Page 3 - Fifth Report on Human Biomonitoring of Environmental Chemicals in Canada

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9 Summaries and results for self-care and consumer product chemicals

9.1 Bisphenol A

Bisphenol A (BPA) (CASRN 80-05-7) is a synthetic chemical used as a monomer in the production of some polycarbonate plastics and as a precursor for monomers of certain epoxy-phenolic resins (EFSA, 2007). Polycarbonate plastics have wide application in consumer products, including storage containers for foods and beverages; they were also used in infant bottles in Canada prior to 2010. Epoxy resins are used as an interior protective lining for food and beverage cans. Additional end-use products containing polycarbonate plastics and resins include medical devices, some dental fillings and sealants, sporting and safety equipment, electronics, and automotive parts (EFSA, 2007; NTP, 2007). BPA is also used in the paper industry to produce thermal paper used for various products, including receipts, prescription labels, airline tickets, and lottery tickets (Geens et al., 2011).

BPA does not occur naturally in the environment (Environment Canada and Health Canada, 2008a). Entry into the environment may occur from industrial sources or from product leaching, disposal, and use (CDC, 2009).

The primary route of exposure to BPA for the general public is through dietary intake as a result of various sources, including migration from food packaging and repeat-use polycarbonate containers (Health Canada, 2008). Health Canada updated its dietary exposure estimates for BPA after completing a number of surveys in which BPA concentrations were measured in various foods, including canned foods and beverages, liquid infant formula, and samples from the Total Diet Study (Health Canada, 2012). Dermal exposure through handling of thermal printing paper is considered an important secondary route of exposure (EFSA CEF Panel, 2015). Oral exposure can also result from leaching of BPA from dental materials; however, the contribution to total BPA exposure is likely negligible (Becher et al., 2018; SCENIHR, 2015). Exposure can also occur from contact with environmental media, including ambient and indoor air, drinking water, soil, and dust, and from the use of consumer products (Environment Canada and Health Canada, 2008a).

In humans, BPA is readily absorbed and undergoes extensive metabolism in the gut wall and the liver (WHO, 2011). Studies have also suggested that it may be absorbed and metabolized by the skin following dermal exposure to free BPA in products such as those made from thermal printing papers (Mielke et al., 2011; Zalko et al., 2011). Glucuronidation has been recognized as a major metabolic pathway for BPA, occurring primarily in the liver and resulting in the BPA-glucuronide conjugate metabolite (EFSA, 2008; FDA, 2008). Conjugation of BPA to BPA-sulphate has been shown to be a minor metabolic pathway (Dekant and Völkel, 2008). The BPA-glucuronide metabolite is not considered to be biologically active and is rapidly excreted in urine with a half-life of less than two hours (WHO, 2011). Urinary levels of total BPA, including both conjugated and free unconjugated forms, are commonly used as biomarkers to assess recent exposures (Arbuckle et al., 2015a; Ye et al., 2005).

Characterization of the potential risk to human health from exposure to BPA includes key effects on the liver and kidneys as well as effects on reproduction, development, neurodevelopment and behaviour (EFSA CEF Panel, 2015; Environment Canada and Health Canada, 2008a; EU, 2010). In 2018, the U.S. National Toxicology Program published results of a comprehensive investigation of BPA toxicity and concluded that early-life and long-term exposures are unlikely to pose a health risk at low doses (NTP, 2018). However, the potential role of BPA and other environmental estrogens in the prevalence of obesity and related metabolic diseases, as well as certain types of cancer, remains under investigation (Heindel et al., 2015; Seachrist et al., 2016).

The Government of Canada conducted a science-based screening assessment under phase one of the Chemicals Management Plan to determine whether BPA presents or may present a risk to the environment or human health as per the criteria set out in section 64 of the Canadian Environmental Protection Act, 1999 (CEPA 1999) (Canada, 1999; Environment Canada and Health Canada, 2008a). Based on information available at that time, the assessment concluded that BPA is toxic under CEPA 1999, as it is considered harmful to the environment and human health (Environment Canada and Health Canada, 2008a). Because of the uncertainty raised by the results of some laboratory animal studies relating to the potential effects of low levels of BPA, a precautionary approach was applied when characterizing risk. Considering the highest potential exposure and subpopulations with potential vulnerability due to potential differences in the toxicokinetics and metabolism of BPA identified in the assessment, the risk management strategy for health focused on decreasing exposure to newborns and infants (Environment Canada and Health Canada, 2008b).

Health Canada has concluded that current dietary exposure to BPA through food packaging is not expected to pose a health risk to the general population, including newborns and young children (Health Canada, 2012). However, exposure to BPA should be as low as reasonably achievable (ALARA) and efforts should continue to limit BPA exposure in infants and newborns from food packaging applications — specifically pre-packaged infant formula products as a sole-source food. As part of the ALARA approach, Health Canada committed to supporting industry to reduce levels of BPA in infant formula can linings (Health Canada, 2014). Health Canada's findings confirm that alternative packaging materials for liquid infant formula products manufactured without BPA have been adopted by industry (Health Canada, 2014). As of March 2010, under the Canada Consumer Product Safety Act, Health Canada has prohibited the manufacturing, advertisement, sale, or import of polycarbonate baby bottles that contain BPA (Canada, 2010). The removal of bisphenol A in polycarbonate baby bottles and liquid infant formula can linings has led Health Canada to conclude that there has been significant progress toward meeting the human health objective for BPA set out in 2008 (Health Canada, 2018a). BPA is also identified as being prohibited on the List of Prohibited and Restricted Cosmetic Ingredients (more commonly referred to as the Cosmetic Ingredient Hotlist or simply the Hotlist), an administrative tool that Health Canada uses to communicate to manufacturers and others that certain substances, when present in a cosmetic, may not be compliant with requirements of the Food and Drugs Act or the Cosmetic Regulations (Health Canada, 2018b). Risk management actions also have been developed under CEPA 1999 with the objective of minimizing releases of BPA in industrial effluents (Canada, 2012).

The Maternal–Infant Research on Environmental Chemicals (MIREC) study is a national-level prospective biomonitoring study carried out in pregnant women aged 18 years and older from 10 sites across Canada (Arbuckle et al., 2013). In the MIREC study of 1,936 participants in their first trimester of pregnancy, the geometric mean and 95^th percentile for total BPA in urine were 0.80 µg/L and 5.40 µg/L, respectively (Arbuckle et al., 2014). The Plastics and Personal-care Products use in Pregnancy (P4) study is a targeted biomonitoring study carried out in 80 pregnant women aged 18 years and older from the Ottawa area. The geometric mean and 95^th percentile for total BPA in urine were 1.1 µg/L and 6.4 µg/L, respectively, based on analyses of multiple samples per woman (Arbuckle et al., 2015b). The First Nations Biomonitoring Initiative (FNBI) is a nationally representative biomonitoring study of adult First Nations peoples living on reserves south of the 60° parallel (AFN, 2013). It comprises 13 randomly selected First Nations communities in Canada with 503 First Nations participants aged 20 years and older. The geometric mean and 95^th percentile for total BPA in urine were 1.55 µg/L and 11.27 µg/L, respectively.

Urinary total BPA (including both free and conjugated forms) was analyzed in the urine of Canadian Health Measures Survey (CHMS) participants aged 6–79 years in cycle 1 (2007–2009), and 3–79 years in cycle 2 (2009–2011), cycle 3 (2012–2013), cycle 4 (2014–2015), and cycle 5 (2016–2017). Data from these cycles are presented as both µg/L and µg/g creatinine. Finding a measurable amount of BPA in urine is an indicator of exposure to BPA and does not necessarily mean that an adverse health effect will occur.

