Per-/polyfluorinated compounds

Per- and polyfluoroalkyl substances (PFASs) have been in use since the 1950ies as ingredients of intermediates of surfactants and surface protectors for assorted industrial and consumer applications. Within the past decade, several long-chain perfluoroalkyl acids have been recognized as extremely persistent, bioaccumulative and toxic. Many have been detected globally in the environment, biota, food items, and in humans (OECD, 2015). It has been recognised more recently that shorter chain PFASs increasingly used as alternatives are also very persistent and thus very mobile in the environment, presumably leading to ground water contamination in future. To date many known and unknown alternatives of the so far regulated PFASs are used worldwide leading to environmental contamination und increasing human body burdens.

Based on the huge amount of available PFAS on the market and the knowledge gaps on identity, toxicity and uses (of the alternatives), the listing of chemicals in categories A-E is an attempt to categorize possibly relevant substances that contribute to the overall PFAS burden in humans. Several substances are listed in category A due to their restriction as PFOS- and PFOA-related substances, although limited or no HBM data are available. Efforts should be made to improve the methods to detect the broader spectrum of Category A substances. However, the priority for future HBM research should cover PFOS and PFOA alternatives with high production volume, wide dispersive use and identified or suspected hazardous properties which qualifies for Substance of Very High Convern (SVHC) identification.

Category A: PFOA, PFOS, PFNA, PFDA, PFU(n)DA, PFDoDA, PFTrDA, PFTeDA, PFHxS, FOSA, PFOSA, n-MeFOSA, N-Et-FOSAA, Et-PFOSA-AcOH, Et-FOSAA, N-EtFOSA, SULFLURAMID, N-EtFOSE, N-MeFOSE, 8:2 diPAP, 6:2/8:2 diPAP, 8:2 monoPAP

Category B: PFBA, PFPeA, PFHxA, PFHpA, PFBS, PFHpS, PFDS, N-Me-PFOSA-AcOH, Me-FOSAA, 6:2 FTSA, H4PFOS, THPFOS, 8:2 FTSA, PFODA, PfHxDA

Category C: PFBA, PFPeA, PFHxA, PFHpA, PFBS, PFHpS, PFDS, N-Me-PFOSA-AcOH, Me-FOSAA, 6:2 FTSA, H4PFOS, THPFOS, 8:2 FTSA, PFODA, PfHxDA

Category D: HFPO, PFCHS, 6:2/8:2 diPAP, 8:2 monoPAP, PFOPA, Perfluorinated Siloxane, FL16.119

Category E: 6:2 FTCA, 8:2 FTCA, 10:2 FTCA, PFECA, FBSA, MeFBSE, 6:2 PAP, 6:2 diPAP, PFHxPA, PFDPA, C8/C10 PFPiA, Denum SH, Krytox, Fomblin Z-DIAC, C3; C15-C20 PFCA, C3, C15-C20 PFSA, TFEE-5, polymers: PTFE, PVDF, PVF, TFE, HFP

One of the major societal concerns is the irreversibility of contamination, together with endocrine disrupting effects, carcinogenicity, toxicity to reproduction, effects on immune system and on lipid metabolism for a broad range of PFASs.

A recent briefing provided by Chemtrust points out that children are currently not sufficiently being protected from chemicals that can disrupt brain development; they list per- and polyfluorinated compounds as one of the chemicals substance groups of concern (Chemtrust, 2017a). Chemtrust also raises the issue of use of PFASs as food contact materials and refers to a report on the implementation of the Food Contact Materials Regulation of the European Parliament which states that action at EU level is needed to address the lack of EU specific measures and the gaps in risk assessment, traceability, compliance and control (EU parliament, 2016) and an assessment of the Joint Research Center on the regulatory and market situation of the non-harmonised food contact materials in the EU (Simoneau et al., 2016). Also, European consumer organisations call for action on fluorinated compounds in fast food packaging (BEUC, 2017).

