Document 6ROK8RXwwJ7EjmqjG1XZk6Y0d

EC Number CAS Number Substance Name (ECHA's Acronym Chain length Registration log Kaw logKow logKlipw Critical micelle Water solubility Vapor pressure logKoa Volatilization pKa Relations with pH logKoc Other physicochemical properties Interactions with organic matter Fate dissem.site) concentration (CMC) 207-021-6 422-64-0 Pentafluoropropionic acid PFPrA C3 Pre-registered Degradability Persistence BCF/BMF/BAF Bioaccumulation Toxicity [63] Perfluoroalkyl and perfluoroether moieties are highly persistent under natural conditions => Even though some PFASs may partially degrade in the environment and biota, they will all ultimately transform into highly stable end products, wich are usually perfluoroalkyl or perfluoroalkyl(poly)ether acids such as PFCAs, PFSAs, PFECAs and PFESAs [63] the high solubility and proteinbinding characteristics of ionic PFAAs challenge the conventional assessment of bioaccumulation potential that is through either bioconcentration factor in aquatic species (usually fish) or models based on octanol-water partition coefficients (Kow) => even when PFAAs are not characterized as bioaccumulative under the current regulatory frameworks, the accumulation of these PFAAs in the environment will lead to increasing external exposure Reproductional/developmental toxicity PBT Exposure Elimination from body Toxicological information [63] the very high persistence of PFAAs leads to poorly reversible exposure to these substances in the global environment and some local/regional environments including groundwater Blood concentrations Other concentrations (semen, Accumulation in tissues Elimination via urine, feces, liver Elimination half-life T1/2 Volumes of distribution Vd Elimination rate constant ke Clearance Cl AUC urine, ...) [47] PFPrA in human blood samples: - Detection frequency: 100% - Mean +/- SD: 0.82 +/- 0.71 ng/mL - 5th percentile: 0.21 ng/mL - Median: 0.62 ng/mL - 95th percentile: 2.1 ng/mL => negatively (-0.245) associated with BMI (p<0.05) [47] PFPrA in human semen samples: - Detection frequency: 100% - Mean +/- SD: 1.6 +/- 2.3 ng/mL - 5th percentile: 0.29 ng/mL - Median: 0.95 ng/mL - 95th percentile: 4.1 ng/mL => Much higher concentrations were found in semen, suggesting that blood-testis barrier may not work for these chemicals => positively (0.279) associated with BMI (p<0.01) LOEL/LOEC/NOEC Quality Standards (QS) Air Food and beverages Groundwater Surface water River / lake water Sea- and coastal water Drinking water source areas Tap water/ mineral water Near wastewater treatment plant Near fluorochemical production plant (or near sites where fluorochemicals are commonly 206-786-3 375-22-4 Heptafluorobutyric acid PFBA C4 Pre-registered [79] log Kaw: 0.30 (calculated from equations) [1] (neutral form) = 2.82 - 4.6 [29] PFBA: logKow ranges from -0.52 to 2.82 (exper. + modelled) [41] PFBA ChemBioOffice: logKow = 3.39 [65] PFBA: EPI pred logKow: 2.43 [79] log Kow: -0.62 (calculated from equations) 206-793-1 375-73-5 1,1,2,2,3,3,4,4,4- PFBS C4 nonafluorobutane-1- sulphonic acid Registered [79] log Kaw: 1.02 (calculated from equations) [1] (neutral form) = 2.82 - 4.6 [29] PFBS: logKow ranges from 2.41 to 3.90 (exper. + modelled) [35] PFBS: logKow = 3.9 [79] log Kow: 0.14 (calculated from equations) [65] PFBA: - logCMC Exp: -0.12 - logCMC Pred: -0.20 - 3: 4.93 [1] > 20 g/l [65] PFBA: logpL (subcooled vapor pressure): - Y-Exp: 0.81 mm Hg - Y-Pred: 0.83 mm Hg - EPI Suite Pred: 1.17 mm Hg - Y-Exp: 2.93 Pa - Y-Pred: 2.12 Pa [1] 6-6.7 [9] PFCAs and PFSAs have low vapour pressures which decrease with increasing carbon chain length => Low potential for volatilization [1] < 1 [65] PFBA: logAqS (aqueous solubility): - Y-Pred: 2.65 mg/L - EPI Suite Pred: 2.88 mg/L [79] Vapor pressure (PL): log (PL/Pa): 2.40 (calculated from equations) [79] log Koa: 4.13 (calculated from equations) [79] Water solubility (SL): log (SL/mol L^-1): -0.64 (calculated from equations) [1] > 20 g/l [79] Vapor pressure (PL): log (PL/Pa): 2.12 (calculated from equations) [1] 6-6.7 [9] PFCAs and PFSAs have low vapour pressures which decrease with increasing carbon chain length => Low potential for volatilization [1] < 1 [79] Water solubility (SL): log (SL/mol L^-1): -1.64 (calculated from equations) [79] log Koa: 4.58 (calculated from equations) [9] Calculated pKa values indicate that both PFCAs and PFASs are strong acids which will predominantly be in their dissociated negatively-charged form at environmentally relevant pH values [1] 2.7-3.6 [79] - Molar mass (MM): 214.0 g/mol - Molar volume (VM): 127.5 cm^3/mol (by ACD/Labs' ACD/PhysChem Suite) - Total surface area (TSA): 201.9 A^2 (by ChemAxon) - Tm (experimental melting point): 256 K (data obtained from ChemSpider) [17] shorter-chained PFAS (C<=5): electrostatic interactions have been hypothesized to be of more importance than hydrophobic interactions for sorption to sludge => partition to sludge increases with decreasing chain length [1] drinking water resources are highly sensitive to contamination [9] PFAAs will likely not be biodegraded under typical drinking water treatment conditions [1] high energy carbon-fluorine bond => extremely persistent [29] no BCF, BMF1 and BMF2 available for PFBA [3] protein binding affinity appears to increase with chainlength => short chain: faster elimination and lower distribution to liver [23] in most cases (hepatotoxicity and reproductive toxicity) shortchain PFAAs are less toxic than long-chain homologues [42] FRESHWATER: Brachionus calyciflorus - rate of population increase was 16.5% lower in PFBA than in the control (p<0.05) => decline of population growth rate poses a chronic and longterm risk on rotifer population - mictic ratios increased from 0.553 in the control to 0.953 in PFBA (p<0.05) => increase of mictic ratio was due to the change of rotifers reproduction => under exposure of environmental stress rotifers transform from asexual reproduction to sexual reproduction => fertilized eggs can resist poor outside environment and keep balance in population dynamics => PFBA may interfere with endocrine signaling in B. calyciflorus , causing a reduction of population growth rate and an increase of mictic ratio [23] Short-chain PFAAs: - P - not B - maybe T - LRTP [1] no degradation => permanent and irreversible exposure of organisms [29] PFBA: logKoc: 1.8 (EPIsuite) [28] PCA showed that PFBA was significantly associated with DOC (p<0.00001) => PFASs may bind to and be co-transported by DOC in water => PFCAs and PFSAs are negatively charged in natural waters by proton dissociation of the acid => Short-chains (more hydrophilic) could readily bind to positively charged ions that are complex bound to DOC (such as Ca2+), while longerchained PFASs rather partition to even more hydrophobic phases in the water, such as the organic carbon fraction of suspended particulate matter [9] - high water solubility, simultaneous hydrophobic/ hydrophilic properties, and low volatility of most PFAA contribute to their presence in all aquatic environments and even rain water - Removing PFASs from drinking water: * Granular Activated Carbon Absorption, however PFBA may pass through or reach breakthrough very quickly * Ion exchange/non-ion exchange resins may be useful for removing PFASs * High pressure membranes will achieve high rejection of most PFASs, however lower molecular weight PFASs, such as PFBA, may be less well rejected [23] short-chain PFAAs are similarly persistent as their long-chain homologues [14] PFBA has shorter serum half-lives in humans and biota than its longerchain homologues [29] ammonium perfluorobutanoate (NH4+ PFBA) showed effects in males such as increased liver weight, slight to minimal hepatocellular hypertrophy, decreased serum total cholesterol and reduced serum thyroxin with no change in serum thyrotropin [42] FRESHWATER: Brachionus calyciflorus - body size values were increased from 0.125 in the control to 0.136 m3 in PFBA (p<0.05) - egg size values were increased from 0.514 in the control to 0.580 m3 in PFBA (p<0.05) => females produce larger and fewer eggs under the condition of the increasing cost of producing larger eggs => large eggs produce large larvae and large offspring which exert a strong influence to the ability of survival => response to toxicity [29] PFAS with carbon chain C<=6: - low bioaccumulative potential - low acute and chronic aquatic toxicity => no direct concern for aquatic environment [2] human exposure primarily via dietary intake => contribution of drinking water [74] PFBA: * R (retardation factor) (40 cm): 1.0 * R (80 cm): 2.8 * log Kd (40 cm): -2.4 * log Kd (80 cm): -0.43 * log Koc (40 cm): 0.8 * log Koc (80 [29] Limited accumulation has been measured in freshwater and coastal sediments PFBA: 0.1-61.2 ng/g dw [10] - volatile PFASs (FTOHs, FASAs, FASEs) enter the atmosphere, where they can degrade, form intermediates during atmospheric oxidation, or transform into more persistent PFASs, such as PFSAs and PFCAs, which may finally end up in the aquatic environment - PFASs have the potential to affect cell membranes of algae => can alter nutritious quality of biofilms => potential effects on physiological fitness of invertebrates [62] PFBA has been detected in surface water and in public and private wells [77] - the extreme persistence of all PFAAs with the increasingly higher aqueous solubility (and thus PFAA mobility) as the chain length shortens, also creates increasing difficulty for them to be treated by many conventional remediation and water treatment technologies - Remediation of PFASs: * Biological treatment: potential for anaerobic degradation of PFASs seems limited, as no organisms have been isolated which can utilize PFAAs as terminal electron acceptors * GAC (granular activated carbon) can be inefficient at removing PFOA and becomes progressively less effective for removing shorter chain PFCAs such as PFHxA, PFPeA, PFBS and PFBA as the chain length diminishes [77] - Remediation of PFASs: * in the same manner that GAC is less effective at removing shorter chain PFAAs and PFAA precursors from water, the injected particulate-activated carbon is also expected to be less effective at removing these PFASs * Ion-exchange resins: long-chain PFAA and precursors are likely to be removed from water significantly more strongly than short-chain PFASs * Osorb is similarly effective as GAC at removing both PFOS and PFOA and more effective than GAC at removing PFBA * Precipitation removal rates for shorter chain PFAAs improve when higher doses of the coagulant are [77] Remediation of PFASs: * Ozofractionation: can convert polyfluorinated precursors to PFAAs via chemical oxidation which assist their removal via foam fractionation and can remove short-chain PFASs in addition to long chain * In situ foam fractionation: limited application * RO and NF have been shown to be extremely effective in removing PFASs regardless of chain length and are also expected to be effective at removing many types of PFAA precursors * Concerns regarding chemical oxidation for in situ application include the generation of sign. concentrations of more mobile short-chain PFASs * Chemical reduction: application at field-scale for in situ PFAA remediation is challenging from a practicality perspective * Sonolysis has been demonstrated at laboratory scale for PFASs but not scaled up for commercial use [77] Remediation of PFASs: * Stabilization: managing PFASs source zone soils in situ provides a sustainable approach that does not create a concentrated waste requiring offsite management or destruction * High energy electron beam: further evaluation necessary * Low/high temperature thermal desorption: a high degree of licensing and stakeholder engagement will likely be required due to the operation of a thermal waste recovery facility and the associated off-gasses * Vapor energy generator process: no full-scale application of VEG for PFASs has been implemented * Ball milling: viability of ball milling has not been studied in detail with respect to the complications of PFASs (short-chain compounds and [63] Perfluoroalkyl and perfluoroether moieties are highly persistent under natural conditions => Even though some PFASs may partially degrade in the environment and biota, they will all ultimately transform into highly stable end products, wich are usually perfluoroalkyl or perfluoroalkyl(poly)ether acids such as PFCAs, PFSAs, PFECAs and PFESAs [23] short-chain PFAAs are less bioaccumulative in animals and humans than the long-chain homologues, but may show higher uptake into the leaves, stems and fruits of plants [39] Relative response of mouse and human PPAR to PFAAs in transiently transfected COS-1 cells, measured by C20max: PFBA: - Mouse: 51 M (Most active to least active: ranked 10/13 PFASs) - Human: 75 M (Most active to least active: ranked 7/13 PFASs) => A lower C20 max value indicates a stronger induction of PPAR alpha activity by that PFAA [106] Study investigated the relationship between serum concentrations of perfluoroalkyl compounds and growth parameters in 2-year-old Korean children. Study included 361 children aged 2 years (192 boys and 169 girls; 22-27 months). PFBA concentrations in serum from children at 2 years of age: - LOD: 0.028 ng/mL - Detection frequency: 87 => detected in 24.1% of all serum samples - Range (min-max): 0.019 - 9.570 ng/mL - Mean SD: 0.230 1.082 ng/mL [29] - max. concentration in biota: 7.62 ng/g ww - PFBA can be considered not bioaccumulable or biomagnificable in the aquatic food web [39] => PPAR is activated by PFAAs of 4-12 carbons in chain length => Increasing activity of PPAR with increasing chain length of the PFAA up to C9, and lower activity with longer chain PFAAs (C>9) with both mouse and human PPAR => PPAR has been identified as a key player in the mode of action for PFAA-induced toxicity [108] Analysis of 424 mother-fetus pairs from the Maoming Birth Cohort, China. * Concentrations of PFBA in maternal and cord serum: - Cord serum: Detection rate: 94.34%, median: 1.45 ng/mL, mean SD: 1.63 1.42 ng/mL - Maternal serum: Detection rate: 77.12%, median: 0.70 ng/mL, mean SD: 0.90 0.91 ng/mL * PFAS in cord serum was positively correlated with PFAS in maternal serum for all PFAS (p<0.05) [63] the high solubility and proteinbinding characteristics of ionic PFAAs challenge the conventional assessment of bioaccumulation potential that is through either bioconcentration factor in aquatic species (usually fish) or models based on octanol-water partition coefficients (Kow) => even when PFAAs are not characterized as bioaccumulative under the current regulatory frameworks, the accumulation of these PFAAs in the environment will lead to increasing external exposure [41] Progesterone content in mLTC-1 (mouse Leydig tumor cells) decreased with the increase in PFBA treatment concentration => This relationship was illustrated by a standardized S-shape curve (dose-response curve) => Predicted IC50 (half-maximal inhibitory effect concentration): 1930 mol/L => A similar dose-response curve was seen for cell viability (decreasing cell viability with higher PFBA treatment concentration), but progesterone content was more sensitive than cell viability to PFBA exposure [108] * Transplacental transfer efficiency (TPT) of PFBA = concentration cord serum (ng/mL)/ concentration maternal serum (ng/mL): - TPT (n=305): Median: 1.41, mean SD: 2.88 4.47 - U-shaped pattern for TPT in PFCAs: TPT decreased from PFBA to PFDA and then increased to PFTrDA * Higher TPT in PFAS alternatives than PFASs, such as PFBA vs. PFOA (median: 1.41 vs. 0.73, p<0.001) => PFBA with its shortchain length may be transferred to the fetus more efficiently than PFOS and PFOA => PFAS alternatives may be more easily transported from mother to infant than conventional PFASs in uterus * PFBA and 6:2 Cl-PFESA are the second and fourth highest PFAS measured, respectively, in cord serum [77] short-chain PFAAs have been found to concentrate in the edible portion of some crops, such as fruits and vegetables, wheras longer chains tend to be retained more in the shoots and roots of plants [41] Sign. difference in mitochondrial membrane potential (MMP) between: 5000 mol/L PFBA and control (p<0.05) => Inhibition effect of PFBA on progesterone production might be due, in part, to decrease in MMP in mLTC-1 [44] All tested PFCs induced hPXR reporter activites in a dosedependent manner => In general, longer carbon chain length seems to be favorable for the hPXR activity of PFAAs with chain length shorter than 10, as chain length is longer than 10, increasing the chain length of PFAA seems to render the chemicals weaker potency to induce hPXR activity => PFBA: EC50 = 30.1 M, maximum induction = 15.9 % => Activation of hPXR: All docked compounds showed a similar pose, in which the polar terminal group (carbonyl group, hydroxyl group or sulfo group) of the PFCs interacted with Ser247 through a hydrogen bonding, and fluorinated carbon chain inserted itself into a hydrophobic domain comprising Met243, Met246, Phe288, Trp299, Tyr306 and Val211 => Agonistic activity depends on interactions of PFCs with hydrophobic residues in the binding pocket => PFBA forms a 4.2 A hydrophobic contact with Phe288 ==> PXR induction is postulated to be associated with lipid homeostasis, atherosclerosis, carcinogenesis and endocrine- [46] PFASs can compete with thyroxine (T4) for binding to the human thyroid hormone transport protein transthyretin (TTR), which may lead to reduced thyroid hormone levels leading to endocrine disrupting adverse effects => PFBA: T4-TTR Binding (%): 106 => Experimental Classification: Inactive (N) => Predicted Classification: Active (Y) => Binding energy: -3.28 kcal/mol => PFASs can compete with thyroxine (T4) for binding to the human thyroid hormone transport protein transthyretin (TTR), which may lead to reduced thyroid hormone levels leading to endocrine disrupting adverse effects [46] Discriminating properties between active (Y) and inactive (N) classes: - Hydrophobicity of the H atom attached to heteroatom - Contributions from the H-050 descriptor => PFASs containing acid functional groups: - If the carbon chain length is between 6 and 10, PFASs will be highly toxic or active ones - Toxicity will be lower or non-existent for PFASs containing a C chain length greater than 10 or below 6 => PFBA is not toxic/inactive [47] negative correlations were significantly observed between sperm motility and PFBA in human semen, therefore exposure to PFAAs may result in a decline in semen mobility [50] NIH-3T3 cells: PFBA led to increased PPAR activity (at 30 M, p<0.05 and 100 M, p<0.001) => short-chain PFCAs generally activate both PPAR and PPAR with similar potency and efficacy as long-chain PFCAs (except PPAR [51] - IC50 value and dissociation constant (Kd) with hPPAR-LBD: PFBA: IC50: ND, Kd: ND - PFBA induced no significant hPPAR activation in testing concentrations - Docking interactions of PFBA with hPPAR-LBD: Hydrogen bonds: [52] human embryonal kidney cell line HEK293, activation of human PPAR, PPAR and PPAR - PFBA triggered activation of PPAR: c20% value: 27.4 M - Most PFCAs also activated PPAR although higher PFCA concentrations were required for PPAR activation => PFBA did not activate PPAR up to a concentration of 1000 M (c20% value) - In case of PPAR only high concentrations of PFHpA and PFOA activated this nuclear receptor => PFBA: c20% value: >1000 M => PPAR seems to be the primary molecular target of PFCA [81] - no sign. cytotoxicity was observed for the shortest chain length PFCs (PFBA, PFHxA, PFBS and PFHxS) after 24h incubation with the human placental choriocarcinoma cell line JEG-3 - differential bioavailability of PFCs in the in-vitro system was assessed by measuring the fraction retained in the cells after exposure: the concentration of PFBA, PFBS and PFHxS was below dection limit under our assay conditions - inhibition of P450 aromatase (CYP19) activity in JEG-3 cells: 13% inhibition by PFBA when tested at 500 M, IC50: N.D. => sulfonates (PFBS, PFHxS) were stronger inhibitors of aromatase [82] Study of inhibition of cellular viability in human liver cell line (HL-7702): * PFBA single: - IC-20 (20% stimulatory effect concentration): 1.53 x 10^(-3) mol/L - IC0 (0% inhibitory concentration): 7.03 x 10^(-3) mol/L - IC10 (10% inhibitory concentration): 7.45 x 10^(-3) mol/L - IC50 (50% inhibitory concentration): 9.23 x 10^(-3) mol/L * PFBA-PFBS mixture: - all IC0, IC10 and IC50 values in the binary mixture were less than that of individual PFSA or PFCA values - a synergistic action took place under the effective concentrations of IC0, IC10 and IC50 with various concentration ratios - for the effective concentration of IC-20, only the higher proportion of PFSA in the mixtures (PFBS > 57.7%) showed a synergistic effect for IC-20 [82] - Mixtures of eleven PFAAs (Mix0, Mix10 and Mix50, PFBA included) presented a partial addition effect - Mixtures of nine PFAAs (Mix0, Mix10, Mix50, PFBA included) with J-shaped curves only presented a synergistic effect with MTI values ranging from 1.16 to 1.25, however Mix-20 showed partial [84] PFCs can compete with thyroxine (T4, the transport form of thyroid hormone), for binding to the human thyroid hormone transport protein transthyretin (TTR) => such competitive capacity may lead to decreased thyroid hormone levels as previously reported for animals exposed to PFCs => PFCs do not affect the regulatory functions of the thyroid hormone system itself, but it is the competitive binding to transport proteins that alters the free thyroxine (T4) levels in blood => TTR is the main T4 carrier in cerebrospinal fluid, and also important in serum of most mammalian species and birds => Binding potency is clearly associated with the degree of fluorination of the alkyl chain, with a maximum potency at a chain length of eight carbons (PFOA) for PFCAs => PFCAs with a carbon chain length longer than eight have low TTR binding potencies => TTR is both in humans and in rodents the most important carrier protein for thyroid hormone to the developing fetus and the brain [11] estimated total daily exposure humans: - low exposure scenario: 6.3 pg/kg/d - intermediate exposure scenario: 19 pg/kg/d - high exposure scenario: 190 pg/kg/d => daily exposures originate almost entirely from direct intakes regardless of the scenario (75%-99%) => direct exposure to PFBA via drinking water consumption is estimated to be the primary exposure pathway in all exposure scenarios (88%-99%) [23] on-going (increasing) emissions and high persistence => short-chain PFAAs, PFECAs and PFESAs accumulate in the environment => increasing exposure to these substances over time => even if emissions cease in the future it will take decades to centuries to reverse global contamination of short-chain PFAAs, PFECAs and PFESAs, due to high persistence and environmental mobility [34] significant correlation with other PFAS-analytes detected in groundwater (p<0.05), indicating a common or similar source of exposure [63] the very high persistence of PFAAs leads to poorly reversible exposure to these substances in the global environment and some local/regional environments including groundwater [109] Study on residential exposure: Vacuum cleaner contents from 39 homes were tested for perfluoroalkyl chemicals: * PFBA: - Detection frequency: 97 %, Median: 3.5 (ND-64) ng/g - Correlation between PFBA and PFBS (0.43, p<0.05) - Correlation between PFBA and PFHpS (0.60, p<0.05) - Correlation between PFBA and PFDS (0.32, p<0.05) * Occupants of all ages can be exposed to indoor contaminants, but preschool-aged children have the greatest risk due to hand to mouth activity and the amount of time they spend playing on or near the floor [111] PFAAs were investigated in home produced eggs and commercially produced eggs surrounding a fluorochemical industrial park in China: PFBA concentrations: * Home produced eggs (n=4, sites 1-4): Egg yolks: - 2 km distance: 110 ng/g - 5 km distance: 16.0 ng/g - 10 km distance: 23.8 ng/g - 20 km distance: 1.75 ng/g Egg whites: - 2 km distance: 6.97 ng/g - 5 km distance: 1.15 ng/g - 10 km distance: 2.43 ng/g - 20 km distance: 0.24 ng/g Whole eggs: - 2 km distance: 22.5 ng/g - 5 km distance: 5.40 ng/g - 10 km distance: 12.0 ng/g [111] * Commercially produced eggs (n=12, sites 5-16): Egg yolks: - Min: 0.31 ng/g - Max: 5.64 ng/g - Mean: 1.72 ng/g - Median: 1.02 ng/g - n > LOD (%): 12 (100) Egg whites: - Min: 0.16 ng/g - Max: 0.52 ng/g - Mean: 0.33 ng/g - Median: 0.31 ng/g - n > LOD (%): 12 (100) [111] Whole eggs: - Min: 0.25 ng/g - Max: 3.82 ng/g - Mean: 0.85 ng/g - Median: 0.49 ng/g [111] PFBA was the second most abundant congener in yolks of home produced eggs with an average contribution of 22% to PFAAs [111] => Unlike in the environmental media, egg yolks contained much lower proportions of PFPeA, PFHxA and PFHpA than of PFBA => PFBA concentrations are higher in the diet of chickens, for example in maize from the farmland around the fluorochemical industrial park [111] * In yolks of commercially produced eggs, the detection frequency of PFBA, PFOA and PFOS was 100% whereas other PFAAs were less frequently detected + contributions of PFBA (27-62%, with a mean 39%), PFOA (6-51%, 38%) and PFOS (9-56%, 30%) to PFAAs were comparable with each other [111] * PFBA was the predominant form in egg whites, contributing to 77% of PFAAs * Concentrations of PFBA in whites of home produced eggs declined with increasing distance from the fluorochemical industrial park * Averagely 20% of PFBA (6-48%) was found in egg whites, although most of it was still distributed in egg yolks [111] * The EDI of PFBA via egg consumption ranged from 1.26 to 73.6 ng/kg.bw/day for adults and 3.54 to 207 ng/kg.bw/day for children * PFBA was reported in eggs and detected in all egg samples * PFOA and PFBA were the predominant forms in homeproduced eggs, while PFOA, PFBA and PFOS dominated in commercially produced eggs * Human exposure pathways of PFBA are worthy of attention and health effects of PFBA exposure to local residents around the fluorochemical industrial park via egg consumptions need urgent further investigation [9] Calculated pKa values indicate that both PFCAs and PFASs are strong acids which will predominantly be in their dissociated negatively-charged form at environmentally relevant pH values [1] 2.7-3.6 [79] - Molar mass (MM): 300.1 g/mol - Molar volume (VM): 162.3 cm^3/mol (by ACD/Labs' ACD/PhysChem Suite) - Total surface area (TSA): 282.1 A^2 (by ChemAxon) - Tm (experimental melting point): 310 K (data obtained from ChemSpider) [17] shorter-chained PFAS (C<=5): electrostatic interactions have been hypothesized to be of more importance than hydrophobic interactions for sorption to sludge => partition to sludge increases with decreasing chain length [1] drinking water resources are highly sensitive to contamination [29] PFBS: logKoc: 2.3 (EPIsuite) [28] Linear regression showed significant positive correlations of a.o. PFBS to DOC (p<0.05) => PFASs may bind to and be co-transported by DOC in water => PFCAs and PFSAs are negatively charged in natural waters by proton dissociation of the acid => Short-chains (more hydrophilic) could readily bind to positively charged ions that are complex bound to DOC (such as Ca2+), while longerchained PFASs rather partition to even more hydrophobic phases in the water, such as the organic carbon fraction of suspended particulate matter [9] Removing PFASs from drinking water: * Granular Activated Carbon Absorption, however PFBS may pass through or reach breakthrough very quickly * Ion exchange/non-ion exchange resins may be useful for removing PFASs * High pressure membranes will achieve high rejection of most PFASs [1] high energy carbon-fluorine bond => extremely persistent [29] PFBS: BCF <1 L/kg BMF <1 <=> other study: BAF 0.3-1736 L/kg [3] protein binding affinity appears to increase with chainlength => short chain: faster elimination and lower distribution to liver [23] short-chain PFAAs are similarly persistent as their long-chain homologues [14] PFBS has shorter serum half-lives in humans and biota than its