Page 3: Sixth 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 for various products, including receipts, prescription labels, airline tickets and lottery tickets (Geens et al., 2011).

BPA does not occur naturally in the environment (EC and HC, 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 from various sources, including migration from food packaging and repeat-use polycarbonate containers (HC, 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 (HC, 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 (EC and HC, 2008a).

In humans, BPA is readily absorbed and undergoes extensive metabolism in the gut wall and 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). There continues to be some uncertainty as to whether the BPA-glucuronide metabolite is biologically active. However, it is rapidly excreted in urine with a half-life of less than 2 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., 2015; 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; EC and HC, 2008a; EU, 2010). In 2018, the U.S. National Toxicology Program published the 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). The potential role of BPA and other environmental estrogens in the prevalence of obesity, related metabolic diseases and certain types of cancer continues to be investigated (Heindel et al., 2015; Seachrist et al., 2016).

The Government of Canada conducted a science-based screening assessment under phase 1 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; EC and HC, 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 (EC and HC, 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 possible 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 (EC and HC, 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 (HC, 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 (HC, 2014). Health Canada's findings confirm that alternative packaging materials for liquid infant formula products manufactured without BPA have been adopted by industry (HC, 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 BPA 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 (HC, 2018). BPA is on the List of Ingredients that are Prohibited for Use in Cosmetic Products (HC, 2019). Risk management actions also have been developed under CEPA 1999 with the objective of minimizing releases of BPA in industrial effluents (Canada, 2012).

BPA concentrations in urine have been measured in a number of biomonitoring studies conducted in Canada, including the Maternal–Infant Research on Environmental Chemicals study (Arbuckle et al., 2014) and the First Nations Biomonitoring Initiative (AFN, 2013).

Urinary total BPA (including both free and conjugated forms) was analyzed in the urine of Canadian Health Measures Survey (CHMS) participants aged 6–79 in cycle 1 (2007–2009) and aged 3–79 in cycle 2 (2009–2011), cycle 3 (2012–2013), cycle 4 (2014–2015), cycle 5 (2016–2017) and cycle 6 (2018–2019). 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 March 1, 2021.
• 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., Marro, L., Davis, K., Fisher, M., Ayotte, P., Bélanger, P., Dumas, P., LeBlanc, A., Bérubé, R., Gaudreau, É., et al. (2015). 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.
• 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 March 1, 2021.
• Canada (2010). Canada Consumer Product Safety Act. SC 2010, c. 21. Retrieved March 1, 2021.
• 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 March 1, 2021.
• 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 1, 2021.
• 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.
• EC and HC (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 March 1, 2021.
• EC and HC (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 March 1, 2021.
• EU (European Union) (2010). Updated risk assessment of 4,4'-Isopropylidenediphenol (bisphenol-A), CAS No: 80-05-7, EINECS No: 201-245-8, Human Health Addendum of February 2008. Publications Office of the European Union, Luxembourg. Retrieved March 1, 2021.
• 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. March 1, 2021.
• 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.
• HC (Health Canada) (2008). Health Risk Assessment of Bisphenol A from Food Packaging Applications. Minister of Health, Ottawa, ON. Retrieved March 1, 2021.
• HC (Health Canada) (2012). Health Canada's Updated Assessment of Bisphenol A (BPA) Exposure from Food Sources. Minister of Health, Ottawa, ON. Retrieved March 1, 2021.
• HC (Health Canada) (2014). Bisphenol A: Update on the Food Directorate's Risk Management Commitments for Infant Formula. Minister of Health, Ottawa, ON. Retrieved March 1, 2021.
• HC (Health Canada) (2018). Bisphenol A (BPA) risk management approach: performance evaluation for BPA-HEALTH component. Minister of Health, Ottawa, ON. Retrieved March 1, 2021.
• HC (Health Canada) (2019). List of Ingredients that are Prohibited for Use in Cosmetic Products (Hotlist). Minister of Health, Ottawa, ON. Retrieved March 1, 2021.
• 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 1, 2021.
• 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 March 1, 2021.
• SCENIHR (Scientific Committee on Emerging and Newly Identified Health Risks) (2015). Safety of the use of bisphenol A in medical devices. European Commission, DG Health and Food Safety, Luxembourg. Retrieved March 1, 2021.
• 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 Report of Stakeholder Meeting on Bisphenol A. World Health Organization, Geneva. Retrieved March 1, 2021.
• 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, 4 of which were measured in cycles 3 through 6 of the Canadian Health Measures Survey (CHMS): methyl, ethyl, propyl and butyl paraben (Table 9.2.1).