PFASs bind to proteins and partition to phospholipids. The elimination kinetics are highly species dependent, with humans showing the longest half-lives of up to e.g. 8.5 years for perfluorohexane sulfonic acid (PFHxS). A recent publication reports an estimated elimination range of 10.1 to 56.4 years – median 15.3 years for chlorinated polyfluoroalkyl ether sulfonic acids [Cl-PFESAs] (Shi et al., 2016). The CLP human health hazard classifications of the different substances are depicted in table 1. Substances which are best-known – perfluorooctane sulfonate (PFOS) and perfluorooctanoic acid (PFOA) – are classified as carcinogenic (Carc. 2, suspected human carcinogens, such as kidney and testicular), toxic for reproduction (Repr. 1B, presumed human reproductive toxicants; Lact., may cause harm to breast-fed children), toxic to specific target organs (STOT RE 1, specific target organ toxicity – repeated exposure) and acute toxic (Acute Tox. 3-4) for different exposure routes. PFOS and PFOA belong to the so called long-chain perfluorinated compounds, which refers to perfluorocarboxylic acids with carbon chain lengths of 8 and higher, including PFOA; perfluoroalkyl sulfonates with carbon chain lengths of 6 and higher, including PFHxS and PFOS; and precursors of these substances that may be produced or may be present in products. A recent report showed that in product samples the detected individual PFAS constituted only a very minor part of the total organic fluorine (TOF), illustrating large data gaps in the current knowledge which PFASs that are being used in these products (Borg, 2017).

Several long-chain compounds beside PFOS and PFOA have also been identified as toxic to reproduction; further endpoints concern carcinogenicity, liver toxicity, neurotoxicity and immunotoxicity. Whether numerous other non-regulated PFASs show similar toxicity is currently less well established. In many cases data availability is poor and therefore no classification is possible. However, persistence is assumed to concerns largely all PFASs by reason of the extreme strength and stability of the carbon-fluorine bonds.

For PFOS and PFOA adverse effects on thyroid metabolism and lipid metabolism have been reported in a multitude of epidemiological studies suggesting endocrine disrupting potential (Barry et al., 2013).

Additional concerns include increased risk of miscarriage, reduced birth weight, increased weight in adult life, and reduced fertility among offspring as a result of early life exposures (Halldorsson et al., 2012; Jensen et al., 2015; Joensen et al., 2013; Timmermann et al., 2014). Postnatal exposures have also been associated with thyroid hormone imbalances and reduced immune response to vaccination (Grandjean and Budtz-Jørgensen, 2013). The US National Toxicology programme has listed both, PFOA and PFOS, as presumed to be an immune hazard to humans (NTP, 2016).

Grandjean and Clapp (2015) documented carcinogenicity, immunotoxicity and developmental toxicity of PFOA and highlighted the endocrine disrupting effects. A recent publication describes prenatal exposure to perfluoroalkyl substances and reduction in anogenital distance in girls at 3 months of age in a Danish mother-child cohort (Lind et al., 2017).

Since PFOS and PFOA can still be measured in highest concentrations in biota and in humans, exerting similar toxic effects along with and similar to a range of long-chain PFASs measured in blood, together with a range of unidentified PFASs the possibility of mixture effects is very high.

Since PFOS and PFOA can still be measured in highest concentrations in biota and in humans, exerting similar toxic effects along with and similar to a range of long-chain PFASs measured in blood, together with a range of unidentified PFASs the possibility of mixture effects is very high.

Trends in production volume/environmental concentrations

A minor part of the family of PFASs are perfluoroalkyl acids (PFAA), perfluoroalkylcarboxylic acids (PFCA), perfluoroalkane sulfonic acids (PFSA), compounds derived from perfluoroalkane sulfonyl fluoride (PASF), fluorotelomer (FT)-based compounds and per- and polyfluoroalkylether (PFPE)-based compounds. Another presumably major part are polymers (fluoropolymers (FPs), side-chain fluorinated polymers and perfluorpolyethers (PFPEs) (OECD, 2013). According to KEMI (2017) there are 2,817 PFASs on the market. For only 15 % of them adequate data are available; whereas for 40% data are missing (KEMI, 2017). Many fluorinated substances enter the EU through the import of articles (e.g. textiles) and for the most part these are not monitored (KEMI, 2015) providing an indirect exposure source. The lack of data concerns identification, use and exposure beside from toxicity and ecotoxicity.