References

• AFN (Assembly of First Nations) (2013). First Nations Biomonitoring Initiative: National results (2011). Assembly of First Nations, Ottawa, ON. Retrieved February 16, 2018.
• Arbuckle, T.E., Davis, K., Marro, L., Fisher, M., Legrand, M., LeBlanc, A., Gaudreau, E., Foster, W.G., Choeurng, V., and Fraser, W.D. (2014). Phthalate and bisphenol A exposure among pregnant women in Canada – Results from the MIREC study. Environment International, 68, 55–65.
• Arbuckle, T.E., Fraser, W.D., Fisher, M., Davis, K., Liang, C.L., Lupien, N., Bastien, S., Velez, M.P., von Dadelszen, P., Hemmings, D.G., et al. (2013). Cohort Profile: The Maternal-Infant Research on Environmental Chemicals Research Platform. Paediatric and Perinatal Epidemiology, 27(4), 415–425.
• Arbuckle, T.E., Marro, L., Davis, K., Fisher, M., Ayotte, P., Bélanger, P., Dumas, P., LeBlanc, A., Bérubé, R., Gaudreau, É., et al. (2015a). Exposure to free and conjugated forms of bisphenol A and triclosan among pregnant women in the MIREC cohort. Environmental Health Perspectives, 123(4), 277–84.
• Arbuckle, T.E., Weiss, L., Fisher, M., Hauser, R., Dumas, P., Bérubé, R., Neisa, A., LeBlanc, A., Lang, C., Ayotte, P., et al. (2015b). Maternal and infant exposure to environmental phenols as measured in multiple biological matrices. Science of the Total Environment, 508, 575–84.
• Becher, R., Wellendorf, H., Sakhi, A.K., Samuelsen, J.T., Thomsen, C., Bølling, A.K., and Kopperud, H.M. (2018). Presence and leaching of bisphenol a (BPA) from dental materials. Acta biomaterialia odontologica Scandinavica, 4(1), 56–62.
• Canada (1999). Canadian Environmental Protection Act, 1999. SC 1999, c. 33. Retrieved February 16, 2018.
• Canada (2010). Canada Consumer Product Safety Act. SC 2010, c. 21. Retrieved February 16, 2018.
• Canada (2012). Notice requiring the preparation and implementation of pollution prevention plans with respect to bisphenol A in industrial effluents. Canada Gazette, Part I: Notices and Proposed Regulations, 146(15). Retrieved February 16, 2018.
• CDC (Centers for Disease Control and Prevention) (2009). Fourth national report on human exposure to environmental chemicals. Department of Health and Human Services, Atlanta, GA. Retrieved February 16, 2018.
• Dekant, W., and Völkel, W. (2008). Human exposure to bisphenol A by biomonitoring: Methods, results and assessment of environmental exposures. Toxicology and Applied Pharmacology, 228(1), 114–134.
• EFSA (European Food Safety Authority) (2007). Opinion of the Scientific Panel on Food Additives, Flavourings, Processing Aids and Materials in Contact with Food on a request from the Commission related to 2,2-Bis(4-hydroxyphenyl)propane (Bisphenol A), Question number EFSA-Q-2005-100. European Food Safety Authority Journal, 428, 1–75.
• EFSA (European Food Safety Authority) (2008). Toxicokinetics of Bisphenol A. Scientific Opinion of the Panel on Food additives, Flavourings, Processing aids and Materials in Contact with Food (AFC), Question number EFSA-Q-2008-382. European Food Safety Authority Journal, 759, 1–10.
• EFSA CEF Panel (European Food Safety Authority Panel on Food Contact Materials, Enzymes, Flavourings and Processing Aids) (2015). Scientific Opinion on the risks to public health related to the presence of bisphenol A (BPA) in foodstuffs: Part I – Exposure assessment. European Food Safety Authority Journal, 13 (1), 3978.
• Environment Canada and Health Canada (2008a). Screening Assessment for the Challenge: Phenol, 4,4'-(1-methylethylidene)bis-(Bisphenol A). Minister of the Environment, Ottawa, ON. Retrieved February 16, 2018.
• Environment Canada and Health Canada (2008b). Proposed risk management approach for phenol, 4,4'-(1-methylethylidene) bis (bisphenol A). Minister of the Environment, Ottawa, ON. Retrieved February 16, 2018.
• EU (European Union) (2010). Updated risk assessment of 4,4'-Isopropylidenediphenol (bisphenol-A), CAS No: 80-05-7, EINECS number: 201-245-8, human health addendum of February 2008. Publications Office of the European Union, Luxembourg. Retrieved September 5, 2019.
• FDA (U.S. Food and Drug Administration) (2008). Draft assessment of bisphenol A for use in food contact applications. U.S. Department of Health and Human Services, Washington, DC. September 16, 2019.
• Geens, T., Goeyens, L., and Covaci, A. (2011). Are potential sources for human exposure to bisphenol-A overlooked? International Journal of Hygiene and Environmental Health, 214(5), 339–347.
• Health Canada (2008). Health Risk Assessment of Bisphenol A from Food Packaging Applications. Minister of Health, Ottawa, ON. Retrieved February 16, 2018.
• Health Canada (2014). Bisphenol A. Minister of Health, Ottawa, ON. Retrieved February 16, 2018.
• Health Canada (2012). Health Canada's Updated Assessment of Bisphenol A (BPA) Exposure from Food Sources. Minister of Health, Ottawa, ON. Retrieved February 16, 2018.
• Health Canada (2014). Bisphenol A: Update on the Food Directorate's Risk Management Commitments for Infant Formula. Minister of Health, Ottawa, ON. Retrieved February 16, 2018.
• Health Canada (2018a). Bisphenol A (BPA) risk management approach: performance evaluation for BPA-HEALTH component. Minister of Health, Ottawa, ON. Retrieved February 13, 2019.
• Health Canada (2018b). List of Ingredients that are Prohibited for Use in Cosmetic Products (Hotlist). Minister of Health, Ottawa, ON. Retrieved November 28, 2018.
• Heindel, J.J., Newbold, R.R., Bucher, J.R., Camacho, L., Delclos, K.B., Lewis, S.M., Vanlandingham, M., Churchwell, M.I., Twaddle, N.C., McLellen, M., et al. (2015) NIEHS/FDA CLARITY-BPA research program. Reproductive Toxicology, 58, 33–44.
• Mielke, H., Partosch, F., and Gundert-Remy, U. (2011). The contribution of dermal exposure to the internal exposure of bisphenol A in man. Toxicology Letters, 204, 190–198.
• NTP (National Toxicology Program) (2007). NTP-CERHR expert panel report on the reproductive and developmental toxicity of bisphenol A. Department of Health and Human Services, Research Triangle Park, NC. Retrieved March 26, 2015.
• NTP (National Toxicology Program) (2018). NTP Research Report on the CLARITY-BPA Core Study: A Perinatal And Chronic Extended-Dose-Range Study of Bisphenol A in Rats. Department of Health and Human Services, Research Triangle Park, NC. Retrieved February 13, 2019.
• SCENIHR (Scientific Committee on Emerging and Newly Identified Health Risks) (2015). SCENIHR opinion on the safety of the use of bisphenol A in medical devices. European Commission, DG Health and Food Safety, Luxembourg. Retrieved February 13, 2019.
• Seachrist, D.D., Bonk, K.W., Ho, S.M., Prins, G.S., Soto, A.M., and Keri, R.A. (2016) A review of the carcinogenic potential of bisphenol A. Reproductive Toxicology, 59, 167–182.
• WHO (World Health Organization) (2011). Toxicological and Health Aspects of Bisphenol A: Report of Joint FAO/WHO Expert Meeting and Stakeholder Meeting on Bisphenol A. World Health Organization, Geneva. Retrieved February 16, 2018.
• Ye, X., Kuklenyik, Z., Needham, L., and Calafat, A. (2005). Quantification of urinary conjugates of bisphenol A, 2,5-dichlorophenol, and 2-hydroxy-4-methoxybenzophenone in humans by online solid phase extraction-high performance liquid chromatography-tandem mass spectrometry. Analytical and Bioanalytical Chemistry, 383(4), 638–644.
• Zalko, D., Jacques, C., Duplan, H., Bruel, S., and Perdu, E. (2011). Viable skin efficiently absorbs and metabolizes bisphenol A. Chemosphere, 82, 424–430.

9.2 Parabens

Parabens are a group of para-hydroxybenzoic (p-hydroxybenzoic) acid esters, four of which were measured in cycle 3, cycle 4, and cycle 5 of the Canadian Health Measures Survey (CHMS): methyl, ethyl, propyl, and butyl paraben.

