Supporting document: Ecological state of the science report on Short-chain (C4–C7) Perfluorocarboxylic Acids (SC-PFCAs) Short-chain (C4–C7) Perfluorosulfonic Acids (SC-PFSAs) Long-chain (C9–C20) Perfluorosulfonic Acids (LC-PFSAs)

Official title: Supporting document: Ecological state of the science report on

Short-chain (C4–C7) Perfluorocarboxylic Acids (SC-PFCAs)

Short-chain (C4–C7) Perfluorosulfonic Acids (SC-PFSAs)

Long-chain (C9–C20) Perfluorosulfonic Acids (LC-PFSAs)

Information in Support of the Draft State of Per- and Polyfluoroalkyl Substances (PFAS) Report

Environment and Climate Change Canada

May 2023

Abbreviation List^{Footnote 1} 

Preface

This document contains additional information that is summarized or referenced in the Draft State of Per-and Polyfluoroalkyl Substances (PFAS) Report. Relevant data were identified up to March 2022.

In the Draft State of PFAS Report, this document is referenced as:

[ECCC] Environment and Climate Change Canada. 2023. Supporting Document: Ecological State of the Science Report on Short-chain PFCAs, Short-chain PFSAs, and Long-chain PFSAs. Gatineau (QC): Government of Canada.

In-text reference: (ECCC 2023)

The supporting working documentation for the figures in this document is available upon request by email from substances@ec.gc.ca.

1. Introduction

Since the 1950s, per- and polyfluoroalkyl substances (PFAS) have been widely used in industrial and consumer applications such as aqueous film-forming foams (AFFF) and the surface treatment of textiles, carpets, and papers where there is a need for extremely low surface energy, surface tension, and/or durable water- and oil-repellency. Their presence in the environment is due to anthropogenic activities and no natural sources exist.

For over a decade, PFOS and PFOA have attracted significant attention as contaminants of global concern. PFOS and PFOA are persistent, bioaccumulative, widespread in the global environment, found in remote areas (due to long-range transport of PFOA and PFOS, and their precursors), and can cause various adverse effects on wildlife at relevant environmental concentrations. Both substances have undergone various regulatory actions in many countries. In Canada, ecological risk assessments for perfluorooctane sulfonate (PFOS), perfluorooctanoic acid (PFOA), and long-chain perfluorocarboxylic acids (LC-PFCAs; C9–C21) conducted under the Canadian Environmental Protection Act, 1999 (CEPA) have concluded that these substances are harmful to the environment (EC, HC 2012; Environment Canada 2006; Environment Canada 2012). Consequently, PFOS, PFOA, and the LC-PFCAs, their salts, and their precursors were placed on the List of Toxic Substances (Schedule 1) under CEPA. Since 2008, Canada’s regulations prohibit the manufacture, use, sale, offer for sale, and import of PFOS and products containing PFOS, with limited exemptions. Since 2016, Canada’s regulations also prohibit the manufacture, use, sale, offer for sale, and import of PFOA and LC-PFCAs and products containing PFOA and LC-PFCAs, with limited exemptions. On May 14, 2022, Canada proposed regulations that would further restrict PFOS, PFOA, and LC-PFCAs by removing or providing time limits for most of the remaining exemptions.

Due in part to various regulatory actions worldwide (including the listing of PFOS and PFOA as Persistent Organic Pollutants under the Stockholm Convention), other perfluorinated substances (for example, SC-PFCAs, SC-PFSAs, and potentially LC-PFSAs) have been used as replacements for PFOA, PFOS, and LC-PFCAs. Initially, the shorter-chain replacement substances were thought to be alternatives with an overall lower bioaccumulation and toxicity potential on the basis of standard toxicity test results for freshwater aquatic test species such as fish, daphnia, and algae. However, it has been more recently recognized—for example, in statements from a range of experts—that these shorter chain PFAS may have impacts similar to those of PFOS and PFOA (Helsingør Statement, Scheringer et al. 2014; Madrid Statement, Blum et al. 2015; Zurich Statement 2018, Ritscher et al. 2018).

Many SC-PFCAs, SC-PFSAs, and LC-PFSAs, along with their precursors, have been detected in the Canadian environment and biota, including in the Canadian Arctic and the Great Lakes. The presence of these substances in Canada may be the result of releases from imported products or manufactured items containing these substances, which make their way into the Canadian environment. In remote environments in Canada, their presence may be due to the long-range transport (LRT) of precursors from within Canada or from international sources that may be transported and subsequently transformed to the acids.

SC-PFCAs, SC-PFSAs, and LC-PFSAs are considered to be as persistent as PFOS, PFOA, and the LC-PFCAs due to the carbon-fluorine bond, which is one of the strongest covalent bonds (about 108–120 kcal/mole). This bond results in these substances being extremely stable and generally resistant to degradation by acids, bases, oxidants, reductants, photolytic processes, microbes, and metabolic processes. Some SC-PFCAs/PFSAs and LC-PFSAs have also been shown to biomagnify in upper trophic level wildlife to a degree comparable to the substances they are meant to replace (that is, PFOS, PFOA, and LC-PFCAs). The time frame over which these substances remain in the environment is expected to be extremely long and has not been meaningfully quantified. It is often difficult to quantify the ecological risks associated with persistent and bioaccumulative substances in the environment, but they are generally acknowledged to have the potential to cause serious, irreversible impacts on wildlife populations in the long term (Environment Canada 2006; MacLeod et al. 2014). In addition, SC-PFCAs/PFSAs and LC-PFSAs may have the capacity to cause various adverse effects in wildlife similar to those of the PFAS that they replaced.

This report provides a summary of data available from the ecological-related science literature (up to March 2022) for PFAS in the following three subgroups:

1. Short-chain (C4–C7) perfluorocarboxylic acids (SC-PFCAs), their salts, and precursors
2. Short-chain (C4–C7) perfluorosulfonic acids (SC-PFSAs), their salts, and precursors
3. Long-chain (C9–C20) perfluorosulfonic acids (LC-PFSAs), their salts, and precursors

This includes data and information on environmental persistence, bioaccumulation and trophic magnification potential, mobility, Canadian environmental monitoring data, and potential for adverse effects in the environment. Most of the content of this report focuses on SC-PFCAs/PFSAs, some of which have been used as replacements for PFAS that have been subject to restrictions in Canada and/or internationally. LC-PFSAs are believed to have more limited use as substitutes for these restricted substances, and are covered to a lesser extent in this document largely due to the lack of information on them available in the literature (except for limited information on PFDS and PFNS).

This report refers to long-chain and short-chain PFAS, where long-chain substances have a carbon chain length of 8 (C8) or higher and short-chain substances have a carbon chain length of 7 (C7) or lower. PFOS and PFOA, which are both C8, are sometimes discussed separately from other long-chain PFAS in this report. Since PFOS and PFOA are well studied, data for these substances are sometimes presented for purposes of comparison. PFOS and PFOA are also discussed separately from long-chain PFAS with regard to past regulatory activities. Reports by other authors (for example, the OECD) may refer to perfluorinated alkyl sulfonates that have 6 (C6) or more fully fluorinated carbons (for example, PFHxS) as long-chain PFAS; however, the definitions of short-chain and long-chain PFAS used in this report are consistent with other Government of Canada publications.

For the purpose of this report, PFAAs refer to the perfluoroalkyl acids (for example, PFCAs and PFSAs). It is these stable forms that are referred to as the moieties of interest in this document. The abbreviations for individual PFCAs and PFSAs could represent either the acid or anionic forms of the chemicals; however, under environmental conditions, PFAAs exist predominantly in their anionic form.

2. Substance identity

By definition, PFAS contain at least one fully fluorinated methyl or methylene carbon (without any H/Cl/Br/I atom attached to it), that is, with a few noted exceptions, any chemical with at least a perfluorinated methyl group (–CF3) or a perfluorinated methylene group (–CF2–) is a PFAS (OECD 2021). The length of the fluorinated carbon chain of most of the perfluorinated substances that are monitored or detected in the environment ranges from 3 to 20 fluorinated carbons. These alkyl chains are attached to various functional groups. For example, perfluorocarboxylic acids (PFCAs) and perfluorosulfonic acids (PFSAs) are organic acids comprised of a fluorinated carbon chain terminated by a carboxylate or sulfonate functional group, respectively. The conjugate anion chemical structures for perfluorooctanoic acid (PFOA) and perfluorooctane sulfonic acid (PFOS) are provided as examples in Figure 1. A number of different configurations are possible for each perfluorinated acid, including the simple linear configuration and branched isomers.

Figure 1. Conjugate anion structures for PFOA (top) and PFOS (bottom)

In terms of composition and nomenclature, the term “perfluorinated” indicates that all hydrogen atoms directly attached to the carbon chain have been replaced by fluorine atoms. The term “poly-fluorinated” indicates that only some of the hydrogen atoms have been replaced with fluorine atoms. The expression “C#” is generally used to define the total number of fluorinated carbons present in a PFSA molecule or the total number of carbons present in a PFCA molecule (which includes both fluorinated carbons and the carbonyl group). For example, C9 PFSA has 9 carbons, all of which are fluorinated. However, C9 PFCA has 9 carbons, of which 8 are fluorinated and one is part of the carbonyl functional group.

In 2006, Canada’s Domestic Substance List (DSL, circa 1984 to 1986)^{Footnote 2} was reviewed to identify PFAS substances (including salts and precursors) included on it at the time. This review included consideration of their potential to transform to moieties of interest on the basis of expert judgement, chemical structures, and the biodegradation estimation model, CATABOL (c2004 to 2008)^{Footnote 3} (Jaworska et al. 2002; Dimitrov et al. 2004, 2007, 2011). CATABOL predicted metabolites by simulating the OECD 302C 28-day biodegradation test and was trained on the basis of MITI (Japan’s Ministry of International Trade and Industry) biodegradation test results. Although this model was designed to accommodate perfluorinated substances, some degradation products predicted by CATABOL may be of lower reliability or relevance in the environment, given the very limited perfluorinated degradation data in its training set. Appendices A, B, and C list the salts and precursors to SC-PFCAs/PFSAs and LC-PFSAs identified on the DSL on the basis of CATABOL modelling, expert judgement, and chemical structure as of 2006. Appendix D provides a list of precursors where empirical evidence is available, corroborating the modelling regarding their degradation potential. These lists are not considered exhaustive. It is also noted that other jurisdictions (for example, Organisation of Economic Co-operation and Development [OECD]; United States Environmental Protection Agency [US EPA]) have developed lists that may differ from those included in this report (OECD 2018).

Where suitable, the PFAS on the DSL were assigned to the following subgroups (denoted by boxes highlighted in grey in Figure 2):

1. PFOS (C8)
2. PFOA (C8)
3. LC (C9–C21) PFCAs
4. SC (C4–C7) PFCAs
5. SC (C4–C7) PFSAs
6. LC (C9–C20) PFSAs

Canada undertook an assessment of and identified the ecological risk for three groups of PFAS: PFOS, PFOA, and LC-PFCAs, and their salts and precursors (Environment Canada 2006, 2012; EC, HC 2012). As PFOS, PFOA, and the LC-PFCAs are subject to restrictions in Canada, this report is focused on the following three chemical sub-groupings (with data for the more heavily studied PFOS and PFOA substances included for comparative purposes):

1. SC-PFCAs (C4–C7), their salts and precursors
2. SC-PFSAs (C4–C7), their salts and precursors
3. LC-PFSAs (C9–C20), their salts and precursors

Consistent with Canada’s definitions for precursors from past ecological risk assessments for PFOS, PFOA, and LC-PFCAs (Environment Canada 2006, 2012; EC, HC 2012), the precursors for the SC-PFCAs/PFSAs and LC-PFSAs are defined as follows:

1. Precursors to the SC-PFSAs: Compounds that contain the C_nF_2n+1SO₂ moiety where 4≤n≤7
2. Precursors to the LC-PFSAs: Compounds that contain the C_nF_2n+1SO₂ moiety where 9≤n≤20
3. Precursors to the SC-PFCAs: Compounds containing the perfluoroalkyl moiety (C_nF_2n+1) where 3≤n≤6 that is directly bonded to any chemical moiety other than a fluorine, chlorine, or bromine atom

The approach to identifying precursors taken in this report is consistent with that used in past reports (EC, HC 2012; Environment Canada 2006; Environment Canada 2012), where precursors can be any substance that contains the moiety of interest and that can ultimately transform through reactions such as oxidation (for example, of precursors such as volatile alcohols in the atmosphere, or of precursors in wastewater treatment systems), metabolism, or hydrolysis to the final transformation product, that is, the moiety of interest. Additionally, to be consistent with ECCC’s past reports for PFOS, PFOA, and LC-PFCAs (EC, HC 2012; Environment Canada 2006; Environment Canada 2012), this report does not directly consider the salts or precursors on the basis of their unique identities or properties. Rather, the contribution of precursors and salts to the environmental presence of SC-PFCAs/PFSAs and LC-PFSAs is recognized as salts and precursors contribute to the total presence of SC-PFCAs/PFSAs and LC-PFSAs in the Canadian environment, including biota, following transformation or dissolution.

3. Fate and behaviour in the aquatic and terrestrial environment

3.1 General characteristics

In general, most PFAS share a combination of properties such as thermal and chemical inertness, persistence, low solubility in polar and non-polar organic solvents, high density, fluidity, compressibility, and high dielectric constants (Lehmler et al. 2001). Appendices E and F show the available empirical physical-chemical data for the SC-PFCAs/PFSAs. No empirical data were found for the LC-PFSAs, and modelling was not applicable.

Chemical structure, functional groups, pKa, partitioning coefficients, and surfactant or surface-active properties impact the fate and behaviour of perfluorinated substances in the environment and have impacts on the traditional approaches to evaluating bioaccumulation and toxicity. Some key points on general characteristics for the SC-PFCAs/PFSAs and LC-PFSAs are identified in the box below. Specific information on the physical-chemical properties for precursors are not part of the scope of this report as precursors are considered only on the basis of their contribution to the total loading of moieties of interest (SC-PFCAs/PFSAs and LC-PFSAs) to the environment.

3.1.1 Influence of length of carbon chain and functional groups

The length of the fluorinated carbon chain, conformations, and the functional group attached to the perfluorinated chain (for example, a charged moiety such as carboxylate or sulfonate) can result in different physicochemical properties that can influence the substance’s behaviour in the environment and in organisms. As a result, PFAS generally have combined properties of lipophobicity, hydrophobicity, and hydrophilicity over different portions of the molecule. In general, the functional group attached to the perfluorinated chain (for example, a charged moiety such as a carboxylate or sulfonate anion) imparts hydrophilicity to that end of the molecule (Key et al. 1997). The hydrophilic portion of the molecule can be neutral, positively charged, or negatively charged. For example, both the carboxylate and sulfonate anionic part of the molecule are considered hydrophilic, while the perfluoroalkyl portions of the molecule can be hydrophobic as well as lipophobic. The anionic functional group and the dipolar nature of the carbon-fluorine bonds of the fluorinated carbon backbone contribute to the surfactant properties, creating a hydrophobic and lipophobic molecular surface that imparts the desired properties of water/oil repellency and stain resistance (Conder et al. 2008). The resulting substances can be non-ionic, cationic, or anionic surface-active as a consequence of their amphiphilic character. Examples of neutral functional groups include fluorotelomer alcohols (–CH₂CH₂OH) and sulfonamides (–SO₃NH₂). Examples of anionic functional groups include carboxylates (–COO^-), sulfonates (–SO₃^-), and phosphates (–OPO₃^-). In cationic PFAS, the functional group can be a quaternary ammonium group (Parsons et al. 2008). Linear isomers appear to be predominant for PFAS detected in biota as they may have significantly slower elimination rates and/or may be present at higher exposure concentrations than branched isomers (Conder et al. 2008).

3.1.2 pKa (acid dissociation constant)

Using the COSMOtherm model, Wang et al. (2011a) derived pKa estimates for the neutral form of PFCAs, PFSAs, PFSiAs and PFPIAs, some branched isomers for C4 to C8 PFCAs, and FTOHs. However, Wang et al. (2011a), cautioned that these values have high and unquantifiable uncertainties due to the estimates being highly dependent on the chosen conformations of the neutral and anionic forms. For example, Wang et al. (2011a) had two predicted pKa values (2.897 and 0.897) for two different conformers of the PFOA anion. Available modelled pKa values for SC-PFSAs range from -0.58 to 0.33 (Wang et al. 2011a; Stockholm Convention 2018). Recent studies point to a pKa of between 0 and 1 for PFCAs (Wang et al. 2011a; Inoue et al. 2012) and even lower pKa values for PFSAs. Empirical pKa values for SC-PFCAs range from 0.43 to 0.7 (Appendix E). Empirical pKa values for SC-PFSAs and LC-PFSAs were not found.

3.1.3 Lipid and protein partitioning coefficients

PFAS transport into cells is likely controlled by a combination of passive diffusion and active facilitation by transporter proteins such as organic anion transporter proteins (Ng and Hungerbühler 2013). Many of these proteins, or other proteins with similar function, have been identified in both rats and fish (De Smet et al. 1998; Manera and Britti 2006), indicating that fatty acid transporter proteins may be conserved between mammals and fish and can result in similar interactions (Jones et al. 2003). However, PFAS may also display substantial interspecies variability in tissue distribution and clearance rates, as well as gender-specific differences in elimination rates (Lee and Schultz 2010; Zhang et al. 2012b; Ng and Hungerbühler 2013). For example, PFHxS has serum elimination half-lives that vary considerably between species (Sundström et al. 2012; Numata et al. 2014) and between genders within species (Hundley et al. 2006; Sundström et al. 2012). In the study by Sundström et al. (2012), the species-specific and sex-specific elimination of PFHxS was highly expressed where male and female rats were investigated in terms of serum elimination. Results showed that females eliminated PFHxS more efficiently than did male rats. Furthermore, rats and mice appeared to be more effective at eliminating PFHxS than did monkeys (Sundström et al. 2012). Dassuncao et al. (2019) demonstrated that partitioning to phospholipids and binding to proteins are both mechanisms for the bioaccumulation of LC-PFCAs in the North Atlantic pilot whale (Globicephala melas). It is therefore worthwhile to consider both protein- and lipid-partitioning for marine and terrestrial mammals and birds. Protein-partitioning (characterized by coefficients such as K_pw) can be an additional representative mechanism governing the partitioning of PFAS along with phospholipid partitioning. However, there are no wildlife protein-partitioning coefficients available for the SC-PFCAs/PFSAs and LC-PFSAs.

Available empirical protein- or membrane-partitioning coefficients are found in Table 1. Bischel et al. (2011) observed that SC-PFAS and LC-PFCAs bind at different locations on bovine serum albumin and that affinity for bovine serum albumin decreased from C8 PFCA to C12 PFCA, which is likely due to steric hindrances associated with longer, more rigid perfluoroalkyl chains. PFBS exhibited increased affinity relative to PFBA. Association constants determined for PFBS and PFPeA with bovine serum albumin are similar to those for LC-PFCAs, suggesting that the physiological implications of strong binding to albumin may be important for shorter-chain PFAS.

Abbreviations: NA, not available; Log K_mw, distribution of a chemical between membrane and water phospholipids; Log K_pw, distribution of a chemical between protein and water
^a Bischel et al. 2011
^b Xie et al. 2010; Lehmler et al. 2006
^c Droge 2019
^d Ebert et al. 2020

3.2 Persistence

The presence of fluorine instead of hydrogen on the carbon chain alters the thermal, chemical, and biological characteristics of the PFAS molecule. The carbon-fluorine bond is one of the strongest in nature (3M Company 1999), making the bond extremely stable and generally resistant to degradation by acids, bases, oxidants, reductants, photolytic processes, microbes, and metabolic processes. The strong C-F bond and dense coating of electron-rich fluorine atoms protects the carbon backbone and results in an inertness to heat and chemical reagents (Hakli et al. 2008; Colomban et al. 2014). This contributes to a high ionization potential, low polarizability, low inter- and intra-molecular interactions, and low surface tension.

A number of studies show that degradation of C4 to C7 PFCAs and C4 and C6 PFSAs does not occur under environmentally relevant conditions (Hurley et al. 2004; Hori et al. 2005; Dillert et al. 2007; Hori et al. 2008; Saez et al. 2008; Park et al. 2009; Quinete et al. 2010). As a result, SC-PFCAs/PFSAs are expected to accumulate in the environment over time. Taniyasu et al. (2013) showed that some photodegradation of PFDS occurs under high altitude conditions and prolonged exposure; however, photodegradation did not occur for PFBS, PFHxS, and PFBA. Conditions achievable in an incinerator (that is, high temperatures at 900 or 1200 K) are considered capable of destroying substances that contain a carbon-fluorine bond (Tsang et al. 1998).

3.3 Bioaccumulation

3.3.1 Summary of the available bioaccumulation metrics (BCF, BAF, BMF, and TMF) for SC-PFCAs/PFSAs and LC-PFSAs

Canada’s past ecological risk assessments for PFOS, PFOA, and most LC-PFCAs have shown that bioconcentration factors (BCFs) and bioaccumulation factors (BAFs), and even food web biomagnification factors (BMF) and trophic magnification factors (TMF), specifically in water-breathing aquatic organisms (that is, fish, daphnia) and algae, were generally indicative of a lower bioaccumulation potential (with the exception of some PFOS BCF/BAFs in certain aquatic species—see Figure 3). However, food web BMF and TMF values in air-breathing marine mammals (for example, polar bears, dolphins), terrestrial mammals (for example, wolves), and birds (for example, Arctic birds) suggest a much higher bioaccumulation potential (EC, HC 2012; Environment Canada 2006; Environment Canada 2012). Thus, in the characterization of the bioaccumulation potential of SC-PFCAs/PFSAs and LC-PFSAs in this report, multiple metrics of bioaccumulation that include BCF, BAF, TMF, and BMF in aquatic, terrestrial, and avian species are considered, when available. In addition, there may be a particular relevance to tissue-specific accumulation (for example, liver, blood, kidney) as these are often the sites of significant toxicological action for PFAS.

Aquatic organisms, marine mammals and aquatic birds

In a laboratory study, Martin et al. (2003a) showed that C5 to C7 PFCAs did not bioaccumulate in any rainbow trout tissues (BAFs <0.1). PFHxA and PFHpA could not be detected in most rainbow trout tissues despite higher exposure concentrations, whereas PFBS was detectable at the last three uptake sampling intervals and at the first sampling time of the depuration phase, which allowed an estimation of depuration but not assimilation (that is, BAF <1). The Martin et al. (2003b) laboratory study showed that PFHxS had a carcass^{Footnote 4} BCF of 9.6 in rainbow trout, whereas the Goeritz et al. (2013) laboratory study calculated a BAF of less than 0.02 and 0.18 for PFBS and PFHxS, respectively, in rainbow trout.

