.1 and 3.4 μg/kg bw/day, respectively.

CHIBP, BCHP and BIOP

Given the absence of reporting to the s. 71 industry survey (CHIBP, BCHP and BIOP), non-detection in dust (CHIBP and BCHP), non-detection in products used by consumers (BCHP; emission chamber study, NRC 2012), negligible modelled indoor air concentrations (CHIBP), and the absence of information as to their presence in product databases (CHIBP, BCHP and BIOP), general population exposure to CHIBP, BCHP and BIOP from environmental media and products used by consumers is expected to be negligible (Environment Canada, Health Canada 2015b).

BBP

Biomonitoring

Monobenzylphthalate (MBzP), the main monoester metabolite of BBP, was monitored in the CHMS Cycle 1 (2007-2009) and Cycle 2 (2009-2011), with 100% detection in all samples (Health Canada 2011b, 2013). Additionally, MBzP was measured as part of the First Nations Biomonitoring Initiative (FNBI) of the Assembly of First Nations (n=492 for on-reserve and Crown-land populations aged 20 years and over). Urine concentrations for MBzP (geometric mean) were reported to be statistically higher in FNBI samples than those observed in the CHMS population (AFN 2013).

MBzP was also monitored by Health Canada in three cohort surveys: P4 (n = 31 women and their infants, 542 individual spot samples, women provided multiple urine samples over two visits), MIREC (n = 1742 pregnant women, spot urine samples), and MIREC-CD Plus (197 children aged 2 to 3 years, 1 spot sample per individual). All three surveys reported high detection frequencies of MBzP (100%, 99% and 97%, respectively) (personal communication from Environmental Health Sciences and Radiation Directorate [EHSRD], Health Canada, to Existing Substances Risk Assessment Bureau [ESRAB], Health Canada, October 2013, 2014, unreferenced; Arbuckle et al. 2014).

Finally, in the United States, the NHANES also monitored MBzP in urine during survey years 1999 to 2012 and reported high detection frequencies (CDC 2014).

Using the CHMS, P4, MIREC 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; Frederiksen et al. 2013; Christensen et al. 2014; US CPSC CHAP 2014). Daily creatinine excretion rates for participants were estimated using the Mage equation. Biomonitoring intakes are presented in Table 9-1 below (see Appendix C for further information on the methodology).

The highest exposed group according to biomonitoring data (all sources, CHMS) is children aged 3 to 5 years with median and 95^th percentile intakes of 0.76 and 4.5 μg/kg bw/day, respectively. For older populations, the highest exposed group (all sources, P4) is pregnant women (aged 19 years and over) with median and 95^th percentile intakes of 0.31 and 3.01 μg/kg bw/day, respectively.

Environmental media and food

The subpopulation with the highest exposure to BBP from environmental media and food consisted of children (aged 0.5 to 4 years ), with total daily intakes of 0.58 and 2.99 μg/kg bw/day estimated on the basis of central tendency and upper-bounding concentrations, respectively (Appendix D, Table D-1a).

Indoor air and dust

One study measured BBP in indoor air in Canadian homes; all samples were less than the limit of detection (LOD) (LOD was not reported; Zhu et al. 2007). Recently, BBP has also been measured in indoor air in homes in the United States (detected in 100% of 20 samples from homes in Albany, NY, method quantification limit = 0.20 ng/m³, median: 2.99 ng/m³, maximum: 24.7 ng/m³; Tran and Kannan 2015) and internationally (Fromme et al. 2004; Rudel et al. 2010; Bergh et al. 2011a; Pei et al. 2013; Blanchard et al. 2014; Lin et al. 2014; Takeuchi et al. 2014). As no Canadian indoor air survey reported BBP above the LOD, median (2.99 ng/m³) and maximum (24.7 ng/m³) concentrations from the US study (Tran and Kannan 2015) were used to estimate the general population daily intake of BBP from indoor air (Appendix D, Table D-1a).

BBP was surveyed in the Canadian House Dust Study (CHDS) and was detected in 100% of homes (range: 0.6 to 944 μg/g, median: 42.3 μg/g, 95^th percentile: 512 μg/g) (Kubwabo et al. 2013). Internationally, BBP has also been reported in house dust (Fromme et al. 2004; Bornehag et al. 2005; Kolarik et al. 2008; Langer et al. 2010; Bergh et al. 2011a; Guo and Kannan 2011; Hsu et al. 2012; Kang et al. 2012; Gevao et al. 2013; Orecchio et al. 2013; Papadopoulos et al. 2013; Blanchard et al. 2014; Lin et al. 2014; Dodson et al. 2015; Luongo and Östman 2015).

BBP has applications as a plasticizer in the manufacturing of automobiles and automobile parts (ECHA 2012; 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 draft screening assessment.

The Canadian survey (Kubwabo et al. 2013) was identified as the key study for exposure characterization, and median (42.3 μg/g) and 95^th percentile (512 μg/g) concentrations were used to estimate the Canadian general population daily intake of DBP from dust (Appendix D, Table D-1a).

Food, infant formula and breast milk

In Canada, BBP presence in food was measured in samples from the 2013 Canadian TDS (Cao et al. 2015). BBP was detected in 32 of 159 food composite samples (average MDL = 3.10 ng/g), with a mean concentration of 12.4 ng/g and a range of 1.86 to 82.7 ng/g (Cao et al. 2015). Milk, soft drinks, and fruit juice are the major sources of intake for all population groups. In addition, BBP presence in food was monitored as part of the Canadian Food Inspection Agency's (CFIA) 2013-2014 and 2014-2015 Food Safety Action Plan (FSAP) surveys (personal communication from the Food Directorate, Health Canada, to ESRAB, Health Canada, April 2014; unreferenced). BBP was detected in less than 1% of 1518 (LOD: 0.1 μg/g) packaged and processed food samples at concentrations ranging from 0.3 to 1.9 μg/g, with a mean of the positive samples of 0.75 μg/g.

Additionally, breast milk was monitored as part of the MIREC study. BBP was not detected in any samples (n=305; MDL=0.00741 μg/g; personal communication from the Food Directorate, Health Canada, to ESRAB, Health Canada, October 2014; unreferenced). Internationally, BBP was detected in breast milk from one study in Sweden (Högberg et al. 2008) and not detected in another study in Germany (Fromme et al. 2011). However, BBP readily metabolizes to MBzP in the human body (Koch and Calafat 2009; Frederiksen et al. 2011), and as a result, BBP, as the parent compound, may not be found in breast milk in high quantities.

MBzP was measured in breast milk as part of the P4 study (n = 31 women, 56 breast milk samples collected from study participants). It was detected in 34% of breast milk samples, with the median value reported as below the LOD (0.018 μg/L) and the maximum value reported as 0.16 μg/L (personal communication from EHSRD, Health Canada, to ESRAB, Health Canada, October 2013; unreferenced). Internationally, MBzP has been detected in breast milk (Mortensen et al. 2005; Högberg et al. 2008).

In the P4 study, MBzP was detected in less than 5% of samples of infant formula (n=23; LOD=0.018 μg/L) (personal communication from EHSRD, Health Canada, to ESRAB, Health Canada, October 2013; unreferenced). The CFIA surveillance data from the 2013-2014 and 2014-2015 FSAP surveys did not detect BBP (LOD = 0.1 μg/g) in any samples of infant foods (n=44), infant formula (n=59) or infant cereals (n=19). Internationally, BBP was not detected in infant formulas in two European studies (Sørensen 2006; Bradley et al. 2013a).

Using Canadian TDS data, probabilistic estimates of dietary intakes were derived for BBP and results are outlined in Appendix D, Table D-1b (for methodology for estimating probabilistic intakes, see Appendix E). Results for BBP in breast milk and infant formula that were below the LOD (i.e., 0.018 μg/L) were set to half the LOD (Appendix D, Table D-1a).

Ambient air, drinking water and soil

No Canadian data were identified for BBP in ambient air. Rudel et al. (2010) measured outdoor air outside homes in the United States (detected in 5% of 43 samples [method reporting limit = 6 ng/m³], median: not reported, maximum: 8.5 ng/m³). Internationally, BBP has been reported in ambient air (Li and Wang 2015). Rudel et al. (2010) was identified as the relevant study for exposure characterization (sampling from North America), and half the method reporting limit (method reporting limit [MRL] = 6 ng/m³) and maximum (8.5 ng/m³) concentrations were used to estimate potential exposures to BBP via ambient air (Appendix D, Table D-1a).

No Canadian data were identified for BBP in drinking water. BBP was analyzed for, but not detected in, Canadian bottled water (MDL: 0.085 μg/L; Cao 2008). Similarly, BBP was not detected in any samples of well water intended to be bottled in Spain (Bono-Blay et al. 2012). Internationally, BBP has been detected in bottled water (Jeddi et al. 2015, Lv et al. 2015), drinking water (Liu et al. 2015), and surface waters (Net et al. 2015, Selvaraj et al. 2015). It was not detected in any samples of tap water or river water in Spain (Dominguez-Morueco et al. 2014). In the absence of Canadian or North American data on levels of BBP in tap water, half the MDL (i.e., 0.085 μg/L) for BBP in bottled water was used to estimate the general population daily intake from drinking water (Cao et al. 2008) in Canada (Appendix D, Table D-1a).

Khosravi and Price (2015) reported various phthalates in agricultural soils in both control and biosolid amended soils collected in Nova Scotia, Canada. BBP was reported at 0.13 ng/g (control soil) and 2.4 ng/g (biosolid amended soil) (Khosravi and Price 2015). Internationally, BBP has been detected in soils (Cheng et al. 2015; Tran et al. 2015; Wang et al. 2015). Conversely, Hongjun et al. (2014) did not detect BBP in any samples of urban, suburban or rural soils in China. The reported concentration of BBP in control agricultural soils (0.13 ng/g; Khosravi and Price 2015) was used to estimate potential exposures to BBP via soil in Canada (Appendix D, Table D-1a).

Products used by consumers

Globally, BBP may also be present in a wide variety of manufactured items, including childcare articles, children's toys, do-it-yourself (DIY) products, paints and exercise balls (HPD 1993- ; ECHA 2012; Korfali et al. 2013;AGDH 2015). In Canada, BBP has also been reported as a plasticizer used in various types of manufactured items (NRC 2012; See Table 2-2; Environment Canada 2014).

Toys and childcare articles

Health Canada has surveyed soft vinyl toys and childcare articles over a number of years for multiple phthalates, including BBP, and it was detected in only one sample from the 2008 survey (below the restricted limit) and has not been detected since (Health Canada 2007, 2009, 2012, 2014). BBP was also monitored in toys purchased in Canada, but made elsewhere; it was detected at 0.001% to 0.02% (Stringer et al. 2000). Internationally, BBP was monitored in toys in India. It was detected in 3 of 24 toy samples, all at less than 0.1% (Johnson et al. 2011).

Currently, Canada (like the United States and the European Union) has regulations in place limiting the amount of certain phthalates (including BBP) in toys and childcare articles (Phthalates Regulations under the Canada Consumer Product Safety Act [Canada 2016]). The European Union (EU)'s Rapid Alert System for Non-Food (RAPEX) database also shows low reporting of toys tested in violation of the regulation for BBP (RAPEX 2015), and the United States Consumer Product Safety Commission (US CPSC) recently stated that less than 10% of exposures to BBP for infants and children would result from mouthing of toys and childcare articles (US CPSC CHAP 2014).

Given the absence of BBP in toys and childcare articles reported in Canadian and international monitoring studies, exposure is expected to be negligible and oral intakes from mouthing toys and/or childcare articles were not estimated.

Cosmetics

On the basis of notifications submitted under the Cosmetic Regulations, BBP is not expected to be present in cosmetics in Canada (personal communication from the Consumer Product Safety Directorate [CPSD], Health Canada, to ESRAB, Health Canada, July 2015; unreferenced). Internationally, BBP has been detected in various types of cosmetics and personal care products^{Footnote 7} (Guo and Kannan 2013; Guo et al. 2013; Bao et al. 2015). In contrast, Liang et al. (2013) did not detect BBP in cosmetics in China. This presence may be due to potential migration from packaging. A summary of recent studies measuring concentrations of BBP in cosmetics and personal care products reported in North America is outlined in Table 9-2.

BBP was not reported as used in Canada. The United States reported low detection frequencies (5% to 13%), and the majority of the concentrations in all studies were in the sub-ppm range. Exposure to BBP from personal care products and cosmetics is therefore not considered significant. In addition, NICNAS (AGDH 2015) and US CPSC CHAP (2014) reported that the cosmetic use of BBP is likely to be rare and they did not assess exposure from this source. For that reason, exposure estimates from this source were not generated.

Other products used by consumers

Globally, BBP may be found in paint products (US CPSC CHAP 2014; AGDH 2015). This use has also been reported in Canada (Environment Canada 2014). The US CPSC CHAP (2014) reported that aerosol paints may contribute to greater than 10% of exposure to BBP for adults, infants, toddlers and children.