Environmental behaviour: half-lives in environment/transport

Perfluoroalkyl and perfluoroether moieties of PFASs are highly persistent under environmental conditions. All PFASs ultimately degrade into highly persistent end products. PFASs are ubiquitously detected in the environment. Contamination of the drinking water resources as environmental health thread has been reported for PFASs e.g from the Veneto Region in Italy (WHO, 2017) but also from Sweden (Banzhaf et al, 2017) and other European countries. Whereas most data are available for the small group of long-chain PFASs, non-reversible environmental exposure has to be considered for a by far larger group. Recent data demonstrate considerable exposure of alternatives such as GenX in the drinking water (e.g. Gebbink et al., 2017).

There are also concerns about short-chain PFASs, which are assumed to be less bioaccumulative but very persistent and mobile contaminants found in drinking water and food, including vegetables (Hedlund, 2016, Danish EPA, 2015).

Human-related exposure sources and uses, human exposure routes

Humans can be exposed directly (via diet, drinking water, consumer products, etc.) and indirectly through transformation of «precursor substances» such as polyfluoroalkyl phosphate esters (PAPs), fluorotelomer alcohols (FTOHs), fluorotelomer iodides (FTIs) and fluorotelomer acrylate monomers (FTAcs). These fluorotelomer-based substances biotransform to yield PFCAs, yet also form bioactive intermediate metabolites, which have been observed to be more toxic than their corresponding PFCAs (e.g. Rand et al., 2017).The precursor contribution to PFASs daily exposures was recently estimated for a high exposure scenario to contribute up to >50% to individual PFCAs like PFOA or PFDA, whereas it is considerable lower up to 10% for e.g. PFOS for a low exposure scenario (Gebbink et al., 2015).

Human biomonitoring (HBM) data availability

Human exposures to PFASs have been reported in numerous studies in Europe and worldwide. Most of these studies were focused on blood or breast milk concentrations of PFOS and PFOA, while others also included PFBS, PFHxS, PFDS, PFBA, PFPeA, PFHxA, PFHpA, PFNA, PFDA, PFUdA, PFDoA, PFTrDA, PFTeDA, FOSA, MeFOSA, N-EtFOSA, N-EtFOSAA and diPAP. Contrary, human exposure to e.g. 8:2 diPAP, 6:2 diPAP, 8:2 PAP, 6:2 PAP, PFDPA, PFOPA, PFHxPA or ADONA has been addressed to a small extent only; the majority of new fluorinated compounds that enter the market as replacements has not been measured in human matrices yet. Concerning PFOS, the effectiveness evaluation under the UNEP Stockholm Convention concluded that for human matrices from Western Europe, Canada, Australia and Asia-Pacific countries levels seem gradually declining. Although PFOS is measured at low concentrations in human breast milk and is detected in higher concentrations in human blood, there are good correlations between the measurement results in these two matrices (UNEP, 2016).

In several studies, time trends of PFAS exposure in European countries were investigated. According to Axmon et al. (2014) investigating plasma samples from 1987-2007 in Sweden there was a peak in PFOS and PFOA blood concentrations around 2000 and increasing PFHxS, PFNA, PFDA and PFUnDA concentrations within the overall study period. Also Glynn et al. (2012) reported increasing concentrations of PFBS, PFHxS, PFNA and PFDA in Swedish breast milk samples between 1996 and 2010. This is also in line with the study from Gebbink et al. (2015) reporting increasing trends in pooled serum samples from Sweden for PFHxS, PFNA, PFDA, PFUnDA, PFODA and PFTrDA. Analyses of serum samples from Norway from 1979 to 2007 documented decreasing concentrations of PFOS and PFOA from 2001 onwards, whereas PFNA, PFDA, PFUnDA were increasing, and for PfHxS and PfHpS no trend could be observed (Nøst et al., 2014). In Denmark, concentrations of seven PFASs (PfHXS, PfHpS, PFOS, PFOA, PfNA, PfDA, PfUnDA) decreased in the period 2008-2013 (Bjerregaard-Olsen et al., 2016). Schröter-Kermani et al. (2013) reported decreasing concentrations from 2001 onwards for PFOS, from 2008 for PFOA and from 2005 for PfHxS and stable concentrations for PFNA in samples from Germany from 1982 to 2010. Also Yeung et al. (2013a, 2013b) observed decreasing concentrations for PFOA after 2000, and increasing concentrations for PFNA, PFDA and PFUnDA. No significant trend was observed for PFHXS and 8:2 di-PAPs. Isomer profiling of perfluorinated substances can be used as a tool for source tracking.