Parabens are widely used as preservatives in personal care products owing to their antibacterial and antifungal properties, and are also used as fragrance ingredients (ECCC and HC, 2020; HC, 2020). These products include makeup, moisturizers, sunscreens, hair-care products, facial and skin cleansers, shaving products and toothpaste. Methyl, propyl, butyl and ethyl parabens are the most common ones used in cosmetic products (FDA, 2020). Concentrations of parabens in cosmetic products are typically 0.5% or less (HC, 2020). Methyl and propyl paraben are permitted for use as food additives (preservatives) in certain foods sold in Canada. Parabens are also used in food packaging materials, prescription and non-prescription drugs and pest control products (ECCC and HC, 2020). Methyl, ethyl, propyl and butyl paraben are listed as non-medicinal ingredients in natural health products (HC, 2021).

Some parabens may occur naturally in certain fruits and vegetables, such as blueberries and carrots (HC, 2020). The synthetic production and use of paraben-containing products can result in their release to the environment through various waste streams. Parabens are expected to degrade and not persist in water, air, sediment or soil (ECCC and HC, 2020).

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 methyl paraben 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). 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). In humans, parabens are eliminated rapidly in urine and are recovered predominantly as hydrolyzed and conjugated isoforms (ECCC and HC, 2020). In a study of orally dosed volunteers, Moos et al. (2016) found that the fraction of parent paraben excreted in urine generally decreased with increasing molecular weight (i.e., increasing length of the alkyl side chain). The concentration of parabens in urine (parent and metabolites) can be used as a biomarker of exposure to parabens. Certain metabolites, such as p-hydroxybenzoic acid and p-hydroxyhippuric acid, are common to all parabens; as such, they may not be optimal biomarkers of exposure for specific parabens. Conversely, parent parabens are specific biomarkers in urine. It should be noted that because parent parabens are widely used as preservatives, their use as biomarkers may be susceptible to contamination during sample collection and analysis (Aylward et al., 2017).

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 low toxicity of parabens (Cosmetic Ingredient Review Expert Panel, 2008). Animal studies have reported various health effects depending on the specific paraben, route of exposure and dose. These include depression and decreased motor activity after ethyl paraben exposure, and delayed onset of puberty, altered morphology of reproductive organs and reduced sperm count and motility in offspring following gestational exposure to butyl paraben (ECCC and HC, 2020). Animal studies have found parabens to be non-allergenic; however, sporadic human cases of allergic reactions have been reported following paraben exposure (Cosmetic Ingredient Review Expert Panel, 2008). 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). 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). Parabens have not been found to be carcinogenic in chronic animal studies. The International Agency for Research on Cancer has not evaluated parabens with respect to human carcinogenicity.

The Government of Canada has conducted a screening assessment to determine whether 7 substances referred to collectively as the Parabens Group 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 (CEPA 1999) (Canada, 1999; ECCC and HC, 2020). The draft assessment released in 2020 proposed to conclude that methyl paraben, propyl paraben and butyl paraben are toxic under CEPA 1999, given that they are considered harmful to human health, but that ethyl paraben does not meet the criteria to be considered toxic (ECCC and HC, 2020).

A limited number of biomonitoring studies conducted in Canada have measured the concentrations of parabens in urine (e.g. Genuis et al., 2013; Fisher et al., 2017).