In addition, BBP may have applications in the production of articles that may come in contact with skin (ECHA 2012; AGDH 2015). However, BBP was reported in only 1 of 35 samples of children's clothing (Brigden et al. 2013). The US CPSC CHAP (2014) did not evaluate dermal exposure to items containing BBP.

However, since exposure to BBP from these uses is considered to be captured by the available Canadian biomonitoring data estimates of exposure were not calculated.

DBP

Biomonitoring

Mono-n-butyl phthalate (MnBP), the major monoester metabolite of DBP, was monitored in the CHMS Cycle 1 (2007-2009) and Cycle 2 (2009-2011) with 100% detection in all samples (Health Canada 2011b, 2013). Additionally, MnBP was measured as part of the First Nations Biomonitoring Initiative (n=492 for on-reserve and Crown-land populations aged 20 years and over). Urine concentrations for MnBP (geometric mean) were reported to be similar in FNBI samples to those reported in the CHMS (AFN 2013).

MnBP was also monitored by Health Canada in three cohort surveys: P4, MIREC and the MIREC-CD Plus. All three surveys reported high detection frequencies of MnBP (100%, 99% and 100%, respectively; personal communication from EHSRD, Health Canada, to ESRAB, Health Canada, October 2013, 2014, unreferenced; Arbuckle et al. 2014). An additional metabolite of DBP (MHBP) was also monitored and detected at high frequency (92% and 100%, respectively) in the P4 and MIREC-CD Plus studies (personal communication from EHSRD, Health Canada, to ESRAB, Health Canada, October 2013, 2014; unreferenced).

Using the CHMS, P4, MIREC and MIREC-CD Plus datasets, reverse dosimetry intake estimates were generated as previously described. Biomonitoring intakes are presented in Table 9-3 below (see Appendix C for further information on the methodology).

The highest exposed group according to biomonitoring data (all sources, CHMS) is children aged 3 to 5 years, with median and 95^th percentile intakes of 1.7 and 5.3 μg/kg bw/day, respectively. For older populations, the highest exposed group (all sources, P4) is pregnant women (aged 19 years and over), with median and 95^th percentile intakes of 0.55 and 4.11 μg/kg bw/day, respectively.

Environmental media and food

The subpopulation with the highest exposure to DBP from environmental media and food consisted of children (aged 0.5 to 4 years), with total daily intakes of 0.88 and 2.96 μg/kg bw/day estimated on the basis of central tendency and upper-bounding concentrations, respectively (Appendix D, Table D-2a).

Indoor air and dust

One study measured DBP in indoor air in Canadian homes (LOD was not reported; Zhu et al. 2007). They reported air concentrations of 200 ng/m³ (median) and a range of 130 to 1100 ng/m³ (Zhu et al. 2007). Recently, DBP has also been measured in indoor air in homes in the United States (detected in 100% of 20 samples from homes in Albany, NY, method quantification limit = 0.10 ng/m³, median: 22.6 ng/m³, maximum: 111 ng/m³; Tran and Kannan 2015) and internationally (Blanchard et al. 2014; Lin et al. 2014; Takeuchi et al. 2014). The Canadian indoor air survey concentrations (median: 200 ng/m³ and maximum: 1100 ng/m³; Zhu et al. 2007) were used to estimate the general population daily intake of DBP from indoor air in Canada (Appendix D, Table D-2a).

DBP was surveyed in the CHDS and was detected in 99% of homes (range: ND-1392 μg/g, median: 16.8 μg/g, 95^thpercentile: 95.4 μg/g; Kubwabo et al. 2013). Internationally, DBP has also been reported in house dust (Dodson et al. 2015; Luongo and Östman 2015).

DBP has applications as a plasticizer in the manufacturing of automobiles and automobile parts (ECHA 2012; AGDH2013a). 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 draft screening assessment.

The Canadian survey (Kubwabo et al. 2013) was identified as the key study for exposure characterization, and median (16.8 μg/g) and 95^th percentile (95.4 μg/g) concentrations were used to estimate the Canadian general population daily intake of DBP from dust (Appendix D, Table D-2a).

Food, infant formula and breast milk

In Canada, DBP presence in food was monitored in samples from the 2013 Canadian TDS (Cao et al. 2015). DBP was detected in 44 of 159 food composite samples (average MDL=16.6 ng/g) with a mean concentration of the positive samples of 23.2 ng/g and a range of 6.21 - 208 ng/g (Cao et al. 2015). Milk, soft drinks, bread, and ice cream are the major sources of intake for the overall population. In addition, DBP presence in food was monitored as part of the CFIA 2013-2014 and 2014-2015 FSAP surveys (personal communication from the Food Directorate, Health Canada, to ESRAB, Health Canada, April 2014; unreferenced). DBP was detected in 14% of 1518 (LOD: 0.1 μg/g) packaged and processed food samples. DBP concentrations ranged from 0.26 to 4.3 μg/g with a mean concentration of the positive samples of 0.76 μg/g.

MnBP was measured in breast milk as part of the P4 study (n = 31 women, 56 breast milk samples collected from study participants, LOD=0.057 μg/L). It was detected in 100% of breast milk samples, with median and maximum values reported as 0.656 and 5.18 μg/L, respectively (personal communication from EHSRD, Health Canada, to ESRAB, Health Canada, October 2013; unreferenced). Internationally, MnBP has been measured in breast milk in Europe (Fromme et al. 2011).

Breast milk was also monitored as part of the MIREC study. DBP was detected in 21 samples (n=305, median=0.0129 μg/g, range= less than MDL-0.030 μg/g, MDL=0.0149 μg/g; personal communication from Food Directorate, Health Canada, to ESRAB, Health Canada, October 2014; unreferenced). Internationally, DBP was detected in breast milk from one study in Sweden (Högberg et al. 2008) and Germany (Fromme et al. 2011). However, DBP readily metabolizes to MnBP in the human body (Koch and Calafat 2009) and would therefore not be expected in breast milk in high frequency.

Recently, an analysis of 23 infant formula samples in the P4 study showed 80% detection of MnBP (LOD: 0.057 μg/L, median: 0.299 μg/L, maximum: 1.16 μg/L) (personal communication from EHSRD, Health Canada, to ESRAB, Health Canada, October 2013; unreferenced). However, measurement of this metabolite in infant formula was affected by high field blank levels (possible contamination); and therefore, was not used to quantify intakes. The CFIA surveillance data did detect DBP in 3 samples of infant foods (n=20, median: ND, 95^th percentile: 0.675 μg/g), 12 samples of infant formula (n=32, median: ND, 95^thpercentile: 1.12 μg/g) and 4 samples of infant cereals (n=7, median: 0.42 μg/g, 95^th percentile: 1.79 μg/g; personal communication from the Food Directorate, Health Canada, to ESRAB, Health Canada, April 2014; unreferenced). Internationally, DBP has been detected in infant formulas in Italy and Europe (Cirillo et al. 2015), but not in the United Kingdom (Bradley et al. 2013a).

Using TDS data, probabilistic estimates of dietary intakes for DBP were derived for the Canadian general population and results are outlined in Appendix B, Table B-2b (the methodology for estimating probabilistic intakes is provided in Appendix E). Median (0.656 μg/L) and maximum (5.18 μg/L) values of MnBP (metabolite of DBP) in breast milk measured in the P4 study was used for risk characterization (Appendix D, Table D-2a).

Ambient air, drinking water and soil

No Canadian data were identified for DBP in ambient air. Rudel et al. (2010) measured outdoor air outside homes in the United States (detected in 35% of 43 samples [MRL = 7 ng/m³], maximum: 32 ng/m³). Internationally, DBP has been reported in ambient air (Li and Wang 2015). Rudel et al. (2010) was identified as the relevant study for exposure characterization (sampling from North America), and half the MRL (7 ng/m³) and maximum (32 ng/m³) concentrations were used to estimate potential exposures to DBP via ambient air (Appendix D, Table D-2a).

No Canadian data were identified for DBP in drinking water. In Canada, Cao et al. (2008) surveyed phthalates in bottled carbonated and non-carbonated water and detected and quantified DBP in all 11 samples (range: 0.075 - 1.72 μg/L). Internationally, DBP was also detected in bottled water samples (Jeddi et al. 2015; Lv et al. 2015), surface waters (Li et al. 2015; Net et al. 2015; Selvaraj et al. 2015), and tap water (Liu et al. 2015). In the absence of Canadian and North American data on levels of DBP in tap water, mean (0.357 μg/L) and maximum (1.72 μg/L) concentrations of DBP in bottled non-carbonated water were used to estimate the general population daily intake from drinking water (Cao et al. 2008) in Canada (Appendix D, Table D-2a).

Khosravi and Price (2015) reported various phthalates in agricultural soils in both control and biosolid amended soils collected in Nova Scotia, Canada. DBP was reported at 0.14 ng/g (control soil) and 1.1 ng/g (biosolid amended) (Khosravi and Price 2015). Internationally, DBP has been detected in soils (Cheng et al. 2015; Tran et al. 2015; Wang et al. 2015). The reported concentration of DBP in control agricultural soils (0.14 ng/g; Khosravi and Price 2015) was used to estimate potential exposures to DBP via soil in Canada (Appendix D, Table D-2a).

Products used by consumers

Globally, DBP may also be present in a wide variety of manufactured items, including childcare articles, children's toys, DIY products, gloves and exercise balls (HPD 1993- ; Stringer et al. 2000; ECHA 2012; Chao et al. 2013; Korfali et al. 2013; AGDH 2013a). In Canada, DBP has also been reported as a plasticizer used in various types of manufactured items (NRC 2012; See Table 2-2; Environment Canada 2014).

Toys and childcare articles

Numerous studies have examined the concentrations of DBP in childcare articles and toys (Stringer et al. 2000; Biedermann-Brem et al. 2008; Johnson et al. 2011; Korfali et al. 2013). A summary of DBP reported in toys and childcare articles available in Canada are presented in Table 9-4.

Currently, Canada (like the United States and the European Union) has regulations in place limiting the amount of certain phthalates (including DBP) in toys and childcare articles (Phthalates Regulations under the Canada Consumer Product Safety Act [Canada 2016]). The EU's RAPEX database also shows low reporting of toys tested in violation of the regulation for DBP (RAPEX 2015), and the US CPSC recently stated that less than 10% of exposures to DBP for infants and children would result from mouthing of toys and childcare articles (US CPSC CHAP 2014).

Given the absence of DBP in toys and childcare articles reported in Canadian monitoring studies and the low frequency and low percent content of DBP reported in international monitoring studies, exposure is expected to be negligible and oral intakes from mouthing toys and/or childcare articles were not estimated.

Cosmetics

On the basis of notifications submitted under the Cosmetic Regulations, DBP is likely to be present in cosmetics in Canada, specifically nail polishes (personal communication from the CPSD, Health Canada, to ESRAB, Health Canada, July 2015; unreferenced).

Koniecki et al. (2011) reported DBP in cosmetics purchased in Canada, including hair sprays, mousses, nail polishes, skin cleansers and baby shampoos. Internationally, DBP has been detected in various types of cosmetics and personal care products (Guo and Kannan 2013; Guo et al. 2013; Liang et al. 2013). A summary of recent studies measuring concentrations of DBP in cosmetics and personal care products reported in North America is outlined in Table 9-5.

Estimates of exposure to DBP from nail polishes were 0.16 and 0.83 μg/kg bw/day for mean and maximum concentrations, respectively.

Other products used by consumers

Globally, DBP may be found in paint products (AGDH 2013a). This use has also been reported in Canada (Environment Canada 2014). In addition, DBP may have applications in the production of articles that may come in contact with skin (ECHA 2012; AGDH 2013a; Environment Canada 2014). DBP was reported in 23 of 35 samples (less than 3.0-120 mg/kg) of children's clothing (Brigden et al. 2013). DBP is also found as a non-medicinal ingredient in 1 topical natural health product in Canada (LNHPD modified 2014). The US CPSC CHAP (2014) did not evaluate dermal exposure to items containing DBP.

Exposure to DBP from these uses is considered to be captured by the available Canadian biomonitoring data, therefore estimates of exposure were not calculated.

DEHP

Biomonitoring

Several metabolites of DEHP^{Footnote 8} were monitored in the CHMS Cycle 1 (2007-2009) and Cycle 2 (2009-2011) with greater than 99% detection in all samples (Health Canada 2011b, 2013). Additionally, these metabolites were measured as part of the First Nations Biomonitoring Initiative (n=492 for on-reserve and Crown-land populations aged 20 years and over). Urine concentrations for MEHP, MEHHP and MEOHP (geometric means) were all reported to be statistically lower in FNBI samples than those reported in CHMS (AFN 2013).