Available bioconcentration and bioaccumulation data were taken from literature and graphed according to chain length, functional group, and organism class (Figure 3 and Figure 4). PFOS and PFOA bioaccumulation data are included for comparative purposes. Figure 3 shows that bioconcentration or bioaccumulation levels of PFHxS in saltwater crabs, gastropods, saltwater fish, and freshwater fish are approaching Canada’s BCF or BAF regulatory numeric criteria for bioaccumulation as set out in the Persistence and Bioaccumulation Regulations of CEPA (Canada 2000). Figure 4 shows that levels of bioconcentration or bioaccumulation of PFHxA in crab, gastropods, and saltwater fish are also approaching Canada’s BCF or BAF regulatory numeric criteria for bioaccumulation (Canada 2000). These data show that read-across between chain lengths can be variable depending on the species and chain lengths chosen. For example, PFHxS has higher rates of uptake than PFOS in bivalves, and PFHxA has higher rates of uptake than either PFOA or PFHpA in freshwater fish, saltwater crab, and gastropods.

Figure 3. Available BCF/BAF (*maximum, **unknown) for PFOS and SC-PFSAs in crab (various species^a), gastropods (various species^b), bivalves (various species^c), fish (various saltwater species^d and freshwater species^e), and eel^{Footnote 5}  

^** Unclear if the study referred to mean, average, or maximum values

^a Hemigrapsus sanguineus, Sesarma pictum, Hemigrapsus penicillatus, Helice tridens tridens, Philyra pisum, and Eriocheir sinensis

^b Littorina brevicula, Monodonta labio, Umbonium thomasi, and Glossaulax didyma

^c Mytilus edulis, Mactra veneriformis, Nuttallia olivacea, and Sinonovacula constricta

^d Acanthogobius flavimanus, Sebastes schlegeli, Tridentiger obscurus, Hexagrammos otakii, and Mugil cephalus

^e Oncorhynchus mykiss, Cyprinus carpio, Misgurnus anguillicaudatus, and Salvelinus namaycush

Figure 4. Available BCF/BAF (*maximum, **unknown) for PFOA and SC-PFCAs in crab (various species^a), bivalves (various species^b), gastropods (various species^c), and fish (various saltwater species^d and freshwater species^e).^{Footnote 6} 

^** Unclear if the study referred to mean, average, or maximum values

^a Hemigrapsus sanguineus, Sesarma pictum, Hemigrapsus penicillatus, Helice tridens tridens, Philyra pisum, and Eriocheir sinensis

^b Mytilus edulis, Mactra veneriformis, Nuttallia olivacea, and Sinonovacula constricta

^c Littorina brevicula, Monodonta labio, Umbonium thomasi, and Glossaulax didyma

^d Acanthogobius flavimanus, Sebastes schlegeli, Tridentiger obscurus, Hexagrammos otakii, and Mugil cephalus

^e Oncorhynchus mykiss, Cyprinus carpio, Misgurnus anguillicaudatus, and Salvelinus namaycush

Figure 5. Comparison of BMFs for PFHxS and PFOS in various trophic levels of various food webs^{Note de bas de page 7} 

Figure 5 shows that food web BMF values for PFHxS and PFOS are greater than 1 (10⁰) in air-breathing organisms such as dolphins, harbour seals, polar bears, and birds despite moderate BCF and BAF values (<3500) in water-breathing organisms for PFHxS. This figure also shows that BMF values for PFHxS are comparable with those for PFOS. Species differences in uptake rates can also be seen. For example, harbour seals have greater uptake rates of PFHxS than any other marine mammal, with BMFs of up to 231. Boisvert et al. (2019) reported a comparison of the arithmetic means of ratios of PFAS concentrations found in polar bear liver with those in ringed seal liver; similarly, concentrations of PFAS in polar bear liver were compared with those in ringed seal blubber from East Greenland (Scores by Sound). PFHxS, PFDS, PFBA, and PFHxA showed bear to seal ratios of >1, reflecting an increase from dietary exposure.

TMF values for PFHxS were 2.2 to 5.4 for lake trout food webs in Lake Ontario and Lake Huron, Canada, and 1.8 for a harbour seal food web in the Westerschelde, The Netherlands (Van den Heuvel-Greve et al. 2009; Ren et al. 2021, 2022). In the freshwater food web of the Yadkin-Pee Dee River (North and South Carolina, United States), TMF values were 1.08 for PFBS and below 1 for PFHpA (Penland et al. 2020). PFHxS had a TMF value of 2.09 in an Antarctic food web (Gao et al. 2020a). However, in a marine food web (Qinzhou Bay, South China Sea), TMF values were below 1 for PFHxA and PFBS (Du et al. 2021). In a temperate macrotidal estuary (Gironde, France), TMF values were below 1 for PFHpA, PFHxS, and PFHpS (Munoz et al. 2017a). Simonnet-Laprade et al. (2019a,b) showed TMF values ranging from 0.65 to 8.3 for PFHxS in a freshwater riverine food web (France). The TMF values for PFHpS ranged from 0.36 to 3.7, and TMF values for PFDS ranged from 0.73 to 17.9. Simonnet-Laprade et al. (2019a) suggested that the high variability of measured TMFs may be related to different metabolic capacities between species as well as to specific exposures in certain regions of the world, and the occurrence of unidentified precursors and their enhanced biotransformation in fish compared to invertebrates. Alternative explanations could include site differences in rates of biotransformation/growth, sediment-water concentration ratios, extent of food web omnivores, and spatial concentration gradients (Mackay et al. 2016).

The studies presented in this section show that the BMF values for the SC-PFCAs and SC-PFSAs can be comparable to those for PFOA and PFOS. The studies also demonstrate that there is a challenge in using empirical or modelled BCF and BAF data for water-breathing organisms (for example, fish) as surrogate data for BMFs/TMFs in air-breathing organisms (for example, polar bears) for PFAS. As a result, when available, empirical data for food web BMFs and TMFs for air-breathing wildlife may be the best indicators for overall bioaccumulation potential in these organisms as data for water-breathing organisms may tend to underestimate overall bioaccumulation potential.

There are several explanations to account for the discrepancy in bioaccumulation between water-breathing organisms and air-breathing organisms. Traditionally, equilibrium partitioning has assumed that if uptake occurs by the same mechanism in both water-breathing organisms (for example, fish) and air-breathing organisms (for example, polar bears), then similar uptake rates would be seen (Mackay and Fraser 2000; Kelly et al. 2004). For example, Kelly et al. (2004) stated that classical organic pollutants (that is, non-polar/non-volatile substances such as PCBs) had low elimination rates to both water and air, resulting in similar bioaccumulation rates for air-breathing and water-breathing organisms. This allowed fish BCF/BAFs to be extrapolated to characterize bioaccumulation in marine mammals and birds for these classical organic pollutants. However, extrapolation of fish-derived bioaccumulation parameters to air-breathing organisms for substances such as PFAS is highly uncertain and not recommended. Gray (2002) states that lower-trophic level organisms may take up contaminants through their body surface or through their respiratory organs by diffusion. For most small organisms (for example, plankton, polychaetes, bivalves, crustaceans), the major route of intake is by respiratory surfaces. Randall et al. (1998) demonstrated that rainbow trout (Oncorhynchus mykiss) had the largest proportion of tetrachlorobenzene taken up via the gills. Randall et al. (1998) also determined that the uptake of toxicants in food plays a minor role in water-breathing animals. Gray et al. (2002) stated that at higher trophic levels, marine birds and mammals do not take up contaminants from their respiratory surfaces as they are air-breathing and the concentrations of contaminants in air are low; thus, the only route for their contaminant uptake is through food.

Many PFAAs will likely be less hydrophobic, with increased water solubility, as their chain length decreases. For water-breathing organisms, this may result in a more rapid elimination of PFAAs to the water phase (via gill exchange) and a reduction in bioaccumulation. For fish, the lamellar blood-water interface of the gills is the major route of clearance (and uptake) of non-metabolizing waterborne substances such as PFAS. However, bioaccumulation in air-breathing organisms is driven primarily by volatility (of the neutral form) rather than polarity (Ankley et al. 2021). Therefore, the non-volatile nature of PFAAs may result in relatively slow elimination to air, resulting in higher bioaccumulation in air-breathing organisms (Kelly et al. 2004). The high water solubility of PFAAs causes their escaping tendency to be relatively high from gills into water, whereas the escaping tendency of PFAA to air, across the alveolar membrane of the lung, would be relatively low because of the low vapor pressure and negative charge of PFAAs. Thus, fish gills provide an additional mode of elimination for PFAAs (that is, “gill exchange”) that birds and terrestrial and marine mammals do not possess. Furthermore, the variability in bioaccumulation potential between different species may be partially related to body size, with larger air-breathing organisms having slower rates of depuration (Ankley et al. 2021).

Additionally, the simultaneous occurrence of different PFAS in wildlife adds further complexity to their bioaccumulation. Wen et al. (2017) demonstrated the inhibitory effect of LC-PFCAs on the bioconcentration of SC-PFCAs/PFSAs. The uptake and elimination rate constants of PFBS, PFBA, PFPeA, PFHxA, and PFHpA declined in all tissues, and their BCF values decreased by between 24% and 89% in the presence of LC-PFCAs (that is, PFNA, PFDA, PFUnA, and PFDoA), PFOS, and PFOA. The inhibitory effect may be attributed to their competition with LC-PFCAs, PFOS, and PFOA for transporters and protein binding sites in zebrafish.

Terrestrial invertebrates, mammals and birds (eagle)

Zhao et al. (2013a) exposed earthworms (Eisenia fetida) to soils artificially contaminated with C6 to C12 PFCAs and C4, C6, and C8 PFSAs. Biota-to-soil accumulation factors (BSAFs) increased with perfluorinated carbon chain length and were greater for PFSAs than for PFCAs of equal perfluoroalkyl chain length. BSAFs were 0.087 g_oc/g_dw (PFHxA), 0.122 g_oc/g_dw (PFHpA), 0.048 g_oc/g_dw (PFBS), and 0.0473 g_oc/g_dw (PFHxS). Higher soil concentrations resulted in lower BSAFs. Grønnestad et al. (2019) determined that whole-body BMF values were below 1 for C4 to C7 PFCAs for earthworm (E. fetida) and bank vole (Myodes glareolus). Rich et al. (2014) exposed earthworms (E. fetida) to field-collected unspiked soils with varying levels of PFAS, including a control soil, an industrially impacted biosolids-amended soil, a municipal biosolids-amended soil, and AFFF-impacted soils. With the exception of the control soil, BAFs were above 1, and PFHxS had the highest BSAF (0.23 g_oc/g_ww) in the municipal soil. Lasier et al. (2011) determined BSAFs for Lumbriculus variegatus in sediments from the Coosa River watershed in Georgia, United States. The mean BSAF_ww values for PFBS and PFHpS were 0.3 and 2.6, respectively. The mean BSAF_ww values for PFHpA and PFHxA were <0.2 and 0.06, respectively. Lasier et al. (2011) indicated that PFHxA had little potential to bioaccumulate or biomagnify in oligochaetes but that PFHpS and PFHxS may be as bioaccumulative as PFOS, which has a mean BSAF_ww value of 0.49.

Huang et al. (2022) suggested that short-chain PFASs (for example, PFBA, PFBS, PFHxS) also possessed high biomagnification potentials in a terrestrial food chain from the Tibetan Plateau involving plants to pika to eagle accumulation. Relatively high TMFs of 5.96, 2.43 and 5.75 were measured for PFBS, PFHxS, and PFOS in the Tibetan Plateau plant−pika−eagle food chain, while eagle (muscle)/pika (whole) BMFs for PFBA, PFBS, and PFHxS were 1.42, 1.34, and 2.29, respectively (Huang et al. 2022).

3.3.2 Special considerations on the mechanism of accumulation for PFAS

The equilibrium partitioning approach (typically used in bioaccumulation models) is usually applied to understand the bioaccumulation of classical organo-halogen pollutants (for example, PCBs) that are neutral, hydrophobic, non-volatile, and slowly metabolized. However, although SC-PFCAs/PFSAs and LC-PFSAs are non-volatile, they can have combined properties of ionization, lipophobicity, hydrophobicity, and hydrophilicity over different portions of the molecule. Therefore, the bioaccumulation of SC-PFCAs/PFSAs and LC-PFSAs in the environment can be quite difficult to predict compared with classical organic pollutants.

Chemical concentrations of neutral organic chemicals are usually lipid-normalized prior to reporting environmental levels of calculation of bioaccumulation metrics. A normalization method for ionic substances that associate with protein/plasma may be more relevant but is not yet routine. Total protein normalization may also lead to confounding issues as different proteins can have varying affinities for PFCAs and PFSAs, and the expression of these proteins may display differences between species and sexes. From a physiological perspective, it is the concentration of a substance at the site of toxic action within the organism that determines whether a response is observed, regardless of the external concentration. In the case of SC-PFCAs/PFSAs and LC-PFSAs, the site of toxic action is often considered to be the liver cells or hepatocytes. Measures of bioaccumulation metrics may be used as indicators of either direct toxicity to organisms that have accumulated PFAS or indirect toxicity to organisms that consume prey containing PFAS (via food chain transfer). Thus, from a toxicological perspective, bioaccumulation metrics that are based on concentrations in individual organs, such as the liver, may be more relevant when predicting the potential for direct organ-specific toxicity (that is, liver toxicity) of PFAS. However, certain metrics (that is, BCFs and, in particular, BMFs/TMFs) that are based on concentrations in whole organisms may provide a useful measure of overall potential for food chain transfer.

An additional consideration associated with the determination of bioaccumulation of PFAS is the exclusion of the un-metabolized precursors, which can lead to an underestimation of the overall bioaccumulation potential. Precursors to PFOS or PFOA have been shown to metabolize in rodents, resulting in the formation of PFOS or PFOA. For example, Nabb et al. (2007) showed that 8:2 fluorotelomer alcohol (8:2 FTOH) can be metabolized to PFOA. Letcher et al. (2014) showed that polar bears can rapidly de-alkylate FOSA (a precursor to PFOS). Therefore, the presence and metabolic transformation of precursors in wildlife can add to the critical body burden of some perfluorinated substances in wildlife.

Traditionally, equilibrium partitioning assumes that the metabolic transformation of a substance in an organism would enable rapid elimination of the substance, thus reducing its levels in the organism (Kelly et al. 2004). However, metabolic transformation for SC-PFCAs/PFSAs and LC-PFSAs is not likely to occur in wildlife. The observed rate of depuration is generally slower than for any previously investigated surfactant in fish, which may be partially attributable to the lack of metabolism or biotransformation (Martin et al. 2003b). The relatively slow depuration half-lives, combined with observations of high blood, liver, and gall bladder concentrations, support the theory that perfluorinated substances can enter into enterohepatic recirculation in fish—the process whereby substances are continuously recycled between blood, liver, gall bladder, and intestines, and where resorption occurs via the portal vein (Martin et al. 2003b). 

4. Environmental occurrence

This section demonstrates that SC-PFCAs/PFSAs and LC-PFSAs are found in the Canadian environment despite the fact that these substances are not known to be manufactured, imported, or used in Canada as pure substances. There is currently a moderate amount of data on the presence of SC-PFCAs/PFSAs in the Canadian environment but few data on LC-PFSAs. Current concentrations of SC-PFCAs/PFSAs and LC-PFSAs may reflect the contribution of precursors and salts that have already transformed to the moiety of interest. Because the salts typically dissociate at environmentally relevant pH, the salts or acids will be in their anionic form in the environment. This report does not directly consider the potential mixture effects amongst the individual moieties of interest and their salts and precursors.

4.1 Sources of exposure

One mechanism that accounts for the presence of SC-PFCAs/PFSAs and LC-PFSAs is the release of consumer, commercial, or industrial products containing SC-PFCAs/SC-PFSAs and LC-PFSAs or their precursors in wastewater treatment systems and landfills, as well as indirectly through the land application of wastewater biosolids. Another mechanism is the long-range transport of precursors that are deposited across Canada, including in remote areas such as the Canadian Arctic. A third mechanism is the potential transformation or biotransformation (that is, metabolism) of precursors in biota (Nabb et al. 2007; Butt et al. 2010a, 2010b; Kim et al. 2012, 2014). Appendix D provides empirical data identifying substances as precursors for SC-PFCAs/PFSAs. No equivalent data were found in publicly available literature for the LC-PFSAs. It is expected that, in the absence of shifts in use patterns or regulatory actions, concentrations of SC-PFCAs/PFSAs in the environment will increase over time due to their resistance to degradation under normal environmental conditions.

4.1.1 Releases from products

The release and degradation of consumer, industrial, or commercial products that contain SC-PFCAs/SC-PFSAs/LC-PFSAs or their precursors is another mechanism that accounts for the presence of these substances in Canada’s populated areas. For example, the presence of residual unbound fluorotelomer alcohols (that is, 4:2 FTOH, 6:2 FTOH, 8:2 FTOH, and 10:2 FTOH) was identified in several commercial consumer products such as stain repellents, paints, polishes, and other coatings (Dinglasan-Panlilio and Mabury 2006). PFCAs can be formed from the atmospheric oxidation of fluorotelomer alcohols.

Another example is the use of aqueous film-forming foams (AFFF), which can be a source of SC-PFAS in the Canadian environment. AFFF products may be used at airports (for example, Moody et al. 2002), during railway incidents (for example, Munoz et al. 2017b) and at military bases (for example, Lescord et al. 2015). Between 2000 and 2015, countries including the United States, Canada, the United Kingdom, Australia, Norway, the Netherlands, Germany, and Sweden introduced regulations and guidelines to phase out and limit the use of PFOS, PFOA, and their precursors in AFFF. While manufacturers have since modified their formulations to eliminate PFOS, AFFF formulations continue to include SC-PFAS. Early alternatives to PFOS-based AFFF that contained longer chain (C8-based) fluorotelomers are being phased out by producers, which has created a shift towards shorter chain (that is, C6-, C4-, and C3-based) perfluoroalkylated chemicals (National Academies of Sciences, Engineering, and Medicine 2017). The most common and most widely used are C6-based fluorotelomer AFFF (National Academies of Sciences, Engineering, and Medicine 2017; Hatton et al. 2018).

A third example of a PFAS source is fluorinated ski waxes. Ski waxes usually contain semi-fluorinated n-alkanes and PFCAs (Chropeňová et al. 2016; Plassmann and Berger 2013). According to Nordic Ecolabelling (2018), product development of ski wax with fluorine is transitioning to the use of SC-PFAS, such as C6 PFCAs. For example, Plassmann and Berger (2013) detected C6 to C22 PFCAs in fluorinated ski waxes. In addition, ski wax was assumed to be a source of PFAS near ski resorts in Slovakia and Norway, where some SC-PFCAs/PFSAs and PFDS were found in pine needles (Chropeňová et al. 2016). Snow meltwater sampled from ski tracks in Norway showed the presence of C6 and C7 PFCAs (Langford et al. 2010 as cited in Plassmann and Berger 2013). Lastly, snow samples taken from a ski area in Sweden after a skiing competition also showed the presence of C6 to C22 PFCAs. No equivalent studies in Canada were identified.

4.1.2 Long-range transport

Long-range atmospheric transport of SC-PFCAs/PFSAs and LC-PFSAs is unlikely given their low volatility. However, the long-range transport of volatile precursors is believed to be a pathway that accounts for the presence of fluorinated acids in the Canadian Arctic. In remote areas including the Canadian Arctic, precursors are presumed to be slowly oxidized by atmospheric radical species to give fluorinated acids that can be deposited by precipitation (Waterland and Dobbs 2007). Fluorotelomer alcohols (FTOHs) (C_nF_2n+1CH₂CH₂OH) were the first molecular class to be proposed as precursors for PFCAs and subsequently detected in low concentrations in the northern hemispheric troposphere (Waterland and Dobbs 2007). Other potential precursors include fluorotelomer olefins (FTOs) and N-alkyl perfluoroalkylsulfonamidoethanols (PFSEs).

The estimated atmospheric lifetimes of FTOHs (12 to 20 days) and FTOs (8 days) permit their transport to remote regions (Waterland and Dobbs 2007). Photoreactor studies have indicated that PFSEs may contribute to the observed environmental burden of both PFSA and PFCA substances (Waterland and Dobbs 2007). N-methylperfluorobutanesulfonamidoethyl alcohol (NMeFBSE) has the potential for atmospheric transformation to PFBS through oxidation by hydroxyl radicals (D'eon et al. 2006) (see Appendix D). In addition, Pickard et al. (2018) collected and sampled a 15-m ice core representing 38 years of deposition (1977 to 2015) from the Devon Ice Cap (Nunavut, Canada). By modelling air mass transport densities and comparing temporal trends in deposition with production changes of possible sources, Pickard et al. (2018) found that continental Asia was the largest PFAS contributor impacting the Devon Ice Cap and that the deposition of PFAS was dominated by atmospheric formation from volatile precursors. PFCAs from C2 to C13 were detected on the Devon Ice Cap with concentrations ranging from 0.00321 ng/L to 0.751 ng/L. PFSAs from C4 to C8 were detected with concentrations ranging from 0.00018 ng/L to 0.391 ng/L. Pickard et al. (2020) also detected PFBA in the Mt. Oxford icefield cores at concentrations ranging from <0.04 ng/L to 1.34 ng/L and 0.003 ng/L to 1.90 ng/L. Pickard et al. (2018) detected neither PFHxS nor PFDS in the Devon Ice Cap. However, MacInnis et al. (2017) detected PFDS but not PFHxS in a snow pit on the Devon Ice Cap representing deposition for the years 1993 to 2007.

Long-range oceanic transport of the acids and their precursors has also been proposed as a potential pathway to account for the presence of fluorinated acids in the Canadian Arctic, since perfluorinated alkyl acids, their salts, and conjugate bases are water soluble and have no appreciable vapour pressure. One hypothesis for the origin of perfluorinated alkyl acids, their salts, and conjugate bases in the atmosphere (and subsequent deposition on land) is transfer from the surface ocean by sea spray, given that the surface active properties of perfluorinated alkyl acids result in their enrichment on the “surface of bursting bubbles” (Reth et al. 2011). More specifically, the water-to-air transfer of C6 to C14 PFCAs and C6, C8, and C10 PFSAs in a laboratory-scale sea spray simulator was studied by Reth et al. (2011). This study found that the sequestration of the perfluorinated alkyl acids, their salts, and conjugate bases out of bulk water to the air-water surface increased exponentially with the length of the perfluorinated alkyl chain. Thus, it is likely that oceanic transport of primary acid emissions plays a role in their transport to the Canadian Arctic. In field tests at two Norwegian coastal sites, C6 and C7 PFCAs and PFSAs were measured in air samples and were positively correlated with Na⁺ ion concentrations. This also suggests that sea spray aerosols are a source of atmospheric PFAAs in coastal areas (Sha et al. 2022).

4.2 Abiotic concentrations in Canada

4.2.1 Surface water

Short-chain PFCAs/PFSAs and LC-PFSAs have been detected in surface freshwater, rain, and snow across Canada, including the Canadian Arctic. Overall, the average concentrations were higher in surface water (which ranged from less than the limit of detection [LOD] to 277 ng/L) than in snow (0.006 ng/L to 8.9 ng/L; Meyer et al. 2011; Bhavsar et al. 2016; MacInnis et al. 2019a). Average concentrations in rain were not reported. PFHxS had the highest measured maximum concentration in surface fresh water at 49 600 ng/L, which was reported in Etobicoke Creek, Ontario, following a spill of AFFF at the Lester B. Pearson Airport (Moody et al. 2001). PFBA had the highest maximum concentrations in both rain and snow at 14 ng/L and 52 ng/L, respectively (Gewurtz et al. 2019; MacInnis et al. 2019a).