Methyl, ethyl, propyl and butyl paraben were analyzed in the urine of CHMS participants aged 3–79 in cycle 3 (2012–2013), cycle 4 (2014–2015), cycle 5 (2016–2017) and cycle 6 (2018–2019). 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.
• Aylward, L.L., Hays, S.M., Kirman, C., and Summit Toxicology, LLP. (2017). Derivation of Biomonitoring Equivalents for Selected Esters of para-Hydroxybenzoic Acid (Parabens). Health Canada Contract No. 4500354002.
• Canada (1999). Canadian Environmental Protection Act, 1999. SC 1999, c. 33. Retrieved March 1, 2021.
• 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 1, 2021.
• 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.
• ECCC and HC (Environment and Climate Change Canada and Health Canada) (2020). Draft Screening Assessment: Parabens Group. Minister of Environment and Climate Change, Ottawa, ON. Retrieved October 22, 2020.
• 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) (2020). Parabens in Cosmetics. U.S. Department of Health and Human Services, Washington, DC. Retrieved March 1, 2021.
• Fisher, M., MacPherson, S., Braun, J.M., Hauser, R., Walker, M., Feeley, M., Mallick, R., Bérubé, R. and Arbuckle, T.E. (2017). Paraben concentrations in maternal urine and breast milk and its association with personal care product use. Environmental Science and Technology, 51, 4009-4017.
• Genuis, S.J., Birkholz, D., Curtis, L., and Sandau, C. (2013). Paraben levels in an urban community of Western Canada. ISRN Toxicology, 1–8.
• HC (Health Canada) (2020). Safety of Cosmetic Ingredients. Minister of Health, Ottawa, ON. Retrieved March 1, 2021.
• HC (Health Canada) (2021). Natural Health Products Ingredients Database. Minister of Health, Ottawa, ON. Retrieved March 1, 2021.
• 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.
• Moos, R.K., Angerer, J., Dierkes, G., Brüning, T., Koch, H.M. (2016). Metabolism and elimination of methyl, iso and n-butyl paraben in human urine after single oral dosage. Archives of Toxicology, 90, 2699–2709.
• NTP (National Toxicology Program) (2005). Butylparaben – Review of Toxicological Literature. U.S. Department of Health and Human Services, Research Triangle Park, NC. Retrieved March 1, 2021.
• 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., 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. Cotinine has a half-life of 10 to 20 hours and can remain in the body at detectable levels for up to 7 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 (HC, 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 (HC, 2013). While cotinine itself may contribute to the neuropharmacological effects of tobacco smoking (Benowitz, 1996; Crooks and Dwoskin, 1997), 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; HC, 2011; IARC, 2004). Levels of cotinine in the blood and urine of non-smokers related to environmental tobacco smoke exposure have been correlated with adverse health effects.

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 sale 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 (HC, 2006). Additional steps include the offer of cessation help along with initiatives to eliminate smoking in workplaces and enclosed public locations (HC, 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 (HC, 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.

Cotinine concentrations in urine have been measured in a limited number of biomonitoring studies conducted in Canada, including the First Nations Biomonitoring Initiative (AFN, 2013).