Metabolites of DEHP were also monitored by Health Canada in three cohort surveys: P4 study (n = 31 women and their infants, 542 individual spot samples, women provided multiple urine samples over two visits, monitored 5 metabolites of DEHP^{Footnote 9}), MIREC study (n = 1742 pregnant women, spot urine samples, monitored 3 metabolites of DEHP^{Footnote 10}) and the MIREC-CD Plus study (n = 197 children aged 2 to 3 years, 1 spot sample per individual, monitored 5 metabolites of DEHP^{Footnote 11}). All three surveys reported high detection frequencies (greater than 90%) of all metabolites monitored (personal communication from EHSRD, Health Canada, to ESRAB, Health Canada, October 2013, 2014, unreferenced; Arbuckle et al. 2014).

Using the CHMS, P4, MIREC and MIREC-CD Plus datasets, reverse dosimetry intake estimates were generated as previously described. Biomonitoring intakes are presented in Table 9-7 below (see Appendix C for further information on the methodology).

The highest exposed group based on biomonitoring data (all sources, CHMS) is children aged 3 to 5 years, with median and 95^th percentile intakes of 4.0 and 12 μg/kg bw/day, respectively. For older populations (19 years and over), the highest exposed group (all sources, MIREC) is pregnant females aged 18 years and over, with median and 95^th percentile intakes of 1.6 and 8.4 μg/kg bw/day, respectively. However, the higher exposure estimates for pregnant women relative to non-pregnant women could be a result of multiple factors (different populations sampled, different sample sizes, etc.) as the results were obtained from different studies. Therefore a correlation between pregnancy and higher DEHP levels cannot be made.

Environmental media and food

The subpopulation with the highest exposure to DEHP from environmental media and food consisted of children (aged 0.5 to 4 years), with total daily intakes of 10.45 and 27.57 μg/kg bw/day estimated on the basis of central tendency and upper-bounding concentrations, respectively (Appendix D, Table D-3a).

Indoor air and dust

One study measured DEHP in indoor air in Canadian homes and reported air concentrations of 88 ng/m³ (median) and a range of 8.8 to 2100 ng/m³ (LOD was not reported; Zhu et al. 2007). Recently, DEHP has also been measured in indoor air in homes in the United States (detected in 100% of 20 samples from homes in Albany, NY, method quantification limit = 0.10 ng/m³, median: 17.4 ng/m³, maximum: 132 ng/m³; Tran and Kannan 2015) and internationally (Blanchard et al. 2014; Lin et al. 2014; Takeuchi et al. 2014). The Canadian indoor air survey concentrations (median: 88 ng/m³ and maximum: 2100 ng/m³; Zhu et al. 2007) were used to estimate the general population daily intake of DEHP from indoor air in Canada (Appendix D, Table D-3a).

DEHP was surveyed in the CHDS and was detected in 100% of homes (range: 35.9-3836 μg/g, median: 462 μg/g, 95^thpercentile: 1880 μg/g) (Kubwabo et al. 2013). Internationally, DEHP has also been reported in house dust (Dodson et al. 2015; Luongo and Östman 2015).

DEHP has applications as a plasticizer in the manufacturing of automobiles and automobile parts (AGDH 2010; ECHA 2012; Environment Canada 2014a). 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 draft screening assessment.

The Canadian survey (Kubwabo et al. 2013) was identified as the key study for exposure characterization, and median (462 μg/g) and 95^th percentile (1880 μg/g) concentrations were used to estimate the Canadian general population daily intake of DBP from dust in Canada (Appendix D, Table D-3a).

Food, infant formula and breast milk

In Canada, DEHP presence in food was measured in samples from the 2013 Canadian TDS (Cao et al. 2015). DEHP was detected in 111 of 159 food composite samples (average MDL=39.0 ng/g) with a range of concentrations from 14.4 to 714 ng/g (Cao et al. 2015). Milk, fruits and vegetables are the major sources of intake for all population groups over 1 year of age. Infant formula is the major source of intake for infants under the age of 1 year.

Regarding fruits and vegetables, detectable levels of DEHP (i.e., concentrations above the LOD) were found in most of the composite samples (Cao et al. 2015). The source of DEHP in fruit and vegetable samples is unclear, but could be linked to their packaging. For example, some fruits or vegetables may have been packaged in plastic. Furthermore, certain composite samples included canned or jarred products where coatings and gaskets on jar lids may be potential sources of DEHP in foods. Environmental contamination and agricultural practices are other possible contributors to DEHP residues in fruits and vegetables. Nonetheless, the DEHP results for composite fruit and vegetable samples from the 2013 TDS are considered unusual and unlikely to represent the frequency and magnitude of DEHP concentrations that are typically found in these types of foods (personal communication from Food Directorate, Health Canada, to ESRAB, Health Canada, April 2016; unreferenced). Health Canada's Food Directorate also analyzed composite fruit and vegetable samples from the 2014 TDS in order to further understand the distribution of DEHP concentrations in fruits and vegetables sold in Canada^Footnote12. Relative to the 2013 TDS data, the 2014 TDS results detected DEHP approximately twofold less frequently in fruit and vegetable composite samples (68% in 2013 versus 30% in 2014). In addition, the mean concentrations in positive samples from 2014 are more than 5 times lower than those from the 2013 TDS samples (61 μg/kg in the 2014 versus 331 μg/kg in 2013). Therefore, the dietary exposures estimates presented herein, which are based only on the 2013 TDS results, are expected to overestimate actual dietary exposure to DEHP (personal communication from Food Directorate, Health Canada, to ESRAB, Health Canada, April 2016; unreferenced).

DEHP presence in food was also monitored as part of the CFIA 2013-2014 and 2014-2015 FSAP surveys (personal communication from Food Directorate, Health Canada, to ESRAB, Health Canada, April 2014; unreferenced). DEHP was detected in 8% of 1518 (LOD: 0.05 μg/g) packaged and processed food samples. DEHP concentrations ranged from 0.27 to 76.2 μg/g, with a mean of the samples of 2.09 μg/g.

Breast milk was monitored as part of the MIREC study. DEHP was detected in 23 samples (n=305, mean of positives=0.0977 μg/g, range= less than MDL-0.236 μg/g, average MDL=0.0668 μg/g; personal communication from Food Directorate, Health Canada, to ESRAB, Health Canada, October 2014; unreferenced). Internationally, DEHP has been detected in breast milk in Sweden (Högberg et al. 2008) and Germany (Fromme et al. 2011), but was not detected in Italy (Guerranti et al. 2013). However, DEHP readily metabolizes to multiple metabolites in the human body (Koch and Calafat 2009) and would therefore not be expected to be found in breast milk in high quantities.

Three metabolites of DEHP^{Footnote 13} were measured in breast milk as part of the P4 study (n = 31 women, 56 breast milk samples collected from study participants). MEHP, MEHHP and MEOHP were detected in 100, 16 and 8% of breast milk samples, respectively (personal communication from EHSRD, Health Canada, to ESRAB, Health Canada, October 2013; unreferenced). LODs were reported to be 0.1, 0.019 and 0.017 μg/L, respectively. The median and maximum reported levels of MEHP were 1.26 and 17.05 μg/L, respectively (personal communication from EHSRD, Health Canada, to ESRAB, Health Canada, October 2013; unreferenced). Internationally, MEHP has also been detected in breast milk in Europe (Mortensen et al. 2005; Högberg et al. 2008; Fromme et al. 2011; Guerranti et al. 2013).

Recently, an analysis of 23 infant formula samples in the P4 study showed 90% detection of MEHP (LOD: 0.1 μg/L) and 8% detection of MEHHP (LOD: 0.019 μg/L). Median and maximum values of MEHP were reported as 0.469 and 2.154 μg/L, respectively. MEOHP was not detected (LOD: 0.017 μg/L) (personal communication from EHSRD, Health Canada, to ESRAB, Health Canada, October 2013; unreferenced). However, measurement of these metabolites in infant formula was affected by high field blank levels (possible contamination) and therefore was not used to quantify intakes. The CFIA surveillance data did detect DEHP in 1 sample of infant cereal (n=7). It was not detected in any samples of infant formula (n=32) or infant foods (n=20) (personal communication from Food Directorate, Health Canada, to ESRAB, Health Canada, April 2014; unreferenced). Internationally, DEHP has been detected in infant formulas in Italy, the UK and Europe (Sørensen 2006; Bradley et al. 2013a; Cirillo et al. 2015).

Using TDS data, probabilistic estimates of dietary intakes for DEHP were derived for the Canadian general population. The results are outlined in Appendix D, Table D-3b and the methodology for estimating probabilistic intakes is provided in Appendix E. Median (1.26 μg/L) and maximum (17.05 μg/L) values of MEHP (metabolite of DEHP) in breast milk measured in the P4 study were used for exposure characterization (Appendix D, Table D-3a). Note that exposure estimates based on food intake are significantly higher than those based on biomonitoring results (see Table 9-7 and Table D-3b). This is expected to be due to detection of frequent and elevated concentrations of DEHP in highly consumed foods from the 2013 TDS, such as certain fruits and vegetables, as previously discussed.

Ambient air, drinking water and soil

No Canadian data were identified for DEHP in ambient air. Internationally, DEHP has been reported in outdoor air (Li and Wang 2015). Rudel et al. (2010) measured outdoor air outside homes in the United States (detected in 14% of 43 samples [MRL= 40 ng/m³], median: not reported, maximum: 230 ng/m³). Rudel et al. (2010) was identified as the relevant study for exposure characterization (sampling from North America), and half the MRL (40 ng/m³) and maximum (230 ng/m³) concentrations were used to estimate potential exposures to DEHP via ambient air (Appendix D, Table D-3a).

No Canadian data were identified for DEHP in drinking water. In Canada, Cao et al. (2008) surveyed phthalates in bottled carbonated and non-carbonated water and detected and quantified DEHP in all 11 samples (range: 0.052 to 0.338 μg/L). Internationally, DEHP was also detected and quantified in bottled water (Jeddi et al. 2015), drinking water (Liu et al. 2015), and surface waters (Li et al. 2015; Net et al. 2015; Selvaraj et al. 2015). In the absence of Canadian and North American data on levels of DEHP in tap water, mean (0.102μg/L) and maximum (0.338 μg/L) DEHP concentrations in bottled non-carbonated water were used to estimate the general population daily intake from drinking water (Cao et al. 2008) in Canada (Appendix D, Table D-3a).

Khosravi and Price (2015) reported various phthalates in agricultural soils in both control and biosolid amended soils collected in Nova Scotia, Canada. DEHP was reported at 0.06 ng/g (control soil) and 4.3 ng/g (biosolid amended) (Khosravi and Price 2015). DEHP has also been detected in soils internationally (Cheng et al. 2015; Tran et al. 2015; Wang et al. 2015a). The reported concentration of DEHP in control agricultural soils (0.06 ng/g; Khosravi and Price 2015) was used to estimate potential exposures to DEHP via soil in Canada (Appendix D, Table D-3a).

Products used by consumers

Globally, DEHP may be present in a wide variety of manufactured items, including childcare articles, children's toys, DIY products, electronics, textiles, and gloves (HPD 1993- ; Stringer et al. 2000; AGDH 2010; US CPSC 2010; ECHA 2012; Chao et al. 2013; Korfali et al. 2013). In Canada, DEHP has also been reported as a plasticizer used in various types of manufactured items (NRC 2012; See table 2-2; Environment Canada 2014).

DEHP may be found in paint products (AGDH 2010). The use has also been reported in Canada (Environment Canada 2014).

Globally, DEHP may have applications in the production of articles that may come in contact with skin (HPD 1993- ; AGDH 2010; ECHA 2012; NRC 2012; Chao et al. 2013; US CPSC CHAP 2014). The US CPSC CHAP (2014) reported that dermal exposure to items containing DEHP (i.e., play pen, change pad) would contribute to greater than 10% of exposure to DEHP for infants and children.

Exposure to DEHP from these uses is considered to be captured by the available Canadian biomonitoring data, therefore estimates of exposure were not calculated.

Toys and childcare articles

Numerous studies have examined the concentrations of DEHP in childcare articles and toys (Stringer et al. 2000; Biedermann-Brem et al. 2008; Johnson et al. 2011; Korfali et al. 2013). A summary of DEHP reported in toys and childcare articles available in Canada is presented in Table 9-8.

Currently, Canada (like the United States and the European Union) has regulations in place limiting the amount of certain phthalates (including DEHP) in toys and childcare articles (Phthalates Regulations under the Canada Consumer Product Safety Act [Canada 2016]). The EU's RAPEX database also shows low reporting of toys tested in violation of the regulation for DEHP (RAPEX 2015), and the US CPSC recently stated that less than 10% of exposures to DEHP for infants and children would result from mouthing of toys and childcare articles (US CPSC CHAP 2014).