There are major knowledge gaps on fluorinated alternatives currently used by industry; these knowledge gaps concern production volumes, use, fate and behaviour, and toxicity (Danish EPA, 2013; Wang et al., 2013, 2016, 2017). Known fluorinated alternatives can be categorized into two groups, namely:

1. shorter-chain homologues of long-chain PFAAs and their precursors,
2. functionalized perfluoropolyethers (PFPEs), in particular perfluoroether carboxylic and sulfonic acids (PFECAs such as ADONA and GenX and PFESAs such as F-53 and F-53B) (Wang et al., 2015). Perfluoroalkyl phosphonic and phosphinic acids are also used as alternatives in certain applications. PFPAs are likely to be persistent and long-range transportable, whereas PFPiAs may be transformed to PFPAs and possibly PFCAs in the environment and in biota (Wang et al., 2016).

In environmental samples fluorotelomer-based substances were identified as the most relevant precursors of PFCAs based on the frequency of detection and the concentration of FTOHs, biotransformation intermediates (e.g. FTUCAs and FTCAs) and persistent biotransformation products (e.g. x:3 acids and PFCAs) (UBA, 2016).

Health based guidance values available for HBM data

In the REACH restriction dossier on PFOA internal DNELS4 were derived based on different endpoints in animal and human studies. The respective values derived for the general population were in the range of 0.3 ng/ml and 277 ng/ml (ECHA, 2014). The Committee for Risk Assessment (RAC) of the European Chemicals Agency (ECHA) has finally derived a DNEL of 800 ng/ml for the general population, arguing that a DNEL cannot be reliably derived from some effects (e.g. on the mammary gland) that may be more sensitive than the animal data currently used in the risk characterisation, as these data are not robust enough (ECHA, 2015). The German Human Biomonitoring Commission has published a re-assessment of the HBM-values of PFOS and PFOA in 2016. The HBM-I-value represents the concentration of a substance in human biological material below which no risk for adverse health effects over life time is expected (HBM Commission, 2014). The respective HBM-I-values are 2 ng PFOA/ml and 5 ng PFOS/ml blood plasma (HBM Commission, 2016). The HBM Commission has decided to use the existing POD ranges of 1 to 10 ng/ml as a basis and selected 2 ng/ml comprising the HBM-I-value for PFOA, pointing to the consistency of results from animal and epidemiological studies.

Within the scientific community discussions on the most sensitive health endpoints are still ongoing, effects on immune system and on cholesterol levels might occur at even lower exposure concentrations.

At the end of 2018, the European Food Safety Authority (EFSA, 2018) released a scientific opinion suggesting the revision of tolerable intake values for two perfluoroalkyl substances: perfluorooctanoic acid (PFOA, CAS 335-67-1) and perfluorooctane sulfonate (PFOS, CAS 1763-23-1). After benchmark modelling of serum levels of PFOS and PFOA, and estimating the corresponding daily intakes, the CONTAM Panel established a tolerable weekly intake (TWI) of 13 ng/kg body weight (bw) per week for PFOS and 6 ng/kg bw per week for PFOA. For both compounds, exposure of a considerable proportion of the population exceeds the proposed TWIs.

Biomarkers available for parent compounds or metabolites in human matrices, and main characteristics of analytical methods (quantitative, semi-quantitative)

Analytical targets for the analysis in biomonitoring studies can include the parent compound, its metabolite(s) and transformation product(s) or other chemical products formed in the body or the environment. Known PFASs were mostly analysed by high performance liquid chromatography coupled with tadem mass-spectrometry (HPLC-MS/MS). FTOH and FTOH precursors (FTMAC and PAPs) and their metabolites can be measured by targeted methods, by low or high resolution mass spectrometry. Methods for possibly cationic PFASs (such as betaines used e.g. in firefighting foams) can be analysed using specific methods used for environmental matrices. Analyses of FTMAC require derivatisation, followed by gas chromatography coupled with mass-spectrometry (GC-MS) analysis (Place and Field, 2012; Trier, pers. comm., 2017). In recent years, several studies on total fluorine (TF), inorganic fluorine (IF), exactable organic fluorine (EOF) and specific known PFASs in environmental and blood samples were conducted. Usually, TF, IF and EOF were fractionated and measured by combustion ion chromatography (CIC). Some of the analytical methods used (please refer to the PFAS scoping document for more details) may not allow distinguishing between PFASs exposure and fluorine based medication. This concern is particularly related to the fact that many pharmaceuticals may contain fluorinated moieties to make them more persistent in human bodies (Wang, pers. Comm., 2017).