longerchain homologues [84] PFBA: - Molecular weight: 214.0 g/mol - T4-TTR binding at maximum concentration (10M): 106% - IC50 (concentration at 50% inhibition): N.D. - Slope of dose-response curve: / - T4-REP (relative potency compared to T4) factor: N.D. - HPLC retention time: 11.7 min - Purity: 99 [85] Mean fold change standard error of mRNA expression after exposure to 100 M of PFBA (in rat H4IIE hepatoma cells): * Thyroid related genes: - Hex: 0.81 0.28 => downregulation of Hex - PAX 8: 6.78 3.30 => upregulation of PAX 8 * Cholesterol related genes: - Per-3-Keto-: 0.11 0.02 - SQSYN: 0.76 0.07 => downregulation of cholesterol related genes * Lipoprotein related genes: - Mito-3-Keto-: 0.12 0.02 - Mito-3-Keto-: 0.44 0.17 => downregulation of Mito-3-Keto- & Mito-3-Keto- - ApoA4: 1.89 0.11 [86] * weak binding of PFBA with the protein TR-LBD (TR = thyroid hormone receptor, LBD = ligand-binding domain) * Binding of PFBA with TR-LBD: - Length: 6.03 A - IC50: ND - RP (relative potency (IC50 T3 / IC50 chemical)): ND - Hydrogen bonding: ARG 228, ARG 262 * all tested PFCs fit into the T3-binding pocket of TR-LBD, with the acid or hydroxyl end group residing toward the inner part and the hydrophobic chain toward the entrance of the binding pocket => all PFCs with an acid end group formed a hydrogen bond with [87] * Binding potency of PFBA to TTR (TTR competitive binding assay): - IC50: 89,252 nM - Kd: 14,280 nM - RP (relative potency (IC50 T4 / IC50 chemical)): 3 x 10^(-4) => perfluoroalkyl acids: Kd values decreased as carbon chain length increased from C4 to C8 => no further increase, but a downward trend was observed for perfluoroalkyl acids with longer chain lengths (C9-C14) * Binding potency of PFBA to TTRmutK15G: - IC50: ND - Kd: ND - Hydrogen bonding: Lys15 => PFASs exhibited much weaker binding affinities to TTRmutK15G compared with wild-type TTR * Docking study: - Similar to T4, PFOA and PFOS could nearly fill the TTR ligandbinding pocket - Perfluoroalkyl acids with carbon chain length less than C8 did not adequately fill the T4 binding pocket - Longer fluorinated carbon chain structures were nearly too large [87] * Binding potency of PFBA to TBG (TBG competitive binding assay): - IC50: ND - Kd: ND - RP (relative potency (IC50 T4 / IC50 chemical)): ND => only PFTA and PFTdA bound to TBG * Binding potency of PFBA to TBGmutR378G: - IC50: ND - Kd: ND * Binding potency of PFBA to TBGmutR381G: - IC50: ND - Kd: ND - Hydrogen bonding: Arg381 => both TBGmut378G and TBGmutR381G exhibited much weaker binding potencies to PFTA and PFTdA when compared with that of wild-type TBG * Docking study: - Perfluoroalkyl acids with carbon chain length less than C12 did not adequately fill the T4 binding pocket - Longer fluorinated carbon chain structures could nearly fill the TBG [23] in most cases (hepatotoxicity and reproductive toxicity) shortchain PFAAs are less toxic than long-chain homologues [103] Analysis of 369 pairs of maternal and umbilical cord serum in a prospective birth cohort in Shandong, China: PFBS: * Maternal serum: - Detection rate: 94.3 % - Geometric mean: 0.16 - Range: <LOD - 0.38 * Cord serum: - Detection rate: 98.9 % - Geometric mean: 0.19 - Range: <LOD - 0.98 * Correlation between maternal and cord levels of PFBS (stat. sign.): 0.159 (p=0.002) * Transplacental transfer efficiency: = Individual PFAS concentration in cord serum (ng/mL) / corresponding PFAS concentration in maternal serum (ng/mL): Range: 0.03 - 60.90 => Short-chain PFASs were transferred more efficiently from maternal serum to cord serum [39] Relative response of mouse and human PPAR to PFAAs in transiently transfected COS-1 cells, measured by C20max: PFBS: - Mouse: 317 M (Most active to least active: ranked 13/13 PFASs) - Human: 206 M (Most active to least active: ranked 10/13 PFASs) => A lower C20 max value indicates a stronger induction of PPAR alpha activity by that PFAA => PPAR is activated by PFAAs of 4-12 carbons in chain length => Increasing activity of PPAR with increasing chain length of the PFAA up to C9, and lower activity with longer chain PFAAs (C>9) with both mouse and human PPAR => PPAR has been identified as a key player in the mode of action for PFAA-induced toxicity [103] * Strong correlations between maternal and cord levels suggest that PFASs could readily pass through the placenta [23] Short-chain PFAAs: - P - not B - maybe T - LRTP [1] no degradation => permanent and irreversible exposure of organisms [2] human exposure primarily via dietary intake => contribution of drinking water [74] PFBS: * R (retardation factor) (80 cm): 2.6 * log Kd (80 cm): -0.47 * log Koc (80 cm): 2.7 [118] PFAS species with high Kd or Koc values were strongly adsorbed by sediments and organic matter, and compounds with very low Kd or Koc values were highly mobile with respect to the sediments (i.e. weak adsorption). [118] As a result, PFHxA, PFHpA, and LPFBS (with lower Kd and Koc values) could be readily transported from sediments to water in comparison with other PFAS compounds with relatively [29] Limited accumulation has been measured in freshwater and coastal sediments PFBS: <1 to 3.5 ng/g dw [10] - volatile PFASs (FTOHs, FASAs, FASEs) enter the atmosphere, where they can degrade, form intermediates during atmospheric oxidation, or transform into more persistent PFASs, such as PFSAs and PFCAs, which may finally end up in the aquatic environment - PFASs have the potential to affect cell membranes of algae => can alter nutritious quality of biofilms => potential effects on physiological fitness of invertebrates [17] - the fate of PFASs sorbed to sludge is generally to be further spread in the environment through application of sludge or sludge-containing products - in 2014: 24% of sludge was applied on agricultural soil, 24% was used to cover mines and dump sites, and 29% was used to manufacture soil intended for less sensitive land use [35] - PFBS also detected in shoots of control plants => due to adsorption of N-EtFOSA from the atmosphere followed by subsequent biodegradation in plants - N-EtFOSA efficiently taken up from solutions by wheat, soybean and pumpkin roots and acropetally translocated to shoots in exposured plants => biotransformed to intermediates FOSAA, PFOSA and terminal PFSAs [77] - the extreme persistence of all PFAAs with the increasingly higher aqueous solubility (and thus PFAA mobility) as the chain length shortens, also creates increasing difficulty for them to be treated by many conventional remediation and water treatment technologies - Remediation of PFASs: * Biological treatment: potential for anaerobic degradation of PFASs seems limited, as no organisms have been isolated which can utilize PFAAs as terminal electron acceptors * GAC (granular activated carbon) can be inefficient at removing PFOA and becomes progressively less effective for removing shorter chain PFCAs such as PFHxA, PFPeA, PFBS and PFBA as the chain length [77] - Remediation of PFASs: * in the same manner that GAC is less effective at removing shorter chain PFAAs and PFAA precursors from water, the injected particulate-activated carbon is also expected to be less effective at removing these PFASs * Ion-exchange resins: long-chain PFAA and precursors are likely to be removed from water significantly more strongly than short-chain PFASs * Osorb is similarly effective as GAC at removing both PFOS and PFOA and more effective than GAC at removing PFBA * Precipitation removal rates for shorter chain PFAAs improve when higher doses of the coagulant are applied [77] Remediation of PFASs: * Ozofractionation: can convert polyfluorinated precursors to PFAAs via chemical oxidation which assist their removal via foam fractionation and can remove short-chain PFASs in addition to long chain * In situ foam fractionation: limited application * RO and NF have been shown to be extremely effective in removing PFASs regardless of chain length and are also expected to be effective at removing many types of PFAA precursors * Concerns regarding chemical oxidation for in situ application include the generation of sign. concentrations of more mobile short-chain PFASs * Chemical reduction: application at field-scale for in situ PFAA remediation is challenging from a practicality perspective * Sonolysis has been demonstrated at laboratory scale for PFASs but not scaled up for commercial [63] Perfluoroalkyl and perfluoroether moieties are highly persistent under natural conditions => Even though some PFASs may partially degrade in the environment and biota, they will all ultimately transform into highly stable end products, wich are usually perfluoroalkyl or perfluoroalkyl(poly)ether acids such as PFCAs, PFSAs, PFECAs and PFESAs [92] - No conclusive evidence for biodegradation of PFAS in sludge under aerobic conditions over a period of up to either 9 or 15 weeks - Although decreases in PFHxA, 6:2 FTOH, and 8:2 FTOH concentrations were observed, their concentrations in the control bottles also decreased, and it is therefore not possible to confirm that they were indeed due to biodegradation - The concentration decrease could also be due to a non-biological degradation process, losses not due to degradation or even to incomplete sterilization of the control bottles [92] - No evidence for degradation of any of the PFAS was observed under anaerobic conditions - This result is consistent with the lack of anaerobic biodegradation of PFAS reported elsewhere in the literature => The PFAS tested in these experiments are non-biodegradable under the experimental conditions used in this study, despite using municipal sewage sludge, which presumably has a history of exposure [15] - urine is a pathway of elimination in rat, monkey and human - PFBS is eliminated at a greater rate from human serum than the higher chain homologues of PFOS and PFHxS [46] PFASs can compete with thyroxine (T4) for binding to the human thyroid hormone transport protein transthyretin (TTR), which may lead to reduced thyroid hormone levels leading to endocrine disrupting adverse effects => PFBS: T4-TTR Binding (%): 69 => Experimental Classification: Active (Y) => Predicted Classification: Active (Y) => Experimental pIC50 (conc. with 50% inhibition of T4): 1.71 mM => Predicted pIC50: 1.44 mM => Binding energy: -4.23 kcal/mol => PFASs can compete with thyroxine (T4) for binding to the human thyroid hormone transport protein transthyretin (TTR), which may lead to reduced thyroid hormone levels leading to endocrine [105] Analysis of breast milk samples (n=293) collected from 127 mothers in Korea. Pairs of pregnant women and fetuses that were recruited from four South Korean cities. Detection rates and concentrations of PFBS in Korean breast milk samples: * No. detected: 4 => detected in 1.4% of all breast milk samples * Range (min-max): <10 - 30.8 ng/L * Average SD: 0.26 2.41 ng/L [23] short-chain PFAAs are less bioaccumulative in animals and humans than the long-chain homologues, but may show higher uptake into the leaves, stems and fruits of plants [46] Discriminating properties between active (Y) and inactive (N) classes: - Hydrophobicity of the H atom attached to heteroatom - Contributions from the H-050 descriptor PFBS fitted exactly in the identical active sites making interactions with Leu17, Ala108, Ala109, Leu110, Thr118, Thr119 and Lys15, but not with Val121 => A hydrogen bond formed with the positively charged Lys15 and negatively charged oxygen atom of sulfonate or acid functional groups PFBS showed high binding energy, though its pIC50 and T4-TTR % binding potency were moderate because of its small chain length, which is responsible for a much higher diverse structural orientation inside the protein compared to other PFASs PFASs containing sulfonate or sulfinate functional groups: - Toxic if chain length is less than C8 - Compounds will be inactive or show little toxicity if the carbon chain length is over 8 atoms => PFBS IS TOXIC/ACTIVE [106] Study investigated the relationship between serum concentrations of perfluoroalkyl compounds and growth parameters in 2-year-old Korean children. Study included 361 children aged 2 years (192 boys and 169 girls; 22-27 months). PFBS concentrations in serum from children at 2 years of age: - LOD: 0.227 ng/mL - Detection frequency: 182 => detected in 50.4% of all serum samples - Range (min-max): 0.160 - 4.610 ng/mL - Mean SD: 0.373 0.351 ng/mL [29] - max. concentration in biota: 10.8 ng/g ww - uncertainty on the bioaccumulation and biomagnification characteristics of PFBS, which requires further monitoring data [47] negative correlations were significantly observed between sperm motility and PFBS in human semen, therefore exposure to PFAAs may result in a decline in semen mobility [107] Analysis of 369 families with matched parental and cord serum samples from a birth cohort in Shandong, one of the regions seriously polluted by PFASs in China * PFBS concentrations in 369 matched serum samples: - Paternal serum: Detection frequency: 99.7%, Geometric mean: 0.20 ng/mL - Maternal serum: Detection frequency: 94.3%, Geometric mean: 0.16 ng/mL - Cord serum: Detection frequency: 98.9%, Geometric mean: 0.19 ng/mL * Positive correlation between maternal & cord serum: 0.16 (p<0.01) [63] the high solubility and proteinbinding characteristics of ionic PFAAs challenge the conventional assessment of bioaccumulation potential that is through either bioconcentration factor in aquatic species (usually fish) or models based on octanol-water partition coefficients (Kow) => even when PFAAs are not characterized as bioaccumulative under the current regulatory frameworks, the accumulation of these PFAAs in the environment will lead to increasing external exposure [48] Change in thyroid hormone TH-responsive genes in neuronal cells of avian species: *Gallus domesticus: D3 mRNA was upregulated following exposure to PFBS: 3 M: Mean fold change +/- SD: 2.34 +/- 0.72 (p<0.05), 10 M: Mean fold change +/- SD: 3.10 +/- 0.41 (p<0.05) RC3 was upregulated following exposure to PFBS: 10 M: Mean fold change +/- SD: 2.02 +/- 0.27 (p<0.05) => short-chained PFCs altered the expression of TH-responsive genes in chicken embryonic neuronal cells to a greater extent than the longchained PFCs => due to bioavailability: could have entered neuronal cells more readily due to their lower binding affinities to extracellular proteins *Larus argentatus: Oct-1 was upregulated following exposure to PFBS: 10 M: Mean fold change +/- SD: 6.47 +/- 1.99 (p<0.05) [107] * Positive correlation between paternal & maternal serum: 0.06 (p>0.05) * Negative correlation between paternal & cord serum: -0.04 (p>0.05) * Positive correlation between PFBS concentrations and vegetables consumption: SE: 0.062 0.029 (p=0.033) * Positive correlation between PFBS concentrations and maternal age: SE: 0.004 0.002 (p=0.040) [77] short-chain PFAAs have been found to concentrate in the edible portion of some crops, such as fruits and vegetables, wheras longer chains tend to be retained more in the shoots and roots of plants [51] - IC50 value and dissociation constant (Kd) with hPPAR-LBD: PFBS: IC50: ND, Kd: ND - PFBS elicited apparent dose-dependent response in Hep G2 cells in terms of hPPAR activation compared to control - Docking interactions of PFBS with hPPAR-LBD: Hydrogen bonds: Ser-289, His-449, Tyr-473 [81] - no sign. cytotoxicity was observed for the shortest chain length PFCs (PFBA, PFHxA, PFBS and PFHxS) after 24h incubation with the human placental choriocarcinoma cell line JEG-3 - differential bioavailability of PFCs in the in-vitro system was assessed by measuring the fraction retained in the cells after exposure: the concentration of PFBA, PFBS and PFHxS was below dection limit under our assay conditions - inhibition of P450 aromatase (CYP19) activity in JEG-3 cells: 97% inhibition by PFBS when tested at 500 M, IC50: 68 +/- 11 M => sulfonates (PFBS, PFHxS) were stronger inhibitors of aromatase activity than the corresponding acidic compounds (PFBA, PFHxA) - PFBS and PFHxS sign. inhibited CYP19 aromatase activity despite the fact that the measured uptake of the compounds by cells was below detection limit => Therefore, both PFBS and PFHxS may exert the inhibitory effect on CYP19 aromatase activity at rather low endogenous cellular concentrations [7] dietary intake = main route of exposure [23] on-going (increasing) emissions and high persistence => short-chain PFAAs, PFECAs and PFESAs accumulate in the environment => increasing exposure to these substances over time => even if emissions cease in the future it will take decades to centuries to reverse global contamination of short-chain PFAAs, PFECAs and PFESAs, due to high persistence and environmental mobility [30] short-chain PFBS has a much shorter half-life than PFOS and PFHxS in humans => thus PFBS should have a lower potential of accumulation in human serum than PFOS and PFHxS => however, serum concentrations of Swedish women show that exposure to PFBS via contaminated drinking water has resulted in increased levels of PFBS [34] significant correlation with other PFAS-analytes detected in groundwater (p<0.05) and surface water (p<0.01), indicating a common or similar source of exposure [35] crops contaminated with PFSAs, including PFOS, PFHxS and PFBS from sulfluramid use may represent a potential source of species exposure and lead to a risk to human health [63] the very high persistence of PFAAs leads to poorly reversible exposure to these substances in the global environment and some local/regional environments including groundwater [11] humans: modelled concentrations in serum: 0.0039 ng/g [47] PFBA in human semen samples: - Detection frequency: 100% - Mean +/- SD: 3.9 +/- 3.5 ng/mL - 5th percentile: 0.93 ng/mL - Median: 3.1 ng/mL - 95th percentile: 8.5 ng/mL => Much higher concentrations were found in semen, suggesting that blood-testis barrier may not work for these chemicals => positively (0.208) associated with age (p<0.05) => positively (0.373) associated with BMI (p<0.01) => negatively (-0.305) correlated with progressive motility (p<0.01) [31] PFBA serum concentrations in exposed (living in area under impact) and not exposed (subjected to background exposure) subjects EXPOSED (n=257) - Minimum: <LOQ - Median: <LOQ - Maximum: 3.59 ng/g - % levels <LOQ: 93 NOT EXPOSED (n=250) - Minimum: <LOQ - Median: <LOQ - Maximum: 0.35 ng/g - % levels <LOQ: 83 => PFBA concentration sign. larger in exposed than in not exposed subjects (p<0.001) [32] Concentration of PFBA in human plasma A samples (1 dataset): - Assigned value: NA - Average: 0.20 ng/ml - Median: 0.20 ng/ml - Min.: NA - Max.: NA - SD: NA - % relative SD: NA [40] Human: * Liver: - Mean: 12.9 ng/g wet weight - Median: 3.0 ng/g wet weight - Range: 128-BDL ng/g wet weight - MLOD: 6.00 ng/g wet weight - % of detection: 10 * Bone: - Mean: BDL - MLOD: 0.03 ng/g wet weight - % of detection: 0 [40] Human: * Brain: - Mean: 13.5 ng/g wet weight - Median: 1.4 ng/g wet weight - Range: 137-BDL ng/g wet weight - MLOD: 2.71 ng/g wet weight - % of detection: 25 * Lung: - Mean: 304.2 ng/g wet weight - Median: 807 ng/g wet weight - Range: 4138-BDL ng/g wet weight - MLOD: 0.01 ng/g wet weight - % of detection: 95 => PFBA was PFAS-compound with highest median concentration in lung [40] Human: * Kidney: - Mean: 464 ng/g wet weight - Median: 263 ng/g wet weight - Range: 4026-BDL ng/g wet weight - MLOD: 0.01 ng/g wet weight - % of detection: 95 => PFBA was PFAS-compound with highest median concentration in kidney [32] Concentration of PFBA in human plasma B samples (1 dataset): - Assigned value: NA - Average: 0.20 ng/ml - Median: 0.20 ng/ml - Min.: NA - Max.: NA - SD: NA - % relative SD: NA [47] PFBA in human blood samples: - Detection frequency: 100% - Mean +/- SD: 3.3 +/- 1.9 ng/mL - 5th percentile: 1.1 ng/mL - Median: 3.0 ng/mL - 95th percentile: 6.4 ng/mL => negatively (-0.343) associated with BMI (p<0.01) => positively (0.248) correlated with progressive motility (p<0.05) [40] - Since PFBA is a short chain compound, its predominance in lung could reflect the inhalation of contaminated dust and the industrial replacement of the eight carbons chain compounds by shorter ones [40] - Human half-life of this compound is much shorter compared to the half-life to other longer chain compounds, thereby accounting for its detection in other tissues such as kidney [117] Surface water and groundwater in a large area of the Veneto Region in northeastern Italy was contaminated with PFAS that were also found in drinking water samples. The health surveillance program is a free-of-charge population-based screening program. Residents who decided to participate in the program completed a structured interview administered by a trained public health nurse, followed by blood pressure measurement, and blood and urine sampling. A total of 18,345 residents were included in the present analysis. Only 3 of the 12 PFAS were quantifiable in at least 80% of serum samples: PFOA (99.9%), PFOS (99.8%) and PFHxS (98.1%) [117] Serum concentrations (ng/mL) of PFBA and percentage of samples above the LOQ in the study population (18,345 subjects, 14-39 years of age): PFBA: - Min: <0.5 - 5th percentile: <0.5 - 25th percentile: <0.5 - Median: <0.5 - 75th percentile: <0.5 - 95th percentile: <0.5 - Max: 23.9 % samples LOQ: 2.0% [40] - Smokers showed smaller accumulation of PFASs - Older people (more than 60 years) showed higher concentrations of PFASs => Clear indication that these compounds accumulate after a long-term exposure [71] Shorter chain PFASs have been shown to absorb into the liver more readily than those with longer chains which are concentrated in the blood proteins [117] PFBA and PFBS were found in high concentrations in drinking water but were detected only in a minority of serum samples at relatively low concentration, whereas PFOS and PFHxS, which were scarcely represented in drinking water, were detected in almost 100% of serum samples. This discrepancy may be explained by the exposure to PFOS and PFHxS from other sources, as demonstrated for the general population, and by the longer human half-lives of PFOS and PFHxS in comparison with PFBA and PFBS. Moreover, the exposure period to PFBA and PFBS was shorter in comparison with that of PFOA and PFOS. [11] humans: T1/2 = 0.0086 y [11] humans: Vd = 220 mL/kg [14] - rats: IV, 30 ppm: female: 1 h, male: 6 h - rats: oral, 30 ppm: female: 2 h, male: 9 h - mice: oral, 10 ppm: female: 3 h, male: 13 +/- 5 h - mice: oral, 30 ppm: female: 3 h, male: 16 +/- 7 h - monkeys: IV, 10 ppm: female: 41 h, male: 40 h - humans: female (n = 2): 87 +/- 31 h, male (n = 7): 68 +/35 h [3] humans: liver AUCSS: 1.9-2.8 liver weight ratio: 1.2-1.25 [3] humans: model liver: LOEL = 30 mg/kg/day liver weight ratio: 1.2-1.3 [29] NOAEL (rat): 6 mg/kg bw/d bw/DFI: 10 kg bw d/kg NOEC: 60 mg/kg [29] QS biota, secpois: Assessment factor: 300 on NOEC: 60 mg/kg => QS biota = 200 g/kg biota ww [29] QS dw,hh (human health via consumption of water): 7 g/L [118] Spatial and seasonal distributions of PFAS in air, water, sediment, soil, and fish samples in the Asan Lake area of South Korea. Estimation of bioaccumulation potential of PFAS using sediment-water partition coefficients and bioaccumulation factors. PFBA concentration in air samples: - Air (unit: pg/m3, n=4): Gaseous: * Range: na * Mean: na * Median: na * DF (%): na [102] * Comparison tap water and hot water NL: - Drinking water treatment plant A: Tap water: 10 ng/L & Hot water: 13 ng/L - Drinking water treatment plant B: Tap water: 17 ng/L & Hot water: 18 ng/L => Concentrations of PFAAs detected in hot water were similar to those in the corresponding tap water * Post-mixed cola NL (n=6): Average SD (Range): 11 4.0 (2.2-14) ng/L => Ratio tap water/post-mixed cola (Ratio of averages): 1.2 => PFAS concentrations in tap water were higher than in cola => A combination of the dilution step and the purification step inside the cola [2] observation well NL max. value: 1200 (10101280) ng/L [118] Particulate: * Range: na * Mean: na * Median: na * DF (%): na Total: * Range: na * Mean: na * Median: na * G (%): na [102] * Coffee NL: - Brewed coffee from coffee machines (n=12): Average SD (Range): Could not be quantified - Manually brewed coffee (n=4): Average SD (Range): Could not be quantified - Procedure blank brewed coffee (n=2): Average SD (Range): 14 1.1 (13-15) ng/L [24] ground water CN, maximum concentration: 3.610 g/L [39] NOEC and LOECs for PFAAs for transactivation of mouse and human PPAR in transiently transfected COS-1 Cells: PFBA: Mouse: - NOEC: 30 M - LOEC: 40 M - LOEC: 9.24 g/mL (p<0.0001) [39] NOEC and LOECs for PFAAs for transactivation of mouse and human PPAR in transiently transfected COS-1 Cells: PFBA: Human: - NOEC: 30 M - LOEC: 40 M - LOEC: 9.24 g/mL (p<0.05) [113] 397 food samples collected from a market in Busan (Korea) in 2011 (n=227) and 2012 (n=170). Food samples of 66 different food types, classified into 7 food categories. 