Given the absence of DEHP in toys and childcare articles reported in Canadian monitoring studies and the low frequency and low percent content of DEHP reported in international monitoring studies, exposure was not quantified and oral intakes from mouthing toys and/or childcare articles were not estimated.

Cosmetics

On the basis of notifications submitted under the Cosmetic Regulations, DEHP is not expected to be present in cosmetics in Canada (personal communication from the CPSD, Health Canada, to ESRAB, Health Canada, July 2015; unreferenced). However, DEHP has been detected in various types of cosmetics and personal care products in Canada (Koniecki et al. 2011) and internationally (Guo and Kannan 2013; Guo et al. 2013). Liang et al. (2013) did not detect DEHP in cosmetics in China. DEHP has also been reported in some fragrance products (0-46ppm; Not too Pretty 2015). This presence may be due to potential migration from packaging. A summary of recent studies measuring concentrations of DEHP in cosmetics and personal care products reported in North America is outlined in Table 9-9.

DEHP is on the List of Prohibited and Restricted Cosmetic Ingredients in Canada (Health Canada 2011a) and was not reported as used in Canada (July 2015 email from the CPSD to ESRAB, Health Canada). Additionally, the Canadian study reported that detection frequencies are low (1% to 5%) and a majority of the concentrations in all studies are in the sub-ppm range. Consequently, exposure to DEHP from personal care products and cosmetics is not considered to be significant. Exposure estimates from this source were not generated.

Adult toys

No Canadian use of DEHP in adult toys was reported under the CEPA section 71 industry survey (Environment Canada 2014). Global use patterns suggest that there is potential for use of DEHP in adult toys. DEHP was reported in 8 of 15 adult toys sampled in the EU, with concentrations ranging from 0.73 to 702 mg/g (Bavarian State Ministry of the Environment and Public Health 2012). The Danish Environmental Protection Agency published a report assessing exposure to DEHP from these products and derived intakes of 1.7 and 47 μg/kg bw/day, for normal-case and worst-case exposures, respectively (Nilsson et al. 2006).

Medical devices

Globally, DEHP is reported to be used in medical applications (ECHA 2012). Medical devices may therefore be a potential source of DEHP exposure. Currently, Canada has guidance in place for DEHP in medical devices. However, given the lack of data and uncertainty in quantifying estimates of exposure from medical devices, this exposure source was not quantified and is currently an uncertainty in this draft screening assessment.

DnHP

No data was identified for DnHP in air or water; therefore, intakes from these sources were not estimated.

DnHP was surveyed in the CHDS and was detected in 98% of homes (range: ND-264 μg/g, median: 3.8 μg/g, 95^thpercentile: 62 μg/g) (Kubwabo et al. 2013). DnHP has also been reported in house dust in the United States (Dodson et al. 2015). The Canadian survey (Kubwabo et al. 2013) was identified as the key study for exposure characterization, and median (3.8 μg/g) and 95^th percentile (62 μg/g) concentrations were used to estimate the Canadian general population daily intake of DnHP from dust (Appendix D, Table D-4a).

No Canadian or North American data were identified for DnHP in soil. DnHP has been detected in 100% of urban (n=17; median=0.018 μg/g, maximum=0.019 μg/g), suburban (n=28; median=0.016 μg/g, maximum=1.227 μg/g) and rural (n=37; median=0.016 μg/g, maximum=0.1 μg/g) soils in China (Hongjun et al. 2014). Given the limited data available on DnHP in soil, exposure intakes from this source were not estimated.

No Canadian data for DnHP in foods were found. One US study looked at many phthalates, including DnHP, in a variety of food items (Schecter et al. 2013). Data from this US study were employed to generate dietary exposure estimates using a probabilistic approach (Appendix D, Table D-4b).

The subpopulation with the highest exposure from dust and food consisted of infants (0 to 6 months old), with total daily intakes of 0.019 and 0.31 μg/kg bw/day estimated on the basis of central tendency and upper-bounding concentrations, respectively (Appendix D, Table D-4a).

Given the absence of reporting to the CEPA section 71 industry survey (Environment Canada 2014), and the lack of information as to DnHP presence in product databases, general population exposure to DnHP from products used by consumers is expected to be negligible.

DIOP

No data was identified for DIOP in air, water, soil or food; therefore, intakes from these sources were not estimated.

DIOP was surveyed in the CHDS and was detected in 87% of homes (range: ND-1165 μg/g, median: 6.6 μg/g, 95^thpercentile: 28.6 μg/g) (personal communication from EHSRD, Health Canada, to ESRAB, Health Canada, August 2014; unreferenced). These data were identified as the key study for exposure characterization, and median (6.6 μg/g) and 95^thpercentile (28.6 μg/g) concentrations were used to estimate the Canadian general population daily intake of DIOP from dust (Appendix D, Table D-5).

In Canada, DIOP has also been reported as a plasticizer used in various types of manufactured items (See Table 5-1; Environment Canada 2014). However, given the low volume of DIOP reported in Canada (See table 4-2), exposure estimates from products used by consumers were not estimated.

The subpopulation with the highest exposure from dust consisted of infants (0 to 6 months old), with total daily intakes of 0.033 and 0.14 μg/kg bw/day estimated on the basis of central tendency and upper-bounding concentrations, respectively (Appendix D, Table D-5).

For detailed information on long-chain phthalates included in this section, please refer to the SOS report on long-chain phthalates esters (Environment Canada, Health Canada 2015d).

DIDP

Exposure estimates were calculated using biomonitoring results as well as data on DIDP occurrence in dust, food and plastic items. They are presented in the LCP SOS report (Environment Canada, Health Canada 2015d) and summarized below. Recently, DIDP has been reported internationally in soils (Tran et al. 2015), indoor air (Takeuchi et al. 2014), settled dust (Luongo and Östman 2015), and tap water (Yang et al. 2014). However, these reported values did not change the exposure estimates presented previously (Environment Canada, Health Canada 2015d).

The highest exposed group according to biomonitoring data (all sources, NHANES) is male children aged 6 to 11 years, with median and 95^th percentile intakes of 1.4 and 4.4 μg/kg bw/day respectively. For older populations, the highest exposed group (all sources, NHANES) is adults over 20 years of age, with median and 95^th percentile intakes of 0.76 and 4.4 μg/kg bw/day, respectively, for males, and 0.65 and 4.9 μg/kg bw/day, respectively, for females.

The subpopulation with the highest exposure from dust and food consisted of infants (6 months to 4 years old), with total daily intakes of 0.514 and 2.87 μg/kg bw/day estimated on the basis of central tendency and upper-bounding concentrations, respectively. For populations aged 12 years and over, the highest exposed group is adolescents (aged 12 to 19 years), with total daily intakes of 0.075 and 0.726 μg/kg bw/day estimated on the basis of central tendency and upper-bounding concentrations, respectively.

Estimated daily intakes for infants from dermal exposure to DIDP from plastic items, based on lower-end and upper-bounding exposure scenarios, were 0.27 and 2.16 μg/kg bw/day, respectively. For adults exposed to plastic items (females aged 20 years and over), estimated daily intakes, based on lower-end and upper-bounding exposure scenarios, were 0.27 and 0.85 μg/kg bw/day, respectively.

DUP

Exposure estimates were calculated using data on DUP occurrence in dust and plastic items. Since the publication of the LCP SOS report, DUP has been analyzed in foods included in the 2013 Canadian Total Diet Study and was not quantified above the method detection limit (average MDL= 6.97 ng/g; Cao et al. 2015). Exposure estimates for DUP are presented in the LCP SOS report (Environment Canada, Health Canada 2015d) and summarized below.

The subpopulation with the highest exposure from dust consisted of infants (0 to 6 months old), with total daily intakes of 0.0198 and 0.349 μg/kg bw/day estimated on the basis of central tendency and upper-bounding concentrations, respectively. For populations aged 12 years and over, the highest exposed group is adolescents (aged 12 to 19 years), with total daily intakes of less than 0.001 and 0.004 μg/kg bw/day estimated on the basis of central tendency and upper-bounding concentrations, respectively.

Estimated daily intakes for infants from dermal exposure to DUP from plastic items, derived on the basis of lower-end and upper-bounding exposure scenarios, were 2.7 and 21.6 μg/kg bw/day, respectively. For adults exposed to plastic items (females aged 20 years and over), estimated daily intakes, based on lower-end and upper-bounding exposure scenarios, were 2.7 and 8.5 μg/kg/day, respectively.

DMP

The SCP SOS (Environment Canada, Health Canada 2015a) summarizes the health effects literature related to DMP and its analogue DEP. No new animal hazard data were identified after the literature cut-off date of the SCP SOS.

Tables 9-10, 9-11 and 9-12 provide critical endpoints and corresponding no-observed-effects level (NOEL) and/or lowest-observed effects level (LOEL) values for DMP, as previously described in SCP SOS (Environment Canada, Health Canada 2015a) that will be used for risk characterization.

Evidence in humans

A literature update was also conducted to identify any recent human data for short-chain phthalates. The updated search was focused on reproductive and developmental endpoints in males because these endpoints were identified as critical health endpoints in the SOS reports. Studies were further evaluated and scored for quality using a consistent evaluation metric (Downs and Black 1998). For the health outcomes evaluated (i.e., sex hormone levels, anogenital distance, birth measures, male infant genitalia, preterm birth and gestational age, altered male puberty, gynecomastia, changes in sem*n parameters, pregnancy loss, and altered time to pregnancy), there was inadequate evidence or no evidence of associations between DMP and the reported outcomes (Table 9-13). More detail is provided in Health Canada (2016a) available upon request.

9.2.2.1 Medium-chain phthalates

DIBP

The MCP SOS (Environment Canada, Health Canada 2015b) summarizes the health effects literature related to DIBP. No new animal hazard studies were identified for this phthalate after the literature cut-off date of the MCP SOS.

Table 9-14 provides critical endpoints and corresponding no-observed-adverse-effects level (NOAEL) and/or lowest-observed-adverse-effects level (LOAEL) values for DIBP, as previously described in MCP SOS (Environment Canada, Health Canada 2015b), that will be used for risk characterization. DIBP has low systemic toxicity as described in MCP SOS (Environment Canada, Health Canada 2015b); hence, no systemic critical effects were established.

DCHP

The MCP SOS (Environment Canada, Health Canada 2015b) summarizes the health effects literature related to DCHP. A literature update identified new developmental and genotoxicity studies, which are summarized in Tables 9-15 and 9-16. In Li et al. (2016), rats were exposed to 0, 10, 100 or 500 mg/kg-bw/day DCHP from GD12-21. The body weight of male pups was significantly reduced in treated animals. DCHP dose-dependently increased abnormal fetal cell aggregation and decreased fetal Leydig cell size, cytoplasmic size and nuclear size at all doses tested. Decreases in AGD and increased incidences of MNGs were noted at 100 mg/kg-bw/day and above. In another study, the authors also examined the effects of in utero DCHP exposure on offspring development (Ahbab and Barlas 2015). Rats were exposed to 0, 20, 100 or 500 mg/kg-bw/day DCHP from GD6-19. Maternal effects were not observed. An increase in resorption was observed in all treated groups. A decrease in AGD and testicular pathological changes were observed starting from 20 mg/kg-bw/day. The LOAELs of theses studies are lower than the lowest NOAEL previously identified in the MCP SOS for DCHP (Environment Canada, Health Canada 2015b). As described in the MCP SOS for DCHP, a decrease in AGD was observed at higher doses in both F1 (at 511 mg/kg-bw/day) and F2 (from 107 mg/kg-bw/day) generations in a two-generation study with a NOAEL of 21 mg/kg-bw/day (Hoshino et al. 2005), suggesting transgenerational effects induced by DCHP. The dose level of 16 to 21 mg/kg-bw/day from this two-generation study was also identified as the NOAEL for developmental effects by the Australian National Industrial Chemicals Notification and Assessment Scheme (NICNAS) (AGDH 2008) and the US CPSC CHAP (2014). Different stains of rats were used by Ahbab and Barlas (2015) than by Hoshino et al. (2005) and Li et al. (2016) (Wistar vs Sprague Dawley, respectively). The lowest LOAELs identified for developmental toxicity were 10 to 20 mg/kg-bw/day estimated on the basis of testicular pathological changes from 10 mg/kg-bw/day, reduced AGD, and increased resorption from 20 mg/kg-bw/day (Ahbab and Barlas 2015; Li et al. 2016).

As described in the MCP SOS for DCHP (Environment Canada, Health Canada 2015b), only one limited repeat-dose oral exposure study with DCHP was identified in sexually immature animals. That study was determined to be limited and was not used to characterize risk of DCHP for this life stage. Therefore, results from the two-generation study (Hoshino et al. 2005) described in the MCP SOS were used, particularly observations reported in the F1 males exposed in uteroto DCHP through lacation and via diet from PND21 for at least 10 weeks until necropsy. The NOAEL for the reproductive toxicity of DCHP at the prepubertal-pubertal life stage was 18 mg/kg-bw/day in F1 males, based on observed decreases in spermatid head counts and testicular atrophy starting from 90 mg/kg-bw/day. Reduced body weight gain and reduced food consumption were observed in F1 males at 90 mg/kg-bw/day and above. It is recognized that this effect level is conservative as animals had been exposed during early development before the prepubertal-pubertal life stage.