D’Agostino and Mabury (2017) measured surface freshwater concentrations from Nunavut, with PFPeA having the highest measured mean concentration at 76 ng/L. Maximum measured concentrations in rain from Nova Scotia, Ontario, British Columbia, and Quebec for SC-PFCAs ranged from 0.9 ng/L to 14 ng/L (Scott et al. 2006a,b; Gewurtz et al. 2019). In the Canadian High Arctic Circle, snowpacks from Lake Hazen had measurable concentrations of PFBA, PFPeA, PFHxA, PFHpA, and PFBS between 2013 and 2014, with the highest concentration measured for PFBA at 52 ng/L (MacInnis et al. 2019a). The snowpacks also had PFBS measurements of up to 0.4 ng/L and PFHxS measurements of up to 0.44 ng/L (MacInnis et al. 2019a). Measurements of PFAS in Lake Hazen also suggest that snowmelt contributed to surface water concentrations of PFCAs.

Concentrations of SC-PFSA/PFCA and LC-PFSAs measured in Canada between 2000 and 2020 are represented using Tukey box plots in Figure 6. Tukey box plots are interpreted as follows: the lower and upper hinges (edges) of the box represent the first and third quantiles (Q1 and Q3), which are the 25th and 75th percentiles, respectively, while the black horizontal line within the box represents the second quantile, also known as the 50th percentile (median). The distance between the 25th and 75th percentile is called the interquartile range (IQR). The lower whisker represents the lowest data that are within the Q1 − 1.5 x IQR threshold, and the upper whisker represents the highest data that are within the Q3 + 1.5 x IQR threshold. Data exceeding these thresholds appear as individual points (for example, circles, triangles, squares). However, if the minimum and maximum are within these thresholds, they represent the lower and upper whiskers, and no outliers are present.

Figure 6. Concentrations of SC-PFSAs, SC-PFCAs, and LC-PFSAs in surface water, rain and, snow from 2000 to 2020 (ng/L).^{Note de bas de page 8}   The numbers above each box represent the number of data points included.

4.2.2 Sediments

Measured concentrations of SC-PFCAs/PFSAs and LC-PFSAs in core and surface sediments were taken from literature and graphed according to chain length and functional group (Figure 7). Only sediment core samples were reported for PFNS and PFDoDS, which were detected in Lake Ontario at a mean concentration of 0.02 ng/g and 0.0019 ng/g, respectively (personal communication, email from the Aquatic Contaminants Research Division, Environment and Climate Change Canada, to the Ecological Assessment Division, Environment and Climate Change Canada, dated March 24, 2020; unreferenced). PFHxS had the highest surface sediment concentration at 96.5 ng/g dw in Lake Niapenco, Ontario (Bhavsar et al. 2016), and PFBA had the highest sediment core concentration at 19.8 ng/g dw in Lake Ontario (Codling et al. 2018).

Figure 7. Concentrations of SC-PFSAs, SC-PFCAs, and LC-PFSAs in surface sediment and sediment core (ng/g dw or ww).^{Note de bas de page 9}   The numbers above each box represent the number of data points included.

4.2.3 Soil

Measured soil concentrations of SC-PFCAs/PFSAs and LC-PFSAs were available from Cabrerizo et al. (2018) and Liu et al. (2022). Soil concentrations were graphed according to chain length and functional group (Figure 8). Soil samples were taken from Cornwallis Island and Melville Island, both located in Nunavut, as well as from firefighting training areas in airports in central and eastern Canada. PFHxS had the highest measured concentration of 203.1 ng/g dw, followed by PFHxA at 43 ng/g and PFNS at 28.9 ng/g (Liu et al. 2022). Cabrerizo et al. (2018) noted that PFBA was correlated with a few other PFCAs, suggesting that the presence of PFBA in Arctic soils may come from sources other than atmospheric transport and the transformation of PFCA precursors. Cabrerizo et al. (2018) indicated that a more likely source of PFBA was chlorofluorocarbon replacement chemicals (including hydrofluoroethers and hydrofluorocarbons that contain the C₄F₉ moiety), which are known to yield PFBA.

Figure 8. Concentrations of SC-PFSAs, SC-PFCAs, and LC-PFSAs in soil (ng/g dw).^{Note de bas de page 10} The numbers above each box represent the number of data points included.

4.2.4 Wastewater treatment plants and landfills

SC-PFCAs/PFSAs were detected in landfill leachate, urban effluent, and Arctic effluent (that is, lagoons) in Canada. Measurements of LC-PFSAs have not yet been reported. Maximum concentrations of SC-PFCAs (C4 to C7) and SC-PFSAs (C4 and C6) in landfill leachate in Ontario ranged from 0.21 µg/L to 1.3 µg/L, with PFHxS having the highest maximum concentration, followed by PFBS (Propp et al. 2021). Higher concentrations of SC-PFCAs/PFSAs have been detected in Arctic lagoon effluent in comparison with urban effluent from temperate wastewater treatment plants (WWTPs). Mean concentrations of SC-PFCAs (C4 to C7) in effluent from a WWTP in Toronto, Ontario, ranged from 2.2 ng/L to 5.3 ng/L (Scott et al. 2006a), and PFHxS had the highest maximum concentration of 7.5 ng/L. However, SC-PFCAs (C4 to C7) and SC-PFSAs (C4 and C6) were also measured in Arctic lagoon influent (below detection up to 20 ng/L) and effluent (below detection up to 18 ng/L), and PFHxA had the highest maximum concentrations (Gewurtz et al. 2020). Gewurtz et al. (2020) hypothesized that low potable water consumption, dry and hostile weather conditions, poor ventilation, small apartment size, and long use times of household items contribute to elevated indoor PFAA concentrations, which would contribute to elevated concentrations found in both influent and effluent concentrations at Arctic lagoons in comparison with temperate WWTPs.

Guerra et al. (2014) sampled influent and effluent from 13 WWTPs in Canada. Mean concentrations of SC-PFSAs/PFCAs (C4 to C7) in influent and effluent ranged from <1.0 ng/L to 453 ng/L and <1.0 ng/L to 419 ng/L, respectively. The highest measured mean concentrations were for PFHxS and were measured at a WWTP servicing 80% industrial airport waste water, as well as a small population of 3000 inhabitants. PFHxS was followed by PFHxA, and then PFPeA (Guerra et al. 2014).

4.2.5 Temporal trends

In addition to the measurement of SC-PFCAs/PFSAs and LC-PFSAs in the Canadian environment, certain temporal trends have been identified. Between 2006 and 2018, Gewurtz et al. (2019) monitored PFAS in Great Lakes precipitation. They determined that PFOA, PFNA, PFDA, and PFOS significantly decreased over the monitoring period likely due to phase-outs and regulatory actions aimed at PFOS, PFOA, and other LC-PFCAs and their precursors. Conversely, concentrations of PFBA and PFHxA appeared to increase in the later years of the study period (Gewurtz et al. 2019). Similar trends were seen in studies of beluga whales (Delphinapterus leucas) in the St. Lawrence River. Barrett et al. (2021) measured PFAS concentrations in beluga whale liver and found that PFOS concentrations were lower in samples taken during the period of 2010 to 2017 compared with samples from 2000 to 2009. SC-PFCAs (C4 to C7) were detected at higher concentrations between 2013 and 2017, whereas they were rarely detected in earlier samples (Barrett et al. 2021).

A review by Muir and Miaz (2021) found that there were significant declining annual median concentrations in the Great Lakes for total PFCAs (C7 to C12), PFOA, total PFSAs, and PFOS between 2004 and 2017, with a significant drop-off in total PFSAs noted beginning around 2010 to 2011. Conversely, there was a dramatic increase in total SC-PFCA (C4 to C6) concentrations between the periods of 2000 to 2009 and 2015 to 2019. In comparisons between these two time periods, median concentrations of PFHxA, PFHpA, and PFOA increased by 3.0-, 16-, and 1.4-fold, respectively, and a large change was also observed for PFBA (870-fold) due to non-detect levels reported from 2000 to 2009. Median PFOS was also higher from 2015 to 2019 (2.6-fold), while PFBS and PFHxS were lower (1.7- and 7.8-fold). Higher concentrations of PFOS, PFPeA, PFHxA, PFHpA, and PFOA were observed in the combined urban and nonurban-influenced sites compared with open lake sites, although the differences were less than 2-fold (Muir and Miaz 2021).

4.3 Concentrations in Canadian biota

4.3.1 Urban and industrial areas

For urban (St. Lawrence River and Great Lakes) and industrial areas (for example, landfills and industrial sites), available measured biota concentrations for SC-PFCAs/PFSAs and LC-PFSAs were obtained from literature and graphed according to chain length and functional group in Figure 9 and Figure 10.

Figure 9. Concentrations of SC-PFCAs in biota in urban areas (ng/g ww).^{Footnote 11}  These include zooplankton, invertebrates, freshwater mussel, freshwater fish, birds, and marine mammals. The numbers above each box represent the number of data points included.

Figure 10. Concentrations of SC- and LC-PFSAs in biota in urban areas (ng/g ww).^{Note de bas de page 12} These include zooplankton, invertebrates, freshwater mussel, turtle, freshwater fish, birds and marine mammals. The numbers above each box represent the number of data points included.

PFCAs and PFSAs were measured in bird eggs of European starling, black guillemot, thick-billed murre, northern fulmar, black-legged kittiwake, gulls, and bald eagle (Gebbink et al. 2011; Braune and Letcher 2013; Letcher et al. 2015; Gewurtz et al. 2016, 2018; Wu et al. 2020). PFDS had the highest maximum concentrations in European starling (Sturnus vulgaris) eggs at 295 ng/g ww, located near a landfill in Calgary, Alberta (Gewurtz et al. 2018). Mean concentrations of up to 50 ng/g ww were reported in bird eggs, with the highest concentration for PFDS found in gull eggs from colony sites in Hamilton Harbour in Lake Ontario (Gewurtz et al. 2016). Bald eagle eggs collected from the Great Lakes region had concentrations of PFCA and PFSA (C4 to C7) of up to 11 ng/g ww, with the highest concentration being for PFHxS, followed by PFHpS at 3.5 ng/g ww (Wu et al. 2020). PFPeS was only reported in one instance in bald eagle eggs in the Great Lakes region (Wu et al. 2020).

Freshwater amphipods (Gammarus or Hyallela sp.) and snapping turtles (Chelydra serpentine) were sampled from Welland River and Lake Niapenco, which are downstream of the John C. Munro International Airport in Hamilton, Ontario (De Solla et al. 2012). Whole body amphipod arithmetic mean concentrations for PFHpS were highest at 40.2 ng/g ww, followed by PFHpA at 25.1 ng/g ww and then PFPeA at 19 ng/g ww. Snapping turtle plasma contained PFHxS with arithmetic mean concentrations that ranged from <0.1 ng/g ww to 8.2 ng/g ww as well as PFDS with arithmetic mean concentrations that ranged from 0.2 ng/g ww to 7.2 ng/g ww (De Solla et al. 2012).

PFCAs and PFSAs were measured in freshwater fish, including yellow perch (Perca flavescens) from the St. Lawrence River, Quebec, and in Lake Huron. Whole body mean concentrations for PFDS were 0.65 ng/g ww upstream of the wastewater treatment plant on the St. Lawrence River, and 0.25 ng/g ww to 0.78 ng/g ww downstream of the wastewater treatment plant. In Lake Huron, PFCAs (C4 to C7) and PFSAs (C4, C6, and C10) ranged from below the detection limit (<0.059 to <0.12) to 0.33, with PFBS having the highest mean concentration, followed by PFHxS at 0.19 ng/g ww (Ren et al. 2022). PFCAs (C4 to C7) and PFSAs (C4, C6, and C10) were measured in lake trout (Salvelinus namaycush) in the Great Lakes. Whole body mean concentrations ranged from below the detection limit (<0.059) to 9.8 ng/g ww, with PFDS having the highest mean concentration in Lake Erie (Furdui et al. 2007; Ren et al. 2021, 2022).

4.3.2 Canadian Arctic

Available measured biota concentrations for SC-PFCAs/PFSAs and LC-PFSAs were obtained from literature and graphed according to chain length and functional group in Figure 11 and Figure 12. PFDS was the only LC-PFSA measured in biota.

Figure 11. Concentrations of SC-PFCAs in Arctic biota (ng/g ww).^{Footnote 13}  Arctic biota include algae, zooplankton, invertebrates, freshwater fish, saltwater fish, birds, terrestrial mammals, and marine mammals. The numbers above each box represent the number of data points included.

Figure 12. Concentrations of SC- and LC-PFSAs in Arctic biota (ng/g ww).^{Footnote 14} Arctic biota include algae, zooplankton, invertebrates, freshwater fish, saltwater fish, birds, terrestrial mammals, and marine mammals. The numbers above each box represent the number of data points included.

Braune and Letcher (2013) measured PFSAs, PFCAs, and precursor compounds in seabird eggs in the Canadian Arctic. Of the birds that were sampled (northern fulmar [Fulmarus glacialis], thick-billed murre [Uria lomvia], black-legged kittiwake [Rissa tridactyla], black guillemot [Cepphus grylle], and glaucous gulls [Larus hyperboreus]), glaucous gulls from Prince Leopold, Nunavut, had the highest mean egg concentration with PFDS measured at 1.75 ng/g ww. Mean measured egg concentrations were generally less than the limit of detection (<0.1 ng/g ww to <0.2 ng/g ww) for PFHxA and PFHpA (Braune and Letcher 2013). Braune et al. (2014) determined that mean measured liver concentrations of PFBS and PFHxS in thick-billed murres and northern fulmars ranged from ND to 0.65 ng/g ww (LOD: 0.1 ng/g ww), with the highest mean concentration found for PFBS in thick-billed murre. Liver concentrations of SC-PFCAs ranged from <0.1 ng/g ww to 0.15 ng/g ww, with PFHxA having the highest concentration at 0.15 ng/g ww.

SC-PFSAs measured in muscle and whole body of Arctic freshwater fish (Arctic char [Salvelinus alpinus]) and saltwater fish (Arctic cod [Boreogadus saida], capelin [Mallotus villosus], and Salmo sp.) were generally below detection limits (<0.03 whole body to <0.2 ng/g ww muscle; Powley et al. 2008; Kelly et al. 2009; Lescord et al. 2015). Kelly et al. (2009) measured concentrations of up to 3.5 ng/g ww and 1.6 ng/g ww PFHxA and PFHpA, respectively, in Salmo sp. in the Hudson Bay. Concentrations of PFHxA in muscle of Arctic char from Nunavut had measurable mean concentrations of 0.058 ng/g ww; however, levels in whole body were below the LOD (0.036 ng/g ww; Lescord et al. 2015).

Overall, SC- and LC-PFSAs had higher concentrations in marine mammals than in terrestrial mammals. Ringed seals (Phoca hispida) across the Canadian Arctic had mean concentrations ranging up to 2.5 ng/g ww in liver (Butt et al. 2007a, 2008). Beluga whales (Delphinapterus leucas) located near Newfoundland did not have measurable levels of PFHxS in their livers for the year 1986 (Reiner et al. 2011). However, beluga whales from Hudson Bay had maximum PFHxS and PFDS concentrations of up to 3.76 ng/g ww and 5.12 ng/g ww, respectively, in their livers from 1999 to 2004 (Kelly et al. 2009).

All SC-PFCAs, with the exception of PFHpA, were below the method detection limits (<0.025 ng/g ww to <0.075 ng/g ww lipid) for measurements in fat of polar bears located in the Hudson Bay during the years 2013 to 2014. PFHpA had a maximum geometric mean lipid concentration of 0.33 ng/g ww (Letcher et al. 2018). PFBS and PFHxS had geometric mean lipid concentrations of 0.04 ng/g ww and 8.28 ng/g ww, respectively (Letcher et al. 2018). The average concentration of PFHxS in the Northwest Territories and Nunavut in 2002 ranged from 35.9 ng/g ww to 71.4 ng/g ww (Smithwick et al. 2005b).

All SC-PFCAs and some SC-PFSAs and LC-PFSAs (C4, C6, C7, C10) were detected in moose (Alces alces) in the Northwest Territories, with PFHxS having the highest maximum concentration in liver of 0.106 ng/g ww, followed by PFBS at 0.087 ng/g ww (Larter et al. 2017). PFHxS, PFDS, and PFHpA were measured in caribou (Rangifer tarandus) from Nunavut between 2002 and 2016. The maximum liver concentrations were up to 0.6 ng/g ww for PFHxS, followed by PFDS at 0.3 ng/g ww (Roos et al. 2021).

Lescord et al. (2015) sampled Arctic char and benthic and pelagic invertebrates in Meretta Lake and Resolute Lake, Nunavut, which are downstream of the local airport. These lakes were likely affected by wastewater discharges (with little treatment) from both the airport and a military base during the years 1949 to 1998. Only PFHxS and PFHxA were analyzed. PFHxS was not detected in benthic invertebrates but was detected in juvenile and adult Arctic char muscle tissue with mean concentrations of up to 2.0 ng/g ww and 1.2 ng/g ww, respectively. PFHxA was detected in benthic invertebrates with mean concentrations of up to 0.38 ng/g ww and in juvenile/adult Arctic char muscle tissue with mean concentrations of up to 0.04 ng/g ww. The total PFAS concentration in Arctic char from Meretta and Resolute Lakes was 100 times higher compared to fish found in nearby reference lakes.

4.3.3 Release events

In June 2000, 22 000 L of AFFF was accidentally released at Lester B. Pearson International Airport (Toronto, Ontario) due to a fire alarm malfunction. The AFFF then entered Etobicoke Creek, a tributary to Lake Ontario (Moody et al. 2002). Fish sampling occurred 21 and 153 days post-incident at Etobicoke Creek upstream and downstream of the release site. With the exception of PFHpA, all analyzed SC-PFCAs and SC-PFSAs (C4 and C6) were detected in common shiner at higher maximum concentrations downstream than upstream of the release site. The source of the upstream concentrations is not known. PFBS was not detected upstream of the release site.

In August 2005, 48 000 L of AFFF entered Etobicoke Creek. The foam was used to douse an aircraft fuselage fire (Oakes et al. 2010). Oakes et al. (2010) analyzed blacknose dace (Rhinichthys atratulus) liver collected 9 and 122 days post-incident at Etobicoke Creek upstream and downstream of the release site. PFHxS, PFHxA, and PFHpA were below the limit of quantification upstream and downstream of the release site. The results by Oakes et al. (2010) suggested that the AFFF that was used likely contained telomerized polyfluorinated materials and that the use of this formulation may be attributed to the phase-out of perfluorinated acids in AFFF.

In July 2013, approximately 33 000 L of AFFF entered Lake Mégantic and the Chaudière River near the municipality of Lac-Mégantic, Quebec (Munoz et al. 2017b). Archived white sucker (Catostomus commersonii) muscle tissue collected two years prior to the incident from Lake Mégantic was used as a reference. White suckers from Lake Mégantic and the Chaudière River were sampled 1, 3, and 12 months post-incident. The reference tissue indicated that PFBA, PFPeA, PFHxS, and PFDS were already present in Lake Mégantic and the Chaudière River, but the sources were not identified. Concentrations of SC-PFCAs and SC-PFSAs (C4, C6, and C7) ranged from less than the limit of detection (LOD) to 0.37 ng/g ww post-incident and from <LOD to 0.36 ng/g ww 12 months post-incident. Maximum PFBA concentrations were lower post-incident. Conversely, maximum PFPeA, PFHxS, and PFDS concentrations increased post-incident. PFHxA, PFHpA, and PFHpS were not detected in reference samples. However, PFHxA and PFHpA were measured post-incident, and PFHpS was detected only 12 months post-incident. PFBS was not detected either pre- or post-incident. The authors stated that it was unlikely that the PFAAs detected near the accident site were from the AFFF formulation. However, the presence of SC-PFCAs could have been due to the environmental transformation of fluorotelomer-based PFAS.

PFAS are resistant to heat and chemical extremes, which makes most conventional treatment technologies ineffective for PFAS removal or destruction. Additionally, different treatment technologies are limited in their ability to be widely used and are thus limited to locations that are economically and logistically feasible. However, PFAS treatment and remediation technologies are rapidly evolving and advancing for potential application at contaminated sites.

4.4 Concentrations in marine wildlife, terrestrial wildlife, and birds worldwide

A review of available literature from 2002 to 2022^{Footnote 15} indicates that measured concentrations of SC-PFCAs/PFSAs in marine mammals, terrestrial wildlife, and birds outside of Canada are generally higher than those measured in Canada.

Antarctic gentoo penguins (Pygoscelis papua) and Adélie penguins (Pygoscelis adeliae) had mean egg concentrations of PFHpA of between 0.5 ng/g ww and 2.5 ng/g ww, whereas PFHxS was not detected for either species of penguins (Schiavone et al. 2009). Schiavone et al. (2009) indicate that Antarctic penguins feed at the top of the polar marine food chain and, due to their non-migratory and non-nomadic species breeding, tissue concentrations of PFAS are an indication of local contamination. In the dung of Papua penguins, PFBS concentrations were between 10.9 ng/g and 45.9 ng/g, between 2.17 ng/g and 3.77 ng/g for PFHxS, and between 19.9 ng/g and 237 ng/g for PFHxA. PFDS, PFBA, PFPeA, and PFHpA were below the method limit of quantification (that is, 0.8 ng/g to 6.36 ng/g for PFCAs, and 2.6 ng/g to 23.9 ng/g for PFSAs; Llorca et al. 2012).

SC-PFCAs and SC-PFSAs (C4 to C7), as well as PFDS, were in marine wildlife in various compartments such as dung, liver, brain, blood, lipid, muscle, plasma, serum, or kidney. Polar bears (Ursus maritimus) from Greenland and Svalbard, Norway, had measured concentrations of PFBS, PFHxS, PFHpS, PFDS, PFHpA, and PFHxA. The highest maximum liver concentration was measured for PFHxS at 4430 ng/g ww (Smithwick et al. 2005b). Maximum liver concentrations ranged from 0.071 ng/g to 390 ng/g ww for SC-PFCAs/PFSAs and LC-PFSAs in other marine wildlife: seals (for example, Phoca vitulina, Leptonychotes weddellii); whales (for example, Delphinapterus leucas, Orcinus orca); dolphins (for example, Tursiops truncates, Pontoporia blainvillei); porpoises (for example, Neophocaena phocaenoides, Phocoena phocoena), shark (Sphyrnidae sp.); saltwater fish (for example, Salmo salar, Oreochromis sp.); turtles (Caretta caretta) (that is, Tseng et al. 2006; Gebbink et al. 2016). Tilapia (Oreochromis sp.; sampled from a fish market in Taiwan) had the highest maximum concentration at 390 ng/g for PFPeA (Tseng et al. 2006).

In birds, SC-PFCAs/PFSAs and PFDS were measured in blood, egg, egg yolk, liver, plasma, serum, whole blood, or whole body. PFPeS and other LC-PFSAs were not analyzed. The highest maximum liver concentrations were measured for PFHxS in the grey heron (Ardea cinerea; sampled in Belgium) at 121 ng/g ww, followed by the Eurasian sparrowhawk (Accipiter nisus; sampled in Belgium) at 41 ng/g ww (Meyer et al. 2009). The highest maximum egg concentrations were found for PFPeA in cormorants (Phalacrocorax carbo) at 17.3 ng/g (Rüdel et al. 2011).