Parabens are widely used as preservatives in personal care products owing to their antibacterial and antifungal properties (Health Canada, 2019). These products include makeup, moisturizers, sunscreens, hair-care products, facial and skin cleansers, and shaving products. Methyl, propyl, butyl, and ethyl paraben are the most common ones used in cosmetic products (FDA, 2018). Typical concentrations of parabens in cosmetic products are generally 0.5% or less (Health Canada, 2019). Methyl paraben and propyl paraben are permitted for use as food additives (preservatives) in foods sold in Canada. Parabens are also used in pharmaceutical drugs (Ye et al., 2008). Propyl paraben and butyl paraben are classified as active ingredients in natural health products as they are used medicinally as a source of p-hydroxybenzoic acid, a major metabolite of parabens (Health Canada, 2018).

Although parabens in commercial use are synthetically produced, some parabens may also occur naturally in certain fruits and vegetables, such as blueberries and carrots (Health Canada, 2019). Production and use of paraben-containing products can result in their release to the environment through various waste streams. A potential route of exposure for the general public is dermal contact with products that contain parabens, such as moisturizers and cosmetics. Approximately 50% of cosmetics in the United States contain parabens, with methylparaben being the most commonly used and lipstick having the highest concentrations (Cosmetic Ingredient Review Expert Panel, 2008; Yazar et al., 2011). Oral exposure to parabens can also occur through consumption of foods or pharmaceuticals containing parabens, ingestion of breast milk, and ingestion of house dust (CDC, 2009; Fan et al., 2010; Ye et al., 2008).

Dermal exposure may result in small amounts of parabens being absorbed. Following oral exposure, parabens are rapidly absorbed from the gastrointestinal tract (NTP, 2005). Once absorbed, parabens are mainly hydrolyzed to p-hydroxybenzoic acid that can then be conjugated with glycine, glucuronide, and sulphate for excretion in urine (Soni et al., 2005). Currently, there is no evidence of bioaccumulation potential in humans. In laboratory animals, complete elimination of orally ingested ethyl and propyl paraben was observed within 72 hours (Soni et al., 2005). Data are limited in humans. One study of premature infants who received intramuscular injections of paraben-containing gentamicin observed that excretion of methyl paraben, primarily in the conjugated form, was variable (approximately 10% to 90%), possibly due to factors such as variation in intramuscular absorption and gestational and postpartum ages (Hindmarsh et al., 1983). A study of parabens in the urine of 100 adults found that parabens in urine appear predominantly in their conjugated forms (Ye et al., 2006). The concentration of parabens in urine (conjugated and free) can be used as a biomarker of exposure to parabens. As p-hydroxybenzoic acid is a nonspecific metabolite of all parabens, it may not be an optimal biomarker of exposure for specific parabens.

Health effects have not been observed as a result of exposures to parabens at concentrations found in cosmetics, with acute, subchronic and chronic experimental animal studies demonstrating a low order of toxicity of parabens (Cosmetic Ingredient Review Expert Panel, 2008). It should be noted that most of the available toxicity data are from single paraben exposure studies, and that the additive and cumulative risks of exposures to multiple parabens are not well studied (Karpuzoglu et al., 2013). Animal studies have found parabens to be non-allergenic; however, sporadic human cases of anaphylactic reactions have been reported following paraben exposure. Parabens have been found to weakly mimic estrogens in vitro, but well-conducted animal studies do not demonstrate estrogenic effects (Sivaraman et al., 2018); human data do not support a link to estrogenic effects because exposure to parabens has not been observed to affect hormone levels or sperm quality (Adoamnei et al., 2018; Meeker et al., 2011). Parabens have not been found to be carcinogenic in chronic animal studies. The International Agency for Research on Cancer (IARC) has not evaluated parabens with respect to human carcinogenicity.

Methyl, ethyl, propyl, and butyl paraben have been identified as priorities for risk assessment based on human health concerns under the Chemicals Management Plan Identification of Risk Assessment Priorities (IRAP) initiative (Environment and Climate Change Canada and Health Canada, 2015). A screening-level risk assessment is currently under way to determine whether parabens present or may present a risk to the environment or human health as per the criteria set out in section 64 of the Canadian Environmental Protection Act, 1999 (Canada, 1999; Environment and Climate Change Canada, 2019). As part of this assessment, Health Canada will present currently available evidence on the use of these parabens in each of the various products regulated by Health Canada as well as the results of its risk assessment. Health Canada continues to monitor and review any new scientific data on parabens (Health Canada, 2019).

Parabens were measured in a 2011 biomonitoring study carried out in Alberta with 39 participants aged 12–67 years who were patients at a primary care clinic specializing in environmental health sciences (Genuis et al., 2013). The 50^th percentile urinary concentrations measured in this study were 25.95, 10.30, 2.80, and 0.32 µg/L for methyl, ethyl, propyl, and butyl paraben, respectively.

Methyl, ethyl, propyl, and butyl paraben were analyzed in the urine of CHMS participants aged 3–79 years in cycle 3 (2012–2013), cycle 4 (2014–2015), and cycle 5 (2016–2017). Data are presented as µg/L and µg/g creatinine. Finding a measurable quantity of parabens in urine is an indicator of exposure to parabens and does not necessarily mean that an adverse health effect will occur.

References

• Adoamnei, E., Mengiola, J., Monino-Garcia, M., Vela-Soria, F., Iribarne-Durán, L.M., Fernández, M.F., Olea, N., Jørgensen, N., Swan, S.H., and Torres-Cantero, A.M. (2018). Urinary concentrations of parabens and reproductive parameters in young men. Science of the Total Environment, 621, 201–209.
• Canada (1999). Canadian Environmental Protection Act, 1999. SC1999, c. 33. Retrieved March 2, 2018.
• CDC (Centers for Disease Control and Prevention) (2009). Fourth national report on human exposure to environmental chemicals. Department of Health and Human Services, Atlanta, GA. Retrieved March 2, 2018.
• Cosmetic Ingredient Review Expert Panel (2008). Final amended report on the safety assessment of Methylparaben, Ethylparaben, Propylparaben, Isopropylparaben, Butylparaben, Isobutylparaben, and Benzylparaben as used in cosmetic products. International Journal of Toxicology, 27(Suppl 4), 1–82.
• Environment and Climate Change Canada and Health Canada (2015). Identification of risk assessment priorities: results of the 2015 review. Minister of the Environment, Ottawa, ON. Retrieved December 12, 2018.
• Environment and Climate Change Canada (2019). List of substances in the third phase of CMP (2016–2021): July 2019 update. Minister of Environment and Climate Change, Ottawa, ON. Retrieved September 6, 2019.
• Fan, X., Kubwabo, C., Rasmussen, P., and Jones-Otazo, H. (2010). Simultaneous quantitation of parabens, triclosan, and methyl triclosan in indoor house dust using solid phase extraction and gas chromatography-mass spectrometry. Journal of Environmental Monitoring, 12(10), 1891–1897.
• FDA (U.S. Food and Drug Administration) (2018). Parabens in Cosmetics. U.S. Department of Health and Human Services, Washington, DC. Retrieved March 2, 2018.
• Genuis, S.J., Birkholz, D., Curtis, L., and Sandau, C. (2013). Paraben levels in an urban community of Western Canada. ISRN Toxicology, 1–8.
• Health Canada (2018). Natural Health Products Ingredients Database. Minister of Health, Ottawa, ON. Retrieved May 8, 2018.
• Health Canada (2019). Safety of Cosmetic Ingredients. Minister of Health, Ottawa, ON. Retrieved March 28, 2019.
• Hindmarsh, K.W., John, E., Asali, L.A., French, J.N., Williams, G.L., and McBride, W.G. (1983). Urinary excretion of methylparaben and its metabolites in preterm infants. Journal of Pharmaceutical Sciences, 72(9), 1039–1041.
• Karpuzoglu, E., Holladay, S.D., and Gogal Jr., R.M. (2013). Parabens: Potential impact of Low-Affinity Estrogen receptor Binding chemicals on Human health. Journal of Toxicology and Environmental Health, Part B, 16(5), 321–335.
• Meeker, J.D., Yang, T., Ye, X., Calafat, A.M., and Huaser, R. (2011). Urinary concentrations of parabens and serum hormone levels, semen quality parameters and sperm DNA damage.Environmental Health Perspectives, 119(2), 252–257.
• NTP (National Toxicology Program) (2005). Butylparaben – Review of Toxicological Literature. U.S. Department of Health and Human Services, Research Triangle Park, NC. Retrieved May 8, 2018.
• Sivaraman, L., Pouliot, L., Wang, B., Brodie, T., and Graziano, M. (2018). Safety assessment of propylparaben in juvenile rats. Regulatory Toxicology and Pharmacology, 92, 370–381.
• Soni, M.G., Carabin, I.G., and Burdock, G.A. (2005). Safety assessment of esters of p-hydroxybenzoic acid (parabens). Food and Chemical Toxicology, 43(7), 985–1015.
• Yazar, K., Johnsson, S., Lind, M., Boman, A., and Lidén, C. (2011). Preservatives and fragrances in selected consumer-available cosmetics and detergents. Contact Dermatitis, 64(5), 265–272.
• Ye, X., Bishop, A.M., Reidy, J.A., Needham, L.L., and Calafat, A.M. (2006). Parabens as urinary biomarkers of exposure in humans. Environmental Health Perspectives, 114(12), 1843–1846.
• Ye, X., Bishop, A.M., Needham, L.L., and Calafat, A.M. (2008). Automated on-line column-switching HPLC-MS/MS method with peak focusing for measuring parabens, triclosan, and other environmental phenols in human milk. Analytica Chimica Acta, 622(1–2), 150–156.