Tables 9-17 and 9-18 provide the critical endpoints and corresponding NOAEL and/or LOAEL values for DCHP that will be used for risk characterization.

DMCHP

As described in the MCP SOS (Environment Canada, Health Canada 2015b), DCHP was used as an analogue as no studies examining the potential health effects of DMCHP were identified for any species or gender. No new literature on DMCHP was identified after the literature cut-off date of the MCP SOS. See the above section for a summary of the critical health effects used for this phthalate.

DBzP

The MCP SOS (Environment Canada, Health Canada 2015b) summarizes the health effects literature related to DBzP and its analogue MBzP. No new literature was identified after the literature cut-off date of the MCP SOS. Table 9-19 provides the critical endpoints and corresponding NOAEL and/or LOAEL values for DBzP, as previously described in MCP SOS (Environment Canada, Health Canada 2015b), that will be used for risk characterization.

B84P

The MCP SOS (Environment Canada, Health Canada 2015b) summarizes the health effects literature related to B84P and its analogues. No new literature was identified for B84P after the literature cut-off date of the MCP SOS. Tables 9-20 and 9-21 provide the critical endpoints and corresponding NOAEL and/or LOAEL values for B84P, as previously described in MCP SOS (Environment Canada, Health Canada 2015b) that will be used for risk characterization.

DIHepP

The MCP SOS (Environment Canada, Health Canada 2015b) summarizes the health effects literature related to DIHepP. No new literature was identified after the literature cut-off date of the MCP SOS. Tables 9-22 and 9-23 provide the critical endpoints and corresponding NO(A)EL and/or LO(A)EL values for DIHepP, as previously described in MCP SOS (Environment Canada, Health Canada 2015b), that will be used for risk characterization.

B79P

The MCP SOS (Environment Canada, Health Canada 2015b) summarizes the health effects literature related to B79P and its analogues. New key studies were identified after the literature cut-off date of the MCP SOS, which are described in the following sections and will be used for risk characterization.

Early development: in utero exposure to B79P

Three oral rat studies were found focusing on effects of B79P during development, including one study that was previously described in the MCP SOS. These studies all examined the effects of B79P when administered during gestation in pregnant rats during the fetal masculinization programming window GD15-17. Summaries of the studies are described in Table 9-24 below.

In an industry submitted developmental study, Tyl et al. (2012) exposed Sprague-Dawley rats to B79P via the diet from GD6-20, at target concentrations of 0, 250, 750 or 3750 ppm (reported intakes: 0, 19, 58 and 288 mg/kg-bw/day). There were no effects on maternal body weight, although food consumption was significantly increased in a dose-related manner in the two highest dose groups for the exposure. At necropsy, both absolute and relative liver weights were significantly increased at the highest dose; no adverse effects were observed at histological examination. There was no effect on the mean number of live fetuses per litter or fetal body weight. There were no differences among groups in the incidence (fetal or litter) of either external malformations/variations or of skeletal malformations. Similarly, there were no differences in the total number of malformations (external, visceral, skeletal). The authors concluded that there had been maternal toxicity at the highest dose only, with increased absolute and relative liver weights and "no adverse findings in the prenatal offspring at any dose." Parameters specific to RPS (AGD, NR, PPS, etc) were not reported.

In a second industry submitted study, Tyl et al. (2013) also conducted a two-generation reproductive toxicity study with Sprague-Dawley rats exposed to B79P at 0, 250, 750, 2500 and 5000 ppm (estimated doses: 0, 19, 56, 188 and 375 for males and 0, 17, 50, 167 and 333 mg/kg-bw/day for females). F0 and F1 male and female rats were exposed for 10 weeks premating and 2 weeks mating. Female rats were also exposed throughout gestation (approximately 3 weeks) and lactation (3 weeks). There were no differences in maternal or offspring lactational parameters or in F1 litter parameters on PND 0-4. No effects were observed in AGD across groups. No differences were observed in acquisition of PPS in F1 males. There was no evidence of effects on mating, production of litters, or litter size. Other parameters specific to RPS, specifically NR, HYP and CRY, were not reported.

In the F2 generation, there were no differences in total litter size, live litter size, live birth ratio, survival ratio, AGD, pup body weight or gross pathology. Relative liver weight was increased in both F2 males. The authors concluded that B79P was "clearly not a reproductive toxicant in either sex" at exposures up to 333-375 mg/kg-bw/day. However, a NOAEL of 50-56 mg/kg-bw/day was assigned for systemic toxicity on the basis of reduced body weights and altered organ weights and organ/body weight ratios, affecting the liver and kidney, which were free from gross and histopathologic changes (Tyl et al. 2013).

Overall, the highest NOAEL identified for developmental toxicity of B79P after gestational exposure was 50 mg/kg-bw/day based on decreased AGD in male pups and increased incidence of epispadias on PND21 observed at the next dose level of 250 mg/kg-bw/day (REACH dossier; ECHA 2013b). A decrease in AGD on PND1 and a slight increase in the incidence of CRY were also observed at 50 mg/kg-bw/day and evidence of retained nipples was observed at 500 mg/kg-bw/day on PND21. It should be noted that there is uncertainty regarding this critical NOAEL as no AGD effect was observed on PND 1 in both F1 and F2 in a recent two-generation reproductive toxicity study up to 333 mg/kg-bw/day (Tyl et al. 2013). Both studies (REACH dossier; ECHA 2013b; Tyl et al. 2013) were conducted using the same strain of rat. No information was provided by Tyl et al. (2013) to explain this discrepancy. The NOAEL for 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 (Tyl et al. 2013).

Exposure to B79P at prepubertal-pubertal life stage

In the MCP SOS, the critical effect level for the reproductive toxicity of B79P was based on read across from closest analogues as no studies were identified for this life stage. Since the publication of the MCP SOS report, an unpublished two-generation reproductive toxicity study was submitted by industry as described above; therefore, results from the two-generation study were considered, particularly observations reported in the F1 males that were exposed in utero, through lactation to B79P and via diet until necropsy at PND126 (Tyl et al. 2013).

There were transient reductions on male body weight gain at 2500 ppm (188 mg/kg-bw/day) on PND56-63, but no effects on terminal body weights or body weight gains at the end of treatment (PND126). No effects on reproductive organ weights were observed at any dose, nor were there any histopathological findings reported. No effect on sperm counts or motility; however, a small, statistically significant increase in abnormal sperm was noted at all dose groups (19 mg/kg-bw/day and above), but there were no effects on reproductive performance of F1 males with the exception of a significantly longer precoital interval (p less than 0.05) at the highest dose (375 mg/kg-bw/day). Changes in abnormal sperm were 1.45, 1.72, 1.96, 1.91, and 2.31% in 0, 19, 56, 188, and 375 mg/kg-bw/day, respectively, which reflect a 16% to 37% increase compared to control. The authors noted these percent abnormal sperm values were within the laboratory's historical control values for this rat strain and supplier. In addition, these values, although showing a dose-responsive nature, are also within typical ranges for untreated males of this strain in other laboratories (Kato et al. 2006; Matsumoto et al. 2008).

Systemic effects were limited to statistically significant increases in relative liver weights at 56 mg/kg-bw/day and 375 mg/kg-bw/day, but not at 188 mg/kg-bw/day with no histopathological findings. There were no changes in absolute or relative adrenal glands, but observations of small, increased incidences of bilateral cortical vacuolation at the highest dose tested (375 mg/kg-bw/day; 9 males).

There were no differences in offspring lactational parameters. There were no differences in F1 litter parameters on PND 0-4. No differences were observed in AGD across groups. No differences were observed in acquisition of PPS in F1 males.

Overall, the NOEAL for the reproductive toxicity of B79P at the prepubertal-pubertal life stage was 375 mg/kg-bw/day since no adverse health effects were observed at the highest dose level tested in F1 males in the two-generation study by Tyl et al. (2013). The lowest LOEL for systemic effects was 375 mg/kg-bw/day based on small, increased incidences of bilateral cortical vacuolation in adrenal glands in the same above cohort.

Oral exposure to B79P at the mature male adult stage

In the MCP SOS, the critical effect level for the reproductive toxicity of B79P was based on read-across from closest analogues as no studies were identified for this life stage. One industry-submitted two-generation study was available after the literature cut-off date of the MCP SOS.

Reproductive effects from a 14-week exposure of adult F0 males to B79P in diet from the two-generation study described in the section above showed no adverse effects in fertility or reproductive organs (Tyl et al. 2013). A small, statistically significant increase in abnormal sperm was noted at 56 mg/kg-bw/day and above, but there were no effects on sperm count, sperm motility or reproductive performance of F0 males (Table 9-25). Changes in abnormal sperm were 1.35, 1.59, 1.73, 1.95, and 2.16% in 0, 19, 56, 188, and 375 mg/kg, respectively, which reflect a 22% to 38% increase compared to control. The authors noted these percent abnormal sperm values were within the laboratory’s historical control values for this rat strain and supplier. In addition, these values, although showing a dose-responsive nature, are within typical ranges for untreated males of this strain in other laboratories (Kato et al. 2006; Matsumoto et al. 2008). One male in the highest dose group (375 mg/kg-bw/day) developed a misshapen bilateral testes. There was no evidence of effects on mating, production of litters, or litter size.

Overall, the NOAEL for reproductive toxicity identified for B79P at the adult life stage was 375 mg/kg-bw/day since no adverse health effects were observed at the highest dose level tested in F0 adult male rats in a two generation diet study after 14-week exposure to B79P (Tyl et al. 2013).

Reproductive Developmental Effects: Oral exposure to B79P in females

In a previously described two-generation study by Tyl et al. (2013), F0 females exhibited a significant increase in absolute and relative liver weights at both 167 and 333 mg/kg-bw/day. There were no differences in maternal or offspring lactational parameters and no differences in F1 litter parameters on PND 0-4. For F1 females, there were no differences in body weight at acquisition of vagin*l patency; reproductive performance, fertility and fecundity were equivalent across all groups. There was no evidence of effects upon evidence of mating, production of litters of litter size for the production of F2 offspring. At weaning and termination of the F2 pups on PND 21, F1 females had equivalent vagin*l cytology parameters and body weights. Absolute and relative liver weights were increased at 167 and 333 mg/kg-bw/day.

Table 9-26 provides a summary of critical effects for the reproductive and/or developmental effects of B79P which will be used for risk characterization.

DINP

The DINP SOS (Environment Canada, Health Canada 2015c) summarizes the health effects literature related to DINP. Since the publication of the DINP SOS report, a recent developmental study was identified (Table 9-28). Li et al. (2015b) examined the effects of in utero DINP exposure on offspring development. Rats were exposed to 0, 10, 100, 500 or 1000 mg/kg-bw/day DINP during GD12-21. No maternal toxicity was observed except in one dam at the highest dose level, which died on GD21.5.

Testicular pathological changes in Leydig cells (clusters) were first observed at 10 mg/kg-bw/day and MNGs were observed starting from 100 mg/kg-bw/day. The LOEL of 10 mg/kg-bw/day of this study is lower than the lowest NOAEL (50 mg/kg-bw/day) and LOAEL (159-395 mg/kg-bw/day) (Waterman et al. 2000; Clewell 2011 cited in ECHA 2013; Clewell et al. 2013) identified in the DINP SOS (Environment Canada, Health Canada 2015c). The NOAEL of 50 mg/kg-bw/day was also established by multiple agencies on the basis of the same studies (AGDH 2012; ECHA 2013b; US CPSC CHAP 2014). It should be noted that similar MNGs and Leydig cell clusters effects were observed in Clewell (2011a, 2013) at 250 mg/kg-bw/day and above. At higher dose levels, other RPS-related parameters include decreased serum testosterone levels, decreased AGD, NR, effects in sperm and other histopathological effects in the testes were observed.

Overall, the lowest LOEL identified for developmental toxicity was 10 mg/kg-bw/day based on testicular pathological changes (i.e., MNGs and Leydig cell clusters) observed at the next dose level of 100 mg/kg-bw/day (Li et al. 2015b).

CHIBP, BCHP and BIOP

As described in the MCP SOS (Environment Canada, Health Canada 2015b), no studies examining the potential reproductive/developmental health effects of CHIBP, BCHP and BIOP were identified. An updated literature search conducted after the literature cut-off date of the MCP SOS did not identify any new studies. As described in the exposure assessment, general population exposure to CHIBP, BCHP and 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.