34 tap water samples were also collected from 16 districts in Busan. PFBA concentrations in each food group (ng/g, beverage unit: ng/L): * Fish and shellfish (n=99): DF (%) = 7.1, Mean = 0.024, Max = 1.19 * Meat and its products (n=39): DF (%) = 7.7, Mean = 0.037, Max = 0.539 [113] * Vegetables and fruit (n=78): DF (%) = 11.5, Mean = 0.030, Max = 0.760 * Processed products (n=90): DF (%) = 7.8, Mean = 0.024, Max = 0.607 * Dairy (n=37): DF (%) = 5.4, Mean = 0.005, Max = 0.098 * Beverage (n=21): DF (%) = 28.6, Mean = 4.45, Max = 48.3 [30] observation wells SE: median water conc. PFBA: 13 ng/L [113] * Others (n=33): DF (%) = 9.1, Mean = 0.037, Max = 0.979 PFBA concentrations in bottled water and tap water samples (ng/L): * Bottled water (n=8): DF (%) = 12.5, Mean = 0.039, Max = 0.311 The short-chain PFCAs such as PFBA and PFPeA were detected more frequently in food samples with high moisture contents, such as beverages, vegetables, and fruit, than in food samples with low moisture contents [115] In September 2011, foods were purchased in 12 representative cities of Catalonia, all with more than 20,000 inhabitants. Food samples were obtained at each locality in 4 shops/stores of different size. Foods selected for PFAS analysis were among the most consumed in Catalonia. * Mean concentrations (in pg/g fw) of PFBA in the groups of analyzed foodstuffs: - Meat and meat products: <90 - Fish and seafood: <110 - Vegetables: <140 - Tubers: <110 - Fruits: <110 [115] - Eggs: <120 - Milk: <150 - Dairy products: 68 - Cereals: <38 - Pulses: <78 - Oils: <43 - Industrial bakery: <33 * PFBA was detected in semi-cured cheese. * Dietary intake of PFBA by the population of Catalonia (in ng/kg body weight/day): - Children: ND = 0: 0.03; ND = 1/2 LOD: 5.42 - Boys adolescents: ND = 0: 0.00; ND = 1/2 LOD: 1.63 - Girls adolescents: ND = 0: 0.00; ND = 1/2 LOD: 1.49 - Male adults: ND = 0: 0.00; ND = 1/2 LOD: 1.23 [115] - Female adults: ND = 0: 0.00; ND = 1/2 LOD: 1.51 - Male seniors: ND = 0: 0.00; ND = 1/2 LOD: 1.41 - Female seniors: ND = 0: 0.00; ND = 1/2 LOD: 1.29 * Concerning human health risks from dietary exposure to PFASs in Catalonia, it is important to note that for any of the age/gender groups of population estimated, the TDIs recommended by the EFSA were not exceeded. [28] Rivers SE: - Detection frequency: 100% - Concentration range: 0.47-3.7 ng/L - Average: 2.1 ng/L => predominant PFAS in rivers: 14% of total PFASs [16] GR near sea shore: - sediment: present in 2 out of 8 samples [2] estimated concentration of PFAA in public supply well field NL: 13 ng/L [1] ex. tap water DE 2 ng/L (19%) [9] ex. river water treatment NL - Raw/influent: 33 ng/L - Finished/tap water: 30 ng/L - Percent removal: 9.1 % [13] - river water NL: detectable concentrations: min. 4.1 ng/L - max. 14 ng/L - drinking water NL: detectable concentrations: 13 ng/L [32] Concentration of PFBA in canal water samples (8 datasets): - Assigned value: 12.2 ng/L - Average: 12.48 ng/L - Median: 12.06 ng/L - Min.: 6.70 ng/L - Max.: 20.05 ng/L - SD: 12.48 - % relative SD: 37 [28] Recipient sea SE: - Detection frequency: 100% - Concentration range: 0.34-0.67 ng/L - Average: 0.51 ng/L [117] In spring 2013, groundwater of a vast area of the Veneto Region (northeastern Italy) was found to be contaminated by perfluoralkyl substances (PFAS) from a PFAS manufacturing plant active since the late 1960s. Residents were exposed to high concentrations of PFAS, particularly perfluorooctanoic acid (PFOA), through drinking water until autumn 2013. [24] tap water CN, maximum concentration: 92.3 ng/L [10] total discharge of PFASs into the aquatic environment ranged between 10 g d-1 and 10 000 g d-1, depending on the water usage in the community connected to the sewage treatment plant [24] river water CN, maximum concentrations in * yearly monitoring: - 2011: 1.350 g/L - 2012: 1.570 g/L - 2013: 3.700 g/L - 2014: 2.580 g/L => significant increase from 2011 to 2014 (one-way ANOVA) * seasonal monitoring: autumn (3.700 g/L) > summer (2.440 g/L) > spring (1.970 g/L) > winter (628 ng/L) => peak river contamination periods: summer & autumn [72] * Direct injection analysis: - Surface water (river Lek, NL): <MQL - Riverbank filtrate (river Oude Rijn, NL): N.D. * Solid-phase extraction analysis: - Riverbank filtrate (river Oude Rijn, NL): N.D. - RO permeate: N.D. [118] Spatial and seasonal distributions of PFAS in air, water, sediment, soil, and fish samples in the Asan Lake area of South Korea. Estimation of bioaccumulation potential of PFAS using sediment-water partition coefficients and bioaccumulation factors. [118] PFBA concentration in water and sediment samples: - Water (unit: ng/L, n=47): * Range: na * Mean: na * Median: na * 75%: na * 95%: na * DF (%): na - Sediments (unit: ng/g dw, n=47): * Range: na * Mean: na * Median: na * 75%: na * 95%: na * DF (%): na [117] Surface water and groundwater in a large area of the Veneto Region in northeastern Italy was contaminated with PFAS that were also found in drinking water samples. A manufacturing plant located in the town of Trissino that produced PFAS since the late 1960s was identified as the only likely source of water contamination. [117] Measurements of 152 drinking water samples collected in July and August 2013 indicated that the main contaminants were PFOA, PFBA and PFBS, followed by PFPeA, PFHxA, PFOS, PFHpA and PFHxS. The longer-chain PFAS congeners (PFNA, PFDA, PFUnA, PFDoA) were detected only in a minority of samples and at lower concentrations. [117] PFBA concentrations (ng/L) in 152 samples of drinking water taken during July and August 2013, before the full implementation of granular activated carbon filters: - Min: <10.0 - 25th percentile: 86.0 - Median: 123.5 - 75th percentile: 173.3 - 95th percentile: 359.4 - Max: 625.0 - % samples LOQ: 88.8% [102] Tap water NL (n=4): Average SD (Range): 15 3.6 (9.3-18) ng/L => Levels of PFAAs observed in four tap water samples closely resembled each other as demonstrated by the relatively low standard deviation of the mean values [113] * Tap water (n=34): DF (%) = 11.8, Mean = 2.02, Max = 23.1 [17] SE - PFBA was the predominant PFCA in filtered effluent water with a mean level of 17 ng/L - Major contributions to the daily discharge Dd connected to the WWTP were from PFBA, ... [18] Effluent waste water: PFOA, PFBA and PFOS in 6 effluents ranged between 15-23 ng/L => TOP (oxidation): largest relative rise for PFBA [119] 3 replicate aqueous and solid samples were collected from each of 19 Australian WWTPs throughout 2017. Samples were taken from various stages within the treatment train from a range of WWTPs to determine the trends in the mass flux and partitioning of PFAS within the sampled WWTPs. Summary statistics for pooled aqueous (n=201, triplicates from 67 individual locations within 19 WWTPs) and pooled solid (n=51, triplicates from 5 primary and secondary sludge locations, 6 lagoon sludges and a lagoon dredge pile) samples. * PFBA concentration in aqueous samples (ng/L): [119] - Median: 5.8 - Mean: 13 - s.d.: 33 - min: <LOQ - max: 370 - Detect (%): 100% * PFBA concentration in solid samples (ng/g dw): - Median: <LOD - Mean: 0.45 - s.d.: 0.91 - min: <LOD - max: 4.1 - Detect (%): 29% [7] humans: concentration in plasma (neonates): geometric mean <LOQ median <LOQ range <LOQ/<LOQ concentration in plasma (adults aged 50-65): geometric mean <LOQ range <LOQ/0.29 [47] PFBS in human semen samples: - Detection frequency: 94% - Mean +/- SD: 0.11 +/- 0.11 ng/mL - 5th percentile: <LOQ - Median: 0.076 ng/mL - 95th percentile: 0.28 ng/mL => Much higher concentrations were found in semen, suggesting that blood-testis barrier may not work for these chemicals => positively (0.328) associated with BMI (p<0.01) => negatively (-0.302) correlated with progressive motility (p<0.01) [40] Human: * Liver: - Mean: 0.9 ng/g wet weight - Median: 0.7 ng/g wet weight - Range: 1.5-BDL ng/g wet weight - MLOD: 1.39 ng/g wet weight - % of detection: 0 * Bone: - Mean: 3.2 ng/g wet weight - Median: 2.4 ng/g wet weight - Range: 17.8-BDL ng/g wet weight - MLOD: 14.41 ng/g wet weight - % of detection: 5 [15] - rats: oral, 30 ppm: % of dose in urine at 24h: male: 68.6 +/- 12.9, female: 74.1 +/- 8.2 % of dose in urine at 96h: male: 0.27 +/- 0.21, female: 0.30 +/- 0.16 % of dose in feces at 24h: male: 0.50 +/- 0.17, female: 0.14 +/- 0.07 % of dose in feces at 96h: male: 0.04 +/- 0.04, female: 0.003 % of dose in liver at 96h: male: 0.03 +/- 0.003, female: 0.05 [14] monkeys: IV, 10 ppm: female: 8 +/- 2 h, male: 15 +/9 h [15] - rats: oral, 30 ppm: male: 676 +/- 55 mL/kg, female: 391 +/- 105 mL/kg - rats: IV, 30 ppm: male: 330 +/- 32 mL/kg, female: 351 +/- 34 mL/kg - monkeys: IV, 10 ppm: male: 254 +/- 31 mL/kg, female: 255 +/- 17 mL/kg [15] - rats: oral, 30 ppm: male: 0.151 +/- 0.015 h- 1, female: 0.095 +/- 0.009 h-1 (difference males p<0.05) - rats: IV, 30 ppm: male: 0.228 +/- 0.075 h-1, female: 0.175 +/- 0.009 h-1 [15] - rats: IV, 30 ppm: male: 119 +/- 34 mL/h, female: 469 +/- 40 mL/h (difference males p<0.01) - monkeys: IV, 10 ppm: male: 511 +/- 141 mL/h, female: 368 +/- 120 mL/h [15] Cmax serum: - rats: oral, 30 ppm: male: 86 +/- 5 g/mL, female: 102 +/- 15 g/mL - rats: IV, 30 ppm: male: 142 +/- 22 g/mL, female: 120 +/- 6 g/mL C serum 2h: - monkeys: male: 32.20 +/- 6.31, female: 55.03 +/- 6.55 g/mL [40] Human: * Brain: - Mean: BDL - MLOD: 0.96 ng/g wet weight - % of detection: 0 * Lung: - Mean: 17.8 ng/g wet weight - Median: 1.1 ng/g wet weight - Range: 9.7-BDL ng/g wet weight - MLOD: 2.10 ng/g wet weight - % of detection: 47 [15] - rats: IV, 30 ppm: % of dose in urine at 24h: male: 66.3 +/- 8.7, female: 74.4 +/- 6.0 % of dose in urine at 96h: male: 0.10 +/- 0.05, female: 0.08 +/- 0.03 % of dose in feces at 24h: male: 0.36 +/- 0.09, female: 0.13 +/- 0.03 % of dose in feces at 96h: male: 0.003, female: N/A % of dose in liver at 96h: male: N/A, female: N/A [15] - rats: oral, 30 ppm: male: 4.68 +/- 0.43 h, female: 7.42 +/- 0.79 h (difference males p<0.05) - rats: IV, 30 ppm: male: 4.51 +/- 2.22 h, female: 3.96 +/- 0.21 h - monkeys: IV, 10 ppm: male: 95.20 +/- 27.09 h, female: 83.20 +/- 41.90 h - humans: male (n = 5): 24.1 d, female (n = 1): 45.7 d [15] - rats: oral, 30 ppm: male: 163 +/- 10 g.h/mL, female: 85 +/- 12 g.h/mL (difference males p<0.01) - rats: IV, 30 ppm: male: 294 +/- 77 g.h/mL, female: 65 +/- 5 g.h/mL - monkeys: IV, 10 ppm: male: 24.3 +/- 8.6 g.h/mL, female: 35.4 +/- 13.3 g.h/mL [3] humans: model liver: no sign. effect up to 600 mg/kg/day [29] NOAEL (rat): 60 mg/kg bw/d bw/DFI: 20 kg bw d/kg NOEC: 1200 mg/kg [29] QS biota, secpois: Assessment factor: 90 on NOEC: 1200 mg/kg => QS biota = 13333 g/kg biota ww [29] QS dw,hh (human health via consumption of water): 3 g/L [118] Spatial and seasonal distributions of PFAS in air, water, sediment, soil, and fish samples in the Asan Lake area of South Korea. Estimation of bioaccumulation potential of PFAS using sediment-water partition coefficients and bioaccumulation factors. L-PFBS concentration in air samples: - Air (unit: pg/m3, n=4): Gaseous: * Range: nd ~ 2.66 * Mean: 0.67 * Median: 2.66 * DF (%): 25.0 [102] * Comparison tap water and hot water NL: - Drinking water treatment plant A: Tap water: 3.2 ng/L & Hot water: 3.3 ng/L - Drinking water treatment plant B: Tap water: 16 ng/L & Hot water: 19 ng/L => Concentrations of PFAAs detected in hot water were similar to those in the corresponding tap water * Post-mixed cola NL (n=6): Average SD (Range): 7.9 5.0 (<0.04-12) ng/L => Ratio tap water/post-mixed cola (Ratio of averages): 2.9 => PFAS concentrations in tap water were higher than in cola => A combination of the dilution step and the purification step inside the cola [2] observation well NL max. value: 91 (62-104) ng/L [118] Particulate: * Range: nd * Mean: * Median: * DF (%): 0.0 Total: * Range: nd ~ 2.66 * Mean: 0.67 * Median: * G (%): 100.0 [102] * Coffee NL: - Brewed coffee from coffee machines (n=12): Average SD (Range): 2.9 2.9 (<0.06-9.8) ng/L - Manually brewed coffee (n=4): Average SD (Range): 1.6 0.30 (1.3-2.0) ng/L - Procedure blank brewed coffee (n=2): Average SD (Range): 4.3 0.21 (4.0-4.5) ng/L [5] LOD = 0.3 ng/L Freq = 15.2 % max = 25 ng/L Average = 0 ng/L med = 0 ng/L Per90 = 1 ng/L [30] levels of PFBS in serum from participating mothers (SE): 1996-1999 (n=132): - MDL: 0.01 ng/g - Median (range): 0.019 (<MDL-0.21) ng/g - Mean (SE): 0.028 (0.003) ng/g - % of levels <MDL: 31 2008-2011 (n=134): - MDL: 0.01 ng/g - Median (range): 0.027 (<MDL-0.80) ng/g - Mean (SE): 0.055 (0.007) ng/g - % of levels <MDL: 19 => sign. higher compared to '96-'99 (p<0.05) [31] PFBS serum concentrations in exposed (living in area under impact) and not exposed (subjected to background exposure) subjects EXPOSED (n=257) - Minimum: <LOQ - Median: <LOQ - Maximum: 4.26 ng/g - % levels <LOQ: 69 NOT EXPOSED (n=250) - Minimum: <LOQ - Median: <LOQ - Maximum: 0.36 ng/g - % levels <LOQ: 82 => PFBS concentration sign. larger in exposed than in not exposed subjects (p<0.001) [40] Human: * Kidney: - Mean: 8.0 ng/g wet weight - Median: 1.7 ng/g wet weight - Range: 80.4-BDL ng/g wet weight - MLOD: 3.40 ng/g wet weight - % of detection: 5 [15] - monkeys: IV, 10 ppm: % of dose in urine from 0 to 24h: male: 48.9 +/- 7.9, female: 44.4 +/- 24.9 % of dose in urine from 144 to 168h: male: 0.15 +/- 0.07, female: 0.18 +/- 0.12 => Urine = major route of elimination in rat and monkey for PFBS => Human: serum concentrations were modestly correlated with paired urine concentrations and both concentrations declined over time such that at end-of-study urine was measured at the LOQ [30] short-chain PFBS has a much shorter half-life (26 days) than PFOS (5 years) and PFHxS (8 years) in humans [40] - Smokers showed smaller accumulation of PFASs - Older people (more than 60 years) showed higher concentrations of PFASs => Clear indication that these compounds accumulate after a long-term exposure [62] PFBS is eliminated in a little over 1 month in humans on average [39] NOEC and LOECs for PFAAs for transactivation of mouse and human PPAR in transiently transfected COS-1 Cells: PFBS: Mouse: - NOEC: 120 M - LOEC: 150 M - LOEC: 50.7 g/mL (p<0.01) [39] NOEC and LOECs for PFAAs for transactivation of mouse and human PPAR in transiently transfected COS-1 Cells: PFBS: Human: - NOEC: 20 M - LOEC: 30 M - LOEC: 10.1 g/mL (p<0.0001) [113] 397 food samples collected from a market in Busan (Korea) in 2011 (n=227) and 2012 (n=170). Food samples of 66 different food types, classified into 7 food categories. 34 tap water samples were also collected from 16 districts in Busan. PFBS concentrations in each food group (ng/g, beverage unit: ng/L): * Fish and shellfish (n=99): DF (%) = 0.0, Mean -, Max * Meat and its products (n=39): DF (%) = 0.0, Mean -, Max - [24] ground water CN, maximum concentration: 865 ng/L [113] * Vegetables and fruit (n=78): DF (%) = 0.0, Mean -, Max * Processed products (n=90): DF (%) = 1.1, Mean = 0.006, Max = 0.561 * Dairy (n=37): DF (%) = 0.0, Mean -, Max * Beverage (n=21): DF (%) = 4.8, Mean = 0.019, Max = 0.398 * Others (n=33): DF (%) = 12.1, Mean = 0.097, Max = 2.32 PFBS concentrations in bottled water and tap water samples (ng/L): * Bottled water (n=8): DF (%) = 0.0, Mean -, Max - [30] observation wells SE: median water conc. PFBS: 100 ng/L [32] Concentration of lin-PFBS in human plasma A samples (2 datasets): - Assigned value: NA - Average: 0.03 ng/ml - Median: 0.03 ng/ml - Min.: 0.01 ng/ml - Max.: 0.06 ng/ml - SD: 0.036 - % relative SD: 105 [32] Concentration of lin-PFBS in human plasma B samples (1 dataset): - Assigned value: NA - Average: 0.02 ng/ml - Median: 0.02 ng/ml - Min.: NA - Max.: NA - SD: NA - % relative SD: NA [47] PFBS in human blood samples: - Detection frequency: 100% - Mean +/- SD: 0.18 +/- 0.12 ng/mL - 5th percentile: 0.056 ng/mL - Median: 0.16 ng/mL - 95th percentile: 0.43 ng/mL => negatively (-0.333) associated with BMI (p<0.01) => positively (0.195) correlated with progressive motility (p<0.05) [117] Surface water and groundwater in a large area of the Veneto Region in northeastern Italy was contaminated with PFAS that were also found in drinking water samples. The health surveillance program is a free-of-charge population-based screening program. Residents who decided to participate in the program completed a structured interview administered by a trained public health nurse, followed by blood pressure measurement, and blood and urine sampling. A total of 18,345 residents were included in the present analysis. Only 3 of the 12 PFAS were quantifiable in at least 80% of serum samples: PFOA (99.9%), PFOS (99.8%) and PFHxS (98.1%) [71] Shorter chain PFASs have been shown to absorb into the liver more readily than those with longer chains which are concentrated in the blood proteins [115] In September 2011, foods were purchased in 12 representative cities of Catalonia, all with more than 20,000 inhabitants. Food samples were obtained at each locality in 4 shops/stores of different size. Foods selected for PFAS analysis were among the most consumed in Catalonia. * Mean concentrations (in pg/g fw) of PFBS in the groups of analyzed foodstuffs: - Meat and meat products: <2.4 - Fish and seafood: 54 [115] - Vegetables: 13 - Tubers: <3.0 - Fruits: <3.0 - Eggs: <3.2 - Milk: <4.1 - Dairy products: <1.8 - Cereals: <1.0 - Pulses: <2.1 - Oils: <1.2 - Industrial bakery: <0.89 * Fish and shellfish was the group in which more PFASs were detected and where the highest PFAS concentrations were found [115] * Dietary intake of PFBS by the population of Catalonia (in ng/kg body weight/day): - Children: ND = 0: 0.10; ND = 1/2 LOD: 0.24 - Boys adolescents: ND = 0: 0.01; ND = 1/2 LOD: 0.05 - Girls adolescents: ND = 0: 0.01; ND = 1/2 LOD: 0.05 - Male adults: ND = 0: 0.02; ND = 1/2 LOD: 0.05 - Female adults: ND = 0: 0.04; ND = 1/2 LOD: 0.08 - Male seniors: ND = 0: 0.03; ND = 1/2 LOD: 0.07 [115] * Concerning human health risks from dietary exposure to PFASs in Catalonia, it is important to note that for any of the age/gender groups of population estimated, the TDIs recommended by the EFSA were not exceeded. [28] Rivers: - Detection frequency: 59% - Concentration range: 0.030-19 ng/L - Average: 9.5 ng/L => predominant PFAS in rivers: 21% of total PFASs [28] Recipient sea: - Detection frequency: 65% - Concentration range: 0.062-0.57 ng/L - Average: 0.32 ng/L [2] estimated concentration of PFAA in public supply well field NL: 13 ng/L [1] ex. tap water DE 2.7 ng/L (42%) [9] ex. river water treatment NL - Raw/influent: 35 ng/L - Finished/tap water: 20 ng/L - Percent removal: 43 % [8] concentrations in water CN: - Dissolved phase: mean: 2.00 ng/L contribution to total PFASs: 0.17% - Separate suspended particulate matter: mean: 3.80 ng/L contribution to total PFASs: 8.68% - Soil: mean: 0.21 ng/L contribution to total PFASs: 2.93% - Leaves: mean: 0.42 ng/L contribution to total PFASs: 0.49% - Bark: mean: 0.07 ng/L contribution to total PFASs: 0.18% [32] Concentration of lin-PFBS in canal water samples (17 datasets): - Assigned value: 8.56 ng/L - Average: 10.10 ng/L - Median: 8.50 ng/L - Min.: 4.40 ng/L - Max.: 31.00 ng/L - SD: 10.10 - % relative SD: 61 [73] PFBS was quantified in 24% of all samples in the Greenland Sea, with concentrations ranging from <51 to 65 pg/L => This result could attribute to the shift of usage from C8 to C4-C6 PFASs after the voluntary phase-out of POSF and PFOA since 2000 [12] SE Drinking water: PFSAs 29% of total PFASs [6] tap water MDL = 0.028 ng/L MLQ = 0.092 ng/L max. value NL: 18.8 ng/L [10] total discharge of PFASs into the aquatic environment ranged between 10 g d-1 and 10 000 g d-1, depending on the water usage in the community connected to the sewage treatment plant [13] - river water NL: detectable concentrations: min. 12 ng/L - max. 27 ng/L - drinking water NL: detectable concentrations: min. 0.54 ng/L - max. 19 ng/L [118] Spatial and seasonal distributions of PFAS in air, water, sediment, soil, and fish samples in the Asan Lake area of South Korea. Estimation of bioaccumulation potential of PFAS using sediment-water partition coefficients and bioaccumulation factors. [73] * North Atlantic Ocean (NAO): - 2007 (n=39): <1.6-45 pg/L - 2008 (n=10): <4.4-20 pg/L - 2010 (n=13): <51 pg/L * Middle Atlantic Ocean (MAO): - 2007 (n=10): <1.6 pg/L - 2008 (n=5): <4.4-17 pg/L - 2010 (n=5): <51 pg/L * South Atlantic Ocean (SAO): - 2007 (n=10): <1.6 pg/L - 2008 (n=13): <4.4-13 pg/L - 2010 (n=16): <51 pg/L [118] L-PFBS concentration in water and sediment samples: - Water (unit: ng/L, n=47): * Range: 1.4 ~ 43.6 * Mean: 17.0 * Median: 12.2 * 75%: 29.6 * 95%: 42.3 * DF (%): 100.0 - Sediments (unit: ng/g dw, n=47): * Range: nd ~ 0.51 * Mean: 0.10 * Median: 0.08 * 75%: 0.15 * 95%: 0.40 * DF (%): 74.5 [118] The predominant PFAS in water was PFPeA, followed by PFHxA. Shortchain PFAS such as PFPeA, PFHxA and LPFBS were detected more frequently in water than in the other environmental matrices. Short-chain PFAS and PFOA were detected during all four seasons. The seasonally dominant PFCAs were PFHxA and PFOA during summer and PFPeA and PFHxA during the other seasons. [66] PFBS levels (ng/L) in outgoing drinking water from two waterworks in Ronneby (SE): - Site 1 (Brantafors): 130 - Site 2 (Krragarden): <2.6 [24] tap water CN, maximum concentration: 97.8 ng/L [17] SE - PFBS was predominantly present in filtered effluent water with a mean level of 1.9 ng/L => highest concentration found was 3.7 ng/L - Other PFSAs detected in sludge were PFBS, ... - Highest concentration found in influent water was 3.2 ng/L [24] river water CN, maximum concentrations in * yearly monitoring: - 2011: 2.900 g/L - 2012: 2.340 g/L - 2013: 2.710 g/L - 2014: 3.780 g/L => significant increase from 2012 to 2014 (one-way ANOVA) * seasonal monitoring: summer (3.870 g/L) > autumn (2.710 g/L) > spring (1.980 g/L) > winter (690 ng/L) => peak river contamination periods: summer & autumn [117] In spring 2013, groundwater of a vast area of the Veneto Region (northeastern Italy) was found to be contaminated by perfluoralkyl substances (PFAS) from a PFAS manufacturing plant active since the late 1960s. Residents were exposed to high concentrations of PFAS, particularly perfluorooctanoic acid (PFOA), through drinking water until autumn 2013. [102] Tap water NL (n=4): Average SD (Range): 16 1.4 (14-17) ng/L => Levels of PFAAs observed in four tap water samples closely resembled each other as demonstrated by the relatively low standard deviation of the mean values [119] 3 replicate aqueous and solid samples were collected from each of 19 Australian WWTPs throughout 2017. Samples were taken from various stages within the treatment train from a range of WWTPs to determine the trends in the mass flux and partitioning of PFAS within the sampled WWTPs. Summary statistics for pooled aqueous (n=201, triplicates from 67 individual locations within 19 WWTPs) and pooled solid (n=51, triplicates from 5 primary and secondary sludge locations, 6 lagoon sludges and a lagoon dredge pile) samples. * PFBS concentration in aqueous samples (ng/L): - Median: 2.5 - Mean: 4.0 - s.d.: 4.9 [117] Surface water and groundwater in a large area of the Veneto Region in northeastern Italy was contaminated with PFAS that were also found in drinking water samples. A manufacturing plant located in the town of Trissino that produced PFAS since the late 1960s was identified as the only likely source of water contamination. [113] * Tap water (n=34): DF (%) = 58.8, Mean = 0.801, Max = 1.80 [119] - min: <LOD - max: 33 - Detect (%): 98% * PFBS concentration in solid samples (ng/g dw): - Median: <LOD - Mean: 0.83 - s.d.: 2.0 - min: <LOD - max: 9.3 - Detect (%): 44% [117] Measurements of 152 drinking water samples collected in July and August 2013 indicated that the main contaminants were PFOA, PFBA and PFBS, followed by PFPeA, PFHxA, PFOS, PFHpA and PFHxS. The longer-chain PFAS congeners (PFNA, PFDA, PFUnA, PFDoA) were detected only in a minority of samples and at lower concentrations. [117] PFBS concentrations (ng/L) in 152 samples of drinking water taken during July and August 2013, before the full implementation of granular activated carbon filters: - Min: <10.0 - 25th percentile: 66.5 - Median: 91.5 - 75th percentile: 183.2 - 95th percentile: 382.1 - Max: 765.0 - % samples LOQ: 89.5% 249-616-3 29420-49-3 Potassium K-PFBS C4 1,1,2,2,3,3,4,4,4- nonafluorobutane-1- sulphonate Registered 816-392-9 30334-69-1 Perfluorobutylsulphonam FBSA C4 ide Not registered [77] Remediation of PFASs: * Stabilization: managing PFASs source zone soils in situ provides a sustainable approach that does not create a concentrated waste requiring offsite management or destruction * High energy electron beam: further evaluation necessary * Low/high temperature thermal desorption: a high degree of licensing and stakeholder engagement will likely be required due to the operation of a thermal waste recovery facility and the associated off-gasses * Vapor energy generator process: no full-scale application of VEG for PFASs has been implemented * Ball milling: viability of ball milling has not been studied in detail with respect to the complications of PFASs (short-chain compounds and polyfluorinated precursors) [82] Study of inhibition of cellular viability in human liver cell line (HL-7702): * PFBS single: - IC-20 (20% stimulatory effect concentration): 3.35 x 10^(-5) mol/L - IC0 (0% inhibitory concentration): 4.81 x 10^(-3) mol/L - IC10 (10% inhibitory concentration): 5.07 x 10^(-3) mol/L - IC50 (50% inhibitory concentration): 6.49 x 10^(-3) mol/L * PFBA-PFBS mixture: - all IC0, IC10 and IC50 values in the binary mixture were less than that of individual PFSA or PFCA values - a synergistic action took place under the effective concentrations of IC0, IC10 and IC50 with various concentration ratios - for the effective concentration of IC-20, only the higher proportion of PFSA in the mixtures (PFBS > 57.7%) showed a synergistic effect for IC-20 [82] - Mixtures of eleven PFAAs (Mix0, Mix10 and Mix50, PFBS included) presented a partial addition effect - Mixtures of nine PFAAs (Mix0, Mix10, Mix50, PFBS included) with J-shaped curves only presented a synergistic effect with MTI values ranging from 1.16 to 1.25, however Mix-20 showed partial addition (MTI = 0.86) - A mixture of seven PFAAs (Mix-20, PFBS included) showed a [84] PFCs can compete with thyroxine (T4, the transport form of thyroid hormone), for binding to the human thyroid hormone transport protein transthyretin (TTR) => such competitive capacity may lead to decreased thyroid hormone levels as previously reported for animals exposed to PFCs => PFCs do not affect the regulatory functions of the thyroid hormone system itself, but it is the competitive binding to transport proteins that alters the free thyroxine (T4) levels in blood => TTR is the main T4 carrier in cerebrospinal fluid, and also important in serum of most mammalian species and birds => Binding potency is clearly associated with the degree of fluorination of the alkyl chain => PFAAs with a carbon chain length longer than eight have low TTR binding potencies => TTR binding potencies were significantly higher for compounds containing a sulfonate functional group than for those containing a carboxylic acid functional group => TTR is both in humans and in rodents the most important carrier protein for thyroid hormone to the developing fetus and the brain [84] PFBS: - Molecular weight: 300.0 g/mol - T4-TTR binding at maximum concentration (10M): 69% - IC50 (concentration at 50% inhibition): 19,460 nM - Slope of dose-response curve: (-1.38) - T4-REP (relative potency compared to T4) factor: 0.003 - HPLC retention time: 24.2 min - Purity: N.V. [85] Mean fold change standard error of mRNA expression after exposure to 100 M of PFBS (in rat H4IIE hepatoma cells): * Thyroid related genes: - Hex: 1.14 0.18 - PAX 8: 4.80 0.42 => upregulation of thyroid related genes * Cholesterol related genes: - Per-3-Keto-: 0.25 0.07 - SQSYN: 0.43 0.04 => downregulation of cholesterol related genes * Lipoprotein related genes: - Mito-3-Keto-: 0.31 0.03 => downregulation of Mito-3-Keto- - Mito-3-Keto-: 1.98 0.31 => upregulation of Mito-3-Keto- - ApoA4: 0.56 0.15 [86] - weak binding of PFBS with the protein TR-LBD (TR = thyroid hormone receptor, LBD = ligand-binding domain) * Binding of PFBS with TR-LBD: - Length: 7.84 A - IC50: >1,000 M - RP (relative potency (IC50 T3 / IC50 chemical)): 0.0003 - Hydrogen bonding: ARG 228 * all tested PFCs fit into the T3-binding pocket of TR-LBD, with the acid or hydroxyl end group residing toward the inner part and the hydrophobic chain toward the entrance of the binding pocket => all PFCs with an acid end group formed a hydrogen bond with ARG 228 [87] * Binding potency of PFBS to TTR (TTR competitive binding assay): - IC50: 13,331 1019 nM - Kd: 2132 163 nM - RP (relative potency (IC50 T4 / IC50 chemical)): 0.002 => perfluoroalkyl acids: Kd values decreased as carbon chain length increased from C4 to C8 => no further increase, but a downward trend was observed for perfluoroalkyl acids with longer chain lengths (C9-C14) * Binding potency of PFBS to TTRmutK15G: - IC50: ND - Kd: ND - Hydrogen bonding: Lys15 => PFASs exhibited much weaker binding affinities to TTRmutK15G compared with wild-type TTR * Docking study: - Similar to T4, PFOA and PFOS could nearly fill the TTR ligandbinding pocket - Perfluoroalkyl acids with carbon chain length less than C8 did not adequately fill the T4 binding pocket [87] * Binding potency of PFBS to TBG (TBG competitive binding assay): - IC50: ND - Kd: ND - RP (relative potency (IC50 T4 / IC50 chemical)): ND => only PFTA and PFTdA bound to TBG * Binding potency of PFBS to TBGmutR378G: - IC50: ND - Kd: ND * Binding potency of PFBS to TBGmutR381G: - IC50: ND - Kd: ND - Hydrogen bonding: Arg381 => both TBGmut378G and TBGmutR381G exhibited much weaker binding potencies to PFTA and PFTdA when compared with that of wild-type TBG * Docking study: - Perfluoroalkyl acids with carbon chain length less than C12 did not adequately fill the T4 binding pocket - Longer fluorinated carbon chain structures could nearly fill the TBG ligand-binding pocket [109] Study on residential exposure: Vacuum cleaner contents from 39 homes were tested for perfluoroalkyl chemicals: * PFBS: - Detection frequency: 59 %, Median: 1.8 (ND-31) ng/g - Correlation between PFBS and PFHxS (0.40, p<0.05) - Correlation between PFBS and PFHpS (0.44, p<0.05) - Correlation between PFBS and PFBA (0.43, p<0.05) - Correlation between PFBS and PFPeA (0.38, p<0.05) - Correlation between PFBS and PFHpA (0.32, p<0.05) - Correlation between PFBS and PFOA (0.36, p<0.05) * Considering data from all households, home age was negatively associated with PFBS levels (p=0.02) * PFBS was detected in 14 of 30 (47%) dust samples from homes built before 1995 and in all of the dust samples from the nine homes built after that date [109] * Dust from two homes built after 2003 only contained trace amounts of PFOS, PFOA and PFHxS, but had higher PFBS levels than found in most of the older homes in our study * Occupants of all ages can be exposed to indoor contaminants, but preschool-aged children have the greatest risk due to hand to mouth activity and the [111] PFAAs were investigated in home produced eggs and commercially produced eggs surrounding a fluorochemical industrial park in China: PFBS concentrations: * Home produced eggs (n=4, sites 1-4): Egg yolks: - 2 km distance: <0.02 ng/g - 5 km distance: <0.02 ng/g - 10 km distance: <0.02 ng/g - 20 km distance: <0.02 ng/g Egg whites: - 2 km distance: <0.02 ng/g - 5 km distance: <0.02 ng/g - 10 km distance: <0.02 ng/g - 20 km distance: <0.02 ng/g Whole eggs: - 2 km distance: <0.02 ng/g - 5 km distance: <0.02 ng/g - 10 km distance: <0.02 ng/g - 20 km distance: <0.02 ng/g [111] * Commercially produced eggs (n=12, sites 5-16): Egg yolks: - Min: <0.02 ng/g - Max: <0.02 ng/g - Mean: <0.02 ng/g - Median: <0.02 ng/g - n > LOD (%): 0 (0) [111] Egg whites: - Min: <0.02 ng/g - Max: <0.02 ng/g - Mean: <0.02 ng/g - Median: <0.02 ng/g - n > LOD (%): 0 (0) Whole eggs: - Min: <0.02 ng/g - Max: <0.02 ng/g - Mean: <0.02 ng/g - Median: <0.02 ng/g - n > LOD (%): 0 (0) [116] In May 2009, indoor dust samples (n=17) were collected from fifteen homes located in nine Korean cities for house dust (n=15), and one shopping mall and one library in Seoul for non-residential indoor dust (n=2). Nine cities for house dust samples were chosen to represent densely populated cities, industrialized cities, and rural area. [116] * Concentrations (ng/g) of PFBS measured in indoor dust samples from Korea (n=17): - House dust (n=15): Range: <LOD - 1.3 Mean ( SD): 0.5 ( 0.5) Median: 0.3 % of < LOD: 27 - Non-residential indoor dust (n=2): Shopping mall: 1.6 Library: 41.1 * There were no significant correlations between PFCAs (or PFSAs) and their precursors in house dust. * In Korean house dust samples, significant correlations (p<0.05) were only found among all the FTOHs and among almost all of the PFCAs. 