10 Summary and results for nicotine

10.1 Nicotine

Cotinine (CASRN 486-56-6) is the major primary metabolite of nicotine, a chemical found naturally in the tobacco plant and present in tobacco products, such as cigarettes, cigars, and smokeless tobacco products (e.g., chewing tobacco and snuff) (Benowitz and Jacob, 1994). Nicotine is also incorporated into nicotine delivery products, such as nicotine gum, patches, lozenges, inhalers, buccal sprays, and vaping products (Etter et al, 2011).

Human exposure to nicotine occurs primarily through the use of tobacco, vaping, and other nicotine delivery products, and from exposure to environmental tobacco smoke (HSDB, 2009). In addition, infants breastfed by women who smoke may be exposed to nicotine in breast milk (HSDB, 2009).

Inhalation is the most effective intake route; on average, 60% to 80% of nicotine is absorbed through the lungs (Iwase et al., 1991). Nicotine absorption through the mouth varies with the pH of the smoke or nicotine delivery product, increasing as alkalinity rises (Benowitz et al., 2009). Nicotine can also be absorbed through the skin and gastrointestinal tract, but at a much lower efficiency compared with inhalation (Karaconji, 2005). Once inside the body, approximately 70% to 80% of nicotine is metabolized into cotinine, primarily by a liver cytochrome P-450 enzyme. It has a half-life of 10 to 20 hours and can remain in the body at detectable levels for up to seven days (Benowitz and Jacob, 1994; Curvall et al., 1990; Hecht et al., 1999). Cotinine is considered to be the most relevant biomarker for exposure to tobacco products and tobacco smoke (Brown et al., 2005; CDC, 2009; Seaton and Vesell, 1993). It has also been shown to be a biomarker of exposure to nicotine via other types of nicotine delivery products, such as e-cigarettes (Schick et al., 2017; Vélez de Mendizábal et al., 2015). It should be noted that there are no validated biomarkers that can differentiate among the use of various combustible products (e.g., cigars, cigarillos, water pipes, and cigarettes), and there are no validated biomarkers that are specific to nicotine-containing or nicotine-free vaping products (Schick et al., 2017).

Nicotine reaches the brain rapidly following inhalation and can cause several reactions in the body, such as: increased heart rate and blood pressure, muscle relaxation, altered brain activity, and constriction of blood vessels leading to a drop in temperature of the hands and feet (Health Canada, 2013). Other effects may include nausea, weakness, stomach cramps, and headache, with symptoms lessening as nicotine tolerance is developed. Nicotine mimics the effects of acetylcholine in the nervous system. Through the release of dopamine and effects on other neurotransmitters, it can activate areas of the brain that are associated with feelings of alertness, calmness, and pleasure (Pandey et al., 2018). As the body builds tolerance to nicotine, the delivery product must continue to be used for the effects to last; use over time may lead to dependence and addiction (Health Canada, 2013). The use of nicotine-containing products is associated with exposure to other chemicals that have their own effects. For example, tobacco smoke contains more than 4,000 chemicals, including at least 70 that cause, initiate, or promote cancer, and others that contribute to adverse health effects, such as emphysema, heart disease, and increased risk of asthma (CDC, 2004; Health Canada, 2011; IARC, 2004). Levels of cotinine in the blood and urine of non-smokers have been correlated with some adverse health effects related to environmental tobacco smoke exposure. Cotinine itself may contribute to the neuropharmacological effects of tobacco smoking (Benowitz, 1996; Crooks and Dwoskin, 1997).

As a result of the adverse health effects associated with tobacco use, the Government of Canada, along with provincial and territorial governments and various municipalities, has taken several steps to reduce the prevalence of tobacco use as well as exposure to tobacco smoke. These steps include prohibitions on the sales of tobacco products and electronic nicotine delivery systems to youth, requirements to apply health warnings on tobacco packaging, and restrictions on the promotion of tobacco products, including the display of tobacco products at retail outlets (Health Canada, 2006). Additional steps include the offer of cessation help along with initiatives to eliminate smoking in workplaces and enclosed public locations (Health Canada, 2006). In 2018, Health Canada enacted the Tobacco and Vaping Products Act, which amends the Canada Consumer Product Safety Act to allow the effective regulation of vaping products as well as the ability to establish plain and standardized appearance requirements for tobacco product packages (Health Canada, 2018). This legislation aims to protect young people and non-smokers from inducements to nicotine addiction and tobacco use, and to enhance public awareness of the health and safety hazards posed by tobacco and vaping products.

The First Nations Biomonitoring Initiative (FNBI) was a nationally representative biomonitoring study of adult First Nations peoples living on reserves south of the 60° parallel (AFN, 2013). It comprised 13 randomly selected First Nations communities in Canada with 503 First Nations participants aged 20 years and older. In 2011, the 50^th percentile concentrations of cotinine in urine from smokers and non-smokers were 315.79 µg/L and <1.1 µg/L, respectively.

Data from cycle 1 (2007–2009) of the Canadian Health Measures Survey (CHMS) demonstrated that a substantial proportion of the Canadian population is exposed to secondhand smoke. The study found detectable cotinine levels (≥1.1 ng/ml) in non-smokers, indicating secondhand smoke exposure, and reported that children and adolescent subpopulations had higher levels compared with adults (Wong et al., 2013). A study of occupationally exposed non-smoking bar workers in the Toronto area examined the effects of a 2004 smoke-free workplace bylaw; the study showed a one-month post-ban decline in the geometric mean of urinary cotinine, from 10.3 µg/L to 3.10 µg/L (Repace et al., 2013). A concentration of 50 µg/L urine for cotinine is recommended for determining smoking status; greater concentrations are attributed to smokers (SRNT Subcommittee on Biochemical Verification, 2002). Using this concentration, a study assessed the validity of self-reported cigarette smoking status among Canadians using urinary cotinine data from cycle 1 (2007–2009) of the CHMS (Wong et al., 2012). Compared with estimates based on urinary cotinine concentration, smoking prevalence based on self-reporting was only 0.3 percentage points lower. This indicates that accurate estimates of the prevalence of cigarette smoking among Canadians can be derived from self-reported smoking status data.