Evidence in humans

A literature update was conducted to identify any recent human data for medium-chain phthalates since the publication of the MCP SOS report. The search was focused on reproductive and developmental endpoints in males as these endpoints were identified as critical health endpoints in the SOS reports. Studies were further evaluated and scored for quality using a consistent evaluation metric (Downs and Black 1998). For the health outcomes evaluated (i.e., sex hormone levels, AGD, birth measures, male infant genitalia, preterm birth and gestational age, altered male puberty, gynecomastia, changes in sem*n parameters, pregnancy loss, and altered time to pregnancy), there was limited evidence of association between DINP and sex hormone levels (Main et al. 2006; Joensen et al. 2012; Mouritsen et al. 2013; Meeker and Ferguson 2014; Specht et al. 2014; Axelsson et al. 2015a; Jensen et al. 2015; Lenters et al. 2015a; Pan et al. 2015) or sem*n parameters (Joensen et al. 2012; Jurewicz et al. 2013; Specht et al. 2014; Axelsson et al. 2015a; Lenters et al. 2015a; Pan et al. 2015). There was inadequate evidence or no evidence of association for the other medium-chain phthalates in the Phthalate Substance Grouping and the outcomes (Table 9-32). More detail is provided in Health Canada (2016a).

9.2.2.2 Additional phthalates

As previously discussed in the cumulative risk assessment (CRA) approach document (Environment Canada, Health Canada 2015e), additional phthalates identified to have common adverse effects within the RPS as a result of alterations will be considered to be included into the CRA. After both the exposure (see section 6.1) and hazard filters were considered, BBP, DBP, DEHP, DnHP and DIOP were included in the CRA. Reproductive and development effects in males (i.e., RPS-related parameters) of these five additional phthalates were evaluated and summarized in the following sections. The RPS-related parameters were previously described in detail in the MCP SOS (Environment Canada, Health Canada 2015b).

BBP

The MCP SOS (Environment Canada, Health Canada 2015b) summarizes the reproductive and development effects of BBP in section 9.2.8.1, as BBP was identified as an analogue to B84P. Since the literature cut-off date of the MCP SOS, a recent developmental study was identified. Ahmad et al. (2014) examined the effects of in utero BBP exposure on offspring development (Table 9-33). Significant decrease in fetal body weight was observed at 4 to 20 mg/kg-bw/day during development from PND1 to 75. However, the percentage decrease in body weight was 2.5% to 5% compared to untreated control animals. Maternal effects were observed at 4 to 20 mg/kg-bw/day with decrease in maternal body weight and increase in gestation length. RPS effects were observed at 100 mg/kg-bw/day with decrease in serum testosterone level, sperm effects and decrease in reproductive organ weights.

DBP

Early development: in utero exposure

The European Commission classified DBP as a Category 1B reproductive toxicant (presumed human reproductive toxicant) as defined in the EU regulation on classification, labelling and packaging of substances and mixtures (ECHA 2015a).

A literature search identified many studies examining the toxicity of DBP during gestation in rodents. For the purposes of characterizing effects during early male development, only studies in which effects of DBP were observed at doses below 250 mg/kg-bw/day in rats and 500 mg/kg-bw/day in mice following in utero exposure during the masculinization programming window are reported here. Summaries of the studies are described below and in Table 9-35.

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

A search of the available literature revealed nine studies examining the effects of gestational exposure of DBP during the masculinization programming window in mice. The majority did not examine the parameters used to describe RPS, with only two reporting any testicular pathology and one examining testicular testosterone levels (Marsman 1995; Gaido et al. 2007, Saffarini et al. 2012).

The reproductive-developmental effects of DBP do not appear to be restricted to rodents, as one study was identified in rabbits exhibiting testicular pathological changes and sperm effects (Higuchi et al. 2003). A full summary of the health effects associated with gestational exposure to DBP is summarized in Health Canada 2016b. Table 9-35 presents a list of key studies with effects from gestational exposure to DBP in male offspring.

Overall, the highest NOAEL identified for developmental toxicity of DBP after gestational exposure was 10 mg/kg-bw/day based on effects on fertility (decreased sperm count and motility and increase levels of abnormal sperm) (Ahmad et al. 2014) and a decrease in male offspring testicular testosterone levels at birth at the next dose of 50 mg/kg-bw/day (Lehmann et al. 2004). This NOAEL is supported by evidence from other studies observing decreases in both the tubular and interstitial cell populations and altered seminiferous tubule morphometry along with other mild effects on spermatocyte development at similar dose levels (Lee et al. 2004; Boekelheide et al. 2009). At lower doses, there was a slight reduction in spermatocyte development (1.5-3 mg/kg-bw/day dose range); however, the severity of this effect at this dose was minimal to slight and only minimal at the second higher dose of 14-19 mg/kg-bw/day (Lee et al. 2004). The lowest LOEL for maternal toxicity of DBP was 509 mg/kg-bw/day based on reductions in body weight gain (less than 10%) in exposed dams (Wine et al. 1997). One study in rabbits indicates that DBP has also effects in other species, but it is unknown if these effects would occur at lower doses (Higuchi et al. 2003).

Exposure at prepubertal/pubertal life stages

A literature search identified many studies examining the reproductive toxicity of DBP in young rodents. For the purposes of brevity, only studies where effects of DBP were observed at doses below 500 mg/kg-bw/day in rodents were evaluated in this screening assessment.

Overall, adverse effects in reproductive parameters observed in (pre)pubertal males after short-termexposure to DBP include changes in serum and testicular testosterone levels, histopathological effects in the testes, and potential effects in fertility (spermatogenesis, sperm motility, and number). The majority of studies available used the rat as a model for evaluation, but only within a relatively high dose range (250 to 1000 mg/kg-bw/day) which limits the interpretation of the potential reproductive toxicity of DBP in this species. One study conducted in mice examined effects at lower dose levels and reported RPS effects at dose levels lower than those in rats (Moody et al. 2013). One study was conducted in rabbits reporting evidence of testicular pathology (Higuchi et al. 2003). A full summary of the health effects associated with propubertal/pubertal exposure to DBP is summarized in Health Canada (2016b). Table 9-36 presents a list of key studies with effects from prepubertal/pubertal exposure to DBP in males.

Overall, the lowest LOEL identified for reproductive toxicity of DBP at the prepubertal-pubertal life stage was 10-50 mg/kg-bw/day based on delayed spermatogenesis in male mice exposed to DBP for 10 days at this dose and above (Moody et al. 2013). At the next dose level of 50 mg/kg-bw/day, significant reduction in absolute AGD was observed on PND14 mice, but the effect was not significant when AGD was measured relative to body weight. An increase in immature Sertoli cell and disorganization was observed at 100 mg/kg-bw/day. No comparison can be made for rats as there were no studies available at similar dose ranges. The lowest dose tested in rats was 250 mg/kg-bw/day with observations of defective spermatogenesis, shrunken tubules, decrease in spermatogenic cells and Leydig cell numbers (Srivastava et al. 1990a; Xiao-feng et al. 2009). The lowest LOEL for systemic toxicity for mice was 500 mg/kg-bw/day based on increased relative heart weight after 3 days of DBP treatment. This effect was lost by 14 days (Moody et al. 2013). One study in rabbits indicates that DBP also effects other species, but it is unknown if these effects would occur at lower doses (Higuchi et al. 2003).

Exposure at the mature male adult stage

The MCP SOS (Environment Canada, Health Canada 2015b) summarizes the health effects of DBP in sexually mature adult rats (PND55+) as DBP was identified as an analogue to DIBP. Since the literature cut-off date of the MCP SOS (Environment Canada, Health Canada 2015b), new literature was identified (Table 9-37). DBP (0, 200, 400 or 600 mg/kg-bw/day) was administered to rats by oral gavage for 15 days (Aly et al. 2015). At the lowest dose level of 200 mg/kg-bw/day, decrease in serum testosterone level, decrease in sperm count, and decrease in sperm motility were observed. In addition, histopathological examination of the testes indicated degeneration with absence of spermatogenic series in the lumen of some seminiferous tubules starting from 200 mg/kg-bw/day. However, the results from this study are limited as systemic effect and clinical signs of the animals were not measured. In another study, adverse effects (testicular pathological changes and effects on sperm) were observed starting from 500 mg/kg bw/day (Nair 2015).

DEHP

Early development: in utero exposure

The European Commission classified DEHP as a Category 1B reproductive toxicant (presumed human reproductive toxicant) as defined in the EU regulation on classification, labelling and packaging of substances and mixtures (ECHA 2015b).

A literature search identified many studies examining the toxicity of DEHP during gestation in rodents. For the purposes of the CRA, only studies covering the masculinization programming window where effects of DEHP were observed at doses at and below 50 mg/kg-bw/day in rats and 100 mg/kg bw/day in mice were evaluated in this draft screening assessment.

Overall, adverse effects in the parameters used to describe RPS in male rat offspring after in utero exposure to DEHP include decreased serum and testicular testosterone levels, delayed PPS, AGD, increase incidences of NR and CRY, gross testicular malformations, and effects in fertility. A full summary of the health effects associated with gestational exposure to DEHP is summarized in Health Canada (2016b). Table 9-39 presents a list of key studies with effects from gestational exposure to DEHP in male offspring.

Overall, the highest NOAEL identified for developmental toxicity of DEHP after gestational exposure was 4.8 mg/kg-bw/day based on small and/or aplastic epididymis, testicular pathology and other RPS effects observed in F1 and F2 at the next dose level of 14 mg/kg-bw/day in a multigenerational reproductive toxicity study (Wolfe and Layton 2003; Blystone et al. 2010). This effect level was also established by other jurisdictions (ECJRC 2008; Danish EPA 2012; US CPSC CHAP 2014; EFSA 2005). At similar dose levels of 10-15 mg/kg-bw/day, decreases in AGD, increased NR, decreased reproductive organ weights, and delayed PPS were observed in rat offspring in other studies (Andrade et al. 2006a,b; Christiansen et al. 2010). The lowest NOEL for maternal toxicity of DEHP was 359 mg/kg-bw/day based on increase in liver weight and food consumption.

Exposure at prepubertal/pubertal life stages

A literature search identified many studies examining the reproductive toxicity of DEHP in young rodents. Results from repeated-dose oral exposure studies in sexually immature rats (PND1-55) have shown that administration of DEHP can cause reproductive effects in male rats.

Overall, adverse effects observed in (pre)pubertal males after short-termexposure to DEHP include changes in serum and testicular testosterone levels, histopathological effects in the testes, and potential effects in fertility (spermatogenesis, sperm motility, and number). Reproductive effects observed in prepubertal or pubertal mice after exposure to DEHP occurred at higher doses than rats. Other species were also found to be less sensitive than rats to DEHP. No significant effects were observed in Cynomolgus monkeys, marmosets and Syrian hamsters treated with DEHP. A summary of the health effects associated with prepubertal/pubertal exposure to DEHP is summarized in Health Canada (2016b). Table 9-40 presents key studies with effects from prepubertal/pubertal exposure to DEHP in males.

Overall, the highest NOAEL identified for reproductive toxicity of DEHP at the prepubertal-pubertal life stage was 10 mg/kg-bw/day based on a significant decrease in absolute testis weight at the next dose level of 100 mg/kg-bw/day in rats exposed to DEHP for 5 days from PND21 (Dostal et al. 1988). At 200 mg/kg-bw/day, a decrease in spermatid heads was observed at 13 weeks of age in rats exposed for 5 days from PND6 (Dostal et al. 1988). Testicular pathological changes were observed at 300 mg/kg-bw/day and higher in rats exposed to DEHP for 21 days from PND3-5 (Cammack et al. 2003) and rats exposed for 40 days from PND10 or PND50 (Tonk et al. 2012).

Exposure at the mature male adult stage

Studies examining the potential reproductive toxicity of DEHP at the adult male life stage (PND55+) were identified in rats, mice, marmosets and ferrets.

In rats, chronic exposure to DEHP (104 weeks to 2 years) resulted in inhibition of spermatogenesis starting from 10-29 mg/kg-bw/day (Ganning 1991; David et al. 2000a). In animals treated for less than a chronic duration (2-13 weeks), inhibition of spermatogenesis, decrease sperm count and decrease in sperm motility was observed at 300-900 mg/kg-bw/day (Poon et al. 1997; Wolfe and Layton 2003; Kwack et al. 2009; Tonk et al. 2012; Abd-Ellah et al. 2016). Decrease in testis weight and testicular pathological changes were generally observed at 300 mg/kg-bw/day and above.