206-792-6 375-72-4 1,1,2,2,3,3,4,4,4- PBSF C4 nonafluorobutane-1- sulphonyl fluoride Registered [23] ultimate transformation products of many fluorinated alternatives (such as PBSF) are short-chain PFAAs and other substances such as COF2 => COF2 is highly toxic 206-782-1 375-01-9 2,2,3,3,4,4,4heptafluorobutan-1-ol 3:1 FTOH C4 (3 fluorinated) Pre-registered [75] logKaw: 1.98 (estimated by COSMOtherm) => lower Kaw than 8:2 FTOH [75] logKow (dry): 1.83 (estimated by COSMOtherm) => lower Kow than 8:2 FTOH [75] logKlipw: 1.96 (estimated with regression equation) [75] logKoa = 3.82 (estimated by COSMOtherm) [75] pKa = 12.05 (estimated by SPARC) [75] - half-life in air with AOPWIN: 2026.75 hour - half-life in water with BIOWIN3: 2880 hour - half-life in soil with BIOWIN3: 2880 hour [75] overall persistence (Pov): 197.70 days => lower Pov than 8:2 FTOH 220-300-7 2706-90-3 Perfluorovaleric acid PFPeA C5 Pre-registered [79] log Kaw: 0.86 (calculated from equations) [29] PFPeA: logKow ranges from 0.09 to 3.43 (exper. + modelled) [65] PFPeA: - logCMC Exp: -0.70 - logCMC Pred: -0.64 - 3: 6.49 [65] PFPeA: logAqS (aqueous solubility): - Y-Pred: 2.08 mg/L - EPI Suite Pred: 1.79 mg/L [65] PFPeA: logpL (sub- cooled vapor pressure): - Y-Pred: 0.41 mm Hg - EPI Suite Pred: 0.71 mm Hg - Y-Pred: 2.53 Pa [79] log Koa: 4.48 (calculated from equations) [9] PFCAs and PFSAs have low vapour pressures which decrease with increasing carbon chain length => Low potential for volatilization [9] Calculated pKa values indicate that both PFCAs and PFASs are strong acids which will predominantly be in their dissociated negatively-charged form at environmentally relevant pH values [29] PFPeA: logKoc: 2.4 (EPIsuite) [79] - Molar mass (MM): 264.1 g/mol - Molar volume (VM): 154.9 cm^3/mol (by ACD/Labs' ACD/PhysChem Suite) - Total surface area (TSA): 244.9 A^2 (by ChemAxon) - Tm (experimental melting point): 413 K (data obtained from ChemSpider) [17] shorter-chained PFAS (C<=5): electrostatic interactions have been hypothesized to be of more importance than hydrophobic interactions for sorption to sludge => partition to sludge increases with decreasing chain length [9] Removing PFASs from drinking water: * Granular Activated Carbon Absorption * Ion exchange/non-ion exchange resins may be useful for removing PFASs * High pressure membranes will achieve high rejection of most PFASs, however lower molecular weight PFASs, such as PFPeA, may be less well rejected [63] Perfluoroalkyl and perfluoroether moieties are highly persistent under natural conditions => Even though some PFASs may partially degrade in the environment and biota, they will all ultimately transform into highly stable end products, wich are usually perfluoroalkyl or perfluoroalkyl(poly)ether acids such as PFCAs, PFSAs, PFECAs and PFESAs [29] PFPeA: BCF insignificant BMF <0.1 [29] - max. concentration in biota: 9.69 ng/g ww - PFPeA can be considered not bioaccumulable or biomagnificable in the aquatic food web [39] Relative response of mouse and human PPAR to PFAAs in transiently transfected COS-1 cells, measured by C20max: PFPeA: - Mouse: 45 M (Most active to least active: ranked 9/13 PFASs) - Human: 52 M (Most active to least active: ranked 6/13 PFASs) => A lower C20 max value indicates a stronger induction of PPAR alpha activity by that PFAA => PPAR is activated by PFAAs of 4-12 carbons in chain length => Increasing activity of PPAR with increasing chain length of the PFAA up to C9, and lower activity with longer chain PFAAs (C>9) with both mouse and human PPAR => PPAR has been identified as a key player in the mode of action for PFAA-induced toxicity [42] FRESHWATER: Brachionus calyciflorus - rate of population increase was 22.4% lower in PFPeA than in the control (p<0.05) => decline of population growth rate poses a chronic and longterm risk on rotifer population - mictic ratios increased by 1.60-fold in PFPeA compared to the control (p<0.05) => increase of mictic ratio was due to the change of rotifers reproduction => under exposure of environmental stress rotifers transform from asexual reproduction to sexual reproduction => fertilized eggs can resist poor outside environment and keep balance in population dynamics => PFPeA may interfere with endocrine signaling in B. calyciflorus , causing a reduction of population growth rate and an increase of mictic ratio [29] PFAS with carbon chain C<=6: - low bioaccumulative potential - low acute and chronic aquatic toxicity => no direct concern for aquatic environment [34] significant correlation with other PFAS-analytes detected in groundwater (p<0.05) and surface water (p<0.01), indicating a common or similar source of exposure [65] PFPeA: EPI pred logKow: 3.40 [79] log Kow: -0.02 (calculated from equations) [79] Water solubility (SL): log (SL/mol L^-1): -1.43 (calculated from equations) [79] Vapor pressure (PL): log (PL/Pa): 2.18 (calculated from equations) [74] PFPeA: * R (retardation factor) (40 cm): 1.3 * log Kd (40 cm): -1.4 * log Koc (40 cm): 1.8 [29] Limited accumulation has been measured in freshwater and coastal sediments PFPeA: <1 ng/g dw [10] - volatile PFASs (FTOHs, FASAs, FASEs) enter the atmosphere, where they can degrade, form intermediates during atmospheric oxidation, or transform into more persistent PFASs, such as PFSAs and PFCAs, which may finally end up in the aquatic environment - PFASs have the potential to affect cell membranes of algae => can alter nutritious quality of biofilms => potential effects on physiological fitness of invertebrates [73] - PFPeA was usually detected in urban rivers and waste water treatment plant effluents, and it could not be effectively removed by common treatment techniques as well as filtration using activated carbon - The abundance of PFPeA could be attributed to the direct release from the urban and industrial regions and the defective water treatment processes [63] the high solubility and proteinbinding characteristics of ionic PFAAs challenge the conventional assessment of bioaccumulation potential that is through either bioconcentration factor in aquatic species (usually fish) or models based on octanol-water partition coefficients (Kow) => even when PFAAs are not characterized as bioaccumulative under the current regulatory frameworks, the accumulation of these PFAAs in the environment will lead to increasing external exposure [44] All tested PFCs induced hPXR reporter activites in a dosedependent manner => In general, longer carbon chain length seems to be favorable for the hPXR activity of PFAAs with chain length shorter than 10, as chain length is longer than 10, increasing the chain length of PFAA seems to render the chemicals weaker potency to induce hPXR activity => PFPeA: EC50 = 19.12 M, maximum induction = 32 % ==> PXR induction is postulated to be associated with lipid homeostasis, atherosclerosis, carcinogenesis and endocrinedisrupting effects [42] FRESHWATER: Brachionus calyciflorus - body size values were increased from 0.125 in the control to 0.137 m3 in PFPeA (p<0.05) - egg size values were increased from 0.514 in the control to 0.613 m3 in PFPeA (p<0.05) => females produce larger and fewer eggs under the condition of the increasing cost of producing larger eggs => large eggs produce large larvae and large offspring which exert a strong influence to the ability of survival => response to toxicity [77] short-chain PFAAs have been found to concentrate in the edible portion of some crops, such as fruits and vegetables, wheras longer chains tend to be retained more in the shoots and roots of plants [47] negative correlations were significantly observed between sperm motility and PFPeA in human semen, therefore exposure to PFAAs may result in a decline in semen mobility [104] Breast milk samples taken from 264 Korean lactating women: * Level of PFPeA in breast milk: - n > LOD: 216 => detected in 81.8% of all breast milk samples - Median (Interquartile range): 0.058 (0.032-0.087) ng/mL - Geometric mean: 0.053 ng/mL * PFPeA was significantly correlated with PFHxA (0.31667, p<0.001) and PFOS (0.17078, p<0.05) [63] the very high persistence of PFAAs leads to poorly reversible exposure to these substances in the global environment and some local/regional environments including groundwater [109] Study on residential exposure: Vacuum cleaner contents from 39 homes were tested for perfluoroalkyl chemicals: * PFPeA: - Detection frequency: 97 %, Median: 5.4 (ND- 32) ng/g - Correlation between PFPeA and PFBS (0.38, p<0.05) - Correlation between PFPeA and PFHxS (0.49, p<0.05) - Correlation between PFPeA and PFOS (0.67, p<0.05) - Correlation between PFPeA and PFHpA (0.91, p<0.05) - Correlation between PFPeA and PFOA (0.78, p<0.05) - Correlation between PFPeA and PFNA (0.54, p<0.05) - Correlation between PFPeA and PFDA (0.46, p<0.05) - Correlation between PFPeA and PFUA (0.48, p<0.05) - Correlation between PFPeA and PFDoA (0.45, p<0.05) - Correlation between PFPeA and PFTrA (0.47, p<0.05) [77] - the extreme persistence of all PFAAs with the increasingly higher aqueous solubility (and thus PFAA mobility) as the chain length shortens, also creates increasing difficulty for them to be treated by many conventional remediation and water treatment technologies - Remediation of PFASs: * Biological treatment: potential for anaerobic degradation of PFASs seems limited, as no organisms have been isolated which can utilize PFAAs as terminal electron acceptors * GAC (granular activated carbon) can be inefficient at removing PFOA and becomes progressively less effective for removing shorter chain PFCAs such as PFHxA, PFPeA, PFBS and PFBA as the chain length [77] - Remediation of PFASs: * in the same manner that GAC is less effective at removing shorter chain PFAAs and PFAA precursors from water, the injected particulate-activated carbon is also expected to be less effective at removing these PFASs * Ion-exchange resins: long-chain PFAA and precursors are likely to be removed from water significantly more strongly than short-chain PFASs * Osorb is similarly effective as GAC at removing both PFOS and PFOA and more effective than GAC at removing PFBA * Precipitation removal rates for shorter chain PFAAs improve when higher doses of the coagulant are applied [77] Remediation of PFASs: * Ozofractionation: can convert polyfluorinated precursors to PFAAs via chemical oxidation which assist their removal via foam fractionation and can remove short-chain PFASs in addition to long chain * In situ foam fractionation: limited application * RO and NF have been shown to be extremely effective in removing PFASs regardless of chain length and are also expected to be effective at removing many types of PFAA precursors * Concerns regarding chemical oxidation for in situ application include the generation of sign. concentrations of more mobile short-chain PFASs * Chemical reduction: application at field-scale for in situ PFAA remediation is challenging from a practicality perspective * Sonolysis has been demonstrated at laboratory scale for PFASs but not scaled up for commercial use [50] NIH-3T3 cells: PFPeA led to increased PPAR (at 30 M, p<0.01 and 100 M, p<0.001) and PPAR activity (at 100 M, p<0.001) => as PPAR is involved in adipocyte differentiation, PFAS affecting this receptor may be involved in development of obesity => short-chain PFCAs generally activate both PPAR and PPAR with similar potency and efficacy as long-chain PFCAs [105] Analysis of breast milk samples (n=293) collected from 127 mothers in Korea. Pairs of pregnant women and fetuses that were recruited from four South Korean cities. Detection rates and concentrations of PFPeA in Korean breast milk samples: * No. detected: 1 => detected in 0.3% of all breast milk samples * Range (min-max): <25 - 12.7 ng/L * Average SD: 0.04 0.74 ng/L => Study showed low detection rates of contaminant short-chain PFCAs [85] Mean fold change standard error of mRNA expression after exposure to 100 M of PFPeA (in rat H4IIE hepatoma cells): * Thyroid related genes: - Hex: 16.03 2.39 - PAX 8: 2.20 0.30 => upregulation of thyroid related genes * Cholesterol related genes: - Per-3-Keto-: 4.55 0.85 - SQSYN: 1.67 0.27 => upregulation of cholesterol related genes * Lipoprotein related genes: - Mito-3-Keto-: 1.09 0.12 - Mito-3-Keto-: 4.47 0.60 => upregulation of Mito-3-Keto- & Mito-3-Keto- - ApoA4: 0.72 0.13 => downregulation of ApoA4 [106] Study investigated the relationship between serum concentrations of perfluoroalkyl compounds and growth parameters in 2-year-old Korean children. Study included 361 children aged 2 years (192 boys and 169 girls; 22-27 months). PFPeA concentrations in serum from children at 2 years of age: - LOD: 0.076 ng/mL - Detection frequency: 122 => detected in 33.8% of all serum samples - Range (min-max): 0.054 - 7.780 ng/mL - Mean SD: 0.209 0.471 ng/mL [109] - Correlation between PFPeA and PFTetraA (0.35, p<0.05) * Among 11 homes built after 1990, home age was positively correlated with PFPeA, PFHpA and PFOA levels at p<0.05, and with PFOS at p<0.10 * Occupants of all ages can be exposed to indoor contaminants, but preschool-aged children have the greatest risk due to hand to mouth activity and the amount of time they spend playing on or near the floor [110] Study on chicken/duck poultry products and eggs in China => 2 or 3 samples were collected from henneries or abattoirs at different locations => Chicken (n=53), duck (n=47), chicken eggs (n=56, each sample being a combination of five eggs), and duck eggs (n=56, each sample being a combination of five eggs) => 3 tissues from chicken and ducks: breast meat, liver and subcutaneous fat => Final sample number: 300 samples from chicken (53*3=159) and duck (47*3=141): * Eggs: Concentrations of PFPeA were highest in chicken eggs * Meat: Concentrations of PFPeA were highest in duck meat (0.31 g/kg) => The current PFC concentrations in chicken, chicken eggs, duck and duck eggs from the Yangtze River Delta and Pearl River Delta are not of concern to adults => Chicken and chicken eggs could be dangerous to [111] PFAAs were investigated in home produced eggs and commercially produced eggs surrounding a fluorochemical industrial park in China: PFPeA concentrations: * Home produced eggs (n=4, sites 1-4): Egg yolks: - 2 km distance: 0.57 ng/g - 5 km distance: 0.11 ng/g - 10 km distance: 0.17 ng/g - 20 km distance: <0.05 ng/g Egg whites: - 2 km distance: <0.05 ng/g - 5 km distance: <0.05 ng/g - 10 km distance: <0.05 ng/g - 20 km distance: <0.05 ng/g Whole eggs: - 2 km distance: 0.22 ng/g - 5 km distance: <0.05 ng/g - 10 km distance: 0.07 ng/g - 20 km distance: <0.05 ng/g [117] Serum concentrations (ng/mL) of PFBS and percentage of samples above the LOQ in the study population (18,345 subjects, 14-39 years of age): PFBS: - Min: <0.5 - 5th percentile: <0.5 - 25th percentile: <0.5 - Median: <0.5 - 75th percentile: <0.5 - 95th percentile: <0.5 - Max: 8.9 % samples LOQ: 3.1% [117] PFBA and PFBS were found in high concentrations in drinking water but were detected only in a minority of serum samples at relatively low concentration, whereas PFOS and PFHxS, which were scarcely represented in drinking water, were detected in almost 100% of serum samples. [117] This discrepancy may be explained by the exposure to PFOS and PFHxS from other sources, as demonstrated for the general population, and by the longer human half-lives of PFOS and PFHxS in comparison with PFBA and PFBS. Moreover, the exposure period to PFBA and PFBS was shorter in comparison with that of PFOA and PFOS. [31] PFPeA serum concentrations in exposed (living in area under impact) and not exposed (subjected to background exposure) subjects EXPOSED (n=257) - Minimum: <LOQ - Median: <LOQ - Maximum: 0.46 ng/g - % levels <LOQ: 74 NOT EXPOSED (n=250) - Minimum: <LOQ - Median: <LOQ - Maximum: 0.22 ng/g - % levels <LOQ: 60 => PFPeA concentration sign. larger in exposed than in not exposed subjects (p<0.001) [47] PFPeA in human semen samples: - Detection frequency: 100% - Mean +/- SD: 2.1 +/- 1.5 ng/mL - 5th percentile: 0.57 ng/mL - Median: 1.7 ng/mL - 95th percentile: 5.1 ng/mL => Much higher concentrations were found in semen, suggesting that blood-testis barrier may not work for these chemicals => positively (0.264) associated with age (p<0.01) => positively (0.372) associated with BMI (p<0.01) => negatively (-0.344) correlated with progressive motility (p<0.01) [40] Human: * Liver: - Mean: 1.4 ng/g wet weight - Median: BDL ng/g wet weight - Range: 27.1-BDL ng/g wet weight - MLOD: 0.001 ng/g wet weight - % of detection: 5 * Bone: - Mean: 0.8 ng/g wet weight - Median: 0.8 ng/g wet weight - Range: 0.8-BDL ng/g wet weight - MLOD: 1.51 ng/g wet weight - % of detection: 0 [32] Concentration of PFPeA in human plasma A samples (2 datasets): - Assigned value: NA - Average: 1.35 ng/ml - Median: 1.35 ng/ml - Min.: 0.09 ng/ml - Max.: 2.60 ng/ml - SD: 1.77 - % relative SD: 132 [40] Human: * Brain: - Mean: BDL - MLOD: 0.59 ng/g wet weight - % of detection: 0 * Lung: - Mean: 44.5 ng/g wet weight - Median: 40.8 ng/g wet weight - Range: 695-BDL ng/g wet weight - MLOD: 6.006 ng/g wet weight - % of detection: 74 [32] Concentration of PFPeA in human plasma B samples (3 datasets): - Assigned value: NA - Average: 1.04 ng/ml - Median: 0.13 ng/ml - Min.: 0.08 ng/ml - Max.: 2.90 ng/ml - SD: 1.61 - % relative SD: 156 [40] Human: * Kidney: - Mean: BDL - MLOD: 0.006 ng/g wet weight - % of detection: 5 [47] PFPeA in human blood samples: - Detection frequency: 100% - Mean +/- SD: 3.2 +/- 1.8 ng/mL - 5th percentile: 0.97 ng/mL - Median: 3.0 ng/mL - 95th percentile: 6.2 ng/mL => negatively (-0.330) associated with BMI (p<0.01) => positively (0.235) correlated with progressive motility (p<0.05) [117] Surface water and groundwater in a large area of the Veneto Region in northeastern Italy was contaminated with PFAS that were also found in drinking water samples. The health surveillance program is a free-of-charge population-based screening program. Residents who decided to participate in the program completed a structured interview administered by a trained public health nurse, followed by blood pressure measurement, and blood and urine sampling. A total of 18,345 residents were included in the present analysis. Only 3 of the 12 PFAS were quantifiable in at least 80% of serum samples: PFOA (99.9%), PFOS (99.8%) and PFHxS (98.1%) [117] Serum concentrations (ng/mL) of PFPeA and percentage of samples above the LOQ in the study population (18,345 subjects, 14-39 years of age): PFPeA: - Min: <0.5 - 5th percentile: <0.5 - 25th percentile: <0.5 - Median: <0.5 - 75th percentile: <0.5 - 95th percentile: <0.5 - Max: 1.3 % samples LOQ: 0.1% [40] - Smokers showed smaller accumulation of PFASs - Older people (more than 60 years) showed higher concentrations of PFASs => Clear indication that these compounds accumulate after a long-term exposure [71] Shorter chain PFASs have been shown to absorb into the liver more readily than those with longer chains which are concentrated in the blood proteins 220-301-2 2706-91-4 Perfluoropentane-1- PFPeS C5 sulphonic acid Pre-registered [77] Remediation of PFASs: * Stabilization: managing PFASs source zone soils in situ provides a sustainable approach that does not create a concentrated waste requiring offsite management or destruction * High energy electron beam: further evaluation necessary * Low/high temperature thermal desorption: a high degree of licensing and stakeholder engagement will likely be required due to the operation of a thermal waste recovery facility and the associated off-gasses * Vapor energy generator process: no full-scale application of VEG for PFASs has been implemented * Ball milling: viability of ball milling has not been studied in detail with respect to the complications of PFASs (short-chain compounds and polyfluorinated precursors) [17] shorter-chained PFAS (C<=5): electrostatic interactions have been hypothesized to be of more importance than hydrophobic interactions for sorption to sludge => partition to sludge increases with decreasing chain length [111] * Commercially produced eggs (n=12, sites 5-16): Egg yolks: - Min: <0.05 ng/g - Max: 0.05 ng/g - Mean: 0.05 ng/g - Median: <0.05 ng/g - n > LOD (%): 1 (8) Egg whites: - Min: <0.05 ng/g - Max: <0.05 ng/g - Mean: <0.05 ng/g - Median: <0.05 ng/g - n > LOD (%): 0 (0) Whole eggs: - Min: <0.05 ng/g - Max: <0.05 ng/g - Mean: <0.05 ng/g - Median: <0.05 ng/g - n > LOD (%): 0 (0) => Unlike in the environmental media, egg yolks contained much lower proportions of PFPeA, PFHxA and PFHpA than of PFBA [100] 54 pooled human serum samples from 4920 individual samples (AU): * PFPeS concentrations in 2011, by age group: - 0-4 y (n=4): FOD: 75%, Min: <LOD, Max: 0.03 ng/mL, Mean: 0.02 ng/mL - 5-15 y (n=4): FOD: 0%, Min: <LOD, Max: /, Mean: / - 16-30 y (n=4): FOD: 25%, Min: <LOD, Max: 0.02 ng/mL, Mean: 0.01 ng/mL - 31-45 y (n=4): FOD: 0%, Min: <LOD, Max: /, Mean: / - 46-60 y (n=4): FOD: 25%, Min: <LOD, Max: 0.02 / [19] 82765-76- Perfluoropentanesulpho FPeSA C5 / 2 namide 206-795-2 375-81-5 Perfluoropentane-1- PPeSF C5 sulphonyl fluoride Pre-registered 671-204-0 151772-58-6 2,2-difluoro-2-[1,1,2,2- / C5 tetrafluoro-2- (trifluoromethoxy)ethoxy] acetic acid Not registered 206-196-6 307-24-4 Undecafluorohexanoic PFHxA C6 acid Pre-registered [79] log Kaw: 1.43 (calculated from equations) [1] (neutral form) = 2.82 - 4.6 [29] PFHxA: logKow ranges from 0.70 to 4.37 (exper. + modelled) [41] PFHxA ChemBioOffice: logKow = 4.90 [65] * PFHxA: - logCMC Exp: -1.05 - logCMC Pred: -1.15 - 3: 8.05 * PFHxA isomer: - logCMC Exp: -1.10 - logCMC Pred: -0.86 - 3: 7.30 [1] > 20 g/l [65] PFHxA: logpL (subcooled vapor pressure): - Y-Pred: -0.04 mm Hg - EPI Suite Pred: 0.30 mm Hg - Y-Pred: 2.08 Pa - Pred: 2.06 Pa [1] 6-6.7 [9] PFCAs and PFSAs have low vapour pressures which decrease with increasing carbon chain length => Low potential for volatilization [1] < 1 [65] PFHxA: logAqS (aqueous solubility): - Y-Pred: 1.47 mg/L - EPI Suite Pred: 0.67 mg/L [79] Vapor pressure (PL): log (PL/Pa): 1.96 (calculated from equations) [79] log Koa: 4.84 (calculated from equations) [9] Calculated pKa values indicate that both PFCAs and PFASs are strong acids which will predominantly be in their dissociated negatively-charged form at environmentally relevant pH values [1] 2.7-3.6 [79] - Molar mass (MM): 314.1 g/mol - Molar volume (VM): 182.4 cm^3/mol (by ACD/Labs' ACD/PhysChem Suite) - Total surface area (TSA): 287.5 A^2 (by ChemAxon) - Tm (experimental melting point): 287 K (data obtained from ChemSpider) [22] - PFASs are known to sorb to particulate matter => sorption increases as carbon chain length increases - significant relationships with TOC (p<0.05) => unsurprising given hydrophobicity of perfluorinated chain and potential for hydrophobic partitioning with organic matter [1] drinking water resources are highly sensitive to contamination [23] 2 studies on hydrolysis with outcomes: - 0% - (almost) 0% [23] Resistant to photolysis, hydroxyl (OH)-radical-mediated reactions, hydrolysis, and biodegradation => known PFECAs and PFESAs are likely to be highly persistent in the environment and not easily metabolized in biota [23] only serum elimination half-lives of two PFECAs in mammals have been reported => they might be as bioaccumulative as the predecessors due to similar physicochemical properties, but this is uncertain [23] three PFECAs have been shown to cause liver damage in rats in repeated- dose tests; one PFECA is suggested by its manufacturer to be classified as T under the REACH regulation => other known PFECAs may have the same mode-of-action and may fulfill the toxicity criteria under REACH [23] Known PFECAs and PFESAs: - P - maybe B - likely T - LRTP [21] PFCAs are not degradable and will only be slowly removed from the oceans by vertical mixing into the deeper oceans => Arctic Ocean will have a very slow response time to changing emissions due to its long water residence time [1] high energy carbon-fluorine bond => extremely persistent [29] PFHxA: BCF <1 L/kg BMF <1 <=> other study: BAF 5.0-120 L/kg [3] protein binding affinity appears to increase with chainlength => short chain: faster elimination and lower distribution to liver [23] PFHxA is reported to be more acutely toxic (3-5 times) to three aquatic species than PFOA => variation in species sensitivity has not yet been studied [37] * Frog embryo teratogenicity assay-Xenopus (FETAX): LC50 (concentration at which 50% of the embryos were killed) = 1523.5 M, EC50 (concentration at which 50% of the embryos were malformed) = 1046.5 M, MCIG (minimum concentration that inhibits growth) = 412.3 M, TI (teratogenic index) = 1.43 => TI was >= 1.2 => compound considered toxic * Exposure to PFHxA or PFHpA greatly increased the mortality and general malformation rates in the developing embryos * PFC exposure induced certain embryonic abnormalities: pericardial edema, dorsal fin blisters, a curved body axis, lack of eye pigment, disturbed facial cells => PFHxA-treated embryos developed mild and general malformations including gut mis-coiling and improper eyeshape [23] Short-chain PFAAs: - P - not B - maybe T - LRTP [23] on-going (increasing) emissions and high persistence => short-chain PFAAs, PFECAs and PFESAs accumulate in the environment => increasing exposure to these substances over time => even if emissions cease in the future it will take decades to centuries to reverse global contamination of short-chain PFAAs, PFECAs and PFESAs, due to high persistence and environmental mobility [1] no degradation => permanent and irreversible exposure of organisms [25] for PFAAs elimination occurs through different mechanisms in aquatic and terrestrial species => in mammalian species: governed by renal and biliary clearance => in fish: PFAAs can also be eliminated by the respiratory system (gills) => elimination in fish is much more efficient than in (some) mammalian species [11] humans: modelled concentrations in serum: 0.014 ng/g [47] PFHxA in human semen samples: - Detection frequency: 100% - Mean +/- SD: 4.3 +/- 3.4 ng/mL - 5th percentile: 1.6 ng/mL - Median: 3.4 ng/mL - 95th percentile: 11 ng/mL => Lower concentration of PFHxA in semen was caused by the bloodtestis barrier, which can protect germ cells from toxic chemicals => positively (0.372) associated with BMI (p<0.01) => negatively (-0.349) correlated with progressive motility (p<0.01) [40] Human: * Liver: - Mean: 115 ng/g wet weight - Median: 68.3 ng/g wet weight - Range: 353-BDL ng/g wet weight - MLOD: 2.73 ng/g wet weight - % of detection: 70 => PFHxA was