Cotinine was analyzed in the urine of all CHMS participants aged 6–79 years in cycle 1 (2007–2009), and 3–79 years in cycle 2 (2009–2011), cycle 3 (2012–2013), cycle 4 (2014–2015), and cycle 5 (2016–2017). Data from these cycles are presented as both µg/L and µg/g creatinine for non-smokers and smokers. Survey participants aged 3–11 years were assumed to be non-smokers. In this survey, a smoker is defined as someone who is a current daily or occasional smoker; a non-smoker is defined as someone who does not currently smoke, and has either never smoked or was previously a daily or occasional smoker. Finding a measurable amount of cotinine in urine is an indicator of exposure to nicotine.

In addition to free cotinine, nicotine and several other metabolites (cotinine-N-glucuronide, nicotine-N-glucuronide, trans-3-hydroxycotinine, trans-3-hydroxycotinine-O-glucuronide, and anabasine) were analyzed in cycle 1 (2007–2009) and cycle 3 (2012–2013) of the CHMS. Free and total 4-(methylnitrosoamino)-1-(3-pyridyl)-1-butanol (NNAL), a metabolite of a tobacco-specific N-nitrosamine found only in tobacco and products derived from tobacco, were also analyzed in cycle 1 (2007–2009) and cycle 3 (2012–2013) of the CHMS. Data on these tobacco chemicals and their metabolites are available from Statistics Canada through the Research Data Centres Program.

References

• AFN (Assembly of First Nations) (2013). First Nations Biomonitoring Initiative: National results (2011). Assembly of First Nations, Ottawa, ON. Retrieved February 27, 2018.
• Benowitz, N.L. (1996). Cotinine as a biomarker of environmental tobacco smoke exposure. Epidemiologic Reviews, 18(2), 188–204.
• Benowitz, N.L., and Jacob, P. (1994). Metabolism of nicotine to cotinine studied by a dual stable isotope method. Clinical Pharmacology and Therapeutics, 56(5), 483–493.
• Brown, K.M., von Weymarn, L.B., and Murphy, S.E. (2005). Identification of N-(hydroxymethyl) norcotinine as a major product of cytochrome P450 2A6, but not cytochrome P450 2A13-catalyzed cotinine metabolism. Chemical Research in Toxicology, 18(12), 1792–1798.
• CDC (Centers for Disease Control and Prevention) (2004). The Health Consequences of Smoking: A Report of the Surgeon General. Department of Health and Human Services, Washington, DC. Retrieved February 27, 2018.
• CDC (Centers for Disease Control and Prevention) (2009). Fourth national report on human exposure to environmental chemicals. Department of Health and Human Services, Atlanta, GA. Retrieved February 27, 2018.
• Crooks, P.A., and Dwoskin, L.P. (1997). Contribution of CNS nicotine metabolites to the neuropharmacological effects of nicotine and tobacco smoking. Biochemical Pharmacology, 54(7), 743–753.
• Curvall, M., Elwin, C.E., Kazemi-Vala, E., Warholm, C., and Enzell, C.R. (1990). The pharmacokinetics of cotinine in plasma and saliva from non-smoking healthy volunteers. European Journal of Clinical Pharmacology, 38(3), 281–287.
• Etter, J.F., Bullen, C., Flouris, A.D., Laugesen, M., and Eissenberg, T. (2010). Electronic nicotine delivery systems: a research agenda. Tobacco Control, 20(3), 243–8.
• Health Canada (2006). The National Strategy: Moving Forward – The 2006 Progress Report on Tobacco Control. Minister of Health, Ottawa, ON. Retrieved February 27, 2018.
• Health Canada (2011). Carcinogens in Tobacco Smoke. Minister of Health, Ottawa, ON. Retrieved February 27, 2018.
• Health Canada (2013). Nicotine addiction. Minister of Health, Ottawa, ON. Retrieved August 23, 2018.
• Health Canada (2018). Tobacco and Vaping Products Act. Minister of Health, Ottawa, ON. Retrieved August 30, 2018.
• Hecht, S.S., Carmella, S.G., Chen, M., Dor Koch, J.F., Miller, A.T., Murphy, S.E., Jensen, J.A., Zimmerman, C.L., and Hatsukami, D.K. (1999). Quantitation of urinary metabolites of a tobacco-specific lung carcinogen after smoking cessation. Cancer Research, 59(3), 590–596.
• HSDB (Hazardous Substances Data Bank) (2009). Nicotine. National Library of Medicine, Bethesda, MD. Retrieved February 27, 2018.
• IARC (International Agency for Research on Cancer) (2004). IARC Monographs on the Evaluation of Carcinogenic Risks to Humans – Volume 83: Tobacco Smoke and Involuntary Smoking. World Health Organization, Lyon. Retrieved February 27, 2018.
• Iwase, A., Aiba, M., and Kira, S. (1991). Respiratory nicotine absorption in non-smoking females during passive smoking. International Archives of Occupational and Environmental Health, 63(2), 139–143.
• Karaconji, I.B. (2005). Facts about nicotine toxicity. Archives of Industrial Hygiene and Toxicology, 56(4), 363–371.
• Pandey, R., Gurumurthy, Narang, P., Remedios, K., Kadam, K., Murugappan, M, Sridhar, A., and Galinski, J. (2018). Nicotinic Receptor Up regulation and Nicotine Addiction: A Review of Mechanisms. Chronicles of Pharmaceutical Science, 2(3), 572–580.
• Repace, J., Zhang, B., Bondy, S.J., Benowitz, N., and Ferrence, R. (2013). Air quality, mortality, and economic benefits of a smoke-free workplace law for non-smoking Ontario bar workers. Indoor Air, 23(2), 93–104.
• Schick, S.F., Blount, B.C., Jacob 3rd, P., Saliba, N.A., Bernert, J.T., El Hellani, A., Jatlow, P., Pappas, R.S., Wang, L., Foulds, J., et al. (2017). Biomarkers of exposure to new and emerging tobacco delivery products. American Journal of Physiology Lung Cellular and Molecular Physiology 313, L425–L452.
• Seaton, M.J., and Vesell, E.S. (1993). Variables affecting nicotine metabolism. Pharmacology and Therapeutics, 60(3), 461–500.
• SRNT Subcommittee on Biochemical Verification (2002). Biochemical verification of tobacco use and cessation. Nicotine and Tobacco Research, 4, 149–159.
• Vélez de Mendizábal, N., Jones, D.R., Jahn, A., Bies, R.R., and Brown, J.W. (2015). Nicotine and Cotinine Exposure from Electronic Cigarettes: A Population Approach. Clinical Pharmacokinetics, 54(6), 615–626.
• WHO (World Health Organization) (2010). Gender, women, and the tobacco epidemic. WHO, Geneva. Retrieved February 27, 2018.
• Wong, S.L., Malaison, E., Hammond, D., and Leatherdale, S.T. (2013). Secondhand smoke exposure among Canadians: Cotinine and self-report measures from the Canadian Health Measures Survey 2007–2009. Nicotine and Tobacco Research, 15(3), 693–700.
• Wong, S.L., Shields, M., Leatherdale, S., Malaison, E., and Hammond, D. (2012). Assessment of validity of self-reported smoking status. Health Reports, 23(1), 1–7.

11 Summary and results for acrylamide

11.1 Acrylamide

Acrylamide (CASRN 79-06-1) is a chemical used primarily in the production of polymers such as polyacrylamides (ATSDR, 2012). Polyacrylamides are used to clarify drinking water and treat effluent from water treatment plants and industrial processes (ATSDR, 2012). They are also used as binding, thickening, or flocculating agents in grout, cement, pesticide formulations, cosmetics, food manufacturing, and soil erosion prevention (Environment Canada and Health Canada, 2009a). Polymers of acrylamide are also used in ore processing, food packaging, and plastic products (Environment Canada and Health Canada, 2009a). In Canada, polyacrylamides are used as coagulants and flocculants for the clarification of drinking water, in potting soils, and as a non-medicinal ingredient in natural health products and pharmaceuticals (Environment Canada and Health Canada, 2009b). Acrylamide can also form in certain foods as a reaction between naturally present components when foods are processed or cooked at high temperatures (Health Canada, 2009a). It is formed mainly in carbohydrate-rich, plant-based foods, such as potatoes and grains. The highest concentrations have been detected in potato chips and french fries, acrylamide has been found in other foods as well (Health Canada, 2012).