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 second-hand smoke. The study found detectable cotinine levels (≥1.1 ng/ml) in non-smokers, indicating second-hand 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 1-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 CHMS participants aged 6–79 in cycle 1 (2007–2009) and cycle 6 (2018–2019), and aged 3–79 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. Cotinine was analyzed in the serum of all CHMS participants aged 6–79 in cycle 6. Data from this cycle are presented in serum as µg/L for non-smokers and smokers. Survey participants aged 3–11 were assumed to be non-smokers. In this survey, a smoker is defined as someone who is a current daily or occasional smoker, while 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. These definitions are based on self-reported data. Finding a measurable amount of cotinine in urine or serum 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 CHMS cycles 1 and 3. Free and total 4-(methylnitrosamino)-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 CHMS cycles 1 and 3. Data on these tobacco-related 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 March 1, 2021.
• 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.
• Benowitz, N.L., Hukkanen, J., and Jacob, P. (2009). Nicotine chemistry, metabolism, kinetics and biomarkers. Handbook of Experimental Pharmacology, 192, 29–60.
• 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 March 1, 2021.
• 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 1, 2021.
• 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. (2011). Electronic nicotine delivery systems: a research agenda. Tobacco Control, 20(3), 243–248.
• HC (Health Canada) (2006). The National Strategy: Moving Forward – The 2006 Progress Report on Tobacco Control. Minister of Health, Ottawa, ON. Retrieved March 1, 2021.
• HC (Health Canada) (2011). Carcinogens in Tobacco Smoke. Minister of Health, Ottawa, ON. Retrieved March 1, 2021.
• HC (Health Canada) (2013). Nicotine addiction. Minister of Health, Ottawa, ON. Retrieved March 1, 2021.
• HC (Health Canada) (2018). Tobacco and Vaping Products Act. Minister of Health, Ottawa, ON. Retrieved March 1, 2021.
• 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 March 1, 2021.
• 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 March 1, 2021.
• 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.
• 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). Polymers of acrylamide are also used in ore processing, food packaging and plastic products (EC and HC, 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 (EC and HC, 2009b). Acrylamide can also form in certain foods as a product of reactions between naturally present components when foods are processed or cooked at high temperatures (HC, 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 (HC, 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; EC and HC, 2009b). Inhalation of tobacco smoke, including second-hand smoke, is also a major source of exposure for the general population; tobacco smoke may be the main source of acrylamide exposure for some smokers (ATSDR, 2012; EC and HC, 2009b; EFSA CONTAM Panel, 2015). Compared with food and cigarettes, exposure from other sources (e.g., drinking water, air and products available to consumers) is very low (EC and HC, 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 (EC and HC, 2009b). Studies with laboratory animals have observed adverse reproductive and developmental effects, and have shown that acrylamide is genotoxic and carcinogenic (EC and HC, 2009b; FAO/WHO, 2006). Reviews of existing epidemiological studies have found inadequate evidence in humans to establish an association between acrylamide exposure and carcinogenicity (HC, 2008; IARC, 1994). However, on the basis of evidence in experimental animal studies, the International Agency for Research on Cancer 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 (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; EC and HC, 2009b). The assessment concluded that acrylamide is toxic under CEPA 1999, as it is considered harmful to human health (EC and HC, 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; 2011). Health Canada's risk management strategy for acrylamide in food is focused on reducing foodborne exposure to acrylamide (HC, 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 (HC, 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 (HC, 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 pursue efforts to reduce acrylamide in processed foods (HC, 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 (HC, 2017). Health Canada has also approved the use of asparaginase in certain food products to reduce the formation of acrylamide during cooking (Canada, 2012; HC, 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, 2019; 2021). Health Canada has also set a maximum level for acrylamide in polyacrylamide-containing formulations used in natural health products in Canada (EC and HC, 2009a; HC, 2021). Acrylamide is on the List of Ingredients that are Prohibited for Use in Cosmetic Products (HC, 2019).

A limited number of biomonitoring studies conducted in Canada have measured the concentrations of haemoglobin adducts of acrylamide and glycidamide in blood (e.g. Brisson et al., 2014).

Acrylamide and its metabolite glycidamide were analyzed as haemoglobin adducts in the whole blood of CHMS participants aged 3–79 in cycle 3 (2012–2013), cycle 4 (2014–2015), cycle 5 (2016–2017) and cycle 6 (2018–2019). 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 March 2, 2021.
• 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 March 2, 2021.
• 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 March 2, 2021.
• Canada (2012). Regulations Amending the Food and Drug Regulations (1652 – Schedule F). Canada Gazette, Part II: Official Regulations, 146(6). Retrieved March 2, 2021.
• 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.
• EC and HC (Environment Canada and Health Canada) (2009a). Proposed Risk Management Approach for 2-Propenamide (Acrylamide). Minister of the Environment, Ottawa, ON. Retrieved March 2, 2021.
• EC and HC (Environment Canada and Health Canada) (2009b). Screening Assessment for the Challenge: 2-Propenamide (Acrylamide). Minister of the Environment, Ottawa, ON. Retrieved March 2, 2021.
• 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 March 2, 2021.
• 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 March 2, 2021.
• HC (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 March 2, 2021.
• HC (Health Canada) (2009a). Acrylamide – What you can do to reduce exposure. Minister of Health, Ottawa, ON. Retrieved March 2, 2021.
• HC (Health Canada) (2009b). Health Canada Updates its Risk Management Strategy for Acrylamide in Food. Minister of Health, Ottawa, ON. Retrieved March 2, 2021.
• HC (Health Canada) (2012). Health Canada's Revised Exposure Assessment of Acrylamide in Food. Minister of Health, Ottawa, ON. Retrieved March 2, 2021.
• HC (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 March 2, 2021.
• HC (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 March 2, 2021.
• HC (Health Canada) (2019). List of Ingredients that are Prohibited for Use in Cosmetic Products (Hotlist). Minister of Health, Ottawa, ON. Retrieved March 2, 2021.
• HC (Health Canada) (2021). Natural Health Products Ingredients Database. Minister of Health, Ottawa, ON. Retrieved March 2, 2021.
• 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 March 2, 2021.
• NSF International (2019). Survey of ASDWA Members on the Use of NSF/ANSI Standards: Standards 60, 61 and 372. NSF International, Ann Arbor, MI. Retrieved March 2, 2021.
• NSF International (2021). NSF/ANSI/CAN Standard 60: Drinking Water Treatment Additives – health effects. NSF International, Ann Arbor, MI. Retrieved March 2, 2021.
• 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 March 2, 2021.
• 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 connected to a functional group. In contrast, polyfluoroalkyl substances are not fully fluorinated and have a hydrogen or oxygen attached to at least 1 carbon in the alkyl chain. Nine perfluoroalkyl substances have been measured in 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 as oil, grease, soil and water repellents and as surfactants in a wide range of industrial and consumer product applications (Bălan et al., 2021). These applications span many sectors of the economy, including building and construction, automotive, chemicals and pharmaceuticals, electronics and semiconductors, first-responder safety and health care (Bălan et al., 2021; 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).