PFAS-compound with highest median concentration in liver [22] - increase in the mobility of PFASs with increasing pH => result of repulsive electrostatic interactions with the anionic functional group head and sediment - significant relationships with pH (p<0.05) [29] PFHxA: logKoc: 3.1 (EPIsuite) [25] if PFASs sorb to solids in the soil or aquifer material, the transport of PFASs will be retarded and their residence time will be even longer than the residence time of groundwater => PFAS contaminated groundwater can act as a long-lasting source, leading to poorly reversible exposure to PFASs in the subterranean environment as well as where the contaminated groundwater resurfaces [10] - volatile PFASs (FTOHs, FASAs, FASEs) enter the atmosphere, where they can degrade, form intermediates during atmospheric oxidation, or transform into more persistent PFASs, such as PFSAs and PFCAs, which may finally end up in the aquatic environment - PFASs have the potential to affect cell membranes of algae => can alter nutritious quality of biofilms => potential effects on physiological fitness of invertebrates [21] - the atmosphere could have faster responses to changing emissions (precursor concentrations are relatively quickly removed by oxidation processes) - relatively fast response times to changing emissions can also be expected for PFCAs in water bodies with short residence times (rivers and coastal waters) [23] short-chain PFAAs are similarly persistent as their long-chain homologues [14] PFHxA has shorter serum half-lives in humans and biota than its longerchain homologues [25] long-chain PFAAs are clearly toxic, but it is unclear if shortchain PFAAs would fulfil the T criterion due to uncertainties in their toxicity [37] * Whole body length was reduced by 14% in tadpoles treated with 1500 M of PFHxA (p<0.05) * Enlarged livers in tadpoles exposed to PFHxA * Hearts from embryos treated with PFHxA and PFHpA showed enlarged atria and a loss of the atrial septum * Hearts from embryos exposed to either PFHxA or PFHpA were smaller and contained no trabeculae + pericardiac cavity was expanded and atrial and ventricular walls were thinner than those observed in the control => PFHxA may cause developmental toxicity and teratogenicity [25] short-chain PFAAs fulfill the P and M criteria (M will decrease with perfluoroalkyl chain length) and are highly bioavailable to humans [2] human exposure primarily via dietary intake => contribution of drinking water [25] Modelling of 2 scenario's: - Scenario 1: emissions to groundwater are stopped after 5 years, but no additional measures are taken to reduce groundwater concentrations => PFHxA will decrease slowly in human serum => residence time of groundwater governs reversibility of human exposure - Scenario 2: PFAS concentrations in groundwater are effectively reduced by proper measures after emissions ceased => humans exposed to contaminated drinking water would be able to eliminate shorter-chain PFHxA within months [40] Human: * Bone: - Mean: 35.6 ng/g wet weight - Median: 1.5 ng/g wet weight - Range: 230-BDL ng/g wet weight - MLOD: 0.001 ng/g wet weight - % of detection: 30 [79] Water solubility (SL): log (SL/mol L^-1): -2.22 (calculated from equations) [74] PFHxA: * R (retardation factor) (40 cm): 4.7 * R (80 cm): 15 * log Kd (40 cm): -0.18 * log Kd (80 cm): 0.46 * log Koc (40 cm): 3.0 * log Koc (80 cm): 3.6 [29] Limited accumulation has been measured in freshwater and coastal sediments PFHxA: <1 ng/g dw [17] - the fate of PFASs sorbed to sludge is generally to be further spread in the environment through application of sludge or sludge-containing products - in 2014: 24% of sludge was applied on agricultural soil, 24% was used to cover mines and dump sites, and 29% was used to manufacture soil intended for less sensitive land use [25] for PFASs an alternative measure of persistence (joint persistence) should be taken into account => because PFASs are, or ultimately degrade to, stable end products, all PFASs possess a joint persistence that exceeds all three of the REACH singlemedia persistence criteria [23] short-chain PFAAs are less bioaccumulative in animals and humans than the long-chain homologues, but may show higher uptake into the leaves, stems and fruits of plants [39] Relative response of mouse and human PPAR to PFAAs in transiently transfected COS-1 cells, measured by C20max: PFHxA: - Mouse: 38 M (Most active to least active: ranked 8/13 PFASs) - Human: 47 M (Most active to least active: ranked 5/13 PFASs) => A lower C20 max value indicates a stronger induction of PPAR alpha activity by that PFAA => PPAR is activated by PFAAs of 4-12 carbons in chain length => Increasing activity of PPAR with increasing chain length of the PFAA up to C9, and lower activity with longer chain PFAAs (C>9) with both mouse and human PPAR => PPAR has been identified as a key player in the mode of action for PFAA-induced toxicity [42] FRESHWATER: Brachionus calyciflorus - rate of population increase was 32.0% lower in PFHxA than in the control (p<0.05) => decline of population growth rate poses a chronic and longterm risk on rotifer population - mictic ratios increased by 1.36-fold in PFHxA compared to the control (p<0.05) => increase of mictic ratio was due to the change of rotifers reproduction => under exposure of environmental stress rotifers transform from asexual reproduction to sexual reproduction => fertilized eggs can resist poor outside environment and keep balance in population dynamics => PFHxA may interfere with endocrine signaling in B. calyciflorus , causing a reduction of population growth rate and an increase of mictic ratio [11] estimated total daily exposure humans: - low exposure scenario: 15 pg/kg/d - intermediate exposure scenario: 63 pg/kg/d - high exposure scenario: 520 pg/kg/d => direct intake is dominant in all scenarios (66%-96%) => direct exposure via food is estimated to be the major exposure pathway in the low (41-88%), intermediate (3886%) and high (42%) exposure scenario [30] levels of PFHxA in serum from participating mothers (SE): 1996-1999 (n=147): - MDL: 0.3 ng/g - % of levels <MDL: 100 2008-2011 (n=134): - MDL: 0.3 ng/g - % of levels <MDL: 100 [40] Human: * Brain: - Mean: 180 ng/g wet weight - Median: 141 ng/g wet weight - Range: 486-10.1 ng/g wet weight - MLOD: 0.72 ng/g wet weight - % of detection: 100 => PFHxA was PFAS-compound with highest median concentration in brain [39] NOEC and LOECs for PFAAs for transactivation of mouse and human PPAR in transiently transfected COS-1 Cells: PFPeA: Mouse: - NOEC: 1 M - LOEC: 5 M - LOEC: 1.32 g/mL (p<0.0001) [29] QS dw,hh (human health via consumption of water): 3 g/L [39] NOEC and LOECs for PFAAs for transactivation of mouse and human PPAR in transiently transfected COS-1 Cells: PFPeA: Human: - NOEC: 0.5 M - LOEC: 1 M - LOEC: 0.26 g/mL (p<0.05) [118] Spatial and seasonal distributions of PFAS in air, water, sediment, soil, and fish samples in the Asan Lake area of South Korea. Estimation of bioaccumulation potential of PFAS using sediment-water partition coefficients and bioaccumulation factors. PFPeA concentration in air samples: - Air (unit: pg/m3, n=4): Gaseous: * Range: nd ~ 2.69 * Mean: 1.58 * Median: 1.95 * DF (%): 75.0 [118] Particulate: * Range: nd * Mean: * Median: * DF (%): 0.0 Total: * Range: nd ~ 2.69 * Mean: 1.58 * Median: 1.81 * G (%): 100.0 [89] Presence of PFASs in beer samples (PFPeA): * All samples (n=93): - n > LOQ: 18 - n > LOQ (%): 19 - median: <LOQ - mean: 2.64 ng/L - maximum: 37.4 ng/L => PFPeA was detected in concentrations greater than the LOQ in 19% of all samples * German samples (Hesse) (n=83): - n > LOQ: 18 - n > LOQ (%): 22 - median: <LOQ - mean: 2.95 ng/L - maximum: 37.4 ng/L => PFPeA was detected in concentrations greater than the LOQ in 22% of German (Hesse) samples [89] Presence of PFASs in beer samples (PFPeA): * German samples (Bavaria) (n=4): - n > LOQ: 0 - n > LOQ (%): 0 - median: <LOQ - mean: <LOQ - maximum: <LOQ * Belgian samples (n=6): - n > LOQ: 0 - n > LOQ (%): 0 - median: <LOQ - mean: <LOQ - maximum: <LOQ => PFPeA was not detected in concentrations above the LOQ in German (Bavaria) and [89] * PFPeA contributes 14% to the total PFAS-concentration in all of the beer samples (n=93) * PFPeA contributes 16% to the total PFAS-concentration in the beer samples from Hesse (n=83) * sources other than just the water may be responsible for the PFAS concentrations detected in beer => PFASs may find their way into the raw materials such as hops, barley or wheat and then finally into beer, all via the exposure path "soil-plant" [12] SE Groundwater: PFCAs ranging from 20% for skiing to 43% for industrial areas (of total PFASs) [30] observation wells SE: median water conc. PFPeA: 27 ng/L [12] SE Surface water: PFCAs ranging from 54% for firefighting sites to 85% for skiing and urban areas (of total PFASs) [18] Urban run-off water samples subjected to TOP oxidation: - before oxidation: sums of PFCA and PFSA in the range 6-42 ng/L and 3-35 ng/L - after oxidation: PFCAs were seen to increase by on average 69% (=14 ng/L) relative to initial PFCA content => approx. rise of ~35% compared to the sum of PFCA and PFSA => largest additions measured for PFHxA and PFPeA => 6:2 telomer based precursors made a significant contribution [32] Concentration of PFPeA in canal water samples (9 datasets): - Assigned value: 3.76 ng/L - Average: 5.06 ng/L - Median: 4.50 ng/L - Min.: 2.62 ng/L - Max.: 10.40 ng/L - SD: 5.06 - % relative SD: 51 [16] GR near sea shore - beached plastic pallets: predominant compound: present in 4 out of 5 samples, with concentrations ranging between 24 and 98 ng/kg - sediment: present in 5 out of 8 samples [12] SE Drinking water: PFCAs 45% of total PFASs [6] tap water MDL = 0.17 ng/L MLQ = 0.53 ng/L max. value NL: 2.69 ng/L [10] total discharge of PFASs into the aquatic environment ranged between 10 g d-1 and 10 000 g d-1, depending on the water usage in the community connected to the sewage treatment plant [8] concentrations in water CN: - Dissolved phase: mean: 29.2 ng/L contribution to total PFASs: 2.41% - Separate suspended particulate matter: mean: 5.47 ng/L contribution to total PFASs: 12.49% - Soil: mean: 1.40 ng/L contribution to total PFASs: 19.53% - Leaves: mean: 9.87 ng/L contribution to total PFASs: 11.58% - Bark: mean: 6.19 ng/L contribution to total PFASs: 16.33% [118] Spatial and seasonal distributions of PFAS in air, water, sediment, soil, and fish samples in the Asan Lake area of South Korea. Estimation of bioaccumulation potential of PFAS using sediment-water partition coefficients and bioaccumulation factors. [73] - The dominant compound in the Bay of Biscay was PFPeA (170 pg/L, accounted for 29% of PFASs) - It was the third abundant compound in the English Channel (120 pg/L, 19% of PFASs) [66] PFPeA levels (ng/L) in outgoing drinking water from two waterworks in Ronneby (SE): - Site 1 (Brantafors): 38 - Site 2 (Krragarden): 10 [89] * mineral water (Germany): PFPeA: <1.00-7.80 ng/L * tap water (France): PFPeA: <4.00-27.0 ng/L [17] SE - Other PFCAs detected in filtered effluent water were PFPeA, ... - Major contributions to the daily discharge Dd connected to the WWTP were from PFPeA, ... - PFPeA predominated in influent water from a domestic-dominated WWTP [13] - river water NL: detectable concentrations: min. 4.1 ng/L - max. 9.2 ng/L - drinking water NL: detectable concentrations: 5.7 ng/L [118] PFPeA concentration in water and sediment samples: - Water (unit: ng/L, n=47): * Range: 2.1 ~ 140 * Mean: 42.5 * Median: 21.5 * 75%: 72.5 * 95%: 128 * DF (%): 100.0 [73] * North Atlantic Ocean (NAO): - 2007 (n=39): ND - 2008 (n=10): 30-74 pg/L - 2010 (n=13): 16-77 pg/L * Middle Atlantic Ocean (MAO): - 2007 (n=10): ND - 2008 (n=5): 21-35 pg/L - 2010 (n=5): <13-32 pg/L * South Atlantic Ocean (SAO): - 2007 (n=10): ND - 2008 (n=13): <14-24 pg/L - 2010 (n=16): <13 pg/L [117] In spring 2013, groundwater of a vast area of the Veneto Region (northeastern Italy) was found to be contaminated by perfluoralkyl substances (PFAS) from a PFAS manufacturing plant active since the late 1960s. Residents were exposed to high concentrations of PFAS, particularly perfluorooctanoic acid (PFOA), through drinking water until autumn 2013. [102] Tap water NL (n=4): Average SD (Range): 2.4 0.50 (1.7-2.8) ng/L => Levels of PFAAs observed in four tap water samples closely resembled each other as demonstrated by the relatively low standard deviation of the mean values [119] 3 replicate aqueous and solid samples were collected from each of 19 Australian WWTPs throughout 2017. Samples were taken from various stages within the treatment train from a range of WWTPs to determine the trends in the mass flux and partitioning of PFAS within the sampled WWTPs. Summary statistics for pooled aqueous (n=201, triplicates from 67 individual locations within 19 WWTPs) and pooled solid (n=51, triplicates from 5 primary and secondary sludge locations, 6 lagoon sludges and a lagoon dredge pile) samples. [24] river water CN, maximum concentrations in * yearly monitoring: - 2011: 82.2 ng/L - 2012: 81.2 ng/L - 2013: 198 ng/L - 2014: 56.1 ng/L * seasonal monitoring: summer (273 ng/L) > autumn (198 ng/L) > winter (139 ng/L) > spring (79.6 ng/L) => peak river contamination periods: summer & autumn [102] Post-mixed cola NL (n=6): Average SD (Range): 2.0 0.55 (0.52-2.2) ng/L => Ratio tap water/post-mixed cola (Ratio of averages): 1.1 => PFAS concentrations in tap water were higher than in cola => A combination of the dilution step and the purification step inside the cola dispenser are responsible for the observed reduction * Coffee NL: - Brewed coffee from coffee machines (n=12): Average SD (Range): Could not be quantified [113] 397 food samples collected from a market in Busan (Korea) in 2011 (n=227) and 2012 (n=170). Food samples of 66 different food types, classified into 7 food categories. 34 tap water samples were also collected from 16 districts in Busan. PFPeA concentrations in each food group (ng/g, beverage unit: ng/L): * Fish and shellfish (n=99): DF (%) = 15.2, Mean = 0.094, Max = 2.58 * Meat and its products (n=39): DF (%) = 12.8, Mean = 0.131, Max = 3.50 * Vegetables and fruit (n=78): DF (%) = 6.4, Mean = 0.019, Max = 0.769 [113] * Processed products (n=90): DF (%) = 3.3, Mean = 0.064, Max = 2.72 * Dairy (n=37): DF (%) = 10.8, Mean = 0.034, Max = 0.493 * Beverage (n=21): DF (%) = 14.3, Mean = 1.28, Max = 20.0 * Others (n=33): DF (%) = 6.1, Mean = 0.020, Max = 0.420 PFPeA concentrations in bottled water and tap water samples (ng/L): * Bottled water (n=8): DF (%) = 12.5, Mean = 0.084, Max = 0.669 [118] - Sediments (unit: ng/g dw, n=47): * Range: nd ~ 0.42 * Mean: 0.08 * Median: 0.00 * 75%: 0.16 * 95%: 0.37 * DF (%): 40.4 [118] The predominant PFAS in water was PFPeA, followed by PFHxA. Shortchain PFAS such as PFPeA, PFHxA and LPFBS were detected more frequently in water than in the other environmental matrices. Short-chain PFAS and PFOA were detected during all four seasons. The seasonally dominant PFCAs were PFHxA and PFOA during summer and PFPeA and PFHxA during the other seasons. [117] Surface water and groundwater in a large area of the Veneto Region in northeastern Italy was contaminated with PFAS that were also found in drinking water samples. A manufacturing plant located in the town of Trissino that produced PFAS since the late 1960s was identified as the only likely source of water contamination. [113] * Tap water (n=34): DF (%) = 100.0, Mean = 1.81, Max = 2.87 [119] * PFPeA concentration in aqueous samples (ng/L): - Median: 5.3 - Mean: 8.3 - s.d.: 8.8 - min: <LOD - max: 47 - Detect (%): 96% * PFPeA concentration in solid samples (ng/g dw): - Median: <LOD - Mean: <LOQ - s.d.: / - min: <LOD - max: 5.2 [117] Measurements of 152 drinking water samples collected in July and August 2013 indicated that the main contaminants were PFOA, PFBA and PFBS, followed by PFPeA, PFHxA, PFOS, PFHpA and PFHxS. The longer-chain PFAS congeners (PFNA, PFDA, PFUnA, PFDoA) were detected only in a minority of samples and at lower concentrations. [119] In aqueous samples: between influent and final effluent, the compounds PFPeA, PFHxA, PFHpA, PFOA, PFNA, and PFDA (all of which are PFCAs) increased significantly. A number of transformation pathways with stable PFCA endproducts are known, this may explain some of the increase in PFCAs from influent to final effluent. [117] PFPeA concentrations (ng/L) in 152 samples of drinking water taken during July and August 2013, before the full implementation of granular activated carbon filters: - Min: <10.0 - 25th percentile: 45.0 - Median: 70.0 - 75th percentile: 100.0 - 95th percentile: 210.6 - Max: 370.0 - % samples LOQ: 90.1% [113] The short-chain PFCAs such as PFBA and PFPeA were detected more frequently in food samples with high moisture contents, such as beverages, vegetables, and fruit, than in food samples with low moisture contents. PFPeA was detected in all of the tap water samples. This indicates that Korean tap water is contaminated with PFAAs, and this is likely to be caused by PFAAs contamination in the original sources of the drinking water. 90% of the drinking water in Korea comes from flowing surface water. [115] In September 2011, foods were purchased in 12 representative cities of Catalonia, all with more than 20,000 inhabitants. Food samples were obtained at each locality in 4 shops/stores of different size. Foods selected for PFAS analysis were among the most consumed in Catalonia. * Mean concentrations (in pg/g fw) of PFPeA in the groups of analyzed foodstuffs: - Meat and meat products: <12 - Fish and seafood: <14 - Vegetables: <19 - Tubers: <15 - Fruits: <15 - Eggs: <16 - Milk: <20 - Dairy products: <9.0 [115] - Cereals: <5.1 - Pulses: <11 - Oils: <5.9 - Industrial bakery: <4.4 * Data corresponding to the 80 individually analyzed composite food samples show that only PFPeA, PFHxDA and PFOcDA could not be detected in any sample. * Concerning human health risks from dietary exposure to PFASs in Catalonia, it is important to note that for any of the age/gender groups of population estimated, the TDIs recommended by the EFSA were not exceeded. [17] SE Other PFSAs detected in filtered effluent water were PFPeS (mean 0.4 ng/L) [119] 3 replicate aqueous and solid samples were collected from each of 19 Australian WWTPs throughout 2017. Samples were taken from various stages within the treatment train from a range of WWTPs to determine the trends in the mass flux and partitioning of PFAS within the sampled WWTPs. Summary statistics for pooled aqueous (n=201, triplicates from 67 individual locations within 19 WWTPs) and pooled solid (n=51, triplicates from 5 primary and secondary sludge locations, 6 lagoon sludges and a lagoon dredge pile) [119] * PFPeS concentration in aqueous samples (ng/L): - Median: <LOQ - Mean: 1.9 - s.d.: 4.1 - min: <LOD - max: 27 - Detect (%): 77% [119] * PFPeS concentration in solid samples (ng/g dw): - Median: <LOD - Mean: <LOQ - s.d.: / - min: <LOD - max: 2.3 - Detect (%): 14% [4] Inhalation exposure rats to 6:2 FTOH: 0.5 ppm PFHxA => Male: 1.3 hrs => Female: 0.5 hrs Inhalation exposure rats tp 6:2 FTOH: 5.0 ppm PFHxA => Male: 1.3 hrs => Female: 0.5 hrs [11] humans: Vd = 200 mL/kg [11] humans: T1/2 = 0.088 y [14] - rats: IV, 10 ppm: female: 0.4 h, male: 1 h - rats: IV, 15 ppm: female: 1.2 h, male: 2.4 h - mice: gastric, 50 ppm: female: <72 h, male: <72 h - monkeys: IV, 10 ppm: female: 2.4 h, male: 5.3 h - humans: male (n = 8): <28 d [4] Inhalation exposure rats to 6:2 FTOH: 0.5 ppm PFHxA => Male: 0.533 h-1 => Female: 1.386 h-1 Inhalation exposure rats to 6:2 FTOH: 5.0 ppm PFHxA => Male: 0.533 h-1 => Female: 1.386 h-1 [4] Inhalation exposure rats to 6:2 FTOH: 0.5 ppm PFHxA => Male: 0.107 lkg-1h-1 => Female: 0.277 lkg-1h-1 Inhalation exposure rats to 6:2 FTOH: 5.0 ppm PFHxA => Male: 0.107 lkg-1h-1 => Female: 0.277 lkg-1h-1 Scaled: average estimated rat clearance: 0.5 ppm 6:2 FTOH PFHxA => 0.192 lkg-1h-1 Scaled: average estimated rat clearance: 5 ppm 6:2 FTOH PFHxA=> 0.192 lkg-1h-1 Scaled: predicted human clearance: 0.5 ppm 6:2 FTOH PFHxA => 4.139 lkg-1h-1 Scaled: predicted human clearance: 5 ppm 6:2 FTOH PFHxA => 4.139 lkg-1h-1 [3] humans: liver AUCSS: 1.9-2.8 liver weight ratio: 1.2-1.25 [3] humans: model liver: LOEL = 200 mg/kg/day liver weight ratio: 1.2-1.3 [29] QS biota, secpois: Assessment factor: 90 on NOEC: 400 mg/kg => QS biota = 4444 g/kg biota ww [4] AUC0- Inhalation exposure rats to 6:2 FTOH: 0.5 ppm PFHxA => Male: 1108.5 nanomoles/l.hr => Female: 490.9 nanomoles/l.hr AUC0- Inhalation exposure rats to 6:2 FTOH: 5.0 ppm PFHxA => Male: 4166.2 nanomoles/l.hr => Female: 990.8 nanomoles/l.hr Occupational exposure to 6:2 FTOH humans: PFHxA => 335.3 ng/ml.d [29] NOAEL (rat): 20 mg/kg bw/d bw/DFI: 20 kg bw d/kg NOEC: 400 mg/kg [29] QS dw,hh (human health via consumption of water): 1 g/L [39] NOEC and LOECs for PFAAs for transactivation of mouse and human PPAR in transiently transfected COS-1 Cells: PFHxA: Mouse: - NOEC: 10 M - LOEC: 20 M - LOEC: 6.28 g/mL (p<0.0001) [118] Spatial and seasonal distributions of PFAS in air, water, sediment, soil, and fish samples in the Asan Lake area of South Korea. Estimation of bioaccumulation potential of PFAS using sediment-water partition coefficients and bioaccumulation factors. PFHxA concentration in air samples: - Air (unit: pg/m3, n=4): Gaseous: * Range: nd ~ 8.30 * Mean: 3.06 * Median: 2.15 * DF (%): 75.0 [89] Presence of PFASs in beer samples (PFHxA): * All samples (n=93): - n > LOQ: 19 - n > LOQ (%): 20 - median: <LOQ - mean: 4.80 ng/L - maximum: 155 ng/L => PFHxA was detected in concentrations greater than the LOQ in 20% of all samples * German samples (Hesse) (n=83): - n > LOQ: 19 - n > LOQ (%): 23 - median: <LOQ - mean: 5.38 ng/L - maximum: 155 ng/L => PFHxA was detected in concentrations greater than the LOQ in 23% of German [2] observation well NL max. value: 570 (506670) ng/L [12] SE Surface water: PFCAs ranging from 54% for firefighting sites to 85% for skiing and urban areas (of total PFASs) [28] Rivers SE: - Detection frequency: 27% - Concentration range: 0.51-4.2 ng/L - Average: 2.4 ng/L [118] Particulate: * Range: 0.48 ~ 0.61 * Mean: 0.52 * Median: 0.49 * DF (%): 100.0 Total: * Range: 0.61 ~ 8.79 * Mean: 3.58 * Median: 2.45 * G (%): 85.5 [89] Presence of PFASs in beer samples (PFHxA): * German samples (Bavaria) (n=4): - n > LOQ: 0 - n > LOQ (%): 0 - median: <LOQ - mean: <LOQ - maximum: <LOQ * Belgian samples (n=6): - n > LOQ: 0 - n > LOQ (%): 0 - median: <LOQ - mean: <LOQ - maximum: <LOQ => PFHxA was not detected in concentrations above the LOQ in German (Bavaria) and Belgian samples [89] * PFHxA contributes 26% to the total PFAS-concentration in all of the beer samples (n=93) * PFHxA contributes 30% to the total PFAS-concentration in the beer samples from Hesse (n=83) * sources other than just the water may be responsible for the PFAS concentrations detected in beer => PFASs may find their way into the raw materials such as hops, barley or wheat and then finally into beer, all via the exposure path "soil-plant" [5] Not detected in ground water [12] SE Groundwater: PFCAs ranging from 20% for skiing to 43% for industrial areas (of total PFASs) [18] Urban run-off water samples subjected to TOP oxidation: - before oxidation: PFHxA present in 5 ng/L with sums of PFCA and PFSA in the range 6-42 ng/L and 3-35 ng/L after oxidation: PFCAs were seen to increase by on average 69% (=14 ng/L) relative to initial PFCA content => approx. rise of ~35% compared to the sum of PFCA and PFSA => largest additions measured for PFHxA and PFPeA => 6:2 telomer based precursors made a significant contribution [34] - Firefighting training sites: PFHxA: 15% of total PFASs - Landfill/waste disposal areas: PFHxA: 22% of total PFASs - Sewage Treatment Plant (STP) effluent: PFHxA: 25% of total PFASs - Unspecific industry: PFHxA: 17% of total PFASs - Urban areas: PFHxA: 19% of total PFASs - Skiing areas: PFHxA: 22% of total PFASs [32] Concentration of PFHxA in canal water samples (22 datasets): - Assigned value: 9.00 ng/L - Average: 8.80 ng/L - Median: 9.00 ng/L - Min.: 0.00 ng/L - Max.: 16.50 ng/L - SD: 8.80 - % relative SD: 36 [34] Mean concentration of 13 ng/L in rivers (SE) [16] GR near sea shore: - sediment: present in 1 out of 8 samples [2] estimated concentration of PFAA in public supply well field NL: 13 ng/L [1] ex. tap water DE 2 ng/L (23%) [21] PFHxA and PFHpA have been reported in fresh and coastal waters at levels comparable to, or much higher than PFOA [12] SE Drinking water: PFCAs 45% of total PFASs [6] tap water MDL = 0.12 ng/L MLQ = 0.38 ng/L max. value NL: 5.15 ng/L [28] Recipient sea SE: - Detection frequency: 0% - Concentration range: NA - Average: NA - Flux of PFHxA into the Baltic Sea (SE) based on discharges from all surrounding countries: modelled value: 50-260 kg/yr, measured/calculated value: 15 kg/yr [18] river water and 3 connected water wells for drinking water: - before oxidation: PFHxA, PFHpA and PFOA ranged between 10-17 ng/L - after oxidation: a rise of +270830% or (+46-127 ng/L) => of this increase 18-53% could be explained by known precursors, in particular 6:2 FTS [89] * mineral water (Germany): PFHxA: <1.00 ng/L * tap water (France): PFHxA: <4.00-125 ng/L [9] ex. treated wastewater AU - raw/influent: 4.4 ng/L - finished/tap water: 6.5 ng/L - percent removal: -48% [8] concentrations in water CN: - Dissolved phase: mean: 827 ng/L contribution to total PFASs: 68.35% - Separate suspended particulate matter: mean: 10.0 ng/L contribution to total PFASs: 22.83% - Soil: mean: 0.28 ng/L contribution to total PFASs: 3.91% - Leaves: mean: 13.7 ng/L contribution to total PFASs: 16.08% - Bark: mean: 5.83 ng/L contribution to total PFASs: 15.38% [10] total discharge of PFASs into the aquatic environment ranged between 10 g d-1 and 10 000 g d-1, depending on the water usage in the community connected to the sewage treatment plant [13] - river water NL: detectable concentrations: min. 4.0 ng/L - max. 6.4 ng/L - drinking water NL: detectable concentrations: min. 2.9 ng/L - max. 5.3 ng/L [17] SE - PFHxA was predominantly present in filtered effluent water with a mean level of 9.7 ng/L - PFHxA was predominantly present in sludge => highest concentration found was 3.3 ng/g - Major contributions to the daily discharge connected to the WWTP were from PFHxA, ... - PFHxA predominated in influent water from WWTPs connected to industries - PFHxA had a net mass increase in all WWTPs with a mean value of 83% => presumably result of degration of PFCAs precursor compounds during the WWTP process [14] discharge of PFHxA from a fluoropolymer manufacturing plant FR to the receiving river: 10 t/yr [65] * PFHxA: EPI pred logKow: 4.37 * PFHxA isomer: EPI pred logKow: 3.66 [79] log Kow: 0.59 (calculated from equations) [94] PFHxA: logKow = 4.37 / [19] 41997-13- Perfluorohexanesulphon FHxSA C6 / 1 amide 206-587-1 355-46-4 Perfluorohexane-1sulphonic acid PFHxS C6 Pre-registered [79] log Kaw: 2.15 (calculated from equations) [35] PFHxS: logKow = 5.17 [41] PFHxS ChemBioOffice: logKow = 5.25 [79] log Kow: 1.35 (calculated from equations) [118] PFAS species with high Kd or Koc values were strongly adsorbed by sediments and organic matter, and compounds with very low Kd or Koc values were highly mobile with respect to the sediments [118] (i.e. weak adsorption). [118] As a result, PFHxA, PFHpA, and LPFBS (with lower Kd and Koc values) could be readily transported from sediments to water in comparison with other PFAS compounds with relatively high log Kd [20] discharge of municipal wastewaters is one of the principal routes of entry of PFASs in the aquatic environment => often PFAS concentrations increase in WWTPs as a result of biodegradation of precursors during the activated sludge process [22] - at the end of their lifetimes, products containing PFASs may be disposed of to landfill - as a result of decomposition processes within landfills, PFASs associated with waste can become mobile and leach into water that enters a landfill to produce potentially contaminated leachate - small and very small sites may pose a local point source of PFAS