Acrylamide may enter the environment during production and industrial use (ATSDR, 2012). The main source of acrylamide in drinking water is through the release of residual monomers from polyacrylamides used as clarifiers in drinking water treatment processes (ATSDR, 2012). Acrylamide is also a component of cigarette smoke and may be released to indoor air as a result of smoking (NTP, 2005; Urban et al., 2006).

Acrylamide exposure in the general population occurs primarily through food (ATSDR, 2012; Environment Canada and Health Canada, 2009b). Inhalation of tobacco smoke, including second-hand smoke, is also a major source of inhalation exposure for the general population; tobacco smoke may be the main source of acrylamide exposure for some smokers (ATSDR, 2012; Environment Canada and Health Canada, 2009b; EFSA CONTAM Panel, 2015). Compared with food and cigarettes, exposure from other sources (e.g., drinking water, air, consumer product use) is very low (Environment Canada and Health Canada, 2009b). Animal studies indicate that acrylamide is readily absorbed via oral and pulmonary routes, and to a lesser degree following dermal exposure (ATSDR, 2012). Once absorbed, acrylamide is widely distributed throughout the body, accumulating in red blood cells (ATSDR, 2012). Acrylamide is metabolized via glutathione conjugation to form a mercapturic acid acrylamide derivative or by oxidation to form the epoxide derivative, glycidamide, which can also undergo conjugation with glutathione. Both acrylamide and glycidamide react with haemoglobin in red blood cells, forming adducts (ATSDR, 2012). Absorbed acrylamide and its metabolites are rapidly eliminated in urine, primarily as mercapturic acid conjugates of acrylamide and glycidamide (ATSDR, 2012). Acrylamide and glycidamide haemoglobin adducts are considered markers of exposure over the previous 120 days, the average life span of red blood cells (ATSDR, 2012).

Exposure to acrylamide has been reported to cause neurotoxicity in humans. Inhalation exposure to acrylamide in occupational settings has been associated with peripheral neuropathy, characterized by muscle weakness and numbness in hands and feet (Environment Canada and Health Canada, 2009b). Studies with laboratory animals have observed adverse reproductive and developmental effects, and have shown that acrylamide is genotoxic and carcinogenic (Environment Canada and Health Canada, 2009b; FAO/WHO, 2006). Reviews of existing epidemiological studies have found inadequate evidence in humans to establish an association between acrylamide exposure and carcinogenicity (Health Canada, 2008; IARC, 1994). However, on the basis of evidence in experimental animal studies, the International Agency for Research on Cancer (IARC) has classified acrylamide as a Group 2A probable carcinogen (IARC, 1994). Further, on the basis of available evidence from animal studies, the Joint Food and Agriculture organization of the United Nations (FAO) and World Health Organization (WHO) Expert Committee on Food Additives determined that the estimated intake of acrylamide from certain foods may be a human health concern (FAO/WHO, 2006; FAO/WHO, 2011). Similarly, an assessment by the European Food Safety Authority (ESFA) concluded that acrylamide in food potentially increases the risk of developing cancer for consumers in all age groups (EFSA CONTAM Panel, 2015).

The Government of Canada has conducted a science-based screening assessment under the Chemicals Management Plan to determine whether acrylamide may present a risk to the environment or human health as per the criteria set out in section 64 of the Canadian Environmental Protection Act, 1999 (CEPA 1999) (Canada, 1999; Environment Canada and Health Canada, 2009b). The assessment concluded, on the basis of carcinogenic potential, that acrylamide is toxic under CEPA 1999, as it is considered harmful to human health (Environment Canada and Health Canada, 2009b). Acrylamide is listed on Schedule 1, List of Toxic Substances, under CEPA 1999. The Act allows the federal government to control the importation, manufacture, distribution, and use of acrylamide in Canada (Canada, 1999; Canada, 2011). Health Canada's risk management strategy for acrylamide in food is focused on reducing foodborne exposure to acrylamide (Health Canada, 2009b). To reduce exposure to acrylamide from food sources, Health Canada suggests following the recommendations provided in Canada's Food Guide, thereby limiting consumption of carbohydrate-rich foods that are high in fat (such as potato chips and French fries), sugar, or salt (Health Canada, 2009a). However, occasional consumption of these products is not likely to be a health concern. Other suggestions for reducing exposure to acrylamide from certain foods include paying careful attention to oil and baking temperatures, following the manufacturer's cooking instructions, storing potatoes at a temperature above 8°C, washing or soaking cut potatoes in water prior to frying, and toasting bread or baked goods to the lightest colour acceptable (Health Canada, 2009a). Health Canada regularly reviews data on the concentrations of acrylamide in foods sold on the Canadian market; these results may be shared with industry, particularly if elevated levels of acrylamide are identified in certain products. Health Canada continues to encourage the food industry to further pursue efforts to reduce acrylamide in processed foods (Health Canada, 2012). Data on the occurrence of acrylamide in foods available for sale in Canada do not demonstrate a decreasing trend in acrylamide concentrations in the food types that can significantly contribute to dietary acrylamide exposure; therefore, continued mitigation efforts are supported (Health Canada, 2017). Health Canada has also amended the Food and Drug Regulations to permit the use of asparaginase in certain food products to reduce the formation of acrylamide during cooking (Canada, 2012; Health Canada, 2013).

Because acrylamide-containing polymers are used in drinking water treatment, most Canadian jurisdictions have requirements to meet health-based standards for additives that limit the amount of acrylamide present in treated drinking water (NSF International, 2015; NSF International, 2016). Health Canada has also set a maximum level for acrylamide in polyacrylamide-containing formulations used in natural health products in Canada (Environment Canada and Health Canada, 2009a; Health Canada, 2018b). Acrylamide is identified as being prohibited on the List of Prohibited and Restricted Cosmetic Ingredients (more commonly referred to as the Cosmetic Ingredient Hotlist or simply the Hotlist), an administrative tool that Health Canada uses to communicate to manufacturers and others that certain substances, when present in a cosmetic, may not be compliant with requirements of the Food and Drugs Act or the Cosmetic Regulations (Health Canada, 2018a).

In a study carried out in Montreal to assess the levels of acrylamide in 195 non-smoking teenagers aged 10–17 years, the geometric mean concentrations of haemoglobin adducts of acrylamide and glycidamide were 45.4 pmol/g globin and 45.6 pmol/g globin, respectively (Brisson et al., 2014).

Acrylamide and its metabolite glycidamide were analyzed as haemoglobin adducts in the whole blood of CHMS participants aged 3–79 years in cycle 3 (2012–2013), cycle 4 (2014–2015), and cycle 5 (2016–2017). Data are presented in blood as pmol/g haemoglobin (Hb). Finding a measurable amount of acrylamide or glycidamide haemoglobin adducts in blood is an indicator of exposure to acrylamide and does not necessarily mean that an adverse health effect will occur.