Long-chain perfluoroalkyl substances include perfluoroalkyl carboxylates with 8 or more fully fluorinated carbons (e.g. PFOA) and perfluorinated sulfonates with 6 or more fully fluorinated carbons (e.g. PFOS and PFHxS), their salts, and precursor compounds capable of forming long-chain perfluoroalkyl substances (ITRC, 2020). 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 has been measured in humans, but not as extensively as PFOS and PFOA, while other PFAS, such as PFBA, PFHxA, PFNA, PFDA, PFUnDA and PFBS, have been measured less frequently in human populations.

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, 2020). PFHxS, a known by-product in the production of PFOS, was also phased out as a result. Manufacturers have been developing replacement technologies, including reformulating longer-chain substances or substituting them with nonfluorinated chemicals, alternate technologies or shorter-chain PFAS (ITRC, 2020). However, these replacement PFAS may also be associated with environmental and/or human health effects.

PFAS enter the environment through releases during manufacturing and transport, use of 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., 2007; 2009; Hölzer et al., 2008). PFAS have been analyzed as part of Health Canada's ongoing Total Diet Study surveys; levels in foods that were commercially sold in Canada were low, similar to levels that have been reported in other countries (HC, 2014; Tittlemier et al., 2006; Tittlemier et al., 2007). The contribution of individual pathways and sources of exposure appears to depend on age, exposure concentration and substance. Generally, ingestion of food, drinking water and house dust are expected to be the main routes of exposure for the general population, while hand-to-mouth contact with consumer products, such as carpets, clothing and upholstery, is a significant contributor for infants, toddlers and children (Sunderland et al., 2019; Trudel et al., 2008).

In general, PFAS are well absorbed in the body, poorly excreted, and not extensively metabolized (Harada et al., 2005; INAC, 2009; Johnson et al., 1984). Half-lives of PFOS, PFOA and PFHxS in humans can range from months to decades (Olsen et al., 2007; Zhang et al., 2013). Shorter-chain PFAS are eliminated much more quickly; for example, the elimination half-life for PFBA is 72 to 81 hours (ATSDR, 2018). 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; Li et al., 2020; Monroy et al., 2008; Motas Guzmàn et al., 2016). In general, 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). The presence of these substances in serum may also be a result of exposure to other PFAS that can be subsequently metabolized to PFOS and PFOA (ATSDR, 2018). Absorbed PFOA and PFOS are ultimately excreted in urine (ATSDR, 2018).

The primary concern with PFAS is their persistence in the environment and potential persistence in the human body (Olsen et al., 2007). Possible associations between exposure to certain PFAS and adverse human health effects have been identified (ATSDR, 2018). For example, 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. Research suggests that exposure to certain PFAS may be associated with other health effects, including increased cholesterol levels, the body's ability to respond to vaccines, decreased fertility in women and an increased risk of conditions like high blood pressure or pre-eclampsia in pregnant women (ATSDR, 2018). In several animal species, the liver has been identified as the primary target organ of toxicity for certain PFAS, while effects on the immune system and development have also been noted (ATSDR, 2018; EPA, 2002; HC, 2006). PFOA has been associated with increased incidence of tumours in rodent bioassays and was classified as possibly carcinogenic to humans (Group 2B) based on limited evidence in humans for a positive association with cancers of the testes and kidneys (IARC, 2017).