In terrestrial wildlife, PFBA, PFPeA, PFHxA, PFHpA, PFBS, PFHxS, and PFDS were measured in liver and serum. The highest maximum concentrations were found for PFHxS in the liver of wild American mink (Neovison vison) sampled in Sweden at 139 ng/g ww (Persson et al. 2013), followed by liver measurements in the wild American mink sampled in the United States at 85 ng/g ww (Kannan et al. 2002a). In blood serum, the average PFHxS concentrations detected in captive African lion (Panthera leo) and Bengal tiger (Panthera tigris tigris) were 0.091 ng/ml and 0.164 ng/ml, respectively. PFBS, PFHpA, and PFHxA were below the limits of quantification (that is, 0.05 ng/ml to 0.25 ng/ml) for both the Bengal tiger and the African lion (Li et al. 2008). The Chinese alligator (Alligator sinensis) had maximum concentrations of PFBA, PFPeA, PFHpA, PFBS, and PFHxS that ranged from 0.03 ng/mL to 1.5 ng/mL, with PFHxS having the highest maximum concentration (Wang et al. 2013a).

5. Toxicity

5.1 Acute and chronic toxicity

Available toxicity data for freshwater aquatic species for SC-PFCAs/PFSAs have endpoint values ranging from 32 mg/L to 20 250 mg/L (Table 2). No data were found for the LC-PFSAs.

Abbreviations: EC₅₀, the concentration of a substance that is estimated to cause some effect on 50% of the test organisms; IC₅₀, the concentration of a substance that is estimated to cause some inhibition on 50% of the test organisms; LC₅₀, median lethal concentration; LOEC, lowest observed effect concentration; NOEC, no observed effect concentration; LOAEL, lowest observed adverse effect level

Perfluorinated substances are persistent and bioaccumulate preferentially in air-breathing marine mammals, terrestrial mammals, and birds (see section 3.0). It is expected that these substances would have greater potential for exposure and adverse effects in air-breathing organisms due to their greater bioaccumulation potential. New approach methodology endpoints such as multi-generational effects and endocrine-related effects, along with consideration of cumulative effects if possible, could help in characterizing the potential toxicity for SC-PFCAs/PFSAs and LC-PFSAs in air-breathing organisms.

Another potential method is the use of data from standard mammalian laboratory studies (for example, rat) as surrogates for toxicity in wildlife species. However, caution is needed when extrapolating from mammalian data to certain wildlife species. For example, Letcher et al. (2014) examined the in vitro hepatic metabolism of a fluoroalkyl sulfonamide precursor of PFOS (that is, N-EtFOSA) for the polar bear (Ursus maritimus), beluga whale (Delphinapterus leucas), ringed seal (Pusa hispida), and laboratory rat (Rattus rattus). The authors expected that the in vitro incubation parameters for the polar bear, seal, and whale would be equivalent to those for the rat, given that they are all mammalian species. On the contrary, results showed that the extent of in vitro depletion of N-EtFOSA was equivalent for rat and polar bear microsomes (that is, >95%); however, the extent of in vitro depletion was lower for ringed seals (that is, 65%), and there was no significant depletion for the beluga whale in comparison to the rat. As a result, Letcher et al. (2014) indicated that interpretive caution must be exercised when comparing the absolute quantitative differences in the degree of metabolite depletion among different species. Nabb et al. (2007) showed that the clearance rates for 8:2 FTOH (a precursor to PFOA) in liver microsomes and cytosol differed among species, with rat > mouse > human > rainbow trout. Overall, the results indicated that 8:2 FTOH is extensively metabolized in rats and mice and, to a lesser extent, in humans and rainbow trout.

5.2 Mode of action

The toxicokinetics of perfluorinated substances have been studied in mammals (for example, bovine and human serum albumin, rats), where it was observed that perfluorinated substances bind strongly to plasma albumin and that transport into cells is likely controlled by a combination of passive diffusion and active facilitation by transporter proteins such as organic anion transporter proteins. These proteins are renal transporters that facilitate the reabsorption of organic anions from urine back to blood and are thought to be responsible for the long metabolic half-life of some perfluorinated substances (Ng and Hungerbühler 2013, 2014). It was also observed that these proteins are more highly expressed in male rats than in females, which may explain the gender differences in clearance rates for PFOA. Perfluorinated substances have also been observed to bind to cytosolic fatty acid binding proteins, which are ubiquitous in a number of cell types and serve as a sink in some tissues.

Perfluorinated substances affect liver function, including lipid and lipoprotein metabolism. These substances can alter lipid metabolism through peroxisome proliferation, alter xenobiotic metabolism by activating the cytochrome CYP-450 system, and alter serum cholesterol levels by inducing or repressing key genes (Hickey et al. 2009). A review by Sonne (2010) indicated that CYP-450 biotransformation of some PFAS in polar bears can result in highly toxic metabolites that are retained in blood plasma and various tissues. Some PFAS may also induce endocrine disruption indirectly through metabolism of endogenous hormones or vitamins, and CYP-450 activity may act as a biomarker for exposure to perfluorinated substances (Sonne 2010).

In addition, perfluorinated substances are known to activate the peroxisome proliferating receptor (PPAR-α) in wildlife (for example, Lake Baikal seals, Ishibashi et al. 2008b; whales and dolphins, Kurtz et al. 2019; polar bears, Routti et al. 2019b), which increases the abundance of hepatic peroxisomes and induces peroxisomal and mitochondrial enzymes involved with β-oxidation, cytochrome P450 (CYP-450) fatty acid ω-oxidation, and cholesterol homeostasis via ligand-dependent activation of the hepatic PPAR-α (Holden and Tugwood 1999; Bosgra et al. 2005). Prolonged exposure to peroxisome proliferators can result in hepatocarcinogenesis, although marked differences in susceptibility between species have been observed. Although these studies do not specifically analyze SC-PFCAs/PFSAs or LC-PFSAs, it is expected that the mode of action is applicable to all homologues of PFCAs and PFSAs.

5.3 Multi-generational effects

There is a concern for substances that have the potential to harm organisms at low concentrations and/or that have modes of toxic action beyond narcosis (for example, endocrine-related effects). The long-term ecological effects of highly persistent and bioaccumulative substances cannot be accurately predicted; however, these types of substances are acknowledged as having the potential to cause serious, irreversible impacts. Persistent substances, such as perfluorinated substances, remain in the environment for long periods of time, which increases the probability and duration of exposure as well as the potential for long-range transport, resulting in regional or global contamination. A substance that does not naturally occur in the environment may also have an increased potential to cause harm, as organisms may not have evolved specific strategies for mitigating exposures and effects (Macleod et al. 2014). Multi-generational toxicity studies can be used as a tool to determine the long-term ecological effects of substances such as SC-PFCAs/PFSAs and LC-PFSAs as these substances are extremely persistent (see section 4.2). Some studies indicate that these substances can be bioaccumulative in air-breathing organisms (see section 3.3.1). Table 3 presents the studies currently available on multi-generational effects.

5.4 Endocrine-related effects

In some instances, substances may cause adverse effects in living organisms by interfering with the normal functioning of the endocrine system. Adverse effects may include delayed or impaired growth, altered intellectual and sexual development, increased susceptibility to certain cancers, disturbances in immune and nervous system function, and a lowered ability to reproduce or produce healthy offspring. Various studies indicate that endocrine-active substances may have the most pronounced effects during early developmental periods (such as prenatal and early postnatal development), when hormone-sensitive systems are developing (HC 2022). Available laboratory studies related to endocrine-related effects (sub-organismal and organism-level) for SC-PFCAs/PFSAs and one LC-PFSA (that is, PFDS) are provided in Table 4 and Table 5.

The PPAR-α mechanism has been proposed as a mode of action for liver toxicity in mammals. However, there is only one study (Ishibashi et al. 2011) that has examined the relative potency of the PPAR-α mechanism in a marine mammal, that is, the Lake Baikal seals. Other studies examined a variety of endocrine responses such as thyroid response, vitellogenin expression, or chemosensitivity. However, it is unclear whether these sub-organismal responses can be directly linked to an actual gross effect on the whole organism or collectively at the population level. Additionally, the relative potency amongst the various pathways (for example, thyroid, estrogen, chemosensitivity, or PPAR-α) is unclear, and the relevance of these responses amongst the various pathways and various species is also unclear.

Read-across for perfluorinated substances on the basis of chain length can be difficult due to the inconsistent responses for the same endpoint amongst various species and taxonomic groups. Ishibashi et al. (2011) found that, relative to PFOA, the Lake Baikal seal PPAR-α was more strongly activated by PFHxS but not by PFBS. In addition, relative to PFOA, PFPeA had greater activation than PFHxA. No studies comparing relative activation of other SC-PFCAs and SC-PFSAs with PFOS appear to have been conducted. Additionally, a single species may not be representative within a taxonomic group. For example, there was no impact on cell cytotoxicity from PFHxA in white leghorn chicken (Cassone et al. 2012b), but there was an impact on cell cytotoxicity from PFHxA in herring gulls (Vongphachan et al. 2011). Houde et al. (2013) showed that there was no correlation between PFDS and plasma vitellogenin in rainbow trout, but PFDS may be correlated to plasma vitellogenin levels in northern pike. Other laboratory-based endocrine-related studies have shown observed effects for SC-PFCA/PFSAs mixtures and/or PFDS in various wildlife species, including top predators (that is, Nobels et al. 2010; Gorrochategui et al. 2016; Annunziato et al. 2019; Menger et al. 2020; Omagamre et al. 2020; Wang et al. 2020; Rericha et al. 2021; Solé et al. 2021). Some illustrative examples are described in Table 4.

5.5 Cumulative effects

While the range of different PFAS examined in many studies has historically been relatively limited, studies have increasingly noted broad occurrence and co-exposure to a range of PFAS. With respect to field-based wildlife studies, it is difficult to uniquely distinguish effects caused by exposure to SC-PFCAs/PFSAs and LC-PFSAs, given that exposures from other PFAS (for example, PFOS or PFOA) or other contaminants cannot be excluded (Knudsen et al. 2007; Letcher et al. 2010; Liu et al. 2018b; Routti et al. 2019a; Hansen et al. 2020). PFAS (including related substances) are also often summed as a group and statistically correlated with the effect observed.

For example, a mixture of PFAS (i.e, PFHxS, PFOS, PFOA, and C9 to C14 PFCAs) was associated with the disruption of thyroid hormone homeostasis in polar bears (Ursus maritimus) from the Barents Sea (Bourgeon et al. 2017). However, these polar bears also had concentrations of 38 organochlorine substances, including polychlorinated biphenyls (PCBs), polybrominated diphenyl ethers (PBDEs), and 10 phenolic substances as well as 8 other PFAS that may also have contributed to the effect observed. Liu et al. (2018a) analyzed pooled polar bear serum from the Hudson Bay and Beaufort Sea subpopulations in the Canadian Arctic and found 5 classes of PCB metabolites, 4 classes of perfluorinated sulfonates, and 4 classes of other polychlorinated substances (that is, chlorinated aromatics, tetrachloro aromatic sulfate, heptachlorinated hydroxylated nitroaromatics, and hexachlorinated substances). Knudsen et al. (2007) measured insecticides (for example, mirex), PFAS, hexachlorocyclohexanes, toxaphenes, dioxins, furans, PCBs, brominated substances, endosulfans, and mercury in northern fulmars (Fulmarus glacialis) from the Barents Sea. Gao et al. (2020b) measured 3108 substances (388 contaminants and 2720 metabolites) in wild crucian carp (Carassius auratus) from Taihu Lake, China. These studies highlight that field-based mixture studies can be confounding when determining whether a singular substance or group of substances is affecting the health and condition of the wildlife species under investigation. Thus, a direct cause-effect correlation is difficult as statistical correlations do not by themselves imply causal relationships.

Additionally, concentrations of PFOS, PFOA, or long-chain (C9 and longer) PFCAs are usually found at higher levels than the SC-PFCAs/PFSAs due to legacy environmental contamination, which suggests that PFOS and PFOA can have greater contributions to the effect seen than any other substance. For example, Eggers Pedersen et al. (2015) indicated that the average brain total PFSA (ΣPFSA) concentration was 29 ng/g ww where PFOS accounted for 91% of the total concentration, and that the average ΣPFCA concentration was 99 ng/g ww where 3 LC-PFCAs combined accounted for 79% of the total concentration. As a result of their comparatively lower tissue concentrations, SC-PFCAs/PFSAs may appear to contribute little to observed effects. However, field studies are only indicative of current (and not potential) effects, and it is anticipated that SC-PFCAs/PFSAs will comprise an increasingly large proportion of PFAS in the environment due to the increased use of short-chain PFAS in place of those that have been regulated. Furthermore, it is possible that tissue concentrations may increase beyond comparable increases in environmental concentrations if observed inhibitory effects of LC-PFCAs on the uptake of SC-PFCAs/PFSAs are reduced (Wen et al. 2017; see section 3.3.1). Future field studies will likely be better positioned to assess the relative contributions of SC-PFCAs and SC-PFSAs to overall toxicity.

Read-across for perfluorinated substances is also difficult when dealing with mixtures of PFAS due to species and sex differences in effects. For example, the ƩPFAS is correlated with eggshell thinning in great tits (Groffen et al. 2019) but not in ivory gulls (Miljeteig et al. 2012).

Recognizing the associated uncertainty, several field-based wildlife studies have shown statistical correlations with observed effects for SC-PFCA/PFSAs mixtures and PFDS in various wildlife species, including top predators (that is, Grønnestad et al. 2018; Guillette et al. 2020; Hansen et al. 2020; Parolini et al. 2020; Persson and Magnusson 2015; Sun et al. 2020; Sun et al. 2021; Tartu et al. 2014). Some illustrative examples are described in Table 5.

6. Summary

Summaries of the report’s key findings are provided below.

Persistence and bioaccumulation

Owing to the extremely high strength of the C-F bond, which imparts high stability to SC-PFCAs/PFSAs and LC-PFSAs, it is expected that these substances will remain in the environment and within certain biota for a very long time. This is supported by a number of studies that show that SC-PFCAs/PFSAs do not degrade under environmentally relevant conditions. As a result, these substances are expected to persist in the environment and wildlife.

Empirical freshwater aquatic organism BCF/BAF data cannot be used alone to reliably predict the food web bioaccumulation for SC-PFCAs/PFSAs and LC-PFSAs. Results for typically tested model organisms (for example, fish) may underestimate the food web bioaccumulation potential. For SC-PFCAs/PFSAs and LC-PFSAs, air-breathing marine mammals, terrestrial mammals, and birds may have higher food web biomagnification and trophic magnification potential in comparison with the water-breathing organisms (such as fish) that are typically considered in bioaccumulation modelling. In general, it has been observed that species differences can result in inconsistent rates of bioaccumulation, making extrapolations amongst species and between chain lengths difficult. However, despite these difficulties, evidence from multiple studies suggests that BMF values for SC-PFCAs and SC-PFSAs can be comparable to those of PFOA and PFOS. Additionally, there are sufficient food web biomagnification data available for PFHxS to provide an early indication of that substance’s overall bioaccumulation potential in marine mammals (for example, polar bears and dolphins) and birds.

Evidence that a substance is persistent and bioaccumulative may itself be a significant indication of its potential to cause environmental harm. Persistent substances remain in the environment for a very long time, which increases their probability, magnitude, and duration of exposure to wildlife. Persistent substances that are subject to long-range transport can result in regional or global contamination. Consequently, releases of SC-PFCAs/PFSAs and LC-PFSAs can lead to elevated concentrations in organisms across wide areas over a long period of time. These persistent and bioaccumulative substances also can biomagnify through the food chain, resulting in increased internal concentrations for top predators (Environment Canada 2006). As a result of their broad co-occurrence in the environment, many of these persistent and bioaccumulative substances may be present simultaneously in the tissues of organisms, increasing the likelihood and potential severity of harm.

Occurrence of SC-PFCAs/PFSAs and LC-PFSA in Canada

Some measured concentration data are available for most SC-PFCAs/PFSAs in most media within Canada. There are some monitoring data for water, snow, rain, sediment, landfill leachate, WWTP influent/effluent, and biota, although there is a lack of data for soils.

As might be expected given their extreme persistence, mobility, and long-range transport potential, SC-PFCAs/PFSAs and LC-PFSAs have been detected in many media and locations throughout Canada. Overall, average concentrations were higher in surface water (ranging from less than the LOD to 277 ng/L) than in snow (0.006 ng/L to 8.9 ng/L). PFHxS had the highest measured concentration in surface fresh water (49 600 ng/L). PFBA had the highest maximum concentrations in both rain and snow at 14 ng/L and 52 ng/L, respectively. PFHxS (96.5 ng/g dw) and PFBA (19.8 ng/g) had the highest measured maximum concentrations in Canadian sediment. SC-PFCAs/PFSAs have also been detected in landfill leachate, urban WWTPs, and Arctic lagoons in Canada.

Furthermore, SC-PFCAs/PFSAs and LC-PFSAs have also been detected in various Canadian biota, including those from urban and industrial areas, the Canadian Arctic, and in close proximity to release events (that is, from AFFF).

Although there is a growing amount of data demonstrating the broad presence of SC-PFCAs/PFSAs in the Canadian environment, there are few data available on the LC-PFSAs, with the exception of PFDS. It has been noted that current concentrations of SC-PFCAs/PFSAs and LC-PFSAs can reflect the contribution of precursors and salts that have already transformed or dissociated to the moiety of interest.

Ecotoxicity

Median toxicity values for acute and chronic toxicity tests for SC-PFCAs and select SC-PFSAs (PFBS and PFHxS) in freshwater aquatic organisms range from 32 mg/L to 20 250 mg/L. There is no information on acute or chronic toxicity data for the LC-PFSAs, although some endocrine-related effects data are available for PFDS. However, the acute and chronic toxicities in freshwater aquatic organisms likely do not reflect the toxicity and exposure potential in air-breathing marine mammals, terrestrial mammals, and avian species. SC-PFCAs/PFSAs and some LC-PFSAs demonstrate greater food web bioaccumulation in air-breathing marine mammals and avian species compared to freshwater aquatic organisms. The higher potential for food web bioaccumulation coupled with the persistent nature of these substances increases the probability and the duration of exposure and, ultimately, the probability of reaching internal toxicity thresholds. A number of multi-generational studies have demonstrated lethal and sub-lethal effects at much lower concentrations than those observed for acute and chronic toxicity tests. Various endocrine-related studies have also identified effects at concentrations orders of magnitude lower than those observed with traditional toxicity tests.

Due to the large number of PFAS that may be used and the demonstration of broad environmental occurrence and co-exposure to PFAS, it is expected that cumulative toxicity as a result of exposure to various PFAS chain lengths and functional groups will increase the ecological impact of SC-PFCAs/PFSAs and LC-PFSAs in Canada. Because of legacy concentrations of PFOS, PFOA, and LC-PFCAs, current field studies are limited in their ability to assess contributions of SC-PFCAs/PFSAs and LC-PFSAs to cumulative effects; however, future studies should be better positioned to evaluate the relative contributions of these substances as a result of changes in use patterns.

Overall conclusions/summary

Despite the fact that substance-specific information is lacking for many of these PFAS, what is known about their persistence, mobility, bioaccumulation, and toxicity profiles on the basis of the empirical information presented throughout this report suggests that these SC-PFCAs/PFSAs and LC-PFSAs may share similar ecological concerns with PFAS that have been previously assessed and regulated. Given their extreme persistence properties, it is expected that these substances will remain and accumulate in the environment once released. Consequently, it is expected that the potential for adverse effects resulting from continued exposure to SC-PFCAs, SC-PFSAs, and LC-PFSAs will increase with increasing environmental loads.

References

3M Company. 1999. The science of organic fluorochemistry. US EPA. OPPT-2002-0043-0006.

Ahrens L, Siebert U, Ebinghaus R. 2009. Temporal trends of polyfluoroalkyl compounds in harbour seals (Phoca vitulina) from the German Bight, 1999–2008. Chemosphere. 76(2):151-158.

Androulakakis A, Alygizakis N, Gkotsis G, Nika M-C, Nikolopoulou V, Bizani E, Chadwick E, Cincinelli A, Claßen D, Danielsson S, et al. 2022. Determination of 56 per- and polyfluoroalkyl substances in top predators and their prey from Northern Europe by LC-MS/MS. Chemosphere. 287(Pt 2):131775.

Ankley GT, Cureton P, Hoke RA, Houde M, Kumar A, Kurias J, Lanno R, McCarthy C, Newsted J, Salice CJ, et al. 2021. Assessing the ecological risks of per-and polyfluoroalkyl substances: Current state-of-the-science and a proposed path forward. Environ Toxicol Chem. 40(3):564-605.

Annunziato KM, Jantzen CE, Gronske MC, Cooper KR. 2019. Subtle morphometric, behavioural and gene expression effects in larval zebrafish exposed to PFHxA, PFHxS and 6:2 FTOH. Aquat Toxicol. 208:126-137.

Arakaki A, Ishii Y, Tokuhisa T, Murata S, Sato K, Sonoi T, Tatsu H, Matsunaga. 2010. Microbial biodegradation of a novel fluorotelomer alcohol, 1H, 1H, 2H, 2H, 8H, 8H-perfluorododecanol, yields short fluorinated acids. Appl Microbiol Biotechnol. 88(5):1193-1203.

Armitage JM, Arnot JA, Wania F. 2012. Potential role of phospholipids in determining the internal tissue distribution of perfluoroalkyl acids in biota. Environ Sci Technol. 46(22):12285-12286.

Awad E, Zhang X, Bhavsar SP, Petro S, Crozier PW, Reiner EJ, Fletcher R, Tittlemier SA, Braekevelt E. 2011. Long-term environmental fate of perfluorinated compounds after accidental release at Toronto airport. Environ Sci Technol. 45(19):8081-8089.

Bao J, Liu W, Liu L, Jin Y, Ran X, Zhang Z. 2010. Perfluorinated compounds in urban river sediments from Guangzhou and Shanghai of China. Chemosphere. 80(2):123-130.

Bao J, Liu W, Liu L, Jin Y, Dai J, Ran X, Zhang Z, Tsuda S. 2011. Perfluorinated compounds in the environment and the blood of residents living near fluorochemical plants in Fuxin, China. Environ Sci Technol. 45(19):8075-8080.

Barmentlo SH, Stel JM, van Doorn M, Eschauzier C, de Voogt P, Kraak MHS. 2015. Acute and chronic toxicity of short chained perfluoroalkyl substances to Daphnia magna. Environ Pollut. 198:47-53.

Barrett H, Du X, Houde M, Lair S, Verreault J, Peng H. 2021. Suspect and nontarget screening revealed class-specific temporal trends (2000–2017) of poly- and perfluoroalkyl substances in St. Lawrence beluga whales. Environ Sci Technol. 55(3):1659-1671.

Bayat S, Geiser F, Kristiansen P, Wilson SC. 2014. Organic contaminants in bats: Trends and new issues. Environ Int. 63:40-52.

Bengtson Nash S, Rintoul SR, Kawaguchi S, Staniland I, van den Hoff J, Tierney M, Bossi R. 2010. Perfluorinated compounds in the Antarctic region: Ocean circulation provides prolonged protection from distant sources. Environ Pollut. 158(9):2985-2991.