Adult mice, ferrets, and marmosets were found to be less sensitive than rats to DEHP. Systemic effects in mice such as a decrease in relative kidney weight and an increase in absolute liver weight were observed at 99 mg/kg-bw/day whereas effects on sperm and testicular pathological changes were observed at 292 mg/kg-bw/day in a chronic study (David et al. 2000b). In Ferrets, testicular pathological changes, decreases in body weight and increases in liver weight were observed at a much higher dose level of 1200 mg/kg-bw/day (Lake et al. 1976). In marmoset, only peroxisome proliferator-activated receptor (PPAR) effects in liver were observed at 500 mg/kg-bw/day (Kurata et al. 1998). A summary of the health effects associated with adult life stage exposure to DEHP is summarized in Health Canada (2016b). Table 9-41 presents key study with effects from adult life stage exposure to DEHP in males.

Overall, the highest NOAEL identified for reproductive toxicity of DEHP at the adult life stage was 5.8 mg/kg-bw/day based on a decrease in spermatogenesis at the next dose level of 29 mg/kg-bw/day in male adult rats exposed chronically to DEHP for 104 weeks (David et al. 2000a).

DnHP

Early development: in utero exposure

The European Commission classified DnHP as a Category 1B reproductive toxicant (presumed human reproductive toxicant) as defined in the EU regulation on classification, labelling and packaging of substances and mixtures (ECHA 2015c).

A literature search identified a number of recent studies examining the toxicity of DnHP during gestation in rodents. For the purposes of the CRA, only studies covering the masculinization programming window in males were evaluated in this draft screening assessment.

Overall, adverse effects in the parameters used to describe RPS in male rat offspring after in utero exposure to DnHP include decreased serum and testicular testosterone levels, delayed PPS, AGD, retention of nipple/areolae (NR), increase incidences of CRY and hypospadias, gross testicular malformations, and effects in fertility. Two studies were conducted in mice where embryotoxicity and effects on fetal viability were observed. A summary of the health effects associated with gestational exposure to DnHP is summarized in Health Canada (2016b). Table 9-42 presents a list of key studies with effects from gestational exposure to DnHP in male offspring.

Overall, the highest NOAEL identified for developmental toxicity of DnHP after gestational exposure was 5 mg/kg-bw/day based on decreased serum or testicular testosterone levels, decreased AGD at birth in males and testicular pathological changes observed at 20 mg/kg-bw/day or higher in rats (Ahbab and Barlas 2015; Saillenfait et al. 2013a).

Exposure at prepubertal/pubertal life stages

A literature search identified four studies examining the reproductive toxicity of DnHP in young sexually immature rats (PND1-55). These studies were generally tested at high dose levels and did not describe RPS-related parameters other than one study that was tested at lower dose levels using castrated male rats. A summary of the health effects associated with prepubertal/pubertal exposure to DnHP is summarized in Health Canada (2016b). Since available studies were limited, DEHP was identified, using the same chemical categories and read-across approach as that used for other phthalates, as the "closest analogue" phthalate, taking into consideration the similarities in monoester metabolism as well as the length and nature of the ester chains (Health Canada 2015).

As described earlier in the DEHP section, the highest NOAEL identified for reproductive toxicity of DEHP at the prepubertal-pubertal life stage was 10 mg/kg-bw/day based on significant decreases in absolute testis weight at the next dose level of 100 mg/kg-bw/day in rats exposed to DEHP for 5 days from PND21 (Dostal et al. 1988). Therefore, the critical effect level of 10 mg/kg-bw/day will be used to characterize the risk of developmental toxicity of DnHP for this life stage.

Exposure at the mature male adult stage

Two studies examining the potential reproductive toxicity of DnHP at the adult male life stage (PND55+) were identified in rodents. One study was conducted in rats where RPS-related parameters were not measured. Another study was conducted in mice where fertility effects and testicular pathology were examined. A summary of the health effects associated with adult stage exposure to DnHP is summarized in Health Canada (2016b). Table 9-43 presents the key study with effects from exposure to DnHP in adult males.

Overall, no NOAEL was identified and the lowest LOAEL identified for reproductive toxicity of DnHP was 390 mg/kg-bw/day in mice based on adverse effects in fertility (decreases in fertility, the number of litters per pair, the number of live pups per litter and the proportion of pups born alive) treated with DnHP for 105 days during the adult male life stage (Lamb et al. 1987).

DIOP

Early development: in utero exposure

The European Commission classified DIOP as a Category 1B reproductive toxicant (presumed human reproductive toxicant) as defined in the EU regulation on classification, labelling and packaging of substances and mixtures (ECHA 2015d).

Three studies were identified that were all conducted by Saillenfait et al. (2013b) during the masculinization programming window. Adverse effects in the parameters used to describe RPS in male rat offspring after in utero exposure to DIOP included decreased testicular testosterone levels, decreased AGD, increased incidence of NR and CRY, gross testicular malformations, and effects in fertility. A summary of the health effects associated with gestational exposure to DIOP is summarized in Health Canada (2016b). Table 9-44 presents the key study with effects from gestational exposure to DnHP in male offspring.

Overall, the highest NOAEL identified for developmental toxicity of DIOP after gestational exposure was 100 mg/kg-bw/day based on testicular pathological changes, effects on fertility and decreased testis weight at the next dose level of 500 mg/kg-bw/day. Maternal decrease in body weight was observed at 1000 mg/kg-bw/day (Saillenfait et al. 2013b).

Exposure at prepubertal/pubertal life stages

There were no repeated-dose oral exposure studies in sexually immature animals (PND1-55) with DIOP. DIHepP was identified as the appropriate analogue to use for read across. The MCP SOS (Environment Canada, Health Canada 2015b) summarizes the health effects literature related to DIHepP. Table 9-45 provides the critical endpoints and corresponding NOAEL and/or LOAEL values for DIOP.

Oral exposure at the mature male adult stage

There were no repeated-dose oral exposure studies in adult animals with DIOP. DIHepP was identified as the appropriate analogue to use for read across. The MCP SOS (Environment Canada, Health Canada 2015b) summarizes the health effects literature related to DIHepP. Table 9-45, described in the above section, provides the critical endpoints and corresponding NOAEL and/or LOAEL values for DIOP.

Evidence in humans

A literature search was conducted to identify any human data for additional phthalates. The search was focused on reproductive and developmental endpoints in males as these endpoints were identified as critical health endpoints in the SOS reports. Studies were further evaluated and scored for quality using a consistent evaluation metric (Downs and Black 1998).

For the health outcomes evaluated (i.e., sex hormone levels, anogenital distance, birth measures, male infant genitalia, preterm birth and gestational age, altered male puberty, gynecomastia, changes in sem*n parameters, pregnancy loss, and altered time to pregnancy), there was limited evidence of association between DEHP and sex hormone levels (Pan et al. 2006; Meeker et al. 2009; Li et al. 2011; Mendiola et al. 2011, 2012; Joensen et al. 2012; Araki et al. 2014; Ferguson et al. 2014a; Pant et al. 2014; Su et al. 2014; Chang et al. 2015; Fong et al. 2015; Jensen et al. 2015; Pan et al. 2015; Wang et al. 2016), birth measures (Zhang et al. 2009; Philippat et al. 2012; de co*ck et al. 2014; Zhao et al. 2014; Lenters et al. 2015b; Xie et al. 2015; Zhao et al. 2015; Casas et al. 2016), or sem*n parameters (Zhang et al. 2006; Pant et al. 2008; Jurewicz et al. 2013; Huang et al. 2014; Pant et al. 2014; Specht et al. 2014; Axelsson et al. 2015a; Axelsson et al. 2015b; Lenters et al. 2015a; Pan et al. 2015; Wang et al. 2015b; Thurston et al. 2016). There was inadequate evidence or no evidence of association between the remaining phthalates and the outcomes (Table 9-46). More detail is provided in Health Canada (2016a).

DIDP

The LCP SOS (Environment Canada, Health Canada 2015d) summarizes the health effects literature related to DIDP. No new animal hazard studies were identified after the literature cut-off date of the LCP SOS.

Table 9-47 provides critical endpoints and corresponding NOAEL and/or LOAEL values for DIDP, as previously described in LCP SOS (Environment Canada, Health Canada 2015d).

DUP

The LCP SOS (Environment Canada, Health Canada 2015d) summarizes the health effects literature related to DUP. No new literature was identified after the literature cut-off date of the LCP SOS. Table 9-48 provides critical endpoints and corresponding NOAEL and/or LOAEL values for DUP, as previously described in LCP SOS (Environment Canada, Health Canada 2015d).

Evidence in humans

A literature update was also conducted to identify any recent human data for long-chain phthalates. The search was focused on reproductive and developmental endpoints in males as these endpoints were identified as critical health endpoints in the SOS reports. No information is currently available on the potential reproductive-developmental effects of DUP in humans. Studies identified for DIDP were further evaluated and scored for quality using a consistent evaluation metric (Downs and Black 1998).

For the health outcomes evaluated (i.e., sex hormone levels, anogenital distance, birth measures, male infant genitalia, preterm birth and gestational age, altered male puberty, gynecomastia, changes in sem*n parameters, pregnancy loss, and altered time to pregnancy), there was no evidence of association between any of the long-chain phthalates evaluated and the outcomes (Table 9-49). More detail is provided in Health Canada (2016a).

DMP

Table 9-50 provides all relevant exposure and hazard values for DMP, as well as resultant margins of exposure (MOEs), for determination of risk, which were previously described in the SCP SOS (Environment Canada, Health Canada 2015a). Overall, the MOEs for DMP are considered to be adequate to account for uncertainties in the exposure and health effect databases.

The critical effects of concern of medium-chain phthalates consisted of adverse effects on the development of the male reproductive system following gestational exposure, with a particular focus on RPS-related parameters that were identified in the rat, the most sensitive species. These parameters are considered adverse and relevant for characterizing risk of exposure of the general Canadian population to this subgrouping of phthalates. Please see MCP SOS (Environment Canada, Health Canada 2015b) for a general summary and rationale.

In cases where limited rat studies were available, effect levels in other species (i.e., mice) that were lower than in rats were used for risk characterization. Evidence in humans, based on the Downs and Black scoring system (Downs and Black 1998), indicated limited evidence of association between DINP and sex hormone levels or sem*n parameters and between DEHP and sex hormone levels, birth measures or sem*n parameters. This supports the selection of the mode of action for risk characterization.

In the following sections, human health risk from exposure to medium-chain phthalates within the Phthalate Grouping is characterized on an individual basis, followed by a cumulative risk assessment to address the potential risk of concurrent exposure to medium-chain phthalates exhibiting a similar mode of action.

9.3.2.1 Individual risk characterization of the original medium-chain subgroup

DIBP

Table 9-51 provides all relevant exposure and hazard values for DIBP, as well as resultant MOEs, for determination of risk, which were previously described in the MCP SOS (Environment Canada, Health Canada 2015b). Overall, the MOEs for DIBP are considered to be adequate to account for uncertainties in the exposure and health effect databases. From the information available, there is evidence that DIBP elicits effects on the developing male reproductive system, indicative of RPS, which suggests that DIBP has a common mode of action with other phthalates in this grouping.

DCHP

Table 9-52 provides all relevant exposure and hazard values for DCHP, as well as resultant MOEs, for determination of risk, which were previously described in the MCP SOS (Environment Canada, Health Canada 2015b). Overall, the MOEs for DCHP are considered to be adequate to account for uncertainties in the exposure and health effect databases. From the information available, there is evidence that DCHP has effects on the developing male reproductive system, indicative of RPS, which suggests that DCHP has a common mode of action with other phthalates in the grouping.

DMCHP

Table 9-53 provides all relevant exposure and hazard values for DMCHP, as well as resultant MOEs, for determination of risk, which were previously described in the MCP SOS (Environment Canada, Health Canada 2015b). Overall, the MOEs for DMCHP are considered to be adequate to account for uncertainties in the exposure and health effect databases. From the information available, there is evidence that DMCHP has effects on the developing male reproductive system, indicative of RPS, which suggests that DMCHP has a common mode of action with other phthalates in the grouping.

DBzP

Table 9-54 provides all relevant exposure and hazard values for DBzP, as well as resultant MOEs, for determination of risk, which were previously described in the MCP SOS (Environment Canada, Health Canada 2015b). Overall, the MOEs for DBzP are considered to be adequate to account for uncertainties in the exposure and health effect databases. From the information available, there is evidence that DBzP has effects on the developing male reproductive system, indicative of RPS, which suggests that DBzP has a common mode of action with other phthalates in the grouping.

B84P

Table 9-55 provides all relevant exposure and hazard values for B84P, as well as resultant MOEs, for determination of risk, which were previously described in the MCP SOS (Environment Canada, Health Canada 2015b). Overall, the MOEs for B84P are considered to be adequate to account for uncertainties in the exposure and health effect databases. From the information available, there is evidence that B84P has effects on the developing male reproductive system, indicative of RPS, which suggests that B84P has a common mode of action with other phthalates in the grouping.