The Government of Canada evaluated the risks of PFOS, PFOA and long-chain perfluorocarboxylic acids that have the molecular formula CnF2n+1CO2H in which 8 ≤ n ≤ 20, as well as their salts and precursors (LC-PFCAs), and published the findings in screening assessment reports in 2006 and 2012 (EC, 2006; 2012; EC and HC, 2012). These substances were found to be toxic to the environment and were added to the List of Toxic Substances under Schedule 1 of the Canadian Environmental Protection Act, 1999 (CEPA 1999) (Canada, 1999; 2012a). As a result, measures to manage risks from these substances were put in place. In response to risk management measures taken over the past decade by the Government of Canada and other international jurisdictions, industry has shifted from using PFOS, PFOA and LC-PFCAs to using other PFAS as substitutes.

In Canada, risk management actions for PFOS have been in place since 2008. Since 2016, the manufacture, use, sale, offer for sale or import of PFOS, PFOA, LC-PFCAs and products that contain them have been prohibited, with a limited number of exemptions (e.g., manufactured items containing PFOA or LC-PFCAs) under the Prohibition of Certain Toxic Substances Regulations, 2012 (Canada, 2012b). In 2018, a consultation document was published outlining a proposal to further restrict these substances by removing all current exemptions (ECCC, 2018). The proposed regulations are targeted for publication in fall 2021. Internationally, Canada is working with the United Nations to eliminate or restrict the production and use of PFOA, PFOS and related substances through the Stockholm Convention on Persistent Organic Pollutants.

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 (HC, 2018a; 2018b). Health Canada has also developed drinking water screening values for several additional PFAS, including PFBA, PFHxA, PFNA, PFBS and PFHxS (HC, 2019; 2020).

A number of biomonitoring studies conducted in Canada have measured concentrations of PFAS in plasma, including the Maternal–Infant Research on Environmental Chemicals (Fisher et al., 2016) and the First Nations Biomonitoring Initiative (AFN, 2013).

PFOS, PFOA and PFHxS were measured in the plasma of CHMS participants aged 20–79 in cycle 1 (2007–2009). PFOS, PFOA and PFHxS, along with PFBA, PFHxA, PFBS, PFNA, PFDA and PFUnDA, were measured in the plasma of CHMS participants aged 12–79 in cycle 2 (2009–2011) and aged 3–79 in cycle 5 (2016–2017) and cycle 6 (2018–2019). Data for 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.