Benninghoff AD, Bisson WH, Koch DC, Ehresman DJ, Kolluri SK, Williams DE. 2011. Estrogen-Like Activity of Perfluoroalkyl Acids In Vivo and Interaction with Human and Rainbow Trout Estrogen Receptors In Vitro. Toxicol Sci. 120(1): 42–58.

Berger U, Haukås M. 2005. Validation of a screening method based on liquid chromatography coupled to high-resolution mass spectrometry for analysis of perfluoroalkylated substances in biota. J Chromatogr A. 1081(2):210-217.

Bhavsar SP, Fowler C, Day S, Petro S, Gandhi N, Gewurtz SB, Hao C, Zhao X, Drouillard KG, Morse D. 2016. High levels, partitioning and fish consumption based water guidelines of perfluoroalkyl acids downstream of a former firefighting training facility in Canada. Environ Int. 94:415-423.

Bhhatarai B, Gramatica P. 2011. Prediction of aqueous solubility, vapor pressure and critical micelle concentration for aquatic partitioning of perfluorinated chemicals. Environ Sci Technol. 45(19):8120-8128.

Bischel HN, MacManus-Spencer LA, Zhang C, Luthy RG. 2011. Strong associations of short-chain perfluoroalkyl acids with serum albumin and investigation of binding mechanisms. Environ Toxicol Chem. 30(11):2423-2430.

Blum A, Balan SA, Scheringer M, Trier X, Goldenman G, Cousins IT, Diamond M, Fletcher T, Higgins C, Lindeman AE, et al. 2015. The Madrid Statement on poly-and perfluoroalkyl substances (PFASs). Environ Health Perspect. 123(5):A107-A111.

Boisvert G, Sonne C, Rigét FF, Dietz R, Letcher RJ. 2019. Bioccumulation and biomagnification of perfluoroalkyl acids and precursors in East Greenland polar bears and their ringed seal prey. Environ Pollut. 252(Pt B):1335-1343.

Bosgra S, Mennes W. Willem S. 2005. Proceedings in uncovering the mechanism behind peroxisome proliferator-induced hepatocarcinogenesis. Toxicology. 206(3):309-323.

Bossi R, Riget FF, Dietz R. 2005a. Temporal and spatial trends of perfluorinated compounds in ringed seal (Phoca hispida) from Greenland. Environ Sci Technol. 39(19):7416-7422.

Bossi R, Riget FF, Dietz R, Sonne C, Fauser P, Dam M, Vorkamp K. 2005b. Preliminary screening of perfluoroctane sulfonate (PFOS) and other fluorochemicals in fish, birds and marine mammals from Greenland and the Faroe Islands. Environ Pollut. 136(2):323-329.

Bost PC, Strynar MJ, Reiner JL, Zweigenbaum JA, Secoura PL, Lindstrom AB, Dye JA. 2016. U.S. domestic cats as sentinels for perfluoroalkyl substances: Possible linkages with housing, obesity, and disease. Environ Res. 151:145-153.

Boudreau TB, Janutka R, Solomon KR, Sibley PK, Muir DCG, Mabury SA. 2002a. Acute toxicity of perfluorinated surfactants on freshwater primary and secondary producers in laboratory studies [master’s thesis]. Guelph (ON): University of Guelph, Department of Environmental Biology. p. 134.

Boudreau TM. 2002b. Toxicology of perfluorinated organic acids to selected freshwater organisms under laboratory and field conditions [PDF] [master’s thesis]. Guelph (ON): University of Guelph. [accessed 2022 Mar 1].

Bourgeon S, Riemer AK, Tartu S, Aars J, Polder A, Jenssen BM, Routti H. 2017. Potentiation of ecological factors on the disruption of thyroid hormones by organo-halogenated contaminants in female polar bears (Ursus maritimus) from the Barents Sea. Environ Res. 158:94-104.

Braune BM, Letcher RJ. 2013. Perfluorinated sulfonate and carboxylate compounds in eggs of seabirds breeding in the Canadian Arctic: Temporal trends (1975–2011) and interspecies comparison. Environ Sci Technol. 47(1):616-624.

Braune BM, Gaston AJ, Letcher RJ, Grant Gilchrist H, Mallory ML, Provencher JF. 2014. A geographical comparison of chlorinated, brominated and fluorinated compounds in seabirds breeding in the eastern Canadian Arctic. Environ Res. 134:46-56.

Butt CM, Muir DCG, Stirling I, Kwan M, Mabury SA. 2007a. Rapid response of Arctic ringed seals to changes in perfluoroalkyl production. Environ Sci Technol. 41(1):42-49.

Butt CM, Mabury SA, Muir DCG, Braune BM. 2007b. Prevalence of long-chained perfluorinated carboxylates in seabirds from the Canadian Arctic between 1975 and 2004. Environ Sci Technol. 41(10):3521-3528.

Butt CM, Mabury SA, Kwan M, Wang X, Muir DCG. 2008. Spatial trends of perfluoroalkyl compounds in ringed seals (Phoca hispida) from the Canadian Arctic. Environ Toxicol Chem. 27(3):542-553.

Butt CM, Muir DCG, Mabury SA. 2010a. Elucidating the pathways of poly- and perfluorinated acid formation in rainbow trout. Environ Sci Technol. 44(13):4973-4980.

Butt CM, Muir DCG, Mabury SA. 2010b. Biotransformation of the 8:2 fluorotelomer acrylate in rainbow trout.1. In vivo dietary exposure. Environ Toxicol Chem. 29(12):2726-2735.

Bytingsvik J, van Leeuwen SPJ, Hamers T, Swart K, Aars J, Lie E, Nilsen EME, Wiig Ø, Derocher AE, Jenssen BM. 2012. Perfluoroalkyl substances in polar bear mother-cub pairs: A comparative study based on plasma levels from 1998 and 2008. Environ Int. 49:92-99.

Canada. 1995. Toxic Substances Management Policy [PDF] [reprinted 2004]. Ottawa (ON): Her Majesty the Queen in Right of Canada (Environment Canada).

Canada. 1999. Canadian Environmental Protection Act, 1999. S.C., 1999, c. 33, Canada Gazette. Part III, vol. 22, no. 3.

Canada. 2000. Canadian Environmental Protection Act, 1999: Persistence and Bioaccumulation Regulations, P.C. 2000–348, 23 March 2000, SOR/2000-107, Canada Gazette, Part II, vol. 134, no. 7, p. 607-612.

Cabrerizo A, Muir DCG, De Silva AO, Wang X, Lamoureux SF, Lafreniere MJ. 2018. Legacy and emerging persistent organic pollutants (POPs) in terrestrial compartments in the High Arctic: Sorption and secondary sources. Environ Sci Technol. 52(24):14187-14197.

Cassone CG, Taylor JJ, O’Brien JM, Williams A, Yauk CL, Crump D, Kennedy SW. 2012a. Transcriptional profiles in the cerebral hemisphere of chicken embryos following in ovo perfluorohexane sulfonate exposure. Toxicol Sci. 129(2):380-391.

Cassone CG, Vongphachan V, Chiu S, Williams KL, Letcher RJ, Pelletier E, Crump D, Kennedy SW. 2012b. In ovo effects of perfluorohexane sulfonate and perfluorohexanoate on pipping success, development, mRNA expression, and thyroid hormone levels in chicken embryos. Toxicol Sci. 127(1):216-224.

Chen L, Lam JCW, Hu C, Tsui MMP, Wang Q, Giesy JP, Lam PKS. 2018. Perfluorobutanesulfonate exposure causes durable and transgenerational dysbiosis of gut microbiota in marine medaka. Environ Sci Technol Lett. 5(12):731-738.

Chen L, Lam JCW, Hu C, Tsui MMP, Lam PKS, Zhou B. 2019. Perfluorobutanesulfonate exposure skews sex ratio in fish and transgenerationally impairs reproduction. Environ Sci Technol. 53(14):8389-8397.

Chowdhury MI, Sana T, Panneerselvan L, Dharmarajan R, Megharaj M. 2021. Acute toxicity and transgenerational effects of perfluorobutane sulfonate on Caenorhabditis elegans. Environ Toxicol Chem. 40(7):1971-1980.

Chropeňová M, Karásková P, Kallenborn R, Gregušková EK, Čupr P. 2016. Pine needles for the screening of perfluorinated alkylated substances (PFASs) along ski tracks. Environ Sci Technol. 50(17):9487-9496.

Codling G, Sturchio NC, Rockne KJ, Li A, Peng H, Tse TJ, Jones PD, Giesy JP. 2018. Spatial and temporal trends in poly- and per-fluorinated compounds in the Laurentian Great Lakes Erie, Ontario and St. Clair. Environ Pollut. 237:396-405.

Colomban C, Kudrik EV, Afanasiev P, Sorokin AB. 2014. Catalytic defluorination of perfluorinated aromatics under oxidative conditions using N-bridged diiron phthalocyanine. J Am Chem Soc. 136(32):11321-11330.

Conder JM, Hoke RA, de Wolf W, Russell MH, Buck RC. 2008. Are PFCAs bioaccumulative? A critical review and comparison with regulatory criteria and persistent lipophilic compounds. Environ Sci Technol. 42(4):995-1003.

Corsolini S, Kannan K. 2004. Perfluorooctanesulfonate and related fluorochemicals in several organisms including humans from Italy. Organohalogen Compd. 66:4079-4085.

D’Agostino LA, Mabury SA. 2017. Certain perfluoroalkyl and polyfluoroalkyl substances associated with aqueous film forming foam are widespread in Canadian surface waters. Environ Sci Technol. 51(23):13603-13613.

Dassuncao C, Hu XC, Zhang X, Bossi R, Dam M, Mikkelsen B, Sunderland EM. 2017. Temporal shifts in poly-and perfluoroalkyl substances (PFASs) in North Atlantic pilot whales indicate large contribution of atmospheric precursors. Environ Sci Technol. 51(8):4512-4521.

Dassuncao C, Pickard H, Pfohl M, Tokranov AK, Li M, Mikkelesen B, Slitt A, Sunderland EM. 2019. Phospholipid levels predict the tissue distribution of poly- and perfluoroalkyl substances in a marine mammal. Environ Sci Technol Lett. 6(3):119-125.

Dasu K, Lee LS, Turco RF, Nies LF. 2013. Aerobic biodegradation of 8:2 fluorotelomer stearate monoester and 8:2 fluorotelomer citrate triester in forest soil. Chemosphere. 91(3):399-405.

Delinsky AD, Strynar MJ, McCann PJ, Varns JL, McMillan L, Nakayama SF, Lindstrom AB. 2010. Geographical distribution of perfluorinated compounds in fish from Minnesota lakes and rivers. Environ Sci Technol. 44(7):2549-2554.

Dennis NM, Hossain F, Subbiah S, Karnjanapiboonwong A, Dennis ML, McCarthy C, Heron CG, Jackson WA, Crago JP, Field JA, et al. 2021. Chronic reproductive toxicity thresholds for northern bobwhite quail (Colinus virginianus) exposed to perfluorohexanoic acid (PFHxA) and a mixture of perfluorooctane sulfonic acid (PFOS) and PFHxA. Environ Toxicol Chem. 40(9):2601-2614.

De Smet H., Blust R., Moens L. 1998. Absence of albumin in the plasma of the common carp Cyprinus carpio: Binding of fatty acids to high-density lipoprotein. Fish Physiol Biochem. 19(1):71-81.

De Solla SR, De Silva AO, Letcher RJ. 2012. Highly elevated levels of perfluorooctane sulfonate and other perfluorinated acids found in biota and surface water downstream of an international airport, Hamilton, Ontario, Canada. Environ Int. 39(1):19-26.

D’eon JC, Hurley MD, Wallington TJ, Mabury SA. 2006. Atmospheric Chemistry of N-methyl perfluorobutane sulfonamidoethanol, C4F9SO2N(CH3)CH2CH2OH: Kinetics and mechanism of reaction with OH. Environ Sci Technol. 40(6):1862-1868.

Dillert R, Bahnemann D, Hidaka H. 2007. Light-induced degradation of perfluorocarboxylic acids in the presence of titanium dioxide. Chemosphere. 67(4):785-792.

Dimitrov S, Pavlov T, Dimitrova N, Georgieva D, Nedelcheva D, Kesova A, Vasilev R, Mekenyan O. 2011. Simulation of chemical metabolism for fate and hazard assessment. II. CATALOGIC simulation of abiotic and microbial degradation. SAR QSAR Environ Res. 22(7-8):719-755.

Dimitrova NH, Dermen IA, Todorova ND, Vasilev KG, Dimitrov SD, Mekenyan OG, Ikenaga Y, Aoyagi T, Zaitsu Y, Hamaguchi C. 2017. CATALOGIC 301C model – validation and improvement. SAR QSAR Environ Res. 28(6):511-524.

Ding G-H, Frömel T, van den Brandhof E-J, Baerselman R, Peijnenburg WJGM. 2012. Acute toxicity of poly- and perfluorinated compounds to two cladocerans, Daphnia magna and Chydorus sphaericus. Environ Toxicol Chem. 31(3):605-610.

Dinglasan-Panlilio MJA, Mabury SA. 2006. Significant residual fluorinated alcohols present in various fluorinated materials. Environ Sci Technol. 40(5):1447-1453.

Droge STJ. 2019. Membrane-water partition coefficients to aid risk assessment of perfluoroalkyl anions and alkyl sulfates. Environ Sci Technol. 53(2):760-770.

Du D, Lu Y, Zhou Y, Li Q, Zhang M, Han G, Cui H, Jeppesen E. 2021. Bioaccumulation, trophic transfer and biomagnification of perfluoroalkyl acids (PFAAs) in the marine food web of the South China Sea. J Hazard Mater. 405:124681.

Ebert A, Allendorf F, Berger U, Goss K-U, Ulrich N. 2020. Membrane/water partitioning and permeabilities of perfluoroalkyl acids and four of their alternatives and the effects on toxicokinetic behaviour. Environ Sci Technol. 54(8):5051-5061.

[EC, HC] Environment Canada, Health Canada. 2012. Screening assessment: Perfluorooctanoic acid, its salts, and its precursors. Ottawa (ON): Government of Canada.

Eggers Pedersen K, Basu N, Letcher R, Greaves AK, Sonne C, Dietz R, Styrishave B. 2015. Brain region-specific perfluoroalkylated sulfonate (PFSA) and carboxylic acid (PFCA) accumulation and neurochemical biomarker responses in east Greenland polar bears (Ursus maritimus). Environ Res. 138:22-31.

Eggers Pedersen K, Letcher RJ, Sonne C, Dietz R, Styrishave B. 2016. Per- and polyfluoroalkyl substances (PFASs) – New endocrine disruptors in polar bears (Ursus maritimus)? Environ Int. 96:180-189.

Ellis DA, Denkenberger KA, Burrow TE, Mabury SA. 2004. The use of ¹⁹F NMR to interpret the structural properties of perfluorocarboxylate acids: A possible correlation with their environmental disposition. J Phys Chem A. 108(46):10099-10106.

Environment Canada. 2006. Screening assessment: Perfluorooctane sulfonate, its salts and its precursors that contain the C8F17SO2 or C8F17SO3, or C8F17SO2N moiety. Gatineau (QC): Environment Canada.

Environment Canada. 2012. Screening assessment: Long-chain (C9–C20) perfluorocarboxylic acids, their salts and their precursors. Gatineau (QC): Environment Canada.

Erikstad KE, Bustnes JO, Lorentsen S-H, Reiertsen TK. 2009. Sex ratio in lesser black-backed gull in relation to environmental pollutants. Behav Ecol Sociobiol. 63:931-938.

Fair PA, Romano T, Schaefer AM, Reif JS, Bossart GD, Houde M, Muir D, Adams J, Rice C, Hulsey TC, et al. 2013. Associations between perfluoroalkyl compounds and immune and clinical chemistry parameters in highly exposed bottlenose dolphins (Tursiops truncatus). Environ Toxicol Chem. 32(4):736-746.

Falandysz J, Taniyasu S, Gulkowska A, Yamashita N, Schulte-Oehlmann U. 2006. Is fish a major source of fluorinated surfactants and repellents in humans living on the Baltic Coast? Environ Sci Technol. 40(3):748-751.

Flanary JR, Reiner JL, Helke KL, Kucklick JR, Gulland FMD, Becker PR. 2012. Measurements of perfluorinated compounds in plasma of northern fur seals (Callorhinus ursinus) and a preliminary assessment of their relationship to peroxisome proliferation and blood chemistry parameters. Reprod Toxicol. 33(4):601.

Frömel T, Knepper TP. 2010. Fluorotelomer ethoxylates: Sources of highly fluorinated environmental contaminants part I: Biotransformation. Chemosphere. 80(11):1387-1392.

Furdui VI, Stock NL, Ellis DA, Butt CM, Whittle M, Crozier PW, Reiner EJ, Muir DCG, Mabury SA. 2007. Spatial distribution of perfluoroalkyl contaminants in lake trout from the Great Lakes. Environ Sci Technol. 41(5):1554-1559.

Furdui VI, Crozier PW, Reiner EJ, Mabury SA. 2008a. Trace level determination of perfluorinated compounds in water by direct injection. Chemosphere. 73(1 Suppl):S24-S30.

Furdui VI, Helm PA, Crozier PW, Lucaciu C, Reiner EI, Marvin CH, Whittle DM, Mabury SA, Tomy GT. 2008b. Temporal trends of perfluoroalkyl compounds with isomer analysis in lake trout from Lake Ontario (1979-2004). Environ Sci Technol. 42(13):4739-4744.

Gamberg M, Muir DCG, Karrman A, Cuyler C, Rigét F, Bossi R, Carlsson P, Roos A. 2017. Perfluoroalkyl acids in Arctic caribou and reindeer. Organohalogen Compd. 79:795-798.

Gao K, Miao X, Fu J, Chen Y, Li H, Pan W, Fu J, Zhang Q, Zhang A, Jiang G. 2020a. Occurrence and trophic transfer of per- and polyfluoroalkyl substances in an Antarctic ecosystem. Environ Pollut. 257:113383.

Gao S, Liu H, Chang H, Zhang Z, Hu J, Tao S, Wan Y. 2020b. Visualized metabolic disorder and its chemical inducer in wild crucian carp from Taihu Lake, China. Environ Sci Technol. 54(6):3343-3352.

Gebbink WA, Letcher RJ, Hebert CE, Weseloh DVC. 2011. Twenty years of temporal change in perfluoroalkyl sulfonate and carboxylate contaminants in herring gull eggs from the Laurentian Great Lakes. J Environ Monitor. 13(12):3365-3372.

Gebbink WA, Bossi R, Rigét FF, Rosing-Asvid A, Sonne C, Dietz R. 2016. Observation of emerging per- and polyfluoroalkyl substances (PFASs) in Greenland marine mammals. Chemosphere. 144:2384-2391.

Gewurtz SB, Martin PA, Letcher RJ, Burgess NM, Champoux L, Elliott JE, Weseloh DVC. 2016. Spatio-temporal trends and monitoring design of perfluoroalkyl acids in the eggs of gull (Larid) species from across Canada and parts of the United States. Sci Total Environ. 565:440-450.

Gewurtz SB, Martin PA, Letcher RJ, Burgess NM, Champoux L, Elliott JE, Idrissi A. 2018. Perfluoroalkyl acids in European starling eggs indicate landfill and urban influences in Canadian terrestrial environments. Environ Sci Technol. 52(10):5571-5580.

Gewurtz SB, Bradley LE, Backus S, Dove A, McGoldrick D, Hung H, Dryfhout-Clark H. 2019. Perfluoroalkyl acids in Great Lakes precipitation and surface water (2006–2018) indicate response to phase-outs, regulatory action, and variability in fate and transport processes. Environ Sci Technol. 53(15):8543-8552.

Gewurtz SB, Guerra P, Kim MG, Jones F, Urbanic JC, Teslic S, Smyth SA. 2020. Wastewater treatment lagoons: Local pathways of perfluoroalkyl acids and brominated flame retardants to the Arctic environment. Environ Sci Technol. 54(10):6053-6062.

Giesy JP, Newsted JL. 2001. Selected fluorochemicals in the Decatur, Alabama area. Project No. 178401. Report prepared for 3M.

Göckener B, Fliedner A, Rüdel H, Fettig I, Koschorreck J. 2021. Exploring unknown per- and polyfluoroalkyl substances in the German environment – The total oxidizable precursor assay as helpful tool in research and regulation. Sci Total Environ. 782:146825.

Godfrey A, Abdel-Moneim A, Sepúlveda MS. 2017. Acute mixture toxicity of halogenated chemicals and their next generation counterparts on zebrafish embryos. Chemosphere. 181:710-712.

Goeritz I, Falk S, Stahl T, Schäfers C, Schlechtriem C. 2013. Biomagnification and tissue distribution of perfluoroalkyl substances (PFASs) in market-size rainbow trout (Oncorhynchus mykiss). Environ Toxicol Chem. 32(9):2078-2088.

Gorrochategui E, Lacorte S, Tauler R, Martin FL. 2016. Perfluoroalkylated substance effects in Xenopus laevis A6 kidney epithelial cells determined by ATR-FTIR spectroscopy and chemometric analysis. Chem Res Toxicol. 29(5):924-932

Gray JS. 2002. Biomagnification in marine systems: The perspective of an ecologist. Mar Pollut Bull. 45(1-12):46-52.

Greaves AK, Letcher RJ, Sonne C, Dietz R, Born EW. 2012. Tissue-specific concentrations and patterns of perfluoroalkyl carboxylates and sulfonates in East Greenland polar bears. Environ Sci Technol. 46(21):11575-11583.

Greaves AK, Letcher RJ, Sonne C, Dietz R. 2013. Brain region distribution and patterns of bioaccumulative perfluoroalkyl carboxylates and sulfonates in East Greenland polar bears (Ursus maritimus). 2013. Environ Toxicol Chem. Environ Toxicol Chem. 32(3):713-722.

Groffen T, Lasters R, Lopez-Antia A, Prinsen E, Bervoets L, Eens M. 2019. Limited reproductive impairment in a passerine bird species exposed along a perfluoroalkyl acid (PFAA) pollution gradient. Sci Total Environ. 652:718-728.

Grønnestad R, Villanger GD, Polder A, Kovacs KM, Lydersen C, Jenssen BM, Borga K. 2018. Effects of a complex contaminant mixture on thyroid hormones in breeding hooded seal mothers and their pups. Environ Pollut. 240:10-16.

Grønnestad R, Péréz Vázquez B, Arukwe A, Jaspers VLB, Jenssen BM, Karimi M, Lyche JL, Krøkje A. 2019. Levels, patterns, and biomagnification potential of perfluoroalkyl substances in a terrestrial food chain in a Nordic skiing area. Environ Sci Technol. 53(22):13390-13397.

Guelfo JL, Higgins CP. 2013. Subsurface transport potential of perfluoroalkyl acids at aqueous film-forming foam (AFFF)-impacted sites. Environ Sci Technol. 47(9):4164-4171.

Guerra P, Kim M, Kinsman L, Ng T, Alaee M, Smyth SA. 2014. Parameters affecting the formation of perfluoroalkyl acids during wastewater treatment. J Hazard Mater. 272:148-154.