contamination into surrounding groundwater (new and large landfills typically use cell liners to prevent leachate migration into groundwater) [77] - the extreme persistence of all PFAAs with the increasingly higher aqueous solubility (and thus PFAA mobility) as the chain length shortens, also creates increasing difficulty for them to be treated by many conventional remediation and water treatment technologies - Remediation of PFASs: * Biological treatment: potential for anaerobic degradation of PFASs seems limited, as no organisms have been isolated which can utilize PFAAs as terminal electron acceptors * GAC (granular activated carbon) can be inefficient at removing PFOA and becomes progressively less effective for removing shorter chain PFCAs such as PFHxA, PFPeA, PFBS and PFBA as the chain length diminishes [77] - Remediation of PFASs: * in the same manner that GAC is less effective at removing shorter chain PFAAs and PFAA precursors from water, the injected particulate-activated carbon is also expected to be less effective at removing these PFASs * Ion-exchange resins: long-chain PFAA and precursors are likely to be removed from water significantly more strongly than short-chain PFASs * Osorb is similarly effective as GAC at removing both PFOS and PFOA and more effective than GAC at removing PFBA * Precipitation removal rates for shorter chain PFAAs improve when higher doses of the coagulant are applied [77] Remediation of PFASs: * Ozofractionation: can convert polyfluorinated precursors to PFAAs via chemical oxidation which assist their removal via foam fractionation and can remove short-chain PFASs in addition to long chain * In situ foam fractionation: limited application * RO and NF have been shown to be extremely effective in removing PFASs regardless of chain length and are also expected to be effective at removing many types of PFAA precursors * Concerns regarding chemical oxidation for in situ application include the generation of sign. concentrations of more mobile short-chain PFASs * Chemical reduction: application at field-scale for in situ PFAA remediation is challenging from a practicality perspective * Sonolysis has been demonstrated at laboratory scale for PFASs but not scaled up for commercial [77] Remediation of PFASs: * Stabilization: managing PFASs source zone soils in situ provides a sustainable approach that does not create a concentrated waste requiring offsite management or destruction * High energy electron beam: further evaluation necessary * Low/high temperature thermal desorption: a high degree of licensing and stakeholder engagement will likely be required due to the operation of a thermal waste recovery facility and the associated off-gasses * Vapor energy generator process: no full-scale application of VEG for PFASs has been implemented * Ball milling: viability of ball milling has not been studied in detail with respect to the complications of PFASs (short-chain compounds and polyfluorinated precursors) [79] Water solubility (SL): log (SL/mol L^-1): -3.23 (calculated from equations) [79] Vapor pressure (PL): log (PL/Pa): 1.68 (calculated from equations) [79] log Koa: 5.29 (calculated from equations) [9] PFCAs and PFSAs have low vapour pressures which decrease with increasing carbon chain length => Low potential for volatilization [9] Calculated pKa values indicate that both PFCAs and PFASs are strong acids which will predominantly be in their dissociated negatively-charged form at environmentally relevant pH values [74] PFHxS: * R (retardation factor) (40 cm): 1.1 * R (80 cm): 3.0 * log Kd (40 cm): -2.0 * log Kd (80 cm): -0.39 * log Koc (40 cm): 1.2 * log Koc (80 cm): 2.8 [79] - Molar mass (MM): 400.1 g/mol - Molar volume (VM): 217.2 cm^3/mol (by ACD/Labs' ACD/PhysChem Suite) - Total surface area (TSA): 367.2 A^2 (by ChemAxon) - Tm (experimental melting point): 314 K (data obtained from ChemSpider) [22] - PFASs are known to sorb to particulate matter => sorption increases as carbon chain length increases => perfluorosulfonates have increased sorption potential than their carboxylate counterparts - significant relationships with TOC (p<0.05) => unsurprising given hydrophobicity of perfluorinated chain and potential for hydrophobic partitioning with organic matter [9] Removing PFASs from drinking water: * Granular Activated Carbon Absorption * Ion exchange/non-ion exchange resins may be useful for removing PFAS * High pressure membranes will achieve high rejection of most PFASs [22] - increase in the mobility of PFASs with increasing pH => result of repulsive electrostatic interactions with the anionic functional group head and sediment - significant relationships with pH (p<0.05) [25] if PFASs sorb to solids in the soil or aquifer material, the transport of PFASs will be retarded and their residence time will be even longer than the residence time of groundwater => PFAS contaminated groundwater can act as a long-lasting source, leading to poorly reversible exposure to PFASs in the subterranean environment as well as where the contaminated groundwater resurfaces [10] - volatile PFASs (FTOHs, FASAs, FASEs) enter the atmosphere, where they can degrade, form intermediates during atmospheric oxidation, or transform into more persistent PFASs, such as PFSAs and PFCAs, which may finally end up in the aquatic environment - PFASs have the potential to affect cell membranes of algae => can alter nutritious quality of biofilms => potential effects on physiological fitness of invertebrates [28] Linear regression showed significant positive correlations of a.o. PFHxS to DOC (p<0.05) => PFASs may bind to and be co-transported by DOC in water => PFCAs and PFSAs are negatively charged in natural waters by proton dissociation of the acid => Short-chains (more hydrophilic) could readily bind to positively charged ions that are complex bound to DOC (such as Ca2+), while longerchained PFASs rather partition to even more hydrophobic phases in the water, such as the organic carbon fraction of suspended particulate matter [17] - the fate of PFASs sorbed to sludge is generally to be further spread in the environment through application of sludge or sludge-containing products - in 2014: 24% of sludge was applied on agricultural soil, 24% was used to cover mines and dump sites, and 29% was used to manufacture soil intended for less sensitive land use [20] discharge of municipal wastewaters is one of the principal routes of entry of PFASs in the aquatic environment => often PFAS concentrations increase in WWTPs as a result of biodegradation of precursors during the activated sludge process [22] - at the end of their lifetimes, products containing PFASs may be disposed of to landfill - as a result of decomposition processes within landfills, PFASs associated with waste can become mobile and leach into water that enters a landfill to produce potentially contaminated leachate - small and very small sites may pose a local point source of PFAS contamination into surrounding groundwater (new and large landfills typically use cell liners to prevent leachate migration into groundwater) [30] - PFAA pattern with PFHxS as dominating homolog (found in observation wells SE) is comparable to patterns found in groundwater from areas around firefighting training sites => high levels of PFHxS in the aquifer could be a marker of AFFF as a potential contamination source => military airport nearby most probably source area for contamination - Relative enrichment of branched isomers in time => AFFF-contaminated drinking water may contain proportionally more branched PFHxS than other exposure media => branched PFHxS isomers may have a higher water solubility compared to the linear isomer, and are therefore enriched in the dissolved phase [35] - PFHxS also detected in shoots of control plants => due to adsorption of N-EtFOSA from the atmosphere followed by subsequent biodegradation in plants - N-EtFOSA efficiently taken up from solutions by wheat, soybean and pumpkin roots and acropetally translocated to shoots in exposured plants => biotransformed to intermediates FOSAA, PFOSA [63] Perfluoroalkyl and perfluoroether moieties are highly persistent under natural conditions => Even though some PFASs may partially degrade in the environment and biota, they will all ultimately transform into highly stable end products, wich are usually perfluoroalkyl or perfluoroalkyl(poly)ether acids such as PFCAs, PFSAs, PFECAs and PFESAs [92] - No conclusive evidence for biodegradation of PFAS in sludge under aerobic conditions over a period of up to either 9 or 15 weeks - Although decreases in PFHxA, 6:2 FTOH, and 8:2 FTOH concentrations were observed, their concentrations in the control bottles also decreased, and it is therefore not possible to confirm that they were indeed due to biodegradation - The concentration decrease could also be due to a non-biological degradation process, losses not due to degradation or even to incomplete sterilization of the control bottles [92] - No evidence for degradation of any of the PFAS was observed under anaerobic conditions - This result is consistent with the lack of anaerobic biodegradation of PFAS reported elsewhere in the literature => The PFAS tested in these experiments are non-biodegradable under the experimental conditions used in this study, despite using municipal sewage sludge, which presumably has a history of exposure to PFAS [29] - max. concentration in biota: 31.4 ng/g ww - uncertainty on the bioaccumulation and biomagnification characteristics of PFHxA, which requires further monitoring data [41] Progesterone content in mLTC-1 (mouse Leydig tumor cells) decreased with the increase in PFHxA treatment concentration => This relationship was illustrated by a standardized S-shape curve (dose-response curve) => Predicted IC50 (half-maximal inhibitory effect concentration): 213.8 mol/L => A similar dose-response curve was seen for cell viability (decreasing cell viability with higher PFHxA treatment concentration), but progesterone content was more sensitive than cell viability to PFHxA exposure => Sign. difference in mitochondrial membrane potential (MMP) between: 125 mol/L PFHxA and control (p<0.05) + 500 mol/L PFHxA and control (p<0.01) => Inhibition effect of PFHxA on progesterone production might be due, in part, to decrease in MMP in mLTC-1 [42] FRESHWATER: Brachionus calyciflorus - egg size values were increased from 0.514 in the control to 0.570 m3 in PFHxA (p<0.05) => females produce larger and fewer eggs under the condition of the increasing cost of producing larger eggs => large eggs produce large larvae and large offspring which exert a strong influence to the ability of survival => response to toxicity [56] PFHxA shown not to be bioaccumulative [63] the high solubility and proteinbinding characteristics of ionic PFAAs challenge the conventional assessment of bioaccumulation potential that is through either bioconcentration factor in aquatic species (usually fish) or models based on octanol-water partition coefficients (Kow) => even when PFAAs are not characterized as bioaccumulative under the current regulatory frameworks, the accumulation of these PFAAs in the environment will lead to increasing external exposure [44] All tested PFCs induced hPXR reporter activites in a dosedependent manner => In general, longer carbon chain length seems to be favorable for the hPXR activity of PFAAs with chain length shorter than 10, as chain length is longer than 10, increasing the chain length of PFAA seems to render the chemicals weaker potency to induce hPXR activity => PFHxA: EC50 = 20.14 M, maximum induction = 22.7 % ==> PXR induction is postulated to be associated with lipid homeostasis, atherosclerosis, carcinogenesis and endocrinedisrupting effects [45] Study on 141 pregnant woman serum samples: - Albumin: 0.200 (p<0.01) (positively correlated with PFHxA) => PFASs may bind to serum albumin which might be related to anti-inflammatory effects of proliferator-activated receptor alpha (PPAR) activation - Total bilirubin: 0.211 (p<0.05) (positively correlated with PFHxA) => PFCAs could be related to adverse liver or bile effects in pregnant women => Results indicate that exposure to PFAAs at high levels may result in health risks to pregnant women or even to the fetus [43] Danio rerio embryos exposed to mixtures involving PFHxA: * UPREGULATION OF ahr2 (involved in metabolism of xenobiotics): - PCB126+PFHxA: at 48 hpf, 54 hpf, 72 hpf and 78 hpf (sign.) - PCB126+PFOS+PFHxA: at 48 hpf and 54 hpf (sign). * UPREGULATION OF cyp1a (involved in metabolism of xenobiotics): - PCB126+PFHxA: at 72 hpf and 78 hpf (sign.) - PCB126+PFOS+PFHxA: at 72 hpf and 78 hpf (sign.) => cyp1a: known to increase the reactive oxygen species content and to cause oxidative stress * UPREGULATION OF gpx1a (involved in oxidative stress): - PCB126+PFHxA: at 30 hpf and 48 hpf (sign.) - PCB126+PFOS+PFHxA: at 30 hpf and 48 hpf (sign). => Effect was only seen in mixtures with PFHxA => PFHxA is capable of having a synergistic effect with PCB126 [43] Danio rerio embryos exposed to mixtures involving PFHxA: * DOWNREGULATION OF acaa2 (involved in lipids metabolism): - PCB126+PFOS+PFHxA: at 78 hpf and 96 hpf (sign.) => Likely that this regulation is due to non-specific effects that do not involve PCB126 but PFOS and PFHxA because of their abilities to modify properties of lipid biological membranes and interfere with lipid signaling pathways * UPREGULATION OF dnmt1 (involved in epigenetic mechanisms): - PCB126+PFOS+PFHxA: at 78 hpf and 96 hpf (sign.) => Likely that an increase in dnmts expression is decreasing expression of the genes they target ==> Potential of a 'non-toxic' chemical, PFHxA, to be actively involved in mixture toxicity => Ability of PFHxA to affect gene expression patterns induced by PCB126 [77] short-chain PFAAs have been found to concentrate in the edible portion of some crops, such as fruits and vegetables, wheras longer chains tend to be retained more in the shoots and roots of plants [46] PFASs can compete with thyroxine (T4) for binding to the human thyroid hormone transport protein transthyretin (TTR), which may lead to reduced thyroid hormone levels leading to endocrine disrupting adverse effects => PFHxA: T4-TTR Binding (%): 43 => Experimental Classification: Active (Y) => Predicted Classification: Active (Y) => Experimental pIC50 (conc. with 50% inhibition of T4): 2.09 mM => Predicted pIC50: 2.17 mM => Binding energy: -4.24 kcal/mol => PFASs can compete with thyroxine (T4) for binding to the human thyroid hormone transport protein transthyretin (TTR), which may lead to reduced thyroid hormone levels leading to endocrine disrupting adverse effects [104] Breast milk samples taken from 264 Korean lactating women: * Level of PFHxA in breast milk: - n > LOD: 187 => detected in 70.8% of all breast milk samples - Median (Interquartile range): 0.047 (0.019-0.079) - Geometric mean: 0.047 * PFHxA was significantly correlated with PFHpA (0.74264, p<0.001), PFOA (0.67035, p<0.001) * PFHxA, PFHpA and PFOA showed high correlation coefficients, implying that they may share similar sources of exposure [46] Discriminating properties between active (Y) and inactive (N) classes: - Hydrophobicity of the H atom attached to heteroatom - Contributions from the H-050 descriptor PFHxA fitted exactly in the identical active sites making interactions with Leu17, Ala108, Ala109, Leu110, Thr118, Thr119 and Lys15, but not with Val121 => A hydrogen bond formed with the positively charged Lys15 and negatively charged oxygen atom of sulfonate or acid functional groups PFASs containing acid functional groups: - If the carbon chain length is between 6 and 10, PFASs will be highly toxic or active ones - Toxicity will be lower or non-existent for PFASs containing a C chain length greater than 10 or below 6 => PFHxA IS TOXIC/ACTIVE [105] Analysis of breast milk samples (n=293) collected from 127 mothers in Korea. Pairs of pregnant women and fetuses that were recruited from four South Korean cities. Detection rates and concentrations of PFHxA in Korean breast milk samples: * No. detected: 116 => detected in 40% of all breast milk samples * Range (min-max): <10 - 129 ng/L * Average SD: 12.8 20.9 ng/L => Study showed low detection rates of contaminant short-chain PFCAs [47] negative correlations were significantly observed between sperm motility and PFHxA in human semen, therefore exposure to PFAAs may result in a decline in semen mobility [48] Change in thyroid hormone TH-responsive genes in neuronal cells of avian species: *Gallus domesticus: D2 mRNA was upregulated following exposure to PFHxA: 10 M: Mean fold change +/- SD: 2.33 +/- 0.53 (p<0.05) D3 mRNA was upregulated following exposure to PFHxA: 3 M: Mean fold change +/- SD: 3.03 +/- 0.20 (p<0.05) MBP was upregulated following exposure to PFHxA: 3M: Mean fold change +/- SD: 2.01 +/- 0.75 (p<0.05), 10 M: Mean fold change +/- SD: 3.39 +/- 0.36 (p<0.05) => short-chained PFCs altered the expression of TH-responsive genes in chicken embryonic neuronal cells to a greater extent than the longchained PFCs => due to bioavailability: could have entered neuronal cells more readily due to their lower binding affinities to extracellular proteins *Larus argentatus: Oct-1 was upregulated following exposure to PFHxA: 3 M: Mean fold change: 3.37 +/- 0.35 (p<0.05), 10 M: [106] Study investigated the relationship between serum concentrations of perfluoroalkyl compounds and growth parameters in 2-year-old Korean children. Study included 361 children aged 2 years (192 boys and 169 girls; 22-27 months). PFHxA concentrations in serum from children at 2 years of age: - LOD: 0.179 ng/mL - Detection frequency: 86 => detected in 23.8% of all serum samples - Range (min-max): 0.127 - 22.80 ng/mL - Mean SD: 0.205 0.298 ng/mL [50] NIH-3T3 cells: PFHxA led to increased PPAR (at 30 M, p<0.01 and 100 M, p<0.001) and PPAR activity (at 100 M, p<0.001) => as PPAR is involved in adipocyte differentiation, PFAS affecting this receptor may be involved in development of obesity => short-chain PFCAs generally activate both PPAR and PPAR with [51] - IC50 value and dissociation constant (Kd) with hPPAR-LBD: PFHxA: IC50: ND, Kd: ND - PFHxA elicited apparent dose-dependent response in Hep G2 cells in terms of hPPAR activation compared to control - Docking interactions of PFHxA with hPPAR-LBD: Hydrogen bonds: Ser-289, His-449, Tyr-473 [52] human embryonal kidney cell line HEK293, activation of human PPAR, PPAR and PPAR - PFHxA triggered activation of PPAR: c20% value: 12.2 M - Most PFCAs also activated PPAR although higher PFCA concentrations were required for PPAR activation => PFHxA: c20% value: 43.2 M - In case of PPAR only high concentrations of PFHpA and PFOA activated this nuclear receptor => PFHxA: c20% value: >200 M => PPAR seems to be the primary molecular target of PFCA [80] Study designed to determine if PFAAs in mixtures exhibit additivity in activating PPAR: - PFAAs were tested individually and in binary combinations of PFOA + either PFNA, PFHxA, PFOS or PFHxS in an 8x8 factorial grid design and the responses of mouse PPAR in transiently transfected COS-1 cells were evaluated - Binary combinations of PFAAs (PFOA + PFNA, PFOA + PFOS, PFOA + PFHxA or PFOA + PFHxS) appear to behave additively in the lower PFOA concentration range of 1-32 M => at higher PFOA concentrations with PFAAs, the observed responses exceeded the predicted responses using either RA or CA (model equations) - For PFOA + PFHxA, curves for modeled data and observed data fit one global curve only at 1-32 M PFOA and at PFHxA concentrations of 8 M and 32-256 M, excluding the two highest [108] Analysis of 424 mother-fetus pairs from the Maoming Birth Cohort, China. * Concentrations of PFHxA in maternal and cord serum: - Cord serum: Detection rate: 63.68%, median: 0.016 ng/mL, mean SD: 0.05 0.14 ng/mL - Maternal serum: Detection rate: 82.55%, median: 0.02 ng/mL, mean SD: 0.04 0.06 ng/mL * PFAS in cord serum was positively correlated with PFAS in maternal serum for all PFAS (p<0.05) [108] * Transplacental transfer efficiency (TPT) of PFHxA = concentration cord serum (ng/mL)/ concentration maternal serum (ng/mL): - TPT (n=247): Median: 1.07, mean SD: 4.32 30.38 - U-shaped pattern for TPT in PFCAs: TPT decreased from PFBA to PFDA and then increased to PFTrDA * Higher TPT in PFAS alternatives than PFASs => PFAS alternatives may be more easily transported from mother to infant than conventional PFASs in uterus * The TPT in a C6 PFCA, PFHxA was higher than PFHxS (1.07 vs. 0.46, p<0.001) which was the corresponding C6 PFSA [81] - no sign. cytotoxicity was observed for the shortest chain length PFCs (PFBA, PFHxA, PFBS and PFHxS) after 24h incubation with the human placental choriocarcinoma cell line JEG-3 - differential bioavailability of PFCs in the in-vitro system was assessed by measuring the fraction retained in the cells after exposure: concentrations for PFHxA at time 0 (3 pmol/mg cell protein) and 24h later (10 pmol/mg cell protein) - inhibition of P450 aromatase (CYP19) activity in JEG-3 cells: 26% inhibition by PFHxA when tested at 500 M, IC50: N.D. => sulfonates (PFBS, PFHxS) were stronger inhibitors of aromatase activity than the corresponding acidic compounds (PFBA, PFHxA) [82] Study of inhibition of cellular viability in human liver cell line (HL-7702): * PFHxA single: - IC-20 (20% stimulatory effect concentration): 2.60 x 10^(-4) mol/L - IC0 (0% inhibitory concentration): 3.33 x 10^(-3) mol/L - IC10 (10% inhibitory concentration): 3.51 x 10^(-3) mol/L - IC50 (50% inhibitory concentration): 4.40 x 10^(-3) mol/L * PFHxA-PFHxS mixture: - all IC0, IC10 and IC50 values in the binary mixture were less than that of individual PFSA or PFCA values - a synergistic action took place under the effective concentrations of IC0, IC10 and IC50 with various concentration ratios - PFHxA/PFHxS showed the strongest synergistic effect with MTI (mixture toxic index) values from 2.30 to 8.08 for five rays - for the effective concentration of IC-20, only the higher proportion of PFSA in the mixtures (PFHxS > 48.9%) showed a synergistic effect for IC-20 - the mixture of PFHxS and PFHxA showed the highest synergistic effect, especially at the high proportion of PFHxS (70%) and indicated that PFHxS had the highest contribution to the synergistic effect [82] - Mixtures of eleven PFAAs (Mix0, Mix10 and Mix50, PFHxA included) presented a partial addition effect - Mixtures of nine PFAAs (Mix0, Mix10, Mix50, PFHxA included) with J-shaped curves only presented a synergistic effect with MTI values ranging from 1.16 to 1.25, however Mix-20 showed partial [84] PFCs can compete with thyroxine (T4, the transport form of thyroid hormone), for binding to the human thyroid hormone transport protein transthyretin (TTR) => such competitive capacity may lead to decreased thyroid hormone levels as previously reported for animals exposed to PFCs => PFCs do not affect the regulatory functions of the thyroid hormone system itself, but it is the competitive binding to transport proteins that alters the free thyroxine (T4) levels in blood => TTR is the main T4 carrier in cerebrospinal fluid, and also important in serum of most mammalian species and birds => Binding potency is clearly associated with the degree of fluorination of the alkyl chain, with a maximum potency at a chain length of eight carbons (PFOA) for PFCAs => PFCAs with a carbon chain length longer than eight have low TTR binding potencies => TTR is both in humans and in rodents the most important carrier protein for thyroid hormone to the developing fetus and the brain [84] PFHxA: - Molecular weight: 314.0 g/mol - T4-TTR binding at maximum concentration (10M): 43% - IC50 (concentration at 50% inhibition): 8220 nM - Slope of dose-response curve: -1.45 - T4-REP (relative potency compared to T4) factor: 0.007 - HPLC retention time: 28.1 min [85] Mean fold change standard error of mRNA expression after exposure to 100 M of PFHxA (in rat H4IIE hepatoma cells): * Thyroid related genes: - Hex: 23.80 3.62 - PAX 8: 3.73 0.28 => upregulation of thyroid related genes * Cholesterol related genes: - Per-3-Keto-: 2.46 0.25 - SQSYN: 14.96 2.83 => upregulation of cholesterol related genes * Lipoprotein related genes: - Mito-3-Keto-: 0.64 0.09 => downregulation of Mito-3-Keto- - Mito-3-Keto-: 1.48 0.30 - ApoA4: 2.18 0.42 => upregulation of Mito-3-Keto- & ApoA4 [86] - weak binding of PFHxA with the protein TR-LBD (TR = thyroid hormone receptor, LBD = ligand-binding domain) * Binding of PFHxA with TR-LBD: - Length: 7.94 A - IC50: >500 M - RP (relative potency (IC50 T3 / IC50 chemical)): 0.0006 - Hydrogen bonding: ARG 228 * all tested PFCs fit into the T3-binding pocket of TR-LBD, with the acid or hydroxyl end group residing toward the inner part and the hydrophobic chain toward the entrance of the binding pocket => all PFCs with an acid end group formed a hydrogen bond with [87] * Binding potency of PFHxA to TTR (TTR competitive binding assay): - IC50: 3189 1156 nM - Kd: 510 184 nM - RP (relative potency (IC50 T4 / IC50 chemical)): 9 x 10^(-3) => perfluoroalkyl acids: Kd values decreased as carbon chain length increased from C4 to C8 => no further increase, but a downward trend was observed for perfluoroalkyl acids with longer chain lengths (C9-C14) * Binding potency of PFHxA to TTRmutK15G: - IC50: ND - Kd: ND - Hydrogen bonding: Lys15 => PFASs exhibited much weaker binding affinities to TTRmutK15G compared with wild-type TTR * Docking study: - Similar to T4, PFOA and PFOS could nearly fill the TTR ligandbinding pocket - Perfluoroalkyl acids with carbon chain length less than C8 did not adequately fill the T4 binding pocket [87] * Binding potency of PFHxA to TBG (TBG competitive binding assay): - IC50: ND - Kd: ND - RP (relative potency (IC50 T4 / IC50 chemical)): ND => only PFTA and PFTdA bound to TBG * Binding potency of PFHxA to TBGmutR378G: - IC50: ND - Kd: ND * Binding potency of PFHxA to TBGmutR381G: - IC50: ND - Kd: ND - Hydrogen bonding: Arg381 => both TBGmut378G and TBGmutR381G exhibited much weaker binding potencies to PFTA and PFTdA when compared with that of wild-type TBG * Docking study: - Perfluoroalkyl acids with carbon chain length less than C12 did not adequately fill the T4 binding pocket - Longer fluorinated carbon chain structures could nearly fill the TBG [23] on-going (increasing) emissions and high persistence => short-chain PFAAs, PFECAs and PFESAs accumulate in the environment => increasing exposure to these substances over time => even if emissions cease in the future it will take decades to centuries to reverse global contamination of short-chain PFAAs, PFECAs and PFESAs, due to high persistence and environmental mobility [25] Shorter-chain PFAAs: more challenging to remove from drinking water than long-chain analogues => only technical solution to reduce external exposure to short-chain PFASs may be to shift water supply sources => exposure of PFHxA will be poorly reversible until technically and financially feasible exposure mitigation measures are identified and taken [34] significant correlation with other PFAS-analytes detected in groundwater (p<0.05) and surface water (p<0.01), indicating a common or similar source of exposure [63] the very high persistence of PFAAs leads to poorly reversible exposure to these substances in the global environment and some local/regional environments including groundwater [101] - For food items of non-animal origin, PFOA, PFHpA and PFHxA were the most frequently detected homologues - Plant uptake of PFCAs and PFSAs is thought to occur primarily through the roots - It is known that sorption to natural organic matter increases with perfluoroalkyl chain length and is stronger for the sulfonates than the carboxylates => These observations explain why the more water soluble and less sorptive PFCAs (PFOA, PFHpA and PFHxA) are more prevalent in vegetable food samples - Although the bioaccumulation potential of PFSAs and PFCAs increases with chain length in hydroponically grown plants, the uptake of the long-chain PFSAs and PFCAs is limited by the low fraction of freely dissolved ("available") chemical in the