References

• AFN (Assembly of First Nations) (2013). First Nations Biomonitoring Initiative: National results (2011). Assembly of First Nations, Ottawa, ON. Retrieved May 11, 2021.
• ATSDR (Agency for Toxic Substances and Disease Registry) (2018). Toxicological Profile for Perfluoroalkyls. U.S. Department of Health and Human Services, Atlanta, GA. Retrieved March 2, 2021.
• Bălan, S.A., Mathrani, V.C., Guo, D.F., and Algazi, A.M. (2021). Regulating PFAS as a Chemical Class under the California Safer Consumer Products Program. Environmental Health Perspectives, 129 (2), 025001.
• Ballesteros, V., Costa, O., Iñiguez, C., Fletcher, T., Ballester, F., and Lopez-Espinosa, M. (2017). Exposure to perfluoroalkyl substances and thyroid function in pregnant women and children: A systematic review of epidemiologic studies. Environment International, 99, 15–28.
• Butenhoff, J.L., Olsen, G.W., and Pfahles-Hutchens, A. (2006). The applicability of biomonitoring data for perfluorooctanesulfonate to the environmental public health continuum. Environmental Health Perspectives, 114 (11), 1776–1782.
• Canada (1999). Canadian Environmental Protection Act, 1999. SC 1999, c. 33. Retrieved March 2, 2021.
• Canada (2012a). Order Adding Toxic Substances to Schedule 1 to the Canadian Environmental Protection Act, 1999. Canada Gazette, Part I: Notices and Proposed Regulations, 146 (39). Retrieved March 2, 2021.
• Canada (2012b). Regulations Amending the Prohibition of Certain Toxic Substances Regulations, 2012 (for the addition of 5 substances). SOR/2016-252. Retrieved March 2, 2021.
• 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, 2021.
• Dallaire, R., Ayotte, P., Pereg, D., Déry, S., Dumas, P., Langlois, E., and Dewailly, E. (2009). Determinants of plasma concentrations of perfluorooctanesulfonate and brominated organic compounds in Nunavik Inuit adults (Canada). Environmental Science and Technology, 43 (13), 5130–5136.
• ECCC (Environment and Climate Change Canada) (2018). Consultation Document on Proposed Amendments to the Prohibition of Certain Toxic Substances Regulations, 2012 for PFOS, PFOA, LC-PFCAs, HBCD, PBDEs, DP and DBDPE (December 2018). Minister of Environment and Climate Change, Ottawa, ON. Retrieved January 27, 2021.
• EC (Environment Canada) (2006). Ecological screening assessment report on perfluorooctane sulfonate, salts and precursors. Minister of the Environment, Ottawa, ON. Retrieved March 2, 2021.
• EC (Environment Canada) (2012). Ecological Screening Assessment Report: Long-Chain (C9–C20) Perfluorocarboxylic Acids, their Salts and their Precursors. Minister of the Environment, Ottawa, ON. Retrieved March 2, 2021.
• EC and HC (Environment Canada and Health Canada) (2012). Screening Assessment: Perfluorooctanoic Acid, its Salts, and its Precursors. Minister of the Environment, Ottawa, ON. Retrieved March 2, 2021.
• EPA (U.S. Environmental Protection Agency) (2002). Revised draft hazard assessment of perflourooctanoic acid and its salts. U.S. Environmental Protection Agency, Washington DC.
• Fisher, M., Arbuckle, T.E., Liang, C.L., LeBlanc, A., Gaudreau, E., Foster, W.G., Haines, D., Davis, K. and Fraser, W.D. (2016). Concentrations of persistent organic pollutants in maternal and cord blood from the maternal-infant research on environmental chemicals (MIREC) cohort study. Environmental Health, 15, 59.
• Fromme, H., Schlummer, M., Möller, A., Gruber, L., Wolz, G., Ungewiss, J., Böhmer, S., Dekant, W., Mayer, R., Liebl, B., et al. (2007). Exposure of an adult population to perfluorinated substances using duplicate diet portions and biomonitoring data. Environmental Science and Technology, 41 (22), 7928–7933.
• Fromme, H., Tittlemier, S.A., Völkel, W., Wilhelm, M., and Twardella, D. (2009). Perfluorinated compounds – Exposure assessment for the general population in western countries. International Journal of Hygiene and Environmental Health, 212 (3), 239–270.
• Harada, K., Inoue, K., Morikawa, A., Yoshinaga, T., Saito, N., and Koizumi, A. (2005). Renal clearance of perfluorooctane sulfonate and perfluorooctanoate in humans and their species-specific excretion. Environmental Research, 99 (2), 253–261.
• HC (Health Canada) (2006). State of the Science Report for a Screening Health Assessment: Perfluorooctane Sulfonate, Its Salts and Its Precursors that Contain the C8F17SO2 or C 8F17SO3 Moiety. Minister of Health, Ottawa, ON. Retrieved March 2, 2021.
• HC (Health Canada) (2014). Questions and Answers on Perfluorinated chemicals in Food. Minister of Health, Ottawa, ON. Retrieved March 2, 2021.
• HC (Health Canada) (2018a). Guidelines for Canadian Drinking Water Quality: Guideline Technical Document – Perfluorooctane Sulfonate (PFOS). Minister of Health, Ottawa, ON. Retrieved March 2, 2021.
• HC (Health Canada) (2018b). Guidelines for Canadian Drinking Water Quality: Guideline Technical Document – Perfluorooctanoic Acid (PFOA). Minister of Health, Ottawa, ON. Retrieved March 2, 2021.
• HC (Health Canada) (2019). Water Talk –Perfluoroalkylated substances in drinking water. Minister of Health, Ottawa, ON. Retrieved March 2, 2021.
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