Guillette TC, McCord J, Guillette M, Polera ME, Rachels KT, Morgeson C, Kotlarz N, Knappe DRU, Reading BJ, Strynar M, et al. 2020. Elevated levels of per- and polyfluoroalkyl substances in Cape Fear River striped bass (Morone saxatilis) are associated with biomarkers of altered immune and liver function. Environ Int. 136:105358.

Gulkowska A, Falandysz J, Taniyasu S, Bochentin I, So MK, Yamashita N. 2005. Perfluorinated chemicals in blood of fish and waterfowl from Gulf of Gdansk, Baltic Sea. Poster presentation at Fluoros Symposium, An International Symposium on Fluorinated Alkyl Organics in the Environment; 2005 August 18–20; Toronto, Canada.

Gulkowska A, Jiang Q, So MK, Taniyasu S, Lam PKS, Yamashita N. 2006. Persistent perfluorinated acids in seafood collected from two cities of China. Environ Sci Technol. 40(12):3736-3741.

Hakli O, Ertekin K, Ozer MS, Aycan S. 2008. Determination of pK_a values of clinically important perfluorochemicals in nonaqueous media. J Anal Chem. 63(11):1051-1056.

Hansen E, Huber N, Bustnes JO, Herzke D, Bårdsen B-J, Eulaers I, Johnsen TV, Bourgeon S. 2020. A novel use of the leukocyte coping capacity assay to assess the immunomodulatory effects of organohalogenated contaminants in avian wildlife. Environ Int. 142:105861.

Hatton J, Holton C, DiGuiseppi B. 2018. Occurrence and behaviour of per- and polyfluoroalkyl substances from aqueous film-forming foam in groundwater systems. Remediation J. 28(2):89-99.

Haukås M, Berger U, Hop H, Gulliksen B, Gabrielsen GW. 2007, Bioaccumulation of per- and polyfluorinated alkyl substances (PFAS) in selected species from the Barents Sea food web. Environ Pollut. 148(1):360-371.

HC [Health Canada]. 2022. [modified Feb 11]. Consideration of endocrine-related effects in risk assessment. Ottawa (ON): Government of Canada. [accessed 2022 Jul 8].

Herzke D, Nygård T, Berger U, Huber S, Røv N. 2009. Perfluorinated and other persistent halogenated organic compounds in European shag (Phalacrocorax aristotelis) and common eider (Somateria mollissima) from Norway: A suburban to remote pollutant gradient. Sci Total Environ. 408(2):340-348.

Hickey NJ, Crump D, Jones SP, Kennedy SW. 2009. Effects of 18 perfluoroalkyl compounds on mRNA expression in chicken embryo hepatocyte cultures. Toxicol Sci. 111(2):311-320.

Higgins CP, Luthy RG. 2006. Sorption of perfluorinated surfactants on sediments. Environ Sci Technol. 40(23):7251-7256.

Hoke RA, Bouchelle LD, Ferrel BD, Buck RC. 2012. Comparative acute freshwater hazard assessment and preliminary PNEC development for eight fluorinated acids. Chemosphere. 87(7):725-733

Holden PR, Tugwood JD. 1999. Peroxisome proliferator-activated receptor alpha: Role in rodent liver cancer and species differences. J Mol Endocrinol. 22(1):1-8.

Holmström KE, Berger U. 2008. Tissue distribution of perfluorinated surfactants in common guillemot (Uria aalge) from the Baltic Sea. Environ Sci Technol. 42(16):5879-5884.

Holmström KE, Johansson A-K, Bignert A, Lindberg P, Berger U. 2010. Temporal trends of perfluorinated surfactants in Swedish peregrine falcon eggs (Falco peregrinus), 1974-2007. Environ Sci Technol. 44(11):4083-4088.

Hong SH, Reiner JL, Jang M, Schuur SS, Han GM, Kucklick JR, Shim WJ. 2022. Levels and profiles of perfluorinated alkyl acids in liver tissues of birds with different habitat types and trophic levels from an urbanized coastal region of South Korea. Sci Total Environ. 806(Pt 3):151263.

Hoover GM, Chislock MF, Tornabene BJ, Guffey SC, Choi YJ, De Perre C, Hoverman JT, Lee LS, Sepúlveda MS. 2017. Uptake and depuration of four per/polyfluoroalkyl substances (PFASS) in northern leopard frog Rana pipiens tadpoles. Envion Sci Technol Lett. 4(10):399-403.

Hoover G, Kar S, Guffey S, Leszczynski J, Sepúlveda MS. 2019. In vitro and in silico modeling of perfluoroalkyl substances mixture toxicity in an amphibian fibroblast cell line. Chemosphere. 233:25-33.

Hori H, Yamamoto A, Hayakawa E, Taniyasu S, Yamashita N, Kutsuna S, Kiatagawa H, Arakawa R. 2005. Efficient decomposition of environmentally persistent perfluorocarboxylic acids by use of persulfate as a photochemical oxidant. Environ Sci Technol. 39(7):2383-2388.

Hori H, Nagaoka Y, Murayama M, Kutsuna S. 2008. Efficient decomposition of perfluorocarboxylilc acids and alternative fluorochemical surfactants in hot water. Environ Sci Technol. 42(19):7438-7443.

Houde M, Wells RS, Fair PA, Bossart GD, Hohn AA, Rowles TK, Sweeney JC, Solomon KR, Muir DCG. 2005. Polyfluoroalkyl compounds in free-ranging bottlenose dolphins (Tursiops truncatus) from the Gulf of Mexico and the Atlantic Ocean. Environ Sci Technol. 39(17):6591-6598.

Houde M, Bujas TAD, Small J, Wells RS, Fair PA, Bossart GD, Solomon KR, Muir DCG. 2006a. Biomagnification of perfluoroalkyl compounds in the bottlenose dolphin (Tursiops truncatus) food web. Environ Sci Technol. 40(13):4138-4144.

Houde M, Martin JW, Letcher RJ, Solomon KR, Muir DG. 2006b. Biological monitoring of polyfluoroalkyl substances: a review. Environ Sci Technol. 40:3463-3473.

Houde M, Balmer BC, Brandsma S, Wells RS, Rowles TK, Solomon KR, Muir DCG. 2006c. Perfluoroalkyl compounds in relation to life-history and reproductive parameters in bottlenose dolphins (Tursiops truncatus) from Sarasota Bay, Florida, USA. Environ Toxicol Chem 25(9):2405-2412.

Houde M, Douville M, Despatie S-P, De Silva AO, Spencer C. 2013. Induction of gene responses in St. Lawrence northern pike (Esox lucius) environmentally exposed to perfluorinated compounds. Chemosphere. 92(9):1195-1200.

Houde M, Giraudo M, Douville M, Bougas B, Couture P, De Silva AO, Spencer C, Lair S, Verreault J, Bernatchez L, et al. 2014. A multi-level biological approach to evaluate impacts of a major municipal effluent in wild St. Lawrence River yellow perch (Perca flavescens). Sci Total Environ. 497-498:307-318.

Huang K, Li Y, Bu D, Fu J, Wang M, Zhou W, Gu L, Fu Y, Cong Z, Hu B, et al. 2022. Trophic magnification of short-chain per- and polyfluoroalkyl substances in a terrestrial food chain from the Tibetan Plateau. Environ Sci Technol Lett. 9(2):147-152.

Hundley SG, Sarrif AM, Kennedy GL. 2006. Absorption, distribution, and excretion of ammonium perfluorooctanoate (APFO) after oral administration to various species. Drug Chem Toxicol. 29(2):137-145.

Hurley MD, Sulbaek Andersen MP, Wallington TJ, Ellis DA, Martin JW, Mabury SA. 2004. Atmospheric chemistry of perfluorinated carboxylic acids: Reaction with OH radicals and atmospheric lifetimes. J Phys Chem A. 108(4):615-620.

Inoue Y, Hashizume N, Yakata N, Murakami H, Suzuki Y, Kikushima E, Otsuka M. 2012. Unique physiochemical properties of perfluorinated compounds and their bioconcentration in common carp Cyprinus carpio L. Arch Environ Contam Toxicol. 62(4):672-680.

Ishibashi H, Iwata H, Kim E-Y, Tao L, Kannan K, Amano M, Miyazaki N, Tanabe S, Batoev VB, Petrov EA. 2008a. Contamination and effects of perfluorochemicals in Baikal seal (Pusa sibirica). 1. Residue level, tissue distribution, and temporal trend. Environ Sci Technol. 42(7):2295-2301.

Ishibashi H, Iwata H, Kim E-Y, Tao L, Kannan K, Tanabe S, Batoev VB, Petrov EA. 2008b. Contamination and effects of perfluorochemicals in Baikal seal (Pusa sibirica). 2. Molecular characterization, expression level, and transcriptional activation of peroxisome proliferator-activated receptor alpha. Environ Sci Technol. 42(7):2302-2308.

Ishibashi H, Kim E-Y, Iwata H. 2011. Transactivation potencies of the Baikal seal (Pusa sibirica) peroxisome proliferator-activated receptor alpha by perfluoroalkyl carboxylates and sulfonates: Estimation of PFOA induction equivalency factors. Environ Sci Technol. 45(7):3123-3130.

Jackson DA, Wallington TJ, Mabury SA. 2013. Atmospheric oxidation of polyfluorinated amides: Historical source of perfluorinated carboxylic acids to the environment. Environ Sci Technol. 47(9):4317-4324

Jing P, Rodgers PJ, Amemiya S. 2009. High lipophilicity of perfluoroalkyl carboxylate and sulfonate: Implications for their membrane permeability. J Am Chem Soc. 131(6):2290-2296.

Jones PD, Hu W, de Coen W, Newsted JL, Giesy JP. 2003. Binding of perfluorinated fatty acids to serum proteins. Environ Toxicol Chem. 22(11):2639-2649.

Kaboré HA, Goeury K, Desrosiers M, Duy SV, Liu J, Cabana G, Munoz G, Sauvé S. 2022. Novel and legacy per- and polyfluoroalkyl substances (PFAS) in freshwater sporting fish from background and firefighting foam impacted ecosystems in Eastern Canada. Sci Total Environ. 816:151563.

Kannan K, Newsted J, Halbrook RS, Giesy JP. 2002a. Perfluorooctanesulfonate and related fluorinated hydrocarbons in mink and river otters from the United States. Environ Sci Technol. 36(12):2566-2571.

Kannan K, Corsolini S, Falandysz J, Oehme G, Focardi S, Giesy JP. 2002b. Perfluorooctane sulfonate and related fluorinated hydrocarbons in marine mammals, fishes, and birds from coasts of the Baltic and the Mediterranean Seas. Environ Sci Technol. 36(15):3210-3216.

Kannan K, Choi J-W, Iseki N, Senthilkumar K, Kim DH, Masunaga S, Giesy JP. 2002c. Concentrations of perfluorinated acids in livers of birds from Japan and Korea. Chemosphere. 49(3):225-231.

Kannan K, Tao L, Sinclair E, Pastva SD, Jude DJ, Giesy JP. 2005. Perfluorinated compounds in aquatic organisms at various trophic levels in a Great Lakes food chain. Arch Environ Contam Toxicol. 48(4):559-566.

Karnjanapiboonwong A, Deb SK, Subbiah S, Wang D, Anderson TA. 2018. Perfluoroalkylsulfonic and carboxylic acids in earthworms (Eisenia fetida): Accumulation and effects results from spiked soils at PFAS concentrations bracketing environmental relevance. Chemosphere. 199:168-173.

Keller JM, Kannan K, Taniyasu S, Yamashita N, Day RD, Arendt MD, Segars AL, Kucklick JR. 2005. Perfluorinated compounds in the plasma of loggerhead and Kemp’s ridley sea turtles from the southeastern coast of the United States. Environ Sci Technol. 39(23):9101-9108.

Kelly BC, Gobas FAPC, McLachlan MS. 2004. Intestinal absorption and biomagnification of organic contaminants in fish, wildlife, and humans. Environ Toxicol Chem. 23(10):2324-2336.

Kelly BC, Ikonomou MG, Blair JD, Surridge B, Hoover D, Grace R, Gobas FAPC. 2009. Perfluoroalkyl contaminants in an Arctic marine food web: Trophic magnification and wildlife exposure. Environ Sci Technol. 43(11):4037-4043.

Key BD, Howell RD, Criddle CS. 1997. Fluorinated organics in the biosphere. Environ Sci Technol. 31(9):2445-2454.

Kim MH, Wang N, McDonald T, Chu K-H. 2012. Biodefluorination and biotransformation of fluorotelomer alcohols by two alkane-degrading Pseudomonas strains. Biotechnol Bioeng. 109(12):3041-3048.

Kim MH, Wang N, Chu KH. 2014. 6:2 fluorotelomer alcohol (6:2 FTOH) biodegradation by multiple microbial species under different physiological conditions. Appl Microbiol Biotechnol. 98(4):1831-1840

Kim M, Li LY, Grace JR, Yue C. 2015. Selecting reliable physicochemical properties of perfluoroalkyl and polyfluoroalkyl substances (PFASs) based on molecular descriptors. Environ Pollut. 196:462-472.

Kirk-Othmer. 1994. Fluorine compounds, organic (higher acids). In: Kirk-Othmer encyclopedia of chemical technology. vol. 11. New York (NY): Wiley. p. 551.

Knudsen LB, Borgå K, Jørgensen EH, van Bavel B, Schlabach M, Verreault J, Gabrielsen GW. 2007. Halogenated organic contaminants and mercury in northern fulmars (Fulmarus glacialis): Levels, relationships to dietary descriptors and blood to liver comparison. Environ Pollut. 146(1):25-33.

Kosswig K. 2000. Sulfonic acids, aliphatic. In: Ullmann’s encyclopedia of industrial chemistry. John Wiley & Sons, Ltd. [accessed 2022 Jul 7].

Kurtz AE, Reiner JL, West KL, Jensen BA. 2019. Perfluorinated alkyl acids in Hawaiian cetaceans and potential biomarkers of effect: Peroxisome proliferator-activated receptor alpha and cytochrome P450 4A. Environ Sci Technol. 53(5):2830-2839.

Kwadijk CJAF, Korytár P, Koelmans AA. 2010. Distribution of perfluorinated compounds in aquatic systems in the Netherlands. Environ Sci Technol. 44(10):3746-3751.

Langford KH, Kringstad A, Brooks L, Reid M. 2010. Inputs of perfluorinated compounds from skiing activities to the Norwegian recreational environment. Nordic Environmental Chemistry Conference, Longyearbyen, Svalbard, Norway (cited in Plassmann and Berger 2013).

Larter NC, Muir D, Wang X, Allaire DG, Kelly A, Cox K. 2017. Persistent organic pollutants in the livers of moose harvested in the southern Northwest Territories, Canada. ALCES. 53:65-83.

Lasier PJ, Washington JW, Hassan SM, Jenkins TM. 2011. Perfluorinated chemicals in surface waters and sediments from northwest Georgia, USA, and their bioaccumulation in Lumbriculus variegatus. Environ Toxicol Chem. 30(10):2194-2201.

Latala A, Nędzi M, Stepnowski P. 2009. Acute toxicity assessment of perfluorinated carboxylic acids towards the Baltic microalgae. Environ Toxicol Pharmacol. 28(2):167-171.

Leat EHK, Bourgeon S, Eze JI, Muir DCG, Williamson M, Bustnes JO, Furness RW, Borgå K. 2013. Perfluoroalkyl substances in egg and plasma of an avian top predator, great skua (Stercorarius skua), in the North Atlantic. Environ Toxicol Chem. 32(3):569-576.

Lee H, D’eon J, Mabury SA. 2010. Biodegradation of polyfluoroalkyl phosphates as a source of perfluorinated acids to the environment. Environ Sci Technol. 44(9):3305-3310.

Lee JJ, Schultz IR. 2010. Sex differences in the uptake and disposition of perfluorooctanoic acid in fathead minnows after oral dosing. Environ Sci Technol. 44(1):491-496.

Lehmler H-J, Oyewumi MO, Jay M, Bummer PM. 2001. Behavior of partially fluorinated carboxylic acids at the air-water interface. J Fluorine Chem. 107(1):141-146.

Lehmler H-J, Xie W, Bothun GD, Bummer PM, Knutson BL. 2006. Mixing of perfluorooctanesulfonic acid (PFOS) potassium salt with dipalmitoyl phosphatidylcholine (DPPC). Colloids Surf. B Biointerfaces. 51(1):25-29.

Leonel J, Kannan K, Tao L, Fillmann G, Montone RC. 2008. A baseline study of perfluorochemicals in Franciscana dolphin and Subantarctic fur seal from coastal waters of Southern Brazil. Baseline Mar Pollut Bull. 56(4):770-797.

Lescord GL, Kidd KA, De Silva AO, Williamson M, Spencer C, Wang X, Muir DCG. 2015. Perfluorinated and polyfluorinated compounds in lake food webs from the Canadian High Arctic. Environ Sci Technol. 49(5):2694-2702.

Letcher RJ, Bustnes JO, Dietz R, Jenssen BM, Jørgensen EH, Sonne C, Verreault J, Vijayan MM, Gabrielsen GW. 2010. Exposure and effects assessment of persistent organohalogen contaminants in arctic wildlife and fish. Sci Total Environ. 408(15):2995-3043.

Letcher RJ, Chu S, McKinney MA, Tomy GT, Sonne C, Dietz R. 2014. Comparative hepatic in vitro depletion and metabolite formation of major perfluorooctane sulfonate precursors in Arctic polar bear, beluga whale, and ringed seal. Chemosphere. 112:225-231.

Letcher RJ, Su G, Moore JN, Williams LL, Martin PA, de Solla SR, Bowerman WW. 2015. Perfluorinated sulfonate and carboxylate compounds and precursors in herring gull eggs from across the Laurentian Great Lakes of North America: Temporal and recent spatial comparisons and exposure implications. Sci Total Environ. 538:468-477.

Letcher RJ, Morris AD, Dyck M, Sverko E, Reiner EJ, Blair DAD, Chu SG, Shen L. 2018. Legacy and new halogenated persistent organic pollutants in polar bears from a contamination hotspot in the Arctic, Hudson Bay Canada. Sci Total Environ. 610-611:121-136.

Li X, Yeung LWY, Taniyasu S, Lam PKS, Yamashita N, Xu M, Dai J. 2008. Accumulation of perfluorinated compounds in captive Bengal tigers (Panthera tigris tigris) and African lions (Panthera leo Linnaeus) in China. Chemosphere. 73(10):1649-1653.

Li Z, Yu Z, Yin D. 2021. Multi- and trans-generational disturbances of perfluorobutane sulfonate and perfluorohexane sulfonate on lipid metabolism in Caenorhabditis elegans. Chemosphere. 280:130666

Liu J, Lee LS, Nies LF, Nakatsu CH, Turco RF. 2007. Biotransformation of 8:2 fluorotelomer alcohol in soil and by soil bacteria isolates. Environ Sci Technol. 41(23):8024-8030

Liu W, Chen S, Quan X, Jin Y-H. 2008. Toxic effect of serial perfluorosulfonic and perfluorocarboxylic acids on the membrane system of a freshwater alga measured by flow cytometry. Environ Toxicol Chem. 27(7):1597-1604.

Liu J, Wang N, Szostek B, Buck RC, Panciroli PK, Folsom PW, Sulecki LM, Bellin CA. 2010a. 6-2 Fluorotelomer alcohol aerobic biodegradation in soil and mixed bacterial culture. Chemosphere. 78(4):437-444.

Liu J, Wang N, Buck RC, Wolstenholme BW, Folsom PW, Sulecki LM, Bellin CA. 2010b. Aerobic biodegradation of [14C] 6:2 fluorotelomer alcohol in a flow-through soil incubation system. Chemosphere. 80(7):716-723.

Liu W, He W, Wu J, Qin N, He Q, Xu F. 2018a. Residues, bioaccumulations and biomagnification of perfluoroalkyl acids (PFAAs) in aquatic animals from Lake Chaohu, China. Environ Pollut. 240:607-614.

Liu Y, Richardson ES, Derocher AE, Lunn NJ, Lehmler H-J, Li X, Zhang Y, Cui JY, Cheng L, Martin JW. 2018b. Hundreds of unrecognized halogenated contaminants discovered in polar bear serum. Angew Chem Int Ed. 57(50):16401-16406.

Liu M, Munoz G, Duy SV, Sauvé S, Liu J. 2022. Per- and polyfluoroalkyl substances in contaminated soil and groundwater at airports: A Canadian case study. Environ Sci Technol. 56(2):885-895.

Llorca M, Farré M, Tavano MS, Alonso B, Koremblit G, Barceló D. 2012. Fate of a broad spectrum of perfluorinated compounds in soils and biota from Tierra del Fuego and Antarctica. Environ Pollut. 163:158-166.

Loi EIH, Yeung LWY, Taniyasu S, Lam PKS, Kannan K, Yamashita N. 2011. Trophic magnification of poly- and perfluorinated compounds in a subtropical food web. Environ Sci Technol. 45(13):5506-5513.

MacInnis JJ, French K, Muir DCG, Spencer C, Criscitiello A, De Silva AO, Young CJ. 2017. Emerging investigator series: A 14-year depositional ice record of perfluoroalkyl substances in the High Arctic. Environ Sci: Processes Impacts. 19(1):22-30

MacInnis JJ, Lehnherr I, Muir DCG, St. Pierre KA, St. Louis VL, Spencer C, De Silva AO. 2019a. Fate and transport of perfluoroalkyl substances from snowpacks into a lake in the High Arctic of Canada. Environ Sci Technol. 53(18):10753-10762.

MacInnis JJ, Lehnherr I, Muir DCG, Quinlan R, De Silva AO. 2019b. Characterization of perfluoroalkyl substances in sediment cores from High and Low Arctic lakes in Canada. Sci Total Environ. 666:414-422.

MacInnis J, De Silva AO, Lehnherr I, Muir DCG, Pierre KAS, St. Louis VL, Spencer C. 2022. Investigation of perfluoroalkyl substances in proglacial rivers and permafrost seep in a high Arctic watershed. Environ Sci: Processes Impacts. 24(1):42-51.

Mackay D, Fraser A. 2000. Bioaccumulation of persistent organic chemicals: Mechanisms and models. Environ Pollut. 110(3):375-391.

Mackay D, Celsie AKD, Arnot JA, Powell DE. 2006. Processes influencing chemical biomagnification and trophic magnification factors in aquatic ecosystems: Implications for chemical hazard and risk assessment. Chemosphere. 154:99-108.

MacLeod M, Breitholtz M, Cousins IT, de Wit CA, Persson LM, Rudén C, McLachlan MS. 2014. Identifying chemicals that are planetary boundary threats. Environ Sci Technol. 48(19):11057-11063.

Mahapatra CT, Damayanti NP, Guffey SC, Serafin JS, Irudayaraj J, Sepúlveda MS. 2016. Comparative in vitro toxicity assessment of perfluorinated carboxylic acids. J Appl Toxicol. 37(6):699-708.

Manera M, Britti D. 2006. Assessment of blood chemistry normal ranges in rainbow trout. J Fish Biol. 69(5):1427-1434.

Martin JW, Mabury SA, Solomon KR, Muir DCG. 2003a. Dietary accumulation of perfluorinated acids in juvenile rainbow trout (Oncorhynchus mykiss). Environ Toxicol Chem. 22(1):189-195.