DIHepP

Table 9-56 provides all relevant exposure and hazard values for DIHepP, as well as resultant MOEs, for determination of risk, which were previously described in the MCP SOS (Environment Canada, Health Canada 2015b). Overall, the MOEs for DIHepP are considered to be adequate to account for uncertainties in the exposure and health effect databases. From the information available, there is evidence that DIHepP has effects on the developing male reproductive system, indicative of RPS, which suggests that DIHepP has a common mode of action with other phthalates in the grouping.

DINP

Table 9-58 provides all relevant exposure and hazard values for DINP, as well as resultant MOEs, for determination of risk. Overall, the MOEs for DINP are considered to be adequate to account for uncertainties in the exposure and health effect databases. From the information available, there is evidence that DINP has effects on the developing male reproductive system, indicative of RPS, which suggests that DINP has a common mode of action with other phthalates in the grouping.

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).

On the basis of 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 given the evidence of the effects of their analogues; however, there is no exposure at this time. Consequently, the risk to human health for these substances is not expected.

Although the above MOEs for the medium-chain phthalates in the Phthalate Substance Grouping described in this section are considered adequate on an individual basis, this does not address the potential risk of concurrent exposure to these substances and other phthalates exhibiting a similar mode of action. Hence, all 10 medium-chain phthalates in the Phthalate Substance Grouping will be included for risk characterization in a cumulative context.

9.3.2.2 Cumulative risk assessment

The human health approach for a CRA for this Grouping has been described in detail in the "Proposed approach for cumulative risk assessment of certain phthalates under the Chemicals Management Plan" document (Environment Canada, Health Canada 2015e).

Human health-focused approaches for quantifying the cumulative risk of phthalates have been conducted by several national organizations including the Australian Department of Health (AGDH 2012, 2013, 2014a,b), the Danish Environmental Protection Agency (Danish EPA 2011), and the recently completed assessment by the United States Chronic Health Advisory Panel (US CPSC CHAP 2014). This draft screening assessment uses a tiered approach, following the schematic of the World Health Organization (WHO) and the International Program on Chemical Safety (IPCS) Framework for Risk Assessment of Combined Exposures to Multiple Chemicals (WHO 2009; Meek et al. 2011). This framework exercises the default assumption that the substances being assessed act by dose addition and the evaluation of cumulative risk involves multiple tiers where each higher level is increasingly dependent on additional data. The lower tier begins with simple assumptions and/or surrogate data for both hazard and exposure and refinement to a higher tier occurs if required and as the data allows.

For exposure characterization, if there is sufficient evidence of co-occurrence, then a CRA can be considered. The exposure approach is described in the CRA approach document (Environment Canada, Health Canada 2015e).

In summary, the pivotal information that determined substances for inclusion into the CRA were industry data collected under section 71 of CEPA (Environment Canada 2014), detection in North American biomonitoring surveys (CHMS, MIREC, MIREC-CD Plus, P4, NHANEs; Health Canada 2011b; Health Canada 2013; personal communication from EHSRD, Health Canada, to ESRAB, Health Canada, October 2013, 2014, unreferenced; Arbuckle et al. 2014; CDC 2014), and detection in household dust (CHDS; Kubwabo et al. 2013). For phthalate parent compounds (i.e., DEHP, DIBP, DBP, BBP, and DINP), with close to a 100% detection in various biomonitoring surveys, there was sufficient evidence for co-exposure; these substances were therefore assessed for cumulative risk.

A significant number of phthalates within the medium-chain phthalates subgroup have not been monitored in biomonitoring samples, but are found in commerce in Canada. These substances (i.e., DIHepP, B79P, B84P, DCHP, and DIOP) were included in the assessment of cumulative risk on the basis of their in-commerce status coupled with close to 100% detection in dust samples from Canadian homes.

Finally, given reporting limits and difficulties in determining import activity of substances, the CEPA section 71 survey may not have captured all in-commerce activity. As a result, substances that fit the profile of non-reporting to section 71 and close to 100% detection in house dust samples were also assessed for cumulative risk (i.e., DMCHP, DBzP, and DnHP).

There were three populations of interest for which exposure estimates were considered for cumulative risk: pregnant women and women of childbearing age (characterized as women aged twelve and up), infants (covered by age groups 0 to 6 months, 6 months to 4 years and 3 to 5 years^{Footnote 14}), and children (covered by age groups 6 months to 4 years and 5 to 11 years). Adolescents and adult males were covered by this approach as they had generally lower estimates of phthalate exposure and are considered less sensitive to the reproductive effects of medium-chain phthalates compared to younger males (i.e. children) (NAS 2008).

In terms of exposure estimates considered for cumulative risk, biomonitoring data, which generally captures all potential exposure sources and routes (environmental media, food and products available to consumers) was considered a primary source and estimated monitoring data (chronic exposures resulting from environmental media and food) was considered as a supporting source of information. The US CPSC CHAP (2014) also used biomonitoring data as the primary source of exposure in their CRA. Regardless of route of exposure, phthalates, in general, are not considered acute toxicants, with LD50 levels from dermal exposure being at minimum two- to five-fold higher than oral values, which in turn are also high (Draize et al. 1948; David et al. 2001; Monsanto Company 1970 cited in US EPA 2006, 2010). Since phthalates metabolize relatively quickly showing no accumulation, and excretion is rapid, within hours to days (Phokha et al. 2002; Clewell et al. 2009), acute exposures were not considered relevant for a CRA. It should also be noted that no other jurisdictions have addressed phthalate exposures and risk from acute, one-time, exposures (AGDH 2011; ECHA 2013b; US CPSC CHAP 2014).

Upper-bound exposure estimates were used to estimate individual phthalate levels in the CRA to account for the uncertainties associated with the exposure data. The upper-bound exposure estimates used to estimate cumulative risk for relevant populations for the 16 medium-chain phthalates are summarized in Appendix F (Tables F-1, F-2, F-3 and F-4).

On the basis of the available information on the common adverse effects (RPS) and the observed differences in potencies within the medium-chain phthalates, a lower-tiered hazard characterization using the hazard index (HI) was considered to be the most appropriate approach. The HI method was selected because it offers the benefit of being simple and flexible and allows for an indication of which substance or substances in the CRA, or which source and route, may be the predominant contributor(s) to the overall risk. Identification of the substances or sources and routes that are drivers of the CRA is beneficial for informing risk management.

An HI is the summation of the individual hazard quotients (HQs) for each individual substance to determine the overall cumulative risk of the substance group of interest. The HQ of each substance is the ratio of exposure to a reference value (RfV) which is calculated by dividing the critical effect level identified in the hazard database by a defined uncertainty factor [UF]). The HI equation is as follows: HI = ∑HQ = ∑ (exposure/RfV). HI values of the medium-chain phthalates are calculated for the three subpopulations that represented the highest exposure groups.

Critical effect levels have been established at three life stages (i.e., in utero, prepubertal/pubertal, and adult) because of different sensitivity to the adverse effects of phthalates on development and reproduction at different life stages. The critical effect levels identified for the in utero life stage were used to calculate the HI value for pregnant woman and for infants. The critical effect levels identified for the prepubertal/pubertal life stage were used to calculate the HI value for children. The critical effect levels of the in utero and prepubertal/pubertal life stages for the medium-chain phthalates as well as the calculated RfVs are summarized in Appendix F (Table F-5 and F-6).

Given the fact that medium-chain phthalates have similar physical-chemical properties, have similar toxicological effects, and show an overall similarity in strength of their health effect databases (especially in regards to in utero exposure), the same default UF for the two corresponding subpopulations was used to calculate HQ values for each phthalate. A similar approach was adopted by another jurisdiction (US CPSC CHAP 2014). For the calculation of the total HI for pregnant women/women of childbearing age and infants, an UF of 100 (10 for intraspecies and 10 for interspecies differences) was used to calculate the RfV for the critical effect levels identified at the in utero life stage. For children, a default UF of 300 was used to calculate the RfV for the critical effects levels identified for the prepubertal-pubertal life stage. An additional UF of 3 was applied in this case on the basis of the limitations of the health effects database for the prepubertal life stage (quality and quantity of the studies currently available), and taking into consideration the variability in exposure duration in the different studies as well as the possibility that animals might also have been exposed to medium-chain phthalates in utero. The HI values for pregnant women/women of childbearing age, infants, and children are presented in Table 9-59 (also see Appendix F; tables F-7, F-8 and F-9).

Individual phthalates that had the highest contribution to the cumulative risk were identified. Tables F-7, F-8 and F-9 (in Appendix F) show that the same three phthalate compounds (DINP, DBP, DEHP) resulted in the majority of cumulative risk, regardless of the age group or source of exposure data (biomonitoring or environmental media and food). Biomonitoring exposure estimates were generally considered more representative of potential exposures, including products available to consumers (regardless of source, route or duration). As a result, the HI values calculated using biomonitoring were considered more realistic but still conservative, because they were calculated using upper-bound exposure estimates.

Children and infants had higher HI values compared to pregnant women and women of childbearing age (aged 12 years and over). HI values calculated using biomonitoring data were lower than those calculated using environmental and dietary monitoring data for infants and children. In the case of DEHP, the high exposures from food were expected to be as a result of the unexpected presence of DEHP in fruits and vegetables which would have overestimated actual dietary exposure to DEHP. Nonetheless, this conservative, lower-tiered HI approach indicated no concern for potential cumulative risk of medium-chain phthalates for the general Canadian population, specifically the more sensitive subpopulations (pregnant women/women of childbearing age, infants, and children) at current exposure levels.

An HI value greater than 1 would indicate the need for further investigation or refinement. The HI values for the three subpopulations with the highest exposures were all below 1. Therefore, further refinement to a higher-tiered assessment is not necessary at this time. While risk is low at current levels for the cumulative risk of medium-chain phthalates (Phthalate Substance Grouping: DIBP, CHIBP, BCHP, DCHP, DBzP, B79P, DMCHP, DIHepP, BIOP, B84P, DINP; Additional phthalates: DPrP, DBP, BBP, DnHP, 79P, DIOP, DEHP), an increase in exposure levels could represent a potential risk to human health.

DIDP

Table 9-60 provides all relevant exposure and hazard values for DIDP, as well as resultant MOEs, for determination of risk, which were previously described in the LCP SOS (Environment Canada, Health Canada 2015d). Overall, the MOEs for DIDP are considered to be adequate to account for uncertainties in the exposure and health effect databases.

DUP

Table 9-61 provides all relevant exposure and hazard values for DUP, as well as resultant MOEs, for determination of risk, which were previously described in the LCP SOS (Environment Canada, Health Canada 2015d). Overall, the MOEs for DUP are considered to be adequate to account for uncertainties in the exposure and health effect databases.

Uncertainties specific to short-chain, medium-chain, long-chain phthalates and DINP are summarized in the SOS reports (Environment Canada, Health Canada 2015b-e).

The key sources of uncertainty related to the CRA are presented in the table below.

Considering all available lines of evidence presented in this draft screening assessment, there is low risk of harm to organisms and the broader integrity of the environment from 13 of the phthalates in the Phthalate Substance Grouping (DMP, DIBP, CHIBP, BCHP, DCHP, DBzP, DMCHP, DIHepP, BIOP, B84P, DINP, DIDP and DUP). However, there is risk of harm to organisms, but not to the broader integrity of the environment, from 1 phthalate included in the Phthalate Substance Grouping, B79P, and 1 additional phthalate, DEHP. DEHP was previously assessed by Environment Canada and Health Canada in 1994 under the Priority Substances Assessment Program. The assessment concluded that DEHP posed a risk to human health in Canada. However, a conclusion for the environment could not be determined because of insufficient information.

It is proposed to conclude that 13 substances in the Phthalate Substance Grouping do not meet the criteria under paragraphs 64(a) or (b) of CEPA as they are not entering the environment in a quantity or concentration or under conditions that have or may have an immediate or long-term harmful effect on the environment or its biological diversity or that constitute or may constitute a danger to the environment on which life depends. It is proposed to conclude that B79P and DEHP meet the criteria under paragraph 64(a) of CEPA as they are entering or may enter the environment in a quantity or concentration or under conditions that have or may have an immediate or long-term harmful effect on the environment or its biological diversity. However, it is proposed to conclude that B79P and DEHP do not meet the criteria under paragraph 64(b) of CEPA as they are not entering the environment in a quantity or concentration or under conditions that constitute or may constitute a danger to the environment on which life depends.

On the basis of the information presented in this draft screening assessment, it is proposed that the 14 phthalates in the Phthalate Substance Grouping do not meet the criteria under paragraph 64(c) of CEPA as they are not entering the environment in a quantity or concentration or under conditions that constitute or may constitute a danger in Canada to human life or health.

Therefore, it is proposed to conclude that B79P and DEHP meet one or more of the criteria set out in section 64 of CEPA. B79P and DEHP have been determined to not meet the persistence and bioaccumulation criteria as set out in the Persistence and Bioaccumulation Regulations of CEPA.

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