References

• ATSDR (Agency for Toxic Substances and Disease Registry) (2012). Toxicological Profile for Acrylamide. U.S. Department of Health and Human Services, Atlanta, GA. Retrieved February 16, 2018.
• Brisson, B., Ayotte, P., Normandin, L., Gaudreau, E., Bienvenu, J.F., Fennell, T.R., Blanchet, C., Phaneuf, D., Lapointe, C., Bonvalot, Y., et al. (2014). Relation between dietary acrylamide exposure and biomarkers of internal dose in Canadian teenagers. Journal of Exposure Science and Environmental Epidemiology, 24(2), 215–221.
• Canada (1999). Canadian Environmental Protection Act, 1999. SC 1999, c. 33. Retrieved February 16, 2018.
• Canada (2011). Order Adding a Toxic Substance to Schedule 1 to the Canadian Environmental Protection Act, 1999. Canada Gazette, Part II: Official Regulations, 145(5). Retrieved February 16, 2018.
• Canada (2012). Regulations Amending the Food and Drug Regulations (1652 – Schedule F). Canada Gazette, Part II: Official Regulations, 146(6). Retrieved February 16, 2018.
• EFSA CONTAM Panel (European Food Safety Authority Panel on Contaminants in the Food Chain) (2015). Scientific Opinion on acrylamide in food. European Food Safety Authority Journal, 13(6), 4104.
• Environment Canada and Health Canada (2009a). Proposed Risk Management Approach for 2-Propenamide (Acrylamide). Minister of the Environment, Ottawa, ON. Retrieved February 16, 2018.
• Environment Canada and Health Canada (2009b). Screening Assessment for the Challenge: 2-Propenamide (Acrylamide). Minister of the Environment, Ottawa, ON. Retrieved February 16, 2018.
• FAO/WHO (Food and Agriculture Organization of the United Nations/World Health Organization) (2006). Evaluation of certain food contaminants. Sixty-fourth report of the Joint FAO/WHO Expert Committee on Food Additives. World Health Organization, Geneva. Retrieved February 16, 2018.
• FAO/WHO (Food and Agriculture Organization of the United Nations/World Health Organization) (2011). Acrylamide (addendum). Safety evaluation of certain contaminants in food. Seventy-second meeting of the Joint FAO/WHO Expert Committee on Food Additives. World Health Organization, Geneva. Retrieved February 16, 2018.
• Health Canada (2008). Health Canada reviews epidemiological studies on dietary acrylamide intake and the risk of endometrial, ovarian and/or breast cancer. Minister of Health, Ottawa, ON. Retrieved February 16, 2018.
• Health Canada (2009a). Acrylamide – What you can do to reduce exposure. Minister of Health, Ottawa, ON. Retrieved February 16, 2018.
• Health Canada (2009b). Health Canada Updates its Risk Management Strategy for Acrylamide in Food. Minister of Health, Ottawa, ON. Retrieved February 16, 2018.
• Health Canada (2012). Health Canada's Revised Exposure Assessment of Acrylamide in Food. Minister of Health, Ottawa, ON. Retrieved February 16, 2018.
• Health Canada (2013). Notice of Modification to the List of Permitted Food Enzymes to Enable the Use of the Enzyme Asparaginase, Obtained from Aspergillus Oryzae (pCaHj621/BECh2#10), in Green Coffee. Minister of Health, Ottawa, ON. Retrieved February 16, 2018.
• Health Canada (2017). Summary of Comments Received as part of Health Canada's 2013–2014 Call for Data to Assess the Effectiveness of Acrylamide Reduction Strategies in Food. Minister of Health, Ottawa, ON. Retrieved February 16, 2018.
• Health Canada (2018a). List of Ingredients that are Prohibited for Use in Cosmetic Products (Hotlist). Minister of Health, Ottawa, ON. Retrieved November 28, 2018.
• Health Canada (2018b). Natural Health Products Ingredients Database. Minister of Health, Ottawa, ON. Retrieved February 16, 2018.
• IARC (International Agency for Research on Cancer) (1994). IARC Monographs on the Evaluation of Carcinogenic Risks to Humans – Volume 60: Some Industrial Chemicals. World Health Organization, Lyon. Retrieved February 16, 2018.
• NSF International (2015). NSF/ANSI Standard 60: Drinking Water Treatment Additives – health effects. NSF International, Ann Arbor, MI. Retrieved February 16, 2018.
• NSF International (2016). Survey of ASDWA Members on the Use of NSF/ANSI Standards: Standards 60, 61, 223 and 372. NSF International, Ann Arbor, MI. Retrieved February 16, 2018.
• NTP (National Toxicology Program) (2005). NTP-CERHR Monograph on the Potential Human Reproductive and Developmental Effects of Acrylamide. U.S. Department of Health and Human Services, Research Triangle Park, NC. Retrieved April 21, 2015.
• Urban, M., Kavvadias, D., Riedel, K., Scherer, G., and Tricker, A.R. (2006). Urinary mercapturic acids and a hemoglobin adduct for the dosimetry of acrylamide exposure in smokers and nonsmokers. Inhalation Toxicology, 18(10), 831–839.

12 Summary and results for perfluoroalkyl and polyfluoroalkyl substances

12.1 Perfluoroalkyl and polyfluoroalkyl substances

Perfluoroalkyl and polyfluoroalkyl substances (PFAS) are structurally related persistent organic compounds that have a fluorinated alkyl (carbon) chain structure. Perfluoroalkyl substances are characterized by the presence of a fully fluorinated alkyl chain that is typically four to 14 carbons in length. In contrast, polyfluoroalkyl substances are not fully fluorinated and have a hydrogen or oxygen attached to at least one carbon in the alkyl chain. Nine perfluoroalkyl substances were measured in cycle 5 of the Canadian Health Measures Survey (CHMS) (Table 12.1.1).

PFAS are synthetic chemicals with high chemical and thermal stability and the ability to repel both water and oils (Kissa, 2001). These characteristics make them ideal for use in a number of industrial and commercial applications (Kissa, 2001). PFAS are used as stain-repellent, water-repellent, and oil-repellent fabric protectors, in water-repellent and oil-repellent paper coatings, wiper blades, bike-chain lubricant, wire and cable insulation, pharmaceutical packaging, and food packaging (Kissa, 2001). They are also used in engine-oil additives, nail polish, hair curling and straightening products, metal plating and cleaning products, fire retardant foams, inks, varnishes, polyurethane production, and vinyl polymerization (Kissa, 2001). Fluoropolymers manufactured using salts of PFAS are used in many industrial and consumer products, including surface coatings on textiles and carpets, personal care products, and non-stick coatings on cookware (INAC, 2009; Kissa, 2001; Prevedouros et al., 2005).

PFOS and PFOA are the most extensively studied and measured PFAS in humans (Dallaire et al., 2009; Hölzer et al., 2008; Kato et al., 2011). PFHxS is another perfluorinated compound that has been measured in humans, but it has not been examined as extensively as PFOS and PFOA. Other PFAS, such as PFBA, PFHxA, PFNA, PFDA, PFUnDA, and PFBS, have been measured less frequently in the human population.

Worldwide use of PFOS and PFOS-related products has decreased significantly since 2002, when the world's largest producer at the time completed its voluntary phase-out of production (ITRC, 2017). PFHxS, a known by-product in the production of PFOS, was also phased out as a result. In 2008, replacements for PFOA were introduced, resulting in the subsequent phase-out of PFOA in the production of fluoropolymers (ITRC, 2017). Potential replacements for PFOS-based substances include new PFBS-based compounds that are rapidly eliminated from the body with relatively low bioaccumulation potential and toxicity; however, it is not yet clear if long-chain PFAS alternatives can achieve the same performance effectiveness of their predecessors (Chang et al., 2008; Newsted et al., 2008; ITRC, 2017).

PFAS do not occur naturally in the environment. Entry into the environment occurs through releases during manufacturing and transport, use of consumer products, and the disposal and breakdown of larger PFAS. As a result, PFAS have been detected in a wide array of environmental media (Houde et al., 2006).

For the general public, exposure to PFAS is widespread through food, drinking water, consumer products, dust, soil, and air (Fromme et al., 2009; Fromme et al., 2007; Hölzer et al., 2008; Kubwabo et al., 2005). PFAS have been analyzed as part of Health Canada's ongoing Total Diet Study surveys; levels in foods that are commercially sold in Canada are low, similar to levels that have been reported in other countries (Health Canada, 2014; Health Canada, 2016a; Tittlemier et al., 2006; Tittlemier et al., 2007). The contribution of individual pathways and sources of exposure appears to depend on age, dose, and substance. Generally, ingestion of food, drinking water, and house dust are expected to be the main routes of exposure for adults in the general population, whereas oral hand-to-mouth contact with consumer products, such as carpets, clothing, and upholstery, is a significant contributor for infants, toddlers, and children (Trudel et al., 2008).