Martin JW, Mabury SA, Solomon KR, Muir DCG. 2003b. Bioconcentration and tissue distribution of perfluorinated acids in rainbow trout (Oncorhynchus mykiss). Environ Toxicol Chem. 22(1):196-204.

Martin JW, Ellis DA, Mabury SA, Hurley MD, Wallington TJ. 2006. Atmospheric chemistry of perfluoroalkanesulfonamides: Kinetic and product studies of the OH radical and Cl atom initiated oxidation of N-ethyl perfluorobutanesulfonamide. Environ Sci Technol. 40(3):864-872.

Marziali L, Rosignoli F, Valsecchi S, Polesello S, Stefani F. 2019. Effects of perfluoroalkyl substances on a multigenerational scale: A case study with Chironomus riparius (Diptera, Chironomidae). Environ Toxicol Chem. 38(5):988-999.

Menger F, Pohl J, Ahrens L, Carlsson G, Örn S. 2020. Behavioural effects and bioconcentration of per- and polyfluoroalkyl substances (PFASs) in zebrafish (Danio rerio) embryos. Chemosphere. 245:125573.

Meyer J, Jaspers VLB, Eens M, de Coem W. 2009. The relationship between perfluorinated chemical levels in the feathers and livers of birds from different trophic levels. Sci Total Environ. 407(22):5894-5900.

Meyer T, De Silva AO, Spencer C, Wania F. 2011. Fate of perfluorinated carboxylates and sulfonates during snowmelt within an urban watershed. Environ Sci Technol. 45(19):8113-8119.

Miljeteig C, Strøm H, Gavrilo MV, Volkov A, Jenssen BM, Gabrielsen GW. 2009. High levels of contaminants in ivory gull Pagophila eburnea eggs from the Russian and Norwegian Arctic. Environ Sci Technol. 43(14):5521-5528.

Miljeteig C, Gabrielsen GW, Strøm H, Gavrilo MV, Lie E, Jenssen BM. 2012. Eggshell thinning and decreased concentrations of vitamin E are associated with contaminants in eggs of ivory gulls. Sci Total Environ. 431:92-99.

Miller A, Elliott JE, Elliott KH, Lee S, Cyr F. 2015. Temporal trends of perfluoroalkyl substances (PFAS) in eggs of coastal and offshore birds: Increasing PFAS levels associated with offshore bird species breeding on the Pacific coast of Canada and wintering near Asia. Environ Toxicol Chem. 34(8):1799-1808.

Moody CA, Kwan WC, Martin JW, Muir DCG, Mabury SA. 2001. Determination of perfluorinated surfactants in surface water samples by two independent analytical techniques: Liquid chromatography/tandem mass spectrometry and ¹⁹F NMR. Anal Chem. 73(10):2200-2206.

Moody CA, Martin JW, Kwan WC, Muir DCG, Mabury SA. 2002. Monitoring perfluorinated surfactants in biota and surface water samples following an accidental release of fire-fighting foam into Etobicoke Creek. Environ Sci Technol. 36(4):545-551.

Morales L, Martrat MG, Olmos J, Parera J, Vicente J, Bertolero A, Abalos M, Lacorte S, Santos FJ, Abad E. 2012. Persistent organic pollutants in gull eggs of two species (Larus michahellis and Larus audouinii) from the Ebro delta Natural Park. Chemosphere. 88(11):1306-1316.

Moroi Y, Yano H, Shibata O, Yonemitsu T. 2001. Determination of acidity constants of perfluoroalkanoic acids. Bull Chem Soc Jpn. 74(4):667-672.

[MSDS] Material Safety Data Sheet. 2004. Heptafluorobutyric acid [PDF]. Oakville (ON): Sigma-Aldrich. [accessed 2008 Sep 15].

Muir D, Miaz LT. 2021. Spatial and temporal trends of perfluoroalkyl substances in global ocean and coastal waters. Environ Sci Technol. 55(14):9527-9537.

Munoz G, Budzinski H, Babut M, Drouineau H, Lauzent M, Le Menach K, Lobry J, Selleslagh J, Simonnet-Laprade C, Labadie P. 2017a. Evidence for the trophic transfer of perfluoroalkyated substances in a temperate macrotidal estuary. Environ Sci Technol. 51(15):8450-8459.

Munoz G, Desrosiers M, Duy SV, Labadie P, Budzinski H, Liu J, Sauvé S. 2017b. Environmental occurrence of perfluoroalkyl acids and novel fluorotelomer surfactants in the freshwater fish Catostomus commersonii and sediments following firefighting foam deployment at the Lac-Mégantic railway accident. Environ Sci Technol. 51(3):1231-1240.

Murakami M, Adachi N, Saha M, Morita C, Takada H. 2011. Levels, temporal trends, and tissue distribution of perfluorinated surfactants in freshwater fish from Asian countries. Arch Environ Contam Toxicol. 61(4):631-641.

Myers AL, Mabury SA. 2010. Fate of fluorotelomer acids in a soil-water microcosm. Environ Toxicol Chem. 29(8):1689-1695.

Nabb DL, Szostek B, Himmelstein MW, Mawn MP, Gargas ML, Sweeney LM, Stadler JC, Buck RC, Fasano WJ. 2007. In vitro metabolism of 8-2 fluorotelomer alcohol: Interspecies comparisons and metabolic pathway refinement. Toxicol Sci. 100(2):333-344.

Naile JE, Khim JS, Wang T, Chen C, Luo W, Kwon B-O, Park J, Koh C-H, Jones PD, Lu Y, et al. 2010. Perfluorinated compounds in water, sediment, soil and biota from estuarine and coastal areas of Korea. Environ Pollut. 158(5):1237-1244.

Naile JE, Khim JS, Hong S, Park J, Kwon B-O, Ryu JS, Hwang JH, Jones PD, Giesy JP. 2013. Distributions and bioconcentration characteristics of perfluorinated compounds in environmental samples collected from the west coast of Korea. Chemosphere. 90(2):387-394.

Nakata H, Kannan K, Nasu T, Cho H-S, Sinclair E, Takemurai A. 2006. Perfluorinated contaminants in sediments and aquatic organisms collected from shallow water and tidal flat areas of the Ariake Sea, Japan: Environmental fate of perfluorooctane sulfonate in aquatic ecosystems. Environ Sci Technol. 40(16):4916-4921.

National Academies of Sciences, Engineering, and Medicine. 2017. Use and potential Impacts of AFFF containing PFASs at airports. Washington, DC: The National Academies Press.

Ng CA, Hungerbühler K. 2013. Bioconcentration of perfluorinated alkyl acids: How important is specific binding? Environ Sci Technol. 47(13):7214-7223.

Ng CA, Hungerbühler K. 2014. Bioaccumulation of perfluorinated alkyl acids: Observations and models. Environ Sci Technol. 48(9):4637-4648.

Nobels I, Dardenne F, De Coen W, Blust R. 2010. Application of a multiple endpoint bacterial reporter assay to evaluate toxicological relevant endpoints of perfluorinated compounds with different functional groups and varying chain length. Toxicol In Vitro. 24(6):1768-1774.

Nolen RM, Faulkner P, Ross AD, Kaiser K, Quigg A, Hala D. 2022. PFASs pollution in Galveston Bay surface waters and biota (shellfish and fish) following AFFFs use during the ITC fire at Deer Park (March 17th–20th 2019), Houston, TX. Sci Total Environ. 805:150361.

Nordic Ecolabelling. 2018. About Nordic Swan Ecolabelled ski wax [PDF]. Draft for consultation, version 1. Copenhagen (DK): Nordic Ecolabelling.

Nøst TH, Helgason LB, Harju M, Heimstad ES, Gabrielsen GW, Jenssen BM. 2012. Halogenated organic contaminants and their correlations with circulating thyroid hormones in developing Arctic seabirds. Sci Total Environ. 414:248-256.

Nouhi S, Ahrens L, Pereira HC, Hughes AV, Campana M, Gutfreund P, Palsson GK, Vorobiev A, Hellsing MS. 2018. Interactions of perfluoroalkyl substances with a phospholipid bilayer studied by neutron reflectometry. J Colloid Interface Sci. 511:474-481.

Numata J, Kowalczyk J, Adolphs J, Ehlers S, Schafft H, Fuerst P, Müller-Graf C, Lahrssen-Wiederholt M, Greiner M. 2014. Toxicokinetics of seven perfluoroalkyl sulfonic and carboxylic acids in pigs fed a contaminated diet. J Agric Food Chem. 62(28):6861-6870.

Oakes KD, Benskin JP, Martin JW, Ings JS, Heinrichs JY, Dixon DG, Servos MR. 2010. Biomonitoring of perfluorochemicals and toxicity to the downstream fish community of Etobicoke Creek following deployment of aqueous film-forming foam. Aquat Toxicol. 98(2):120-129.

O’Connell SG, Arendt M, Segars A, Kimmel T, Braun-McNeill J, Avens L, Schroeder B, Ngai L, Kucklick JR, Keller JM. 2010. Temporal and spatial trends of perfluorinated compounds in juvenile loggerhead sea turtles (Caretta caretta) along the east coast of the United States. Environ Sci Technol. 44(13):5202-5209.

Odinokov VN, Akhmetova VR, Savchenko RG, Bazunova MV, Fatykhov, Zapevalov AY. 1997. Ozonolysis of perfluoroalkenes and perfluorocycloalkenes. Rus Chem Bull. 46(6):1190-1191.

[OECD] Organisation for Economic Co-operation and Development. 2002. Co-operation on existing chemicals. Hazard assessment of perfluorooctane sulfonate (PFOS) and its salts. Environment Directorate. Joint meeting of the Chemicals Committee and the Working Party on Chemicals, Pesticides and Biotechnology. ENV/JM/RD(2002)17/FINAL (Unclassified).

[OECD] Organisation for Economic Co-operation and Development. 2018. Toward a new comprehensive global database of per- and polyfluoroalkyl substances (PFASs): Summary report on updating the OECD 2007 list of per- and polyfluoroalkyl substances (PFASs). OECD Environment, Health and Safety Publications Series on Risk Management; Paris, France.

[OECD] Organisation for Economic Co-operation and Development. 2021. Reconciling terminology of the universe of per- and polyfluoroalkyl substances: Recommendations and practical guidance [PDF]. Series on Risk Management No. 61. [accessed 2022 Jan 19].

Olivero-Verbel J, Tao L, Johnson-Restrepo B, Guette-Fernández J, Baldiris-Avila R, O’Byrne-Hoyos I, Kannan K. 2006. Perfluorooctanesulfonate and related fluorochemicals in biological samples from the north coast of Colombia. Environ Pollut. 142(2):367-372.

Omagamre EW, Ojo F, Zebelo SA, Pitula JS. 2020. Influence of perfluorobutanoic acid (PFBA) on the developmental cycle and damage potential of the beet armyworm Spodoptera exigua (Hübner) (Insecta: Lepidoptera: Noctuidae). Arch Environ Contam Toxicol. 79(4):500-507.

O’Rourke E, Hynes J, Losada S, Barber JL, Pereira MG, Kean EF, Hailer F, Chadwick EA. 2022. Anthropogenic drivers of variation in concentrations of perfluoroalkyl substances in otters (Lutra lutra) from England and Wales. Environ Sci Technol. 56(3):1675-1687.

Park H, Vecitis CD, Cheng J, Choi W, Mader BT, Hoffmann MR. 2009. Reductive defluorination of aqueous perfluorinated alkyl surfactants: Effects of ionic headgroup and chain length. J Phys Chem A. 113(4):690-696.

Park K, Barghi M, Lim J-E, Ko H-M, Nam H-Y, Lee S-I, Moon H-B. 2021. Assessment of regional and temporal trends in per- and polyfluoroalkyl substances using the Oriental Magpie (Pica serica) in Korea. Sci Total Environ. 793:148513.

Parolini M, Cappelli F, De Felice B, Possenti CD, Rubolini D, Valsecchi S, Polesello S. 2020. Within- and among-clutch variation of yolk perfluoroalkyl acids in a seabird from the Northern Adriatic Sea. Environ Toxicol Chem. 40(3):744-753.

Parsons JR, Sáez M, Dolfing J, de Voogt P. 2008. Biodegradation of perfluorinated compounds. In: Whitacre EM (editor). Reviews of environmental contamination and toxicology. Vol. 196. New York (NY): Springer.

Peden-Adams MM, Romano T, Rice CD, Hesseman L, EuDaly J, Muir D, Houde M, Bossart G, Fair P. 2004a. Immune function and clinical blood parameters correlate with perfluorinated alkyl acid concentrations in bottlenose dolphins. Poster presentation at the Society of Environmental Toxicology and Chemistry (SETAC) 25th Annual Meeting in North America; November 14–18, 2004; Portland, Oregon.

Peden-Adams MM, Kannan K, EuDaly JG, Hessemann M, Kucklick JR, Arendt MD, Maier PP, Segars AI, Whitaker JD, Keller JM. 2004b. Perfluorinated alkyl acids measured in sea turtle blood correlate to modulations in plasma chemistry values and immune function measurements. Poster presentation at the Society of Environmental Toxicology and Chemistry (SETAC) 25th Annual Meeting in North America; November 14–18, 2004; Portland, Oregon.

Penland TN, Cope GW, Kwak TJ, Strynar MJ, Grieshaber CA, Heise RJ, Sessions FW. 2020. Trophodynamics of per- and polyfluoroalkyl substances in the food web of a large Atlantic slope river. Environ Sci Technol. 54(11):6800-6811.

Persson S, Rotander A, Kärrman A, van Bavel B, Magnusson U. 2013. Perfluoroalkyl acids in subarctic wild male mink (Neovison vison) in relation to age, season and geographical area. Environ Int. 59:425-430.

Persson S, Magnusson U. 2015. Environmental pollutants and alterations in the reproductive system in wild male mink (Neovison vison) from Sweden. Chemosphere. 120:237-245.

Picard J-C, Munoz G, Duy SV, Sauvé S. 2021. Longitudinal and vertical variations of waterborne emerging contaminants in the St. Lawrence Estuary and Gulf during winter conditions. Sci Total Environ. 777:146073.

Pickard HM, Criscitiello AS, Spencer C, Sharp MJ, Muir DCG, DeSilva AO, Young CJ. 2018. Continuous non-marine inputs of per- and polyfluoroalkyl substances to the High Arctic: A multi-decadal temporal record. Atmos Chem Phys. 18(7):5045-5058.

Pickard HM, Criscitiello AS, Persaud D, Spencer C, Muir DCG, Lehnherr I, Sharp MJ, De Silva AO, Young CJ. 2020. Ice core record of persistent short-chain fluorinated alkyl acids: Evidence of the impact from global environmental regulations. Geophys Res Lett. 47(10):e2020GL087535.

Plassmann MM, Berger U. 2013. Perfluoroalkyl carboxylic acids with up to 22 carbon atoms in snow and soil samples from a ski area. Chemosphere. 91(6):832-837.

Point AD, Holsen TM, Fernando S, Hopke PK, Crimmins BS. 2021. Trends (2005–2016) of perfluoroalkyl acids in top predator fish of the Laurentian Great Lakes. Sci Total Environ. 778:146151.

Powley CR, George SW, Russell MH, Hoke RA, Buck RC. 2008. Polyfluorinated chemicals in a spatially and temporally integrated food web in the Western Arctic. Chemosphere. 70(4):664-672.

Propp VR, De Silva AO, Spencer C, Brown SJ, Catingan SD, Smith JE, Roy JW. 2021. Organic contaminants of emerging concern in leachate of historic municipal landfills. Environ Pollut. 276:116474.

Quinete N, Wu Q, Zhang T, Yun SH, Moreira I, Kannan K. 2009. Specific profiles of perfluorinated compounds in surface and drinking waters and accumulation in mussels, fish, and dolphins from southeastern Brazil. Chemosphere. 77(6):863-869.

Quinete N, Orata F, Maes A, Gehron M, Bauer K-H, Moreira I, Wilken R-D. 2010. Degradation studies of new substitutes for perfluorinated surfactants. Arch Environ Contam Toxicol. 59(1):20-30.

Randall DJ, Connell DW, Yang R., Wu SS. 1998. Concentrations of persistent lipophilic compounds in fish are determined by exchange across the gills, not through the food chain. Chemosphere. 37(7):1263-1270.

Reiner JL, O’Connell SG, Moors AJ, Kucklick JR, Becker PR, Keller JM. 2011. Spatial and temporal trends of perfluorinated compounds in beluga whales (Delphinapterus leucas) from Alaska. Environ Sci Technol. 45(19):8129-8136.

Ren J, Point A, Fakouri Baygi S, Fernando S, Hopke P. K, Holsen T. M, Lantry B, Weidel B, Crimmins BS. 2021. Bioaccumulation of perfluoroalkyl substances in a Lake Ontario food web. J Great Lakes Res. 48(2):315-325.

Ren J, Point AD, Fakouri Baygi S, Fernando S, Hopke PK, Holsen TM, Crimmins BS. 2022. Bioaccumulation of polyfluoroalkyl substances in the Lake Huron aquatic food web. Sci Total Environ. 819(1):152974.

Rericha Y, Cao D, Truong L, Simonich M, Field JA, Tanguay RL. 2021. Behavior effects of structurally diverse per- and polyfluoroalkyl substances in zebrafish. Chem Res Toxicol. 34(6):1409-1416.

Reth M, Berger U, Broman D, Cousins IT, Nilsson ED, McLachlan MS. 2011. Water-to-air transfer of perfluorinated carboxylates and sulfonates in a sea spray simulator. Environ Chem. 8:381-388.

Rich CD, Blaine AC, Hundal L, Higgins CP. 2014. Bioaccumulation of perfluoroalkyl acids by earthworms (Eisenia fetida) exposed to contaminated soils. Environ Sci Technol. 49(2):881-888.

Rigét F, Bossi R, Sonne C, Vorkamp K, Dietz R. 2013. Trends of perfluorochemicals in Greenland ringed seals and polar bears: Indications of shifts to decreasing trends. Chemosphere. 93(8):1607-1614.

Rijnders J, Bervoets L, Prinsen E, Eens M, Beemster GTS, AbdElgawad H, Groffen T. 2021. Perfluoroalkylated acids (PFAAs) accumulate in field-exposed snails (Cepaea sp.) and affect their oxidative status. Sci Total Environ. 790:148059.

Ritscher A, Wang Z, Scheringer M, Boucher JM, Ahrens L, Berger U, Bintein S, Bopp SK, Borg D, Buser AM, et al. 2018. Zürich statement on future actions on per- and polyfluoroalkyl substances (PFASs). Environ Health Perspect. 126(8): 84502.

Rodríguez-Jorquera IA, Colli-Dula RC, Kroll K, Jayasinghe BS, Parachu Marco MV, Silva-Sanchez C, Toor GS, Denslow ND. 2019. Blood transcriptomics analysis of fish exposed to perfluoro alkyls substances: Assessment of a non-lethal sampling technique for advancing aquatic toxicology research. Environ Sci Technol. 53(3):1441-1452.

Roos AM, Gamberg M, Muir D, Kärrman A, Carlsson P, Cuyler C, Lind Y, Bossi R, Rigét F. 2021. Perfluoroalkyl substances in circum-ArcticRangifer: Caribou and reindeer. Environ Sci Pollut Res Int. 29(16):23721-23735.

Rosal R, Rodea-Palomares I, Boltes K, Fernández-Piñas F, Leganés F, Petre A. 2010. Ecotoxicological assessment of surfactants in the aquatic environment: Combined toxicity of docusate sodium with chlorinated pollutants. Chemosphere. 81(2):288-293.

Rotander A, Kärmann A, van Bavel B, Polder A, Rigét F, Auðunsson GA, Víkingsson G, Gabrielsen GW, Bloch D, Dam M. 2012. Increasing levels of long-chain perfluorocarboxylic acids (PFCAs) in Arctic and North Atlantic marine mammals, 1984–2009. Chemosphere. 86(3):278-285.

Routti H, Krafft BA, Herzke D, Eisert R, Oftedal O. 2015. Perfluoroalkyl substances detected in the world’s southernmost marine mammal, the Weddell seal (Leptonychotes weddellii). Environ Pollut. 197:62-67.

Routti H, Gabrielsen GW, Herzke D, Kovacs KM, Lydersen C. 2016. Spatial and temporal trends in perfluoroalkyl substances (PFASs) in ringed seals (Pusa hispida) from Svalbard. Environ Pollut. 214:230-238.

Routti H, Atwood TC, Bechshoft T, Boltunov A, Ciesielski TM, Desforges J-P, Dietz R, Gabrielsen GW, Jenssen BM, Letcher RJ, et al. 2019a. State of knowledge on current exposure, fate and potential health effects of contaminants in polar bears from the circumpolar Arctic. Sci Total Environ. 664(10):1063-1083.

Routti H, Berg MK, Lille-Langøy R, Øygarden L, Harju M, Dietz R, Sonne C, Goksøyr A. 2019b. Environmental contaminants modulate the transcriptional activity of polar bear (Ursus maritimus) and human peroxisome proliferator-activated receptor alpha (PPARA). Sci Rep. 9:6918.

Rüdel H, Müller J, Jürling H, Bartel-Steinbach M, Koschorreck J. 2011. Survey of patterns, levels, and trends of perfluorinated compounds in aquatic organisms and bird eggs from representative German ecosystems. Environ Sci Pollut Res Int. 18(9):1457-1470.

Sáez M, de Voogt P, Parsons JR. 2008. Persistence of perfluoroalkylated substances in closed bottle tests with municipal sewage sludge. Environ Sci Pollut Res Int. 15(6):472-477.

Scheringer M, Trier X, Cousins IT, de Voogt P, Fletcher T, Wang Z, Webster TF. 2014. Helsingør statement on poly- and perfluorinated alkyl substances (PFASs). Chemosphere. 114:337-339.

Schiavone A, Corsolini S, Kannan K, Tao L, Trivelpiece W, Torres D Jr, Focardi S. 2009. Perfluorinated contaminants in fur seal pups and penguin eggs from South Shetland, Antarctica. Sci Total Environ. 407(12):3899-3904.

Schlabach M, Gabrielsen GW, Herzke D, Hanssen L, Routti H, Borgen A. 2017. Screening of PFAS and dechlorane compounds in selected Arctic top predators. Norway: Norwegian Environment Agency.

Scott BF, Moody CA, Spencer C, Small JM, Muir DCG, Mabury SA. 2006a. Analysis for perfluorocarboxylic acids/anions in surface waters and precipitation using GC-MS and analysis of PFOA from large-volume samples. Environ Sci Technol. 40(20):6405-6410.

Scott BF, Spencer C, Mabury SA, Muir DCG. 2006b. Poly and perfluorinated carboxylates in North American precipitation. Environ Sci Technol. 40(23):7167-7174.

Scott BF, Spencer C, Lopez E, Muir DCG. 2009. Perfluorinated alkyl acid concentrations in Canadian rivers and creeks. Water Qual Res J Can. 44(3):263-277.

[SDS] Safety Data Sheet. 2004a. Perfluorohexanoic acd [PDF]. Derbyshire (UK): Fluorochem Ltd. [accessed 2007 Jan 19].

[SDS] Safety Data Sheet. 2004b. Perfluoroheptanoic acid [PDF]. Derbyshire (UK): Fluorochem Ltd. [accessed 2007 Jan 19].