soil pore water [101] - Dietary intake of PFOS, PFNA, PFDA, PFUnDA, PFDoDA, PFTrDA and PFTeDA was largely dominated by the consumption of fish products (60-100% of the total dietary intake) - In contrast, dietary intake of PFHxA, PFHpA, PFHxS and PFOA originated from a number of food categories with less than 20% of the total dietary intake coming from consumption of fish - While drinking water intake contributes to a major part of the total exposure (36-53%) for shorter chain PFAAs (PFHxA, PFHpA and PFHxS), dust ingestion is estimated to make a significant contribution (27-49%) to the total exposure of PFHxA, PFHpA, PFNA, PFDoDA and PFTeDA - Nevertheless, dietary intake contributed more than 50% of the total exposure for all PFAAs with more than 8 perfluorinated carbon atoms, as dietary exposure to these longer chain PFAAs is primarily due to fish consumption [109] Study on residential exposure: Vacuum cleaner contents from 39 homes were tested for perfluoroalkyl chemicals: * PFHxA: - Detection frequency: 20 %, Median: 0 (ND-180) ng/g [109] * Occupants of all ages can be exposed to indoor contaminants, but preschool-aged children have the greatest risk due to hand to mouth activity and the amount of time they spend playing on or near the floor [110] Study on chicken/duck poultry products and eggs in China => 2 or 3 samples were collected from henneries or abattoirs at different locations => Chicken (n=53), duck (n=47), chicken eggs (n=56, each sample being a combination of five eggs), and duck eggs (n=56, each sample being a combination of five eggs) => 3 tissues from chicken and ducks: breast meat, liver and subcutaneous fat => Final sample number: 300 samples from chicken (53*3=159) and duck (47*3=141): [110] * Eggs: Concentrations of PFHxA were highest in duck eggs * Meat: Concentrations of PFHxA were highest in duck meat (0.20 g/kg) * PFHxA was second highest, after PFHpA, in subcutaneous fat tissues in both chicken and duck (0.350.67 g/kg) => The current PFC concentrations in chicken, chicken eggs, duck and duck eggs from the Yangtze River Delta and Pearl River Delta are not of concern to adults => Chicken and chicken eggs could be dangerous to children! [111] PFAAs were investigated in home produced eggs and commercially produced eggs surrounding a fluorochemical industrial park in China: PFHxA concentrations: * Home produced eggs (n=4, sites 1-4): Egg yolks: - 2 km distance: 0.20 ng/g - 5 km distance: <0.02 ng/g - 10 km distance: 0.08 ng/g - 20 km distance: <0.02 ng/g [111] Egg whites: - 2 km distance: 0.02 ng/g - 5 km distance: <0.02 ng/g - 10 km distance: <0.02 ng/g - 20 km distance: 0.02 ng/g Whole eggs: - 2 km distance: 0.04 ng/g - 5 km distance: <0.02 ng/g - 10 km distance: <0.02 ng/g - 20 km distance: <0.02 ng/g * Commercially produced eggs (n=12, sites 5-16): Egg yolks: - Min: <0.02 ng/g - Max: 0.07 ng/g - Mean: 0.04 ng/g - Median: <0.02 ng/g - n > LOD (%): 3 (25) Egg whites: - Min: <0.02 ng/g - Max: 0.03 ng/g - Mean: 0.03 ng/g - Median: <0.02 ng/g - n > LOD (%): 1 (8) [111] Whole eggs: - Min: <0.02 ng/g - Max: 0.05 ng/g - Mean: 0.03 ng/g - Median: 0.04 ng/g - n > LOD (%): 6 (50) [111] => Unlike in the environmental media, egg yolks contained much lower proportions of PFPeA, PFHxA and PFHpA than of PFBA [116] In May 2009, indoor dust samples (n=17) were collected from fifteen homes located in nine Korean cities for house dust (n=15), and one shopping mall and one library in Seoul for non-residential indoor dust (n=2). Nine cities for house dust samples were chosen to represent densely populated cities, industrialized cities, and rural area. [116] * Concentrations (ng/g) of PFHxA measured in indoor dust samples from Korea (n=17): - House dust (n=15): Range: <LOD - 4.8 Mean ( SD): 1.3 ( 1.3) Median: 1.2 % of < LOD: 27 - Non-residential indoor dust (n=2): Shopping mall: 4.5 Library: 6.6 * There were no significant correlations between PFCAs (or PFSAs) and their precursors in house dust. * In Korean house dust samples, significant correlations (p<0.05) were only found among all the FTOHs and among almost all of the PFCAs. [31] PFHxA serum concentrations in exposed (living in area under impact) and not exposed (subjected to background exposure) subjects EXPOSED (n=257) - Minimum: <LOQ - Median: <LOQ - Maximum: 0.68 ng/g - % levels <LOQ: 80 NOT EXPOSED (n=250) - Minimum: <LOQ - Median: <LOQ - Maximum: 0.26 ng/g - % levels <LOQ: 82 => PFHxA concentration sign. larger in exposed than in not exposed subjects (p<0.001) [32] Concentration of PFHxA in human plasma A samples (3 datasets): - Assigned value: NA - Average: 0.14 ng/ml - Median: 0.07 ng/ml - Min.: 0.03 ng/ml - Max.: 0.33 ng/ml - SD: 0.16 - % relative SD: 114 [32] Concentration of PFHxA in human plasma B samples (3 datasets): - Assigned value: NA - Average: 0.12 ng/ml - Median: 0.07 ng/ml - Min.: 0.04 ng/ml - Max.: 0.26 ng/ml - SD: 0.12 - % relative SD: 97 [45] PFHxA in 141 pregnant woman serum samples: - Mean: 0.49 ng/mL - Max: 1.07 ng/mL - Min: 0.13 ng/mL - Geomean: 0.45 ng/mL - Median: 0.47 ng/mL [47] PFHxA in human blood samples: - Detection frequency: 100% - Mean +/- SD: 37 +/- 22 ng/mL - 5th percentile: 11 ng/mL - Median: 29 ng/mL - 95th percentile: 70 ng/mL [40] Human: * Lung: - Mean: 50.1 ng/g wet weight - Median: 207 ng/g wet weight - Range: 569-BDL ng/g wet weight - MLOD: 9.42 ng/g wet weight - % of detection: 89 => PFHxA was PFAS-compound with second highest median concentration in lung * Kidney: - Mean: 5.6 ng/g wet weight - Median: 2.7 ng/g wet weight - Range: 57.1-BDL ng/g wet weight - MLOD: 5.37 ng/g wet weight - % of detection: 25 [40] - Smokers showed smaller accumulation of PFASs - Older people (more than 60 years) showed higher concentrations of PFASs => Clear indication that these compounds accumulate after a long-term exposure [71] Shorter chain PFASs have been shown to absorb into the liver more readily than those with longer chains which are concentrated in the blood proteins [117] Surface water and groundwater in a large area of the Veneto Region in northeastern Italy was contaminated with PFAS that were also found in drinking water samples. The health surveillance program is a free-of-charge population-based screening program. Residents who decided to participate in the program completed a structured interview administered by a trained public health nurse, followed by blood pressure measurement, and blood and urine sampling. A total of 18,345 residents were included in the present analysis. Only 3 of the 12 PFAS were quantifiable in at least 80% of serum samples: PFOA (99.9%), PFOS (99.8%) and PFHxS (98.1%) [117] Serum concentrations (ng/mL) of PFHxA and percentage of samples above the LOQ in the study population (18,345 subjects, 14-39 years of age): PFHxA: - Min: <0.5 - 5th percentile: <0.5 - 25th percentile: <0.5 [117] - Median: <0.5 - 75th percentile: <0.5 - 95th percentile: <0.5 - Max: 7.1 % samples LOQ: 0.3% [23] short-chain PFAAs are similarly persistent as their long-chain homologues [25] for PFASs an alternative measure of persistence (joint persistence) should be taken into account => because PFASs are, or ultimately degrade to, stable end products, all PFASs possess a joint persistence that exceeds all three of the REACH singlemedia persistence criteria [63] Perfluoroalkyl and perfluoroether moieties are highly persistent under natural conditions => Even though some PFASs may partially degrade in the environment and biota, they will all ultimately transform into highly stable end products, wich are usually perfluoroalkyl or perfluoroalkyl(poly)ether acids such as PFCAs, PFSAs, PFECAs and PFESAs [14] PFHxS has similar or even longer serum half-lives than PFOS in all tested animals (expect female rats), and in humans, which makes PFHxS inappropriate as an alternative to PFOS [23] in most cases (hepatotoxicity and reproductive toxicity) shortchain PFAAs are less toxic than long-chain homologues [98] Beijing prenatal exposure study (human): Correlations between fetal PFHxS and maternal thyroid hormones (adjusted for influential covariates): - Maternal FT3: -0.192 (p<0.05) - Maternal FT4: Not significant - Maternal T3: -0.172 (p<0.05) - Maternal T4: Not significant - Maternal TSH: Not significant => Fetal PFHxS is negatively correlated with maternal FT3 and T3 [23] Short-chain PFAAs: - P - not B - maybe T - LRTP [23] short-chain PFAAs are less bioaccumulative in animals and humans than the long-chain homologues, but may show higher uptake into the leaves, stems and fruits of plants [25] long-chain PFAAs are clearly toxic, but it is unclear if shortchain PFAAs would fulfil the T criterion due to uncertainties in their toxicity [103] Analysis of 369 pairs of maternal and umbilical cord serum in a prospective birth cohort in Shandong, China: PFHxS: * Maternal serum: - Detection rate: 99.7 % - Geometric mean: 0.32 - Range: <LOD - 1.68 * Cord serum: - Detection rate: 100 % - Geometric mean: 0.31 - Range: 0.11 - 1.06 * Correlation between maternal and cord levels of PFHxS (stat. sign.): 0.518 (p=0.000) * Transplacental transfer efficiency: = Individual PFAS concentration in cord serum (ng/mL) / corresponding PFAS concentration in maternal serum (ng/mL): Range: 0.19 - 9.22 => Short-chain PFASs were transferred more efficiently from maternal serum to cord serum * Strong correlations between maternal and cord levels suggest that PFASs could readily pass through the placenta [25] short-chain PFAAs fulfill the P and M criteria (M will decrease with perfluoroalkyl chain length) and are highly bioavailable to humans [56] PFHxS shown to be bioaccumulative [33] serum PFHxS sign. negatively associated with lung function in children with asthma Mean change in pulmonary function with increasing PFHxS concentration: * Forced vital capacity (FVC): -0.075 (-0.148 to -0.002) * Forced expiratory volume in 1 s (FEV1): -0.101 (-0.166 to -0.036) * Forced expiratory flow 25-75% (FEF25-75): -0.195 (-0.325 to 0.065) [105] Analysis of breast milk samples (n=293) collected from 127 mothers in Korea. Pairs of pregnant women and fetuses that were recruited from four South Korean cities. Detection rates and concentrations of PFHxS in Korean breast milk samples: * No. detected: 102 => detected in 35% of all breast milk samples * Range (min-max): <10 - 133 ng/L * Average SD: 7.64 14.8 ng/L * Duration of breastfeeding was positively associated with the serum concentrations of ln PFHxS (p<0.001) [63] the high solubility and proteinbinding characteristics of ionic PFAAs challenge the conventional assessment of bioaccumulation potential that is through either bioconcentration factor in aquatic species (usually fish) or models based on octanol-water partition coefficients (Kow) => even when PFAAs are not characterized as bioaccumulative under the current regulatory frameworks, the accumulation of these PFAAs in the environment will lead to increasing external exposure [38] PFHxS: Repeated study on aging individuals (SE): associations between change in plasma PFAS concentrations and markers of liver function (95% CI) - Biluribin: -1.00 (-1.31, -0.70) mol/L (p=1.10E-10) => Changes in plasma concentrations of PFHxS were inversely associated with changes in circulatory bilirubin => Decreased levels, or highly increased, bilirubin has been associated with adverse health effects, including liver dysfunction, cardiovascular disease, diabetes and the metabolic syndrome => These findings suggest a relationship between low-dose background PFAS exposure and altered liver function in the general population [106] Study investigated the relationship between serum concentrations of perfluoroalkyl compounds and growth parameters in 2-year-old Korean children. Study included 361 children aged 2 years (192 boys and 169 girls; 22-27 months). PFHxS concentrations in serum from children at 2 years of age: - LOD: 0.160 ng/mL - Detection frequency: 359 => detected in 99.4% of all serum samples - Range (min-max): 0.113 - 5.980 ng/mL - Mean SD: 1.398 0.889 * Height at 2 years of age was sign. inversely related to PFHxS concentrations: (95% CI): -0.84 (-1.26, -0.42) (p<0.001) * 2-year height gain sign. decreased with increased PFHxS concentrations: (95% CI): -0.89 (-1.45, -0.33) (p=0.002) [2] human exposure primarily via dietary intake => contribution of drinking water [25] for PFAAs elimination occurs through different mechanisms in aquatic and terrestrial species => in mammalian species: governed by renal and biliary clearance => in fish: PFAAs can also be eliminated by the respiratory system (gills) => elimination in fish is much more efficient than in (some) mammalian species => PFHxS has low BCF FISH due to rapid elimination but PFHxS is only slowly eliminated in humans with half-lives on the scales of years [7] humans: concentration in plasma (neonates): geometric mean = 0.34 g/L median = 0.37 g/L range <LOQ/1.33 g/L concentration in plasma (adults aged 50-65): geometric mean = 1.57 g/L median = 1.61 g/L range 0.10/12.43 g/L [47] PFHS in human semen samples: - Detection frequency: 93% - Mean +/- SD: 0.13 +/- 0.11 ng/mL - 5th percentile: <LOQ - Median: 0.092 ng/mL - 95th percentile: 0.40 ng/mL => Lower concentration of PFHS in semen was caused by the bloodtestis barrier, which can protect germ cells from toxic chemicals => positively (0.361) associated with age (p<0.01) => positively (0.413) associated with BMI (p<0.01) => negatively (-0.244) correlated with progressive motility (p<0.05) [40] Human: * Liver: - Mean: 4.6 ng/g wet weight - Median: 1.8 ng/g wet weight - Range: 20.6-BDL ng/g wet weight - MLOD: 3.00 ng/g wet weight - % of detection: 10 * Bone: - Mean: 1.8 ng/g wet weight - Median: 1.2 ng/g wet weight - Range: 13.8-BDL ng/g wet weight - MLOD: 2.40 ng/g wet weight - % of detection: 5 [7] dietary intake = main route of exposure influencing factors: - age mother: cord blood concentrations 91% higher in mothers older than 35 compared to mothers aged 25 or less (p<0.001) - age: adults 60-65 yrs had 40% higher serum levels compared to adults 50-55 yrs (p<0.01) - consumption potatoes: higher cord blood levels with higher consumption potatoes (p<0.05) - Mediterranean diet: a higher score for Mediterranean diet was associated with higher serum levels (p<0.05) - alcohol consumption before pregnancy: 60% higher cord blood levels for mothers drinking weekly compared to mothers never drinking (p<0.001) -gender: significant gender differences in serum of adults (p<0.001) - parity: women with three or more children had 39% lower serum levels compared to women without children (p<0.05) - duration lactation: mothers with a lactation period longer than six months had lower cord blood levels of 40% (p<0.001) [30] levels of PFHxS in serum from participating mothers (SE): Tot PFHxS: 1996-1999 (n=146): - MDL: 0.01 ng/g - Median (range): 1.8 (0.37-9.5) ng/g - Mean (SE): 2.3 (0.13) ng/g - % of levels <MDL: 0 2008-2011 (n=148): - MDL: 0.01 ng/g - Median (range): 3.7 (0.32-34) ng/g - Mean (SE): 5.4 (0.45) ng/g - % of levels <MDL: 0 => sign. higher compared to '96-'99 (p<0.05) [40] Human: * Brain: - Mean: 3.2 ng/g wet weight - Median: 2.3 ng/g wet weight - Range: 14.4-BDL ng/g wet weight - MLOD: 4.54 ng/g wet weight - % of detection: 5 * Lung: - Mean: 8.1 ng/g wet weight - Median: 5.7 ng/g wet weight - Range: 47.6-BDL ng/g wet weight - MLOD: 3.30 ng/g wet weight - % of detection: 32 [23] on-going (increasing) emissions and high persistence => short-chain PFAAs, PFECAs and PFESAs accumulate in the environment => increasing exposure to these substances over time => even if emissions cease in the future it will take decades to centuries to reverse global contamination of short-chain PFAAs, PFECAs and PFESAs, due to high persistence and environmental mobility [25] Shorter-chain PFAAs: more challenging to remove from drinking water than long-chain analogues => only technical solution to reduce external exposure to short-chain PFASs may be to shift water supply sources [30] levels of PFHxS in serum from participating mothers (SE): % branched PFHxS: 1996-1999 (n= 146): - MDL: 0.01 ng/g - Median (range): 5.3 (0.41-14) ng/g - Mean (SE): 5.6 (0.21) ng/g - % of levels <MDL: 0 2008-2011 (n=148): - MDL: 0.01 ng/g - Median (range): 6.9 (0.24-15) ng/g - Mean (SE): 7.1 (0.23) ng/g - % of levels: <MDL: 0 => sign. higher compared to '96-'99 (p<0.05) => relative enrichment of branched isomers [31] PFHxS serum concentrations in exposed (living in area under impact) and not exposed (subjected to background exposure) subjects EXPOSED (n=257) - Minimum: <LOQ - Median: 2.98 ng/g - Maximum: 43.43 ng/g - % levels <LOQ: 8.9 NOT EXPOSED (n=250) - Minimum: <LOQ - Median: 2.49 - Maximum: 9.14 ng/g - % levels <LOQ: 4.8 => PFHxS concentration sign. larger in exposed than in not exposed subjects (p=0.0022) [40] Human: * Kidney: - Mean: 20.8 ng/g wet weight - Median: 18 ng/g wet weight - Range: 37-BDL ng/g wet weight - MLOD: 4.20 ng/g wet weight - % of detection: 5 [40] - Smokers showed smaller accumulation of PFASs - Older people (more than 60 years) showed higher concentrations of PFASs => Clear indication that these compounds accumulate after a long-term exposure [68] PFHxS: Association constants (Ka) and binding sites (n): - Binding to serum albumin in blood and interstitial fluids: KA Alb (albumin) (low): 1.2 x 10^4 M^(-1) (n = 1.65) - Binding to FABP in the liver: KA FABP (fatty acid binding protein): 1.7 x 10^4 M^(-1) (n = 1) [39] Relative response of mouse and human PPAR to PFAAs in transiently transfected COS-1 cells, measured by C20max: PFHxS: - Mouse: 76 M (ranked most active to least active: 11/13 PFASs) - Human: 81 M (ranked most active to least active: 8/13 PFASs) => A lower C20 max value indicates a stronger induction of PPAR alpha activity by that PFAA => PPAR is activated by PFAAs of 4-12 carbons in chain length => Increasing activity of PPAR with increasing chain length of the PFAA up to C9, and lower activity with longer chain PFAAs (C>9) with both mouse and human PPAR => PPAR has been identified as a key player in the mode of action [106] * Weight at 2 years of age was inversely related to PFHxS concentrations: (95% CI): -0.20 (-0.42, 0.02) (p=0.070) * 2-year weight gain decreased with increased PFHxS concentrations: (95% CI): -0.17 (-0.33, 0.16) (p=0.501) * Results indicated a positive correlation (all p<0.001) between perfluorinated sulfonates (PFHxS and PFOS) and carboxylates (PFHpA, PFOA, PFNA, PFDA and PFUnDA), suggesting common exposure for such perfluoroalkyl compounds * Duration of breastfeeding was positively associated with the serum concentrations of ln PFHxS (p<0.001) [68] PFHxS: Renal clearance/reabsorption facilitated by organic anion transporters (OATs): * Flux for protein-facilitated uptake: - Oat1: 0.41 nmol/mg protein/min - Oat3: 0.44 nmol/mg protein/min - Oatp1a1: 0.13 nmol/mg protein/min * Rate constant (b) for clearance: bclear: 0.023 s-1 * Rate constant (b) for reabsorption: breab: 0.004 s-1 [41] Progesterone content in mLTC-1 (mouse Leydig tumor cells) decreased with the increase in PFHxS treatment concentration => This relationship was illustrated by a standardized S-shape curve (dose-response curve) => Predicted IC50 (half-maximal inhibitory effect concentration): 450.3 mol/L => A similar dose-response curve was seen for cell viability (decreasing cell viability with higher PFHxS treatment concentration), but progesterone content was more sensitive than cell viability to PFHxS exposure [107] Analysis of 369 families with matched parental and cord serum samples from a birth cohort in Shandong, one of the regions seriously polluted by PFASs in China * PFHxS concentrations in 369 matched serum samples: - Paternal serum: Detection frequency: 100.0%, Geometric mean: 0.82 ng/mL - Maternal serum: Detection frequency: 99.7%, Geometric mean: 0.32 ng/mL - Cord serum: Detection frequency: 100.0%, Geometric mean: 0.31 ng/mL * Positive correlation between maternal & cord serum: 0.52 (p<0.01) * Positive correlation between paternal & maternal serum: 0.23 (p<0.01) [41] Sign. difference in mitochondrial membrane potential (MMP) between: 62.5 mol/L PFHxS and control (p<0.05) + 125 mol/L PFHxS and control (p<0.01) + 250 mol/L PFHxS and control (p<0.01) + 500 mol/L and control (p<0.01) => Inhibition effect of PFHxS on progesterone production might be due, in part, to decrease in MMP in mLTC-1 [107] * Positive correlation between paternal & cord serum: 0.14 (p<0.01) * Positive correlation between PFHxS concentrations and maternal age: SE: 0.006 0.002 (p=0.010) * Positive correlation between PFHxS concentrations and milk consumption: SE: 0.041 0.018 (p=0.024) [45] Study on 141 pregnant woman serum samples: - WBC: 0.191 (p<0.05) (positively correlated with PFHxS) => Elevated concentrations of PFHxS may have relationship with some diseases, which are caused by the change of WBC count - Hemoglobin: 0.180 (p<0.05) (positively correlated with PFHxS) => Hemoglobin might be a more sensitive parameter than RBC to indicate the influence of PFAAs on human => Results indicate that exposure to PFAAs at high levels may result [108] Analysis of 424 mother-fetus pairs from the Maoming Birth Cohort, China. * Concentrations of PFHxS in maternal and cord serum: - Cord serum: Detection rate: 89.86%, median: 0.07 ng/mL, mean SD: 0.15 0.38 ng/mL - Maternal serum: Detection rate: 99.76%, median: 0.16 ng/mL, mean SD: 0.20 0.17 ng/mL * PFAS in cord serum was positively correlated with PFAS in [46] PFASs can compete with thyroxine (T4) for binding to the human thyroid hormone transport protein transthyretin (TTR), which may lead to reduced thyroid hormone levels leading to endocrine disrupting adverse effects => PFHxS: T4-TTR Binding (%): 3 => Experimental Classification: Active (Y) => Predicted Classification: Active (Y) => Experimental pIC50 (conc. with 50% inhibition of T4): 3.14 mM => Predicted pIC50: 3.13 mM => Binding energy: -4.37 kcal/mol => PFASs can compete with thyroxine (T4) for binding to the human thyroid hormone transport protein transthyretin (TTR), which may lead to reduced thyroid hormone levels leading to endocrine disrupting adverse effects [46] Discriminating properties between active (Y) and inactive (N) classes: - Hydrophobicity of the H atom attached to heteroatom - Contributions from the H-050 descriptor PFHxS, the most active compound considering its pIC50 and T4-TTR % binding potency values, showed the highest binding energy by fitting in the active protein sites via hydrophobic interactions with Leu17, Ala108, Ala109, Leu110 and Val121; polar interactions with Thr118 and Thr119; and hydrogen bonding with positively charged Lys15 PFASs containing sulfonate or sulfinate functional groups: - Toxic if chain length is less than C8 - Compounds will be inactive or show little toxicity if the carbon chain length is over 8 atoms [108] * Transplacental transfer efficiency (TPT) of PFHxS = concentration cord serum (ng/mL)/ concentration maternal serum (ng/mL): - TPT (n=380): Median: 0.46, mean SD: 1.49 8.49 - U-shaped pattern for TPT in PFCAs: TPT decreased from PFBA to PFDA and then increased to PFTrDA * Higher TPT in PFAS alternatives than PFASs => PFAS alternatives may be more easily transported from mother to infant than conventional PFASs in uterus [47] negative correlations were significantly observed between sperm motility and PFHS in human semen, therefore exposure to PFAAs may result in a decline in semen mobility [48] Change in thyroid hormone TH-responsive genes in neuronal cells of avian species: *Gallus domesticus: D3 mRNA was upregulated following exposure to PFHxS: 3 M: Mean fold change +/- SD: 5.34 +/- 0.95 (p<0.05), 10 M: Mean fold change +/- SD: 4.23 +/- 0.95 (p<0.05) TTR was decreased following exposure to PFHxS: 3 M: Mean fold change +/- SD: 0.22 +/- 0.09 (p<0.05), 10 M: Mean fold change +/- SD: 0.34 +/- 0.23 (p<0.05) RC3 was upregulated following exposure to PFHxS: 3 M: Mean fold change +/- SD: 6.58 +/- 3.63 (p<0.05), 10 M: Mean fold change +/- SD: 10.97 +/- 2.24 (p<0.05) => short-chained PFCs altered the expression of TH-responsive genes in chicken embryonic neuronal cells to a greater extent than the longchained PFCs => due to bioavailability: could have entered neuronal cells more readily due to their lower binding affinities to extracellular proteins * Larus argentatus: Oct-1 was upregulated following exposure to PFHxS: 3 M: Mean fold change +/- SD: 2.58 +/- 0.29 (p<0.05), 10 M: Mean fold change +/- SD: 2.32 +/- 0.10 (p<0.05) [51] - IC50 value and dissociation constant (Kd) with hPPAR-LBD: PFHxS: IC50: 41.2 6.8 M, Kd: 285.3 47.1 M - PFHxS elicited apparent dose-dependent response in Hep G2 cells in terms of hPPAR activation compared to control - Docking interactions of PFHxS with hPPAR-LBD: Hydrogen bonds: Ser-289, His-449, Tyr-473 [30] enrichment of branched PFHxS isomers in serum from residents in areas receiving contaminated drinking water => could potentially be used as a marker of exposure to AFFF contaminated drinking water => PFAAs can be transported over long distances in groundwater aquifers => contaminations at AFFF training sites could be a threat to drinking water quality in many areas around the world [31] PFHxS serum concentrations in exposed (living in area under impact) and not exposed (subjected to background exposure) subjects => significant positive correlation between PFHxS and other PFAS-analytes (in exposed (all) + not exposed (all, except for PFHpA)) => may be due to the presence of a dominant common source of exposure in the area [34] significant correlation with other PFAS-analytes detected in groundwater (p<0.05) and surface water (p<0.01), indicating a common or similar source of exposure [35] crops contaminated with PFSAs, including PFOS, PFHxS and PFBS from sulfluramid use may represent a potential source of species exposure and lead to a risk to human health [63] the very high persistence of PFAAs leads to poorly reversible exposure to these substances in the global environment and some local/regional environments including groundwater [101] - Dietary intake of PFOS, PFNA, PFDA, PFUnDA, PFDoDA, PFTrDA and PFTeDA was largely dominated by the consumption of fish products (60-100% of the total dietary intake) - In contrast, dietary intake of PFHxA, PFHpA, PFHxS and PFOA originated from a number of food categories with less than 20% of the total dietary intake coming from consumption of fish - While drinking water intake contributes to a major part of the total exposure (36-53%) for shorter chain PFAAs (PFHxA, PFHpA and PFHxS), dust ingestion is estimated to make a significant contribution (27-49%) to the total exposure of PFHxA, PFHpA, PFNA, PFDoDA and PFTeDA [101] - Nevertheless, dietary intake contributed more than 50% of the total exposure for all PFAAs with more than 8 perfluorinated carbon atoms, as dietary exposure to these longer chain PFAAs is primarily due to fish consumption [109] Study on residential exposure: Vacuum cleaner contents from 39 homes were tested for perfluoroalkyl chemicals: * PFHxS: - Detection frequency: 100 %, Median: 16 (2.11000) ng/g - Correlation between PFHxS and PFBS (0.40, p<0.05) - Correlation between PFHxS and PFOS (0.87, p<0.05) - Correlation between PFHxS and PFDS (0.38, p<0.05) - Correlation between PFHxS and PFPeA (0.49, p<0.05) - Correlation between PFHxS and PFHpA (0.37, p<0.05) - Correlation between PFHxS and PFOA (0.48, p<0.05) * Four of five homes in our study that had unusually high levels of PFOS and PFHxS in dust were constructed between 1968 and 1995 => mean levels were sign. higher in homes built during this time period compared to levels in dust from other homes (PFHxS means 219 vs 57 ng/g, p<=0.05) => more use of floor coverings, draperies and upholstered furnishings that were treated with these compounds * PFHxS made up 24% of the summed PFC residues in these samples [109] * High correlation between PFHxS and PFOS may suggest a common source of these residues * PFOS, PFOA, PFHpA and PFHxS were the most prominent PFCs found in these dust samples with median concentrations ranging from 16 to 47 ng/g [31] - males sign. higher serum concentrations than females in exposed (p<0.001) and not exposed group (p<0.001) - higher