In best of our knowledge, it is not feasible and reasonable to measure all relevant PFAA precursors due to a lack of an overview on which precursors are being produced and used and to which ones humans are exposed to at the moment. Considering that most precursors would be transformed into acids in human body, it would be an interesting approach to measure the “total oxidisable precursors” in human matrices. The “total oxidisable precursors” methods have been used to reflect the total exposure to PFAAs and PFAA precursors in a number of environmental samples. Due to its nature of radical reactions with a large, complex mixture, the methods may not easily or never be standardised and the results may not be reproducible. However, it might be a semi-quantitative indicator to demonstrate PFAAs exposure stemming from the variety of precursors (Wang, pers. Comm., 2017).

Further analytical methods to simultaneously analyse as many PFASs as possible should be developed (Wang et al., 2016).

Following on the Advisory Board’s advice to strengthen the science-policy interface, HBM4EU developed a strategic and systematic approach to outreach and align science and policy. A legislative mapping exercise was done by RPA Consultants, providing relevant public policy processes that may benefit from the knowledge generated under HBM4EU. The documents are available for consultation here, with the tables presented here.

Due to the extensive list of PFAS covered, please consult Table 1 in the scoping document on PFAS.

As an example, PFOA (Perfluorooctanoic acid (linear and branched isomers) is covered by the following legislation: REACH Annex XVII restriction (Regulation (EU) 2017/1000), SVHC Candidate List (PBT, Repr.), CLH (Carc. 2, Repr. 1B, STOT RE 1, Acute Tox. 4, Eye Dam. 1), and are proposed for inclusion in the Stockholm Convention, Norman 2011.

For PFOS (perfluorooctane sulphonate) Heptadecafluorooctane-1-sulphonic acid (linear and branched isomers), it is covered by REACH Annex XVII restriction, CLH (Carc. 2, Repr. 1B, Lact., STOT RE 1, Acute Tox. 4, Aquatic Chron. 2), PIC regulation, POP Regulation (EG) No. 757/2010, Stockholm Convention, environmental legislation (Seveso, Directive 2012/18/EU; Regulation 649/2012 concerning export and import of hazardous chemicals).

Throughout the duration of this project, the following policy questions will be addressed:

1. What is the current exposure of the EU population to PFASs and do they exceed Guidance values (reference and HBM values), where available?
2. Are there differences in exposure of the EU population to regulated and non-regulated PFASs?
3. Has restriction of PFOS according to the POP Regulation led to a reduction in exposure, especially for children?
4. Is exposure driven by diet, consumer exposure, occupation or environmental contamination?
5. Which areas and environmental media in Europe are contaminated with PFASs?
6. How can this feed into an assessment of the TDI for PFOS and PFOA set by EFSA?
7. What is the impact of a pending 2016 ECHA decision to restrict the manufacturing, marketing and use of PFOA under REACH?
8. Is it important to eliminate legacy PFASs from material cycles (i.e. waste electronic equipment) when implementing a circular economy in order to protect human health?
9. Can differences in PFASs profiles be observed in different population groups and time periods?
10. What are the PFASs levels and health effects in vulnerable population groups?
11. How can mixture effects of environmental and human PFASs mixtures present to date be estimated?
12. How can PFAS substances of relevance for human exposure and health be identified having in mind that more than 3000 substances are at the market?
13. How can identification and assessment (including data on (potential) adverse effects on human health and the environment) of alternatives currently hampered by CBI (Confidential Business Information) be facilitated?
14. How much are HBM values dependent on host characteristics and does this have implications for identifying vulnerable groups?

In the interest of transparency and accountability, HBM4EU invites interested stakeholders to submit comments on the scoping document.

All submitted comments will be made available for download on this webpage and will be taken into consideration by the HBM4EU consortium, where possible.

Please click here to submit your comments.

Please click here to access the Substance Report.

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