Senthilkumar K, Ohi E, Sajwan K, Takasuga T, Kannan K. 2007. Perfluorinated compounds in river water, river sediment, market fish, and wildlife samples from Japan. Bull Environ Contam Toxicol. 79(4):427-431.

Sha B, Johansson JH, Tunved P, Bohlin-Nizzetto P, Cousins IT, Salter ME. 2022. Sea spray aerosol (SSA) as a source of perfluoroalkyl acids (PFAAs) to the atmosphere: Field evidence from long-term air monitoring. Environ Sci Technol. 56(1):228-238.

Sharp S, Sardiña P, Metzeling L, McKenzie R, Leahy P, Menkhorst P, Hinwood A. 2021. Per‐ and polyfluoroalkyl substances in ducks and the relationship with concentrations in water, sediment, and soil. Environ Toxicol Chem. 40(3):846-858.

Shaw SD, Berger ML, Brenner D, Kannan K. 2006. Perfluorooctane sulfonate and related perfluorinated hydrocarbons in harbor seals (Phoca vitulina concolor) from the Northwest Atlantic. Poster presentation at Poster presentation at the Society of Environmental Toxicology and Chemistry (SETAC).

Shi Y, Wang J, Pan Y, Cai Y. 2012. Tissue distribution of perfluorinated compounds in farmed freshwater fish and human exposure by consumption. Environ Toxicol Chem. 31(4):717-723.

Simonnet-Laprade C, Budzinski H, Babut M, Le Menach K, Munoz G, Lauzent M, Ferrari BJD, Labadie P. 2019a. Investigation of the spatial variability of poly- and perfluoroalkyl substance trophic magnification in selected riverine ecosystems. Sci Total Environ. 636:393-401.

Simonnet-Laprade C, Budzinski H, Maciejewski K, Le Menach K, Santos R, Alliot F, Goutte A, Labadie P. 2019b. Biomagnification of perfluoroalkyl acids (PFAAs) in the food web of an urban river: Assessment of the trophic transfer of targeted and unknown precursors and implications. Environ Sci: Processes Impacts. 21(11):1864-1874.

Smithwick M, Muir DCG, Mabury SA, Solomon KR, Martin JW, Sonne C, Born EW, Letcher RJ, Dietz R. 2005a. Perfluoroalkyl contaminants in liver tissue from East Greenland polar bears (Ursus maritimus). Environ Toxicol Chem. 24(4):981-986.

Smithwick M, Mabury SA, Solomon KR, Sonne C, Martin JW, Born EW, Dietz R, Derocher AE, Letcher RJ, Evans TJ, et al. 2005b. Circumpolar study of perfluoroalkyl contaminants in polar bears (Ursus maritimus). Environ Sci Technol. 39(15):5517-5523.

Solé M, Lacorte S, Vinyoles D. 2021. Biochemical aspects of susceptibility to stressors in two small cyprinids Squalius laietanus and Barbus meridionalis from the NW Mediterranean. Comp Biochem Physiol C Toxicol Pharmacol. 242:108940.

Sonne C. 2010. Health effects from long-range transported contaminants in Arctic top predators: An integrated review based on studies of polar bears and relevant model species. Environ Int. 36(5):461-491.

Sonne C, Bossi R, Dietz R, Leifsson PS, Rigét FF, Born EW. 2008. Potential correlation between perfluorinated acids and liver morphology in East Greenland polar bears (Ursus maritimus). Toxicol Environ Chem. 90(2):275-283.

Stevenson CN, MacManus-Spencer LA, Luckenbach T, Luthy RG, Epel D. 2006. New perspectives on perfluorochemical ecotoxicology: Inhibition and induction of an efflux transporter in the marine mussel, Mytilus californianus. Environ Sci Technol. 40(17):5580-5585.

Stockholm Convention. 2018. Perfluorohexane sulfonic acid (PFHxS), its salts and PFHxS-related compounds. DRAFT Risk Profile. Persistent Organics Pollutant Review Committee.

Styler SA, Myers AL, Donaldson DJ. 2013. Heterogeneous photooxidation of fluorotelomer alcohols: A new source of aerosol-phase perfluorinated carboxylic acids. Environ Sci Technol. 47(12):6358-6367.

Sun J, Letcher RJ, Eens M, Covaci A, Fernie KJ. 2020. Perfluoroalkyl acids and sulfonamides and dietary, biological and ecological associations in peregrine falcons from the Laurentian Great Lakes Basin, Canada. Environ Res. 191:110151.

Sun J, Letcher RJ, Waugh CA, Jaspers VLB, Covaci A, Fernie KJ. 2021. Influence of perfluoroalkyl acids and other parameters on circulating thyroid hormones and immune-related microRNA expression in free-ranging nestling peregrine falcons. Sci Total Environ. 770(20):145346.

Sundström M, Chang S-C, Noker PE, Gorman GS, Hart JA, Ehresman DJ, Bergman Å, Butenhoff JL 2012. Comparative pharmacokinetics of perfluorohexanesulfonate (PFHxS) in rats, mice, and monkeys. Reprod Toxicol. 33(4):441-451.

Suski JG, Salice CJ, Chanov MK, Ayers J, Rewerts J, Field J. 2021. Sensitivity and accumulation of perfluorooctanesulfonate and perfluorohexanesulfonic acid in fathead minnows (Pimephales promelas) exposed over critical life stages of reproduction and development. Environ Toxicol Chem. 40(3):811-819.

Szabo D, Nuske MR, Lavers JL, Shimeta J, Green MP, Mulder RA, Clarke BO. 2022. A baseline study of per- and polyfluoroalkyl substances (PFASs) in waterfowl from a remote Australian environment. Sci Total Environ. 812:152528.

Taniyasu S, Kannan K, Horii Y, Yamashita N. 2002. The first environmental survey of perfluorooctane sulfonate (PFOS) and related compounds in Japan. Organohalogen Compd. 59:311-314.

Taniyasu S, Kannan K, Horii Y, Hanari N, Yamashita N. 2003. A survey of perfluorooctane sulfonate and related perfluorinated organic compounds in water, fish, birds, and humans from Japan. Environ Sci Technol. 37(12):2634-2639.

Taniyasu S, Yamashita N, Yamazaki E, Petrick G, Kannan K. 2013. The environmental photolysis of perfluorooctanesulfonate, perfluorooctanoate, and related fluorochemicals. Chemosphere. 90(5):1686-1692.

Tartu S, Gabrielsen GW, Blévin P, Ellis H, Bustnes JO, Herzke D, Chastel O. 2014. Endocrine and fitness correlates of long-chain perfluorinated carboxylates exposure in Arctic breeding black-legged kittiwakes. Environ Sci Technol. 48(22):13504-13510.

Tartu S, Aars J, Andersen M, Polder A, Bourgeon S, Merkel B, Lowther AD, Bytingsvik J, Welker JM, Derocher AE, et al. 2018. Choose your poison—Space-use strategy influences pollutant exposure in Barents Sea polar bears. Environ Sci Technol. 52(5):3211-3221.

Taylor MD, Johnson DD. 2016. Preliminary investigation of perfluoroalkyl substances in exploited fishes of two contaminated estuaries. Mar Pollut Bull. 111(1-2):509-513.

Teunen L, Bervoets L, Belpaire C, De Jonge M, Groffen T. 2021. PFAS accumulation in indigenous and translocated aquatic organisms from Belgium, with translation to human and ecological health risk. Environ Sci Eur. 33(1):39.

Tsang W., Burgess DR Jr, Babushok V. 1998. On the incinerabilty of highly fluorinated organic compounds. Combust Sci Technol. 139(1):385-402.

Tseng C-L, Liu L-L, Chen C-M, Ding W-H. 2006. Analysis of perfluorooctane sulfonate and related fluorochemicals in water and biological tissue samples by liquid chromatography-ion trap mass spectrometry. J Chromatogr A. 1105(1-2):119-126.

Ulhaq M, Orn S, Carlsson G, Morrison DA, Norrgren L. 2013. Locomotor behaviour in zebrafish (Danio rerio) larvae exposed to perfluoroalkyl acids. Aquat Toxicol. 144-145:332-340.

Van de Vijver KI, Hoff P, Das K, Brasseur S, Van Dongen W, Esmans E, Reijnders P, Blust R, De Coen W. 2005. Tissue distribution of perfluorinated chemicals in harbor seals (Phoca vitulina) from the Dutch Wadden Sea. Environ Sci Technol. 39(18):6978-6984.

Van den Heuvel-Greve M, Leonards P, Brasseur S, Kotterman M, Zabel A, Vethaak D. 2009. Bioaccumulation of perfluorinated compounds in a harbor seal food web of the Westerschelde, the Netherlands: A field study. Poster presentation at the Society of Environmental Toxicology and Chemistry (SETAC) 30th Annual Meeting in North America; November 19–23, 2009; New Orleans, LA.

Verreault J, Houde M, Gabrielsen GW, Berger U, Haukås M, Letcher RJ, Muir DCG. 2005. Perfluorinated alkyl substances in plasma, liver, brain, and eggs of glaucous gulls (Larus hyperboreus) from the Norwegian Arctic. Environ Sci Technol. 39(19):7439-7445.

Verreault J, Berger U, Gabrielsen GW. 2007. Trends of perfluorinated alkyl substances in herring gull eggs from two coastal colonies in northern Norway: 1983-2003. Environ Sci Technol. 41(19):6671-6677.

Vongphachan V, Cassone CG, Wu D, Chiu S, Crump D, Kennedy SW. 2011. Effects of perfluoroalkyl compounds on mRNA expression levels of thyroid hormone-responsive genes in primary cultures of avian neuronal cells. Toxicol Sci. 120(2):392-402.

Wang N, Szostek B, Buck RC, Folsom PW, Sulecki LM, Capka V, Berti WR, Gannon JT. 2005. Fluorotelomer alcohol biodegradation – Direct evidence that perfluorinated carbon chains breakdown. Environ Sci Technol. 39(19):7516-7528.

Wang Y, Yeung LWY, Taniyasu S, Yamashita N, Lam JCW, Lam PKS. 2008. Perfluorooctane sulfonate and other fluorochemicals in waterbird eggs from South China. Environ Sci Technol. 42(21):8146-8151.

Wang N, Szostek B, Buck RC, Folsom PW, Sulecki LM, Gannon JT. 2009. 8-2 fluorotelomer alcohol aerobic soil biodegradation: Pathways, metabolites, and metabolite yields. Chemosphere. 75(8):1089-1096.

Wang N, Liu J, Buck RC, Folsom PW, Sulecki LM, Wolstenholme B, Panciroli PK, Bellin CA. 2010. Aerobic soil biotransformation of fluorotelomer alcohols. Poster presentation at PFAA Days III; North Carolina, USA.

Wang Z, MacLeod M, Cousins IT, Scheringer M, Hungerbühler K. 2011a. Using COSMOtherm to predict physicochemical properties of poly- and perfluorinated alkyl substances (PFASs). Environ Chem. 8(4):389-398.

Wang N, Liu J, Buck RC, Korzeniowski SH, Wolstenholme BW, Folsom PW, Sulecki LM. 2011b. 6:2 fluorotelomer sulfonate aerobic biotransformation in activated sludge of waste water treatment plants. Chemosphere. 82(6):853-858.

Wang N, Buck RC, Szostek B, Sulecki LM, Wolstenholme BW. 2012. 5:3 polyfluorinated acid aerobic biotransformation in activated sludge via novel “one-carbon removal pathways.” Chemosphere. 87(5):527-534.

Wang J, Zhang Y, Zhang F, Yeung LWY, Taniyasu S, Yamazaki E, Wang R, Lam PKS, Yamashita N, Dai J. 2013a. Age- and gender-related accumulation of perfluoroalkyl substances in captive Chinese alligators (Alligator sinensis). Environ Pollut. 179:61-67.

Wang Z, Cousins IT, Scheringer M, Hungerbühler K. 2013b. Fluorinated alternatives to long-chain perfluoroalkyl carboxylic acids (PFCAs), perfluoroalkane sulfonic acids (PFSAs) and their potential precursors. Environ Int. 60:242-248.

Wang Y, Niu J, Zhang L, Shi J. 2014. Toxicity assessment of perfluorinated carboxylic acids (PFCAs) towards the rotifer Brachionus calyciflorus. Sci Total Environ. 491-492:266-270.

Wang Y, Yao J, Dai J, Ma L, Liu D, Xu H, Cui Q, Ma J, Zhang H. 2020. Per- and polyfluoroalkyl substances (PFASs) in blood of captive Siberian tigers in China: Occurrence and associations with biochemical parameters. Environ Pollut. 265(Pt B):114805.

Washington JW, Ellington JJ, Jenkins TM, Evans JJ, Yoo H, Hafner SC. 2009. Degradability of an acrylate-linked, fluorotelomer polymer in soil. Environ Sci Technol. 43(17):6617-6623.

Waterland RL, Dobbs KD. 2007. Atmospheric chemistry of linear perfluorinated aldehydes: Dissociation kinetics of CnF2n+1CO radicals. J. Phys Chem A. 111(13):2555-2562.

Wen W, Xia X, Hu D, Zhou D, Wang H, Zhai Y, Lin H. 2017. Long-chain perfluoroalkyl acids (PFAAs) affect the bioconcentration and tissue distribution of short-chain PFAAs in zebrafish (Danio rerio). Environ Sci Technol. 51(21):12358-12368.

Wilkinson BP, Robuck AR, Lohmann R, Pickard HM, Jodice PGR. 2022. Urban proximity while breeding is not a predictor of perfluoroalkyl substance contamination in the eggs of brown pelicans. Sci Total Environ. 803:150110.

Wille K, Vanden Bussche J, Noppe H, De Wulf E, Van Caeter P, Janssen CR, De Brabander HF, Vanhaecke L. 2010. A validated analytical method for the determination of perfluorinated compounds in surface-, sea-, and sewagewater using liquid chromatography coupled to time-of-flight mass spectrometry. J Chromatogr A. 1217(43):6616-6622.

Wille K, Kiebooms JAL, Claessens M, Rappé K, Vanden Bussche J, Noppe H, Van Praet N, De Wulf E, Van Caeter P, Janssen CR, et al. 2011. Development of analytical strategies using U-HPLC-MS/MS and LC-ToF-MS for the quantification of micropollutants in marine organisms. Anal Bioanal Chem. 400(5):1459-1472.

Wu Y, Simon KL, Best DA, Bowerman W, Venier M. 2020. Novel and legacy per- and polyfluoroalkyl substances in bald eagle eggs from the Great Lakes region. Environ Pollut. 260:113811.

Xie W, Bothun GD, Lehmler H-J. 2010. Partitioning of perfluorooctanoate into phosphatidylcholine bilayers is chain length-independent. Chem Phys Lipids. 163(3):300-308.

Yang L, Tian S, Zhu L, Liu Z, Zhang Y. 2012. Bioaccumulation and distribution of perfloroalkyl acids in seafood products from Bohai Bay, China. Environ Toxicol Chem. 31(9):1972-1979.

Ye X, Schoenfuss HL, Jahns ND, Delinsky AD, Strynar MJ, Varns J, Nakayama SF, Helfant L, Lindstrom AB. 2008a. Perfluorinated compounds in common carp (Cyprinus carpio) fillets from the Upper Mississippi River. Environ Int. 34(7):932-938.

Ye X, Strynar MJ, Nakayama SF, Varns J, Helfant L, Lazorchak J, Lindstrom AB. 2008b. Perfluorinated compounds in whole fish homogenates from the Ohio, Missouri, and Upper Mississippi Rivers, USA. Environ Pollut. 156(3):1227-1232.

Yeung LWY, Miyake Y, Wang Y, Taniyasu S, Yamashita N, Lam PKS. 2009. Total fluorine, extractable organic fluorine, perfluorooctane sulfonate and other related fluorochemicals in liver of Indo-Pacific humpback dolphins (Sousa chinensis) and finless porpoises (Neophocaena phocaenoides) from South China. Environ Pollut. 157(1):17-23.

Yeung LWY, De Silva AO, Loi EIH, Marvin CH, Taniyasu S, Yamashita N, Mabury SA, Muir DCG, Lam PKS. 2013. Perfluoroalkyl substances and extractable organic fluorine in surface sediments and cores from Lake Ontario. Environ Int. 59:389-397.

Yoo H, Kannan K, Kim SK, Lee KT, Newsted JL, Giesy JP. 2008. Perfluoroalkyl acids in the egg yolk of birds from Lake Shihwa, Korea. Environ Sci Technol. 42(15):5821-5827.

Young CJ, Hurley MD, Wallington TJ, Mabury SA. 2008. Atmospheric chemistry of 4:2 fluorotelomer iodide (n-C4F9CH2CH2I): Kinetics and products of photolysis and reaction with OH radicals and CI atoms. J Phys Chem A. 112(51):13542-13548.

Young WM, South P, Begley TH, Noonan GO. 2013. Determination of perfluorochemicals in fish and shellfish using liquid chromatography-tandem mass spectrometry. J Agric Food Chem. 61(46):11166-11172.

Zhang T, Sun HW, Wu Q, Zhang XZ, Yun SH, Kannan K. 2010. Perfluorochemicals in meat, eggs and indoor dust in China: Assessment of sources and pathways of human exposure to perfluorochemicals. Environ Sci Technol. 44(9):3572-3579.

Zhang T, Sun H, Lin Y, Wang L, Zhang X, Liu Y, Geng X, Zhao L, Li F, Kannan K. 2011. Perfluorinated compounds in human blood, water, edible freshwater fish, and seafood in China: Daily intake and regional differences in human exposures. J Agric Food Chem. 59(20):11168-11176.

Zhang Y, Meng W, Guo C, Xu J, Yu T, Fan W, Li L. 2012a. Determination and partitioning behaviour of perfluoroalkyl carboxylic acids and perfluorooctanesulfonate in water and sediment from Dianchi Lake, China. Chemosphere. 88(11):1292-1299.

Zhang W, Zhang Y, Zhang H, Wang J, Cui R, Dai J. 2012b. Sex differences in transcriptional expression of FABPs in zebrafish liver after chronic perfluorononanoic acid exposure. Environ Sci Technol. 46(9):5175-5182.

Zhang S, Szostek B, McCausland PK, Wolstenholme BW, Lu X, Wang N, Buck RC. 2013. 6:2 and 8:2 fluorotelomer alcohol anaerobic biotransformation in digester sludge from a WWTP under methanogenic conditions. Environ Sci Technol. 47(9):4227-4235.

Zhao L, Zhu L, Yang L, Liu Z, Zhang Y. 2012. Distribution and desorption of perfluorinated compounds in fractionated sediments. Chemosphere. 88(11):1390-1397.

Zhao S, Zhu L, Liu L, Liu Z, Zhang Y. 2013a. Bioaccumulation of perfluroalkyl carboxylates (PFCAs) and perfluoroalkane sulfonates (PFSAs) by earthworms (Eisenia fetida) in soil. Environ Pollut. 179:45-52.

Zhao L, Folsom PW, Wolstenholme BW, Sun H, Wang N, Buck RC. 2013b. 6:2 fluorotelomer alcohol biotransformation in an aerobic river sediment system. Chemosphere. 90(2):203-209.

Zhao L, Zhang Y, Fang S, Zhu L, Liu Z. 2014. Comparative sorption and desorption behaviors of PFHxS and PFOS on sequentially extracted humic substances. J Environ Sci (China). 26(12):2517-2525.

Zhou Z , Shi Y, Li W, Xu L, Cai Y. 2012. Perfluorinated compounds in surface water and organisms from Baiyangdian Lake in North China: Source profiles, bioaccumulation and potential risk. Bull Environ Contam Toxicol. 89(3):519-524.

Appendix A. Non-exhaustive list of SC-PFCAs, their salts and their precursors

Abbreviations: DSL, Domestic Substances List; NDSL, Non-Domestic Substances List.

^a CAS RN idenitified via DSL search engine and/or US EPA Chemistry Dashboard.

^b Substances not appearing on the DSL are not used commercially in Canada above trigger quantities specified in the New Substances Notification Regulations (Chemicals and Polymers). Substances listed on the NDSL are subject to notification and requirements set out in the New Substances Notification Regulations (Chemicals and Polymers), but with reduced information requirements. Substances on the NDSL are those that are on the US Toxic Substances Control Act (TSCA) inventory.

Appendix B. Non-exhaustive list of SC-PFSAs, their salts and precursors

Abbreviations: DSL, Domestic Substances List; NDSL, Non-Domestic Substances List; TSCA, Toxic Substance Control Act (United States); NA, not available

^a CAS RN idenitified via DSL search engine and/or US EPA Chemistry Dashboard

^b Substances not appearing on the DSL are not used commercially in Canada above trigger quantities specified in the New Substances Notification Regulations (Chemicals and Polymers). Substances listed on the NDSL are subject to notification and requirements set out in the New Substances Notification Regulations (Chemicals and Polymers) but with reduced information requirements. Substances on the NDSL are those found on the US Toxic Substances Control Act (TSCA) inventory.

^c Also within the scope of Canada’s Priority Substance List 1 assessment of “Chromium and its compounds”.

Appendix C. Non-exhaustive list of LC-PFSAs and their salts

Abbreviations: DSL, Domestic Substances List; NDSL, Non-Domestic Substances List.

^a CAS RN idenitified via DSL search engine and/or US EPA Chemistry Dashboard.

^b Substances not appearing on the DSL are not used commercially in Canada above trigger quantities specified in the New Substances Notification Regulations (Chemicals and Polymers). Substances listed on the NDSL are subject to notification and requirements set out in the New Substances Notification Regulations (Chemicals and Polymers) but with reduced information requirements. Substances on the NDSL are those that are on the US Toxic Substances Control Act (TSCA) inventory.

Appendix D. Empirical evidence of precursors for SC-PFCAs and SC-PFSAs

Note: Most available empirical data identifying precursors are limited to the short-chain PFCAs. There is one empirical study identifying a precursor for the short-chain PFSAs. There is no empirical evidence identifying precursors for the long-chain PFSAs. Empirical evidence for precursors includes that for fluorotelomer olefins (FTO) and N-alkyl perfluoroalkylsulfonamidoethanols (PFSE). The estimated “atmospheric lifetimes” of FTOHs (12 to 20 days), FTOs (8 days), and a major degradation product of PFSEs (20 to 50 days) permit their transport to remote regions. Photoreactor studies have indicated that PFSEs may contribute to the observed environmental burden of both PFSAs and PFCAs substances (Waterland and Dobbs 2007).

Appendix E. Available empirical physical-chemical properties for SC-PFCAs

Abbreviations: pK_a, acid dissociation constant; Kow, octanol-water partition coefficient; Koc, organic carbon-water partition coefficient; Kd, equilibrium dissociation constant

† As SC-PFCAs are considered to have surface-active properties, preferentially partition to proteins, and are ionizing, estimates of log K_ow and/or using log K_ow on the basis of the neutral form can result in unreliable predictions for bioaccumulation

Appendix F. Available empirical physical-chemical properties for SC-PFSAs

Abbreviations: pK_a, acid dissociation constant; Kow, octanol-water partition coefficient; Koc, organic carbon-water partition coefficient; Kd, equilibrium dissociation constant

† As SC-PFSAs are considered to have surface-active properties, preferentially partition to proteins, and are ionizing, estimates of log K_ow and/or using log K_ow on the basis of the neutral form can result in unreliable predictions for bioaccumulation.