Longer-chain PFAS are well absorbed in the body, poorly excreted, and not extensively metabolized (Harada et al., 2005; INAC, 2009; Johnson et al., 1984). Average half-lives of PFOS, PFOA, and PFHxS in humans range from 3–9 years (Olsen et al., 2007). Shorter-chain PFAS are eliminated much more quickly; for example, the elimination half-life for PFBA is 72 to 81 hours (ATSDR, 2015). In humans, PFOS and PFOA are found in serum, plasma, kidneys, and the liver (Butenhoff et al., 2006; Fromme et al., 2009; Kärrman et al., 2010). PFAS have also been measured in breast milk and umbilical cord blood (Kärrman et al., 2010; Monroy et al., 2008). PFAS have a strong affinity for the protein fraction in blood and do not typically accumulate in lipids (Kärrman et al., 2010; Martin et al., 2004). Serum levels of PFAS, in particular PFOA and PFOS, can reflect cumulative exposure over several years (CDC, 2009). Although both PFOA and PFOS are biomarkers of exposures to themselves, animal studies have indicated that their presence in serum may also result from exposure to and subsequent metabolism of other PFAS (ATSDR, 2015). Absorbed PFOA and PFOS are ultimately excreted in urine (ATSDR, 2015).

The primary concern with PFAS is their persistence in both the environment and the human body (Olsen et al., 2007). Possible linkages between exposure to PFAS and adverse human health effects have been examined in occupational studies and studies of populations exposed to contaminated drinking water (ATSDR, 2015). Although no definitive links have been established, reports in children and neonates suggest associations between serum PFAS and thyroid effects (Lopez-Espinosa et al., 2012; Wang et al., 2014). A recent review by Ballesteros et al. (2017) also reported a positive association between maternal or teenage male exposure to certain PFAS and thyroid-stimulating hormone levels, despite heterogeneity across studies. In several animal species, the liver has been identified as the primary target organ of toxicity for PFAS regardless of the route of exposure (ATSDR, 2015; EPA, 2002; Health Canada, 2006). PFOA has been associated with increased incidence of tumours in rodent bioassays; following the identification of PFOA and other PFAS as priority agents for International Agency for Research on Cancer (IARC) monographs, PFOA was classified as possibly carcinogenic to humans (Group 2B) based on limited evidence in humans for a positive association with cancers of the testis and kidney (IARC, 2017).

In 2006, Environment Canada and Health Canada concluded that PFOS was not a concern for human health at current levels of exposure (Health Canada, 2006). However, PFOS and its salts were declared toxic to the environment and its biological diversity, and PFOS was added to Schedule 1 of the Canadian Environmental Protection Act, 1999 (CEPA 1999) (Canada, 1999; Environment Canada, 2006). In 2009, PFOS and its salts were added to the Virtual Elimination List under CEPA 1999 (Canada, 2009). In 2016, PFOS was added to the Prohibition of Certain Toxic Substances Regulations, prohibiting most uses aside from exemptions for specific uses (Health Canada, 2016c). Canada is also working through the Convention on Long-Range Transboundary Air Pollution and the Stockholm Convention on Persistent Organic Pollutants under the United Nations to reduce the global production of PFOS (Health Canada, 2016c).

In 2012, Environment Canada and Health Canada published screening assessments for PFOA and long-chain perfluorocarboxylic acids (including PFNA, PFDA, and PFUnDA), along with their salts and their precursors (Environment Canada, 2012; Environment Canada and Health Canada, 2012b). The assessments concluded that the substances are an ecological concern, but that PFOA and its salts and precursors are not a concern for human health (Environment Canada, 2012; Environment Canada and Health Canada, 2012b). Long-chain perfluorocarboxylic acids and their salts and precursors were not determined to be a high priority for assessment of potential risks to human health; as such, no human health assessment was conducted. Based on the assessments, both PFOA and long-chain perfluorocarboxylic acids and their salts and precursors have been added to the List of Toxic Substances in Schedule 1 of CEPA 1999 (Canada, 2012a).

Health Canada, in collaboration with the Federal-Provincial-Territorial Committee on Drinking Water, has also developed guidelines for Canadian drinking water quality that establish maximum acceptable concentrations for PFOS and PFOA in drinking water (Health Canada, 2018a; Health Canada, 2018b). Both guidelines were developed based on liver effects in laboratory animals and are considered to be protective of both non-cancer and cancer effects. Health Canada has also developed drinking water screening values for several additional PFAS, including PFBA, PFHxA, PFNA, PFBS, and PFHxS (Health Canada, 2018c).

A number of risk management measures for perfluorocarboxylic acids and their precursors have been implemented by the Government of Canada. These measures include regulations prohibiting the manufacture, use, sale, offer for sale, and import of four fluorotelomer-based substances found to be precursors to long-chain perfluorinated carboxylic acids, unless present in certain manufactured items (Canada, 2010; Environment Canada and Health Canada, 2012a). A five-year Environmental Performance Agreement that commenced in 2010 resulted in participating companies successfully meeting their commitments to eliminate residual PFOA, long-chain perfluorocarboxylic acids, and their precursors in products (Health Canada, 2016b). Long-chain perfluorocarboxylic acids, PFOA, and PFOS, along with their salts and precursors, are now regulated under the Prohibition of Certain Toxic Substances Regulations, 2012 (Canada, 2012b).

Globally, there is an initiative to reduce PFOA emissions and product content. In 2006, the United States Environmental Protection Agency (EPA) and eight major companies in the industry launched the 2010/15 PFOA Stewardship Program. Under this voluntary effort, several companies exited the PFAS industry altogether, while others stopped the manufacture and import of long-chain PFAS and transitioned to alternative chemicals (EPA, 2018a; EPA, 2018c).The EPA recently released draft toxicity assessments for GenX chemicals and PFBS, members of a larger group of PFAS (EPA, 2018b). Canada's 2010 Environmental Performance Agreement was consistent with the targets and commitments by industry in the United States (Environment Canada, 2010). The European Union and the Australian government have initiated similar policies where PFAS are either prohibited or subject to further toxicity testing for evaluation.

Several human biomonitoring studies in Canada have measured PFAS in serum and plasma (Alberta Health and Wellness, 2008; Fisher et al., 2016; Hamm et al., 2010; Kubwabo et al., 2004; Monroy et al., 2008; Tittlemier et al., 2004; Turgeon O'Brien et al., 2012). In 2002, serum samples from 56 individuals in Ottawa, Ontario, and Gatineau, Québec were analyzed for PFOA and PFOS. Mean concentrations of PFOA and PFOS were 3.4 µg/L and 28.8 µg/L, respectively (Kubwabo et al., 2004). In 2004, PFOS was measured in plasma samples from 883 Nunavik Inuit living in the Canadian Arctic with a geometric mean concentration of 18.68 µg/L (Dallaire et al., 2009). The concentrations of PFAS were measured in 86 Inuit children, 11 months to 4.5 years of age, attending childcare centres in Nunavik between 2006 and 2008 (Turgeon O'Brien et al., 2012). The geometric mean concentrations in plasma for PFOA, PFHxS, and PFOS were 1.62 µg/L, 0.33 µg/L, and 3.37 µg/L, respectively. In the Maternal–Infant Research on Environmental Chemicals (MIREC) study of 1,940 participants, the geometric means (95^th percentile) for PFOA, PFHxS, and PFOS in plasma were 1.65 µg/L (4.1 µg/L), 1.03 µg/L (4.3 µg/L), and 4.56 µg/L (11 µg/L), respectively (Fisher et al., 2016). MIREC is a national-level prospective biomonitoring study carried out in pregnant women aged 18 years and older recruited from 10 sites across Canada between 2008 and 2011 (Arbuckle et al., 2013).

PFOS, PFOA, and PFHxS were measured in the plasma of CHMS participants aged 20–79 years in cycle 1 (2007–2009), 12–79 years in cycle 2 (2009–2011) and 3–79 years in cycle 5 (2016–2017). PFBA, PFHxA, PFBS, PFNA, PFDA, and PFUnDA were measured in the plasma of CHMS participants aged 12–79 years in cycle 2 (2009–2011) and 3–79 years in cycle 5 (2016–2017). Data for the PFAS are presented as μg/L in plasma (Tables 12.1.2 to 12.1.19). Finding a measurable amount of PFAS in plasma is an indicator of exposure to PFAS and does not necessarily mean that an adverse health effect will occur.

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