serum concentrations with BMI in exposed group (p=0.001) - higher serum concentrations in rural environments in exposed group (p<0.001) - raising own livestock sign. higher serum concentrations in exposed group (p=0.029) - sign. and direct correlation with specific food groups in exposed (cereals and derivatives) and not [32] Concentration of lin-PFHxS in human plasma A samples (21 datasets): - Assigned value: 3.33 ng/ml - Average: 3.28 ng/ml - Median: 3.38 ng/ml - Min.: 1.41 ng/ml - Max.: 4.90 ng/ml - SD: 0.80 - % relative SD: 25 [32] Concentration of lin-PFHxS in human plasma B samples (21 datasets): - Assigned value: 3.27 ng/ml - Average: 3.30 ng/ml - Median: 3.15 ng/ml - Min.: 1.39 ng/ml - Max.: 5.60 ng/ml - SD: 0.89 [38] PFHxS: Repeated study on aging individuals SE: Median (Interquartile range): - 2001-2004 (Age 70): 2.08 (1.6, 3.42) - 2006-2009 (Age 75): 3.24 (2.05, 6.22) - 2011-2014 (Age 80): 2.87 (1.8, 10.5) Increasing trend (p<0.00001) [45] PFHxS in 141 pregnant woman serum samples: - Mean: 0.48 ng/mL - Max: 3.65 ng/mL - Min: 0.08 ng/mL - Geomean: 0.38 ng/mL - Median: 0.38 ng/mL [47] PFHS in human blood samples: - Detection frequency: 100% - Mean +/- SD: 4.3 +/- 2.8 ng/mL - 5th percentile: 0.29 ng/mL - Median: 3.9 ng/mL - 95th percentile: 9.7 ng/mL => positively (0.194) associated with age (p<0.05) => positively (0.220) associated with BMI (p<0.05) [62] PFHxS was detected in every U.S. human blood sample from the 1999-2000 National Health and Nutrition Examination Survey (NHANES) that was analyzed for PFAAs [66] PFHxS serum concentrations (ng/mL) - In 106 participants in a panel study 6 months after end of exposure through contaminated drinking water: Mean +/- SD: 353 +/- 260, Min: 12.3, Median: 277, Max: 1660 - Main Ronneby (3418 participants): Mean +/- SD: 228 +/- 232, Min: <0.5 (LOD), Median: 152, Max: 1790 - Karlshamm reference (242 participants): Mean +/- SD: 1.91 +/- 5.27, Min: <0.5 (LOD), Median: 0.84, Max: 60.1 => Median serum level of PFHxS was 180 times higher in the investigated Ronneby population compared with the referents from a neighbouring municipality => Fire drill site at nearby military airport in Ronneby (SE) seemed to be the cause [66] PFHxS: Initial serum levels (ng/mL) (info from other studies): - Study 1: Median (range): 193 (16-1295) - Study 2: Median (range): ~2.0 (<0.1-2.7) - Study 3: GM: 6.4 [71] Shorter chain PFASs have been shown to absorb into the liver more readily than those with longer chains which are concentrated in the blood proteins [25] PFHxA has a shorter elimination half-life in humans (14-49 days in highly exposed populations) than PFOS [9] The reported serum half-life of PFHxS in humans is 8.5 years [14] - rats: IV, 10 ppm: female: 2 d, male: 29 d - mice: oral, 1 ppm: female: 25 d, male: 31 d - mice: oral, 20 ppm: female: 27 d, male: 28 d - monkeys: IV, 10 ppm: female: 87 +/- 27 d, male: 141 +/30 d - humans: female (n = 2): 12.8 +/- 0.6 yr, male (n = 24): 8.2 +/- 5.1 yr [30] short-chain PFBS has a much shorter half-life (26 days) than PFOS (5 years) and PFHxS (8 years) in humans [62] PFHxS is eliminated in 8.5 years in humans on average [66] Excretion rate constant for serum PFHxS concentrations in 106 participants in a panel study after end of exposure through contaminated drinking water: - All: Mean: 0.13, 95% CI: 0.12 to 0.15 - Men aged 15-50: Mean: 0.09, 95% CI: 0.07 to 0.11 - Women aged 15-50: Mean: 0.15, 95% CI: 0.12 to 0.18 (difference between genders in the model for excretion rate, p=0.008) (subgroup aged 15-50 includes 20 men and 30 women) [66] PFHxS: Half-life (years) (info from other studies): - Study 1: Median: 7.1, range: 2.2-27.0, GM: 7.3, 95% CI: 5.8 to 9.2 - Study 2: (Relative reduction 2006-2008, 30% in women, 14% in men) - Study 3: Average: 15.5 [66] Half-life (years) for serum PFHxS concentrations in 106 participants in a panel study after end of exposure through contaminated drinking water: - All: Mean: 5.3, 95% CI: 4.6 to 6.0, Median: 5.5, 95% CI: 3.0 to 9.2 - Men aged 15-50: Mean: 7.4, 95% CI: 6.0 to 9.7 - Women aged 15-50: Mean: 4.7, 95% CI: 3.9 to 5.9 (difference between genders in the model for excretion rate, p=0.008) => Women aged 15-50 had a considerably shorter mean half-life for PFHxS compared with men, with men 1.6-fold longer (subgroup aged 15-50 includes 20 men and 30 women) [100] PFHxS population halving time (time period 20022013, based on human serum samples): 4.6 years [39] NOEC and LOECs for PFAAs for transactivation of mouse and human PPAR in transiently transfected COS-1 Cells: PFHxA: Human: - NOEC: 5 M - LOEC: 10 M - LOEC: 3.14 g/mL (p<0.05) [101] Concentrations of PFHxA in food basket samples (SE): - Dairy products: 1999-2010: <MDL - Meat products: 1999-2010: <MDL - Fats: 1999: <MDL, 2005: <MDL, 2010: 4.3 pg/g - Pastries: 1999: 6.4 pg/g, 2005: 4.2 pg/g, 2010: 4.1 pg/g - Fish products: 1999-2010: <MDL - Egg: 1999: 5.1 pg/g, 2005: 5.1 pg/g, 2010: 3.6 pg/g - Cereal products: 1999: 7.5 pg/g, 2005: 11 pg/g, 2010: 4.4 pg/g - Vegetables, including root vegetables: 1999: 5.2 pg/g, 2005: 3.0 pg/g, 2010: 3.2 pg/g - Fruit: 1999: <MDL, 2005: <MDL, 2010: 2.8 pg/g [101] - Potatoes: 1999: <MDL, 2005: 2.0 pg/g, 2010: 2.6 pg/g - Sugar and sweets: 1999: 4.0 pg/g, 2005: 3.4 pg/g, 2010: 3.2 pg/g - Soft drinks, lemonade: 1999: 1.6 pg/g, 2005: 1.7 pg/g, 2010: 1.4 pg/g => Short-chain PFCAs (PFHxA and PFHpA) were detected at concentrations close to MDL in cereals, vegetables and potatoes [30] observation wells SE: median water conc. PFHxA: 81 ng/L [102] * Comparison tap water and hot water NL: - Drinking water treatment plant A: Tap water: 1.3 ng/L & Hot water: 1.3 ng/L - Drinking water treatment plant B: Tap water: 2.3 ng/L & Hot water: 2.4 ng/L => Concentrations of PFAAs detected in hot water were similar to those in the corresponding tap water * Post-mixed cola NL (n=6): Average SD (Range): 0.89 0.77 (0.16-2.0) ng/L => Ratio tap water/post-mixed cola (Ratio of averages): 2.1 => PFAS concentrations in tap water were higher than in cola => A combination of the dilution step and the purification step inside the cola [102] * Coffee NL: - Brewed coffee from coffee machines (n=12): Average SD (Range): Could not be quantified - Manually brewed coffee (n=4): Average SD (Range): Could not be quantified - Procedure blank brewed coffee (n=2): Average SD (Range): 1.7 0.13 (1.6-1.7) ng/L [113] 397 food samples collected from a market in Busan (Korea) in 2011 (n=227) and 2012 (n=170). Food samples of 66 different food types, classified into 7 food categories. 34 tap water samples were also collected from 16 districts in Busan. PFHxA concentrations in each food group (ng/g, beverage unit: ng/L): * Fish and shellfish (n=99): DF (%) = 8.1, Mean = 0.037, Max = 1.00 * Meat and its products (n=39): DF (%) = 51.3, Mean = 0.515, Max = 4.38 * Vegetables and fruit (n=78): DF (%) = 20.5, Mean = 0.039, Max = 0.444 * Processed products (n=90): DF (%) = 15.6, Mean = 0.148, Max = 5.15 [113] * Dairy (n=37): DF (%) = 8.1, Mean = 0.051, Max = 0.928 [113] * Beverage (n=21): DF (%) = 9.5, Mean = 0.187, Max = 2.63 * Others (n=33): DF (%) = 33.3, Mean = 0.280, Max = 1.05 PFHxA concentrations in bottled water and tap water samples (ng/L): * Bottled water (n=8): DF (%) = 0.0, Mean -, Max PFHxA was the most frequently detected PFAA in the meat and meat products samples, being found in 51% of the samples analyzed, whereas PFOS was only found in 13% of the meat and meat products samples [113] PFHxA was detected in all of the tap water samples. This indicates that Korean tap water is contaminated with PFAAs, and this is likely to be caused by PFAAs contamination in the original sources of the drinking water. [113] 90% of the drinking water in Korea comes from flowing surface water. [115] In September 2011, foods were purchased in 12 representative cities of Catalonia, all with more than 20,000 inhabitants. Food samples were obtained at each locality in 4 shops/stores of different size. Foods selected for PFAS analysis were among the most consumed in Catalonia. * Mean concentrations (in pg/g fw) of PFHxA in the groups of analyzed foodstuffs: [115] - Meat and meat products: <30 - Fish and seafood: 67 - Vegetables: <47 - Tubers: <36 - Fruits: <36 - Eggs: <39 - Milk: <50 - Dairy products: 31 - Cereals: <12 - Pulses: <26 - Oils: <14 - Industrial bakery: <11 [115] * Fish and shellfish was the group in which more PFASs were detected and where the highest PFAS concentrations were found * Dietary intake of PFHxA by the population of Catalonia (in ng/kg body weight/day): - Children: ND = 0: 0.10; ND = 1/2 LOD: 1.87 - Boys adolescents: ND = 0: 0.01; ND = 1/2 LOD: 0.54 - Girls adolescents: ND = 0: 0.01; ND = 1/2 LOD: 0.50 [115] - Male adults: ND = 0: 0.01; ND = 1/2 LOD: 0.41 - Female adults: ND = 0: 0.02; ND = 1/2 LOD: 0.51 - Male seniors: ND = 0: 0.02; ND = 1/2 LOD: 0.48 - Female seniors: ND = 0: 0.01; ND = 1/2 LOD: 0.43 * Concerning human health risks from dietary exposure to PFASs in Catalonia, it is important to note that for any of the age/gender groups of population estimated, the TDIs recommended by the EFSA were not exceeded. [118] Spatial and seasonal distributions of PFAS in air, water, sediment, soil, and fish samples in the Asan Lake area of South Korea. Estimation of bioaccumulation potential of PFAS using sediment-water partition coefficients and bioaccumulation factors. [73] PFHxA was quantified in 56% of all samples in the Greenland Sea, with concentrations ranging from <5.9 to 38 pg/L => This result could attribute to the shift of usage from C8 to C4-C6 PFASs after the voluntary phase-out of POSF and PFOA since 2000 [66] PFHxA levels (ng/L) in outgoing drinking water from two waterworks in Ronneby (SE): - Site 1 (Brantafors): 320 - Site 2 (Krragarden): 3.6 [102] Tap water NL (n=4): Average SD (Range): 2.3 0.10 (2.2-2.4) ng/L => Levels of PFAAs observed in four tap water samples closely resembled each other as demonstrated by the relatively low standard deviation of the mean values [18] Effluent waste water: PFHxA highest concentration in six effluents (mean 24 ng/L) => TOP (oxidation): in absolute numbers PFHxA and PFOA increased the most [21] PFHxA near air force base: 372 g/L [118] PFHxA concentration in water and sediment samples: - Water (unit: ng/L, n=47): * Range: 2.3 ~ 105 * Mean: 36.5 * Median: 20.6 * 75%: 59.7 * 95%: 101 * DF (%): 100.0 [73] * North Atlantic Ocean (NAO): - 2007 (n=39): <5.7-88 pg/L - 2008 (n=10): 38-54 pg/L - 2010 (n=13): 40-79 pg/L * Middle Atlantic Ocean (MAO): - 2007 (n=10): <5.7 pg/L - 2008 (n=5): 20-31 pg/L - 2010 (n=5): 33-38 pg/L * South Atlantic Ocean (SAO): - 2007 (n=10): <5.7 pg/L - 2008 (n=13): <3.0-26 pg/L - 2010 (n=16): <5.9 pg/L [117] In spring 2013, groundwater of a vast area of the Veneto Region (northeastern Italy) was found to be contaminated by perfluoralkyl substances (PFAS) from a PFAS manufacturing plant active since the late 1960s. Residents were exposed to high concentrations of PFAS, particularly perfluorooctanoic acid (PFOA), through drinking water until autumn 2013. [113] * Tap water (n=34): DF (%) = 100.0, Mean = 11.7, Max = 16.7 [20] EU WWTP effluents: PFHxA: - LOQ: 1 ng/L - Freq.: detected in 72 % of samples - Max. concentration: 23.8 g/L - Average concentration: 304 ng/L - Median concentration: 5.7 ng/L - 90th percentile (Per90): 18.2 ng/L [24] river water CN, maximum concentrations in * yearly monitoring: - 2011: 59.2 ng/L - 2012: 180 ng/L - 2013: 795 ng/L - 2014: 196 ng/L * seasonal monitoring: autumn (795 ng/L) > summer (603 ng/L) > winter (183 ng/L) > spring (168 ng/L) => peak river contamination periods: summer & autumn [118] - Sediments (unit: ng/g dw, n=47): * Range: nd ~ 0.25 * Mean: 0.07 * Median: 0.06 * 75%: 0.10 * 95%: 0.17 * DF (%): 78.7 [117] Surface water and groundwater in a large area of the Veneto Region in northeastern Italy was contaminated with PFAS that were also found in drinking water samples. A manufacturing plant located in the town of Trissino that produced PFAS since the late 1960s was identified as the only likely source of water contamination. [21] PFCAs were found in nearly equal concentrations in municipal (14-459 ng/L) and industrial (3-664 ng/L) wastewaters => origin: PFOA in fluorinated polymeric products + degradation of precursors used in household products [118] The predominant PFAS in water was PFPeA, followed by PFHxA. Shortchain PFAS such as PFPeA, PFHxA and LPFBS were detected more frequently in water than in the other environmental matrices. Short-chain PFAS and PFOA were detected during all four seasons. The seasonally dominant PFCAs were PFHxA and PFOA during summer and PFPeA and PFHxA during the other seasons. [117] Measurements of 152 drinking water samples collected in July and August 2013 indicated that the main contaminants were PFOA, PFBA and PFBS, followed by PFPeA, PFHxA, PFOS, PFHpA and PFHxS. The longer-chain PFAS congeners (PFNA, PFDA, PFUnA, PFDoA) were detected only in a minority of samples and at lower concentrations. [117] PFHxA concentrations (ng/L) in 152 samples of drinking water taken during July and August 2013, before the full implementation of granular activated carbon filters: - Min: <10.0 - 25th percentile: 39.0 - Median: 52.0 - 75th percentile: 79.5 - 95th percentile: 185.8 - Max: 330.0 - % samples LOQ: 89.5% [22] Mass of PFHxA annually exported to WWTPs from landfills (AU): - Min: 2.33 g - Mean: 17 g - Max: 62 g => PFASs are only partially removed from wastewater + potential for additional formation during treatment from degradation of precursors => WWTPs are considered major point sources of PFASs release into the aquatic environment [119] 3 replicate aqueous and solid samples were collected from each of 19 Australian WWTPs throughout 2017. Samples were taken from various stages within the treatment train from a range of WWTPs to determine the trends in the mass flux and partitioning of PFAS within the sampled WWTPs. Summary statistics for pooled aqueous (n=201, triplicates from 67 individual locations within 19 WWTPs) and pooled solid (n=51, triplicates from 5 primary and secondary sludge locations, 6 lagoon sludges and a lagoon dredge pile) samples. * PFHxA concentration in aqueous samples (ng/L): - Median: 16 - Mean: 21 [119] - s.d.: 17 - min: 1.4 - max: 92 [119] - Detect (%): 100% * PFHxA concentration in solid samples (ng/g dw): - Median: 0.92 - Mean: 1.9 - s.d.: 2.8 - min: <LOD - max: 13 - Detect (%): 82% In aqueous samples: between influent and final effluent, the compounds PFPeA, PFHxA, PFHpA, PFOA, PFNA, and PFDA (all of which are PFCAs) increased significantly. A number of transformation pathways with stable PFCA endproducts are known, this may explain some of the increase in PFCAs from influent to final effluent. [119] PFHxA had the highest mean concentration from pooled aqueous samples in secondary effluent, final effluent, and recycled water; increasing in concentration from influent to primary, secondary and final effluent and recycled water (11, 16, 28, 28 and 32 ng/L, [119] Estimated yearly mass discharged: PFHxA: 87 kg annually. [39] NOEC and LOECs for PFAAs for transactivation of mouse and human PPAR in transiently transfected COS-1 Cells: PFHxS: Mouse: - NOEC: 10 M - LOEC: 20 M - LOEC: 8.76 g/mL (p<0.05) [39] NOEC and LOECs for PFAAs for transactivation of mouse and human PPAR in transiently transfected COS-1 Cells: PFHxS: Human: - NOEC: 5 M - LOEC: 10 M - LOEC: 4.38 g/mL (p<0.001) [118] Spatial and seasonal distributions of PFAS in air, water, sediment, soil, and fish samples in the Asan Lake area of South Korea. Estimation of bioaccumulation potential of PFAS using sediment-water partition coefficients and bioaccumulation factors. L-PFHxS concentration in air samples: - Air (unit: pg/m3, n=4): Gaseous: * Range: nd ~ 1.77 * Mean: 0.76 * Median: 0.67 * DF (%): 75.0 [101] Concentrations of PFHxS in food basket samples (SE): - Dairy products: 1999: <MDL, 2005: <MDL, 2010: 1.0 pg/g - Meat products: 1999: 8.5 pg/g, 2005: 5.1 pg/g, 2010: 4.5 pg/g - Fats: 1999: <MDL, 2005: 0.9 pg/g, 2010: <MDL - Pastries: 1999: 1.2 pg/g, 2005: 1.3 pg/g, 2010: <MDL - Fish products: 1999: 21.7 pg/g, 2005: 8.8 pg/g, 2010: 9.2 pg/g - Egg: 1999: 39 pg/g, 2005: <MDL, 2010: 2.5 pg/g - Cereal products: 1999-2010: <MDL - Vegetables, including root vegetables: 1999: 1.2 pg/g, 2005: 1.0 pg/g, 2010: 1.2 [2] observation well NL max. value: 99 (89-107) ng/L [118] Particulate: * Range: 0.24 ~ 0.58 * Mean: 0.40 * Median: 0.38 * DF (%): 100.0 Total: * Range: 0.58 ~ 2.19 * Mean: 1.15 * Median: 0.92 * G (%): 65.6 [101] - Potatoes: 1999-2010: <MDL - Sugar and sweets: 1999: <MDL, 2005: 1.3 pg/g, 2010: 1.5 pg/g - Soft drinks, lemonade: 1999: 0.7 pg/g, 2005: <MDL, 2010: <MDL [5] LOD = 0.4 ng/L Freq = 34.8 % max = 19 ng/L Average = 1 ng/L med = 0 ng/L Per90 = 5 ng/L [12] SE Surface water: PFHxS/PFOS = 0.5-1.8 PFHxS levels were similar to PFOS [28] Rivers: - Detection frequency: 77% - Concentration range: 0.051-18 ng/L - Average: 9.0 ng/L - predominant PFAS in rivers: 18% of total PFASs [28] Recipient sea SE: - Detection frequency: 100% - Concentration range: 0.11-1.7 ng/L - Average: 0.91 ng/L [2] estimated concentration of PFAA in public supply well field NL: 13 ng/L [6] tap water MDL = 0.048 ng/L MLQ = 0.16 ng/L max. value SE: 2.50 ng/L [9] ex. river water treatment NL - raw/influent: 2.0 ng/L - finished/tap water: 0.6 ng/L - percent removal: 70% [8] concentrations in water CN: - Dissolved phase: mean: 1.24 ng/L contribution to total PFASs: 0.10% -Separate suspended particulate matter: mean: 0.41 ng/L contribution to total PFASs: 0.94% - Soil: mean: 0.80 ng/L contribution to total PFASs: 11.16% - Leaves: mean: 0.23 ng/L contribution to total PFASs: 0.27% - Bark: mean: 0.06 ng/L contribution to total PFASs: 0.16% [25] Measured conc. in surface water at sites where fire-fighting foams have been used or spilled: - Toronto Airport (CA): <LOD - 134000 ng/L - Air force base F18 (SE): <0.5 - 25.1 ng/L [32] Concentration of lin-PFHxS in canal water samples (26 datasets): - Assigned value: 24.9 ng/L - Average: 23.45 ng/L - Median: 24.85 ng/L - Min.: 0.00 ng/L - Max.: 43.00 ng/L - SD: 23.45 - % relative SD: 35 [73] PFHxS was quantified in 88% of all samples in the Greenland Sea, with concentrations ranging from <6.5 to 45 pg/L [12] SE Drinking water: PFSAs 29% of total PFASs [102] Tap water NL (n=4): Average SD (Range): 0.55 0.10 (0.5-0.6) ng/L => Levels of PFAAs observed in four tap water samples closely resembled each other as demonstrated by the relatively low standard deviation of the mean values [10] total discharge of PFASs into the aquatic environment ranged between 10 g d-1 and 10 000 g d-1, depending on the water usage in the community connected to the sewage treatment plant [13] - river water NL: detectable concentrations: min. 1.5 ng/L - max. 2.2 ng/L - drinking water NL: detectable concentrations: min. 0.02 ng/L - max. 0.43 ng/L [102] * Comparison tap water and hot water NL: - Drinking water treatment plant A: Tap water: <0.38 ng/L & Hot water: <0.38 ng/L - Drinking water treatment plant B: Tap water: 0.55 ng/L & Hot water: 0.74 ng/L => Concentrations of PFAAs detected in hot water were similar to those in the corresponding tap water * Post-mixed cola NL (n=6): Average: <0.63 => Ratio tap water/post-mixed cola (Ratio of averages): / [102] * Coffee NL: - Brewed coffee from coffee machines (n=12): Average SD (Range): Could not be quantified - Manually brewed coffee (n=4): Average SD (Range): Could not be quantified - Procedure blank brewed coffee (n=2): Average SD (Range): <0.38 ng/L [12] SE Groundwater: PFHxS: ranging from 18% for industrial areas to 37% for landfill/ waste desposal areas, no PFHxS detected in skiing areas (of total PFASs) => PFHxS/PFOS = 2.3-3.9 PFHxS levels were higher compared to PFOS => could be due to stronger sorption of PFOS to soil particles compared to PFHxS [18] groundwater at firefighting training site: sum of PFAS before TOP (oxidation): median: ~180 000 ng/L => major PFAS identified was PFHxS: median: 71 000 ng/L => oxidation raised PFCA+PFSA content with approx. 1/3 => ~60% of rise attributed to precursors initially measured, especially 6:2 FTS and PFHxSA [34] branched PFHxS occurred with 7% of total PFHxS in surface water [118] Spatial and seasonal distributions of PFAS in air, water, sediment, soil, and fish samples in the Asan Lake area of South Korea. Estimation of bioaccumulation potential of PFAS using sediment-water partition coefficients and bioaccumulation factors. [73] * North Atlantic Ocean (NAO): - 2007 (n=39): ND - 2008 (n=10): 13-27 pg/L - 2010 (n=13): <6.5-39 pg/L * Middle Atlantic Ocean (MAO): - 2007 (n=10): ND - 2008 (n=5): 8.1-14 pg/L - 2010 (n=5): <6.5-12 pg/L * South Atlantic Ocean (SAO): - 2007 (n=10): ND - 2008 (n=13): <4.1-17 pg/L - 2010 (n=16): <6.5 pg/L [118] L-PFHxS concentration in water and sediment samples: - Water (unit: ng/L, n=47): * Range: 2.9 ~ 24.7 * Mean: 11.4 * Median: 10.6 * 75%: 16.8 * 95%: 24.1 * DF (%): 100.0 [66] PFHxS levels (ng/L) in outgoing drinking water from two waterworks in Ronneby (SE): - Site 1 (Brantafors): 1700 - Site 2 (Krragarden): 4.6 [113] * Tap water (n=34): DF (%) = 100.0, Mean = 0.785, Max = 1.29 [17] SE - PFHxS was predominantly present in filtered effluent water with a mean value of 1.5 ng/L - Other PFSAs detected in sludge were PFHxS, ... - PFHxS had a net mass increase in all WWTPs with a mean value of 37% => presumably result of degradation of PFSAs precursor compounds during the WWTP process [117] In spring 2013, groundwater of a vast area of the Veneto Region (northeastern Italy) was found to be contaminated by perfluoralkyl substances (PFAS) from a PFAS manufacturing plant active since the late 1960s. Residents were exposed to high concentrations of PFAS, particularly perfluorooctanoic acid (PFOA), through drinking water until autumn 2013. [18] in WWTP effluent: PFHxS concentration up to 24 ng/L => oxidation achieved increases of 21% for WWTP effluents [113] 397 food samples collected from a market in Busan (Korea) in 2011 (n=227) and 2012 (n=170). Food samples of 66 different food types, classified into 7 food categories. 34 tap water samples were also collected from 16 districts in Busan. PFHxS concentrations in each food group (ng/g, beverage unit: ng/L): * Fish and shellfish (n=99): DF (%) = 4.0, Mean = 0.012, Max = 0.712 [25] Measured conc. in groundwater at sites where fire-fighting foams have been used or spilled: - Wurtsmith air force base (US): 9000 - 120000 ng/L - Tyndall air force base (US): 107000 - 920000 ng/L - Fallon Naval Air station (US): <LOD - 876000 ng/L - Military base (US): 36000- 360000 ng/L - Military base (US): 81 - 1700 ng/L [113] * Meat and its products (n=39): DF (%) = 0.0, Mean -, Max * Vegetables and fruit (n=78): DF (%) = 0.0, Mean -, Max * Processed products (n=90): DF (%) = 2.2, Mean = 0.007, Max = 0.338 * Dairy (n=37): DF (%) = 16.2, Mean = 0.029, Max = 0.331 * Beverage (n=21): DF (%) = 0.0, Mean -, Max * Others (n=33): DF (%) = 9.1, Mean = 0.028, Max = 0.347 PFHxS concentrations in bottled water and tap water samples (ng/L): [30] observation wells SE: median water conc. PFHxS: 690 ng/L [113] * Bottled water (n=8): DF (%) = 0.0, Mean -, Max PFHxS was detected in all of the tap water samples. [115] In September 2011, foods were purchased in 12 representative cities of Catalonia, all with more than 20,000 inhabitants. Food samples were obtained at each locality in 4 shops/stores of different size. Foods selected for PFAS analysis were among the most consumed in Catalonia. * Mean concentrations (in pg/g fw) of PFHxS [115] - Meat and meat products: 3.2 - Fish and seafood: 45 - Vegetables: 4.5 - Tubers: <1.9 - Fruits: <1.9 - Eggs: <2.0 - Milk: <2.6 - Dairy products: <1.1 - Cereals: <0.6 - Pulses: <1.3 - Oils: <0.7 - Industrial bakery: <0.56 [34] - Firefighting training sites: lin-PFHxS: 29% of total PFASs - Landfill/ waste disposal areas: lin-PFHxS: 37% of total PFASs - Urban areas: lin-PFHxS: 20% of total PFASs => branched PFHxS occured with 5% of total PFHxS in groundwater [115] * Fish and shellfish was the group in which more PFASs were detected and where the highest PFAS concentrations were found * Dietary intake of PFHxS by the population of Catalonia (in ng/kg body weight/day): - Children: ND = 0: 0.08; ND = 1/2 LOD: 0.17 - Boys adolescents: ND = 0: 0.02; ND = 1/2 LOD: 0.04 - Girls adolescents: ND = 0: 0.02; ND = 1/2 LOD: 0.04 - Male adults: ND = 0: 0.02; ND = 1/2 LOD: 0.04 - Female adults: ND = 0: 0.03; ND = 1/2 [115] - Female seniors: ND = 0: 0.02; ND = 1/2 LOD: 0.04 [115] * Concerning human health risks from dietary exposure to PFASs in Catalonia, it is important to note that for any of the age/gender groups of population estimated, the TDIs recommended by the EFSA were not exceeded. * PFOS was the compound found in the highest number of samples (33 out of 80). PFOA, PFHpA, PFHxS, PFDA and PFDS were the compounds that, concurrently with PFOS, were detected in the greatest number of food samples. [118] - Sediments (unit: ng/g dw, n=47): * Range: nd ~ 1.48 * Mean: 0.17 * Median: 0.00 * 75%: 0.25 * 95%: 0.95 * DF (%): 40.4 [118] The predominant PFAS in water was PFPeA, followed by PFHxA. Shortchain PFAS such as PFPeA, PFHxA and LPFBS were detected more frequently in water than in the other environmental matrices. Short-chain PFAS and PFOA were detected during all four seasons. The seasonally dominant PFCAs were PFHxA and PFOA during summer and PFPeA and PFHxA during the other seasons. [117] Surface water and groundwater in a large area of the Veneto Region in northeastern Italy was contaminated with PFAS that were also found in drinking water samples. A manufacturing plant located in the town of Trissino that produced PFAS since the late 1960s was identified as the only likely source of water contamination. [117] Measurements of 152 drinking water samples collected in July and August 2013 indicated that the main contaminants were PFOA, PFBA and PFBS, followed by PFPeA, PFHxA, PFOS, PFHpA and PFHxS. The longer-chain PFAS congeners (PFNA, PFDA, PFUnA, PFDoA) were detected only in a minority of samples and at lower concentrations. [117] PFHxS concentrations (ng/L) in 152 samples of drinking water taken during July and August 2013, before the full implementation of granular activated carbon filters: - Min: <10.0 - 25th percentile: <10.0 - Median: <10.0 [117] - 95th percentile: 26.0 - Max: 66.0 - % samples LOQ: 40.1% [20] EU WWTP effluents: PFHxS: - LOQ: 1 ng/L - Freq.: detected in 71 % of samples - Max. concentration: 922 ng/L - Average concentration: 48.6 ng/L - Median concentration: 3.4 ng/L - 90th percentile (Per90): 34.8 ng/L [22] Mass of PFHxS annually exported to WWTPs from landfills (AU): - Min: 0.81 g - Mean: 11 g - Max: 37 g => PFASs are only partially removed from wastewater + potential for additional formation during treatment from degradation of precursors => WWTPs are considered major point sources of PFASs release into the aquatic environment [119] 3 replicate aqueous and solid samples were collected from each of 19 Australian WWTPs throughout 2017. Samples were taken from various stages within the treatment train from a range of WWTPs to determine the trends in the mass flux and partitioning of PFAS within the sampled WWTPs. [119] Summary statistics for pooled aqueous (n=201, triplicates from 67 individual locations within 19 WWTPs) and pooled solid (n=51, triplicates from 5 primary and secondary sludge locations, 6 lagoon sludges and a lagoon dredge pile) samples. [119] * PFHxS concentration in aqueous samples (ng/L): - Median: 3.1 - Mean: 13 - s.d.: 31 - min: <LOD - max: 200 - Detect (%): 95% [119] * PFHxS concentration in solid samples (ng/g dw): - Median: <LOQ - Mean: 1.1 - s.d.: 2.8 - min: <LOD - max: 17 - Detect (%): 50% 223-393-2 3871-99-6 Potassium perfluorohexane-1sulphonate K-PFHxS C6 Pre-registered / / Perfluorohexane PFHxPA C6 / phosphonic acid 700-242-3 62037-80-3 Ammonium 2,3,3,3tetrafluoro-2(heptafluoropropoxy)prop anoate GenX/HFPO- C6 (5 DA fluorinated) Registered [75] logKaw: 2.13 (estimated by COSMOtherm) => similar Kaw as PFOA [75] logKow (dry): 4.24 (estimated by COSMOtherm) => lower Kow than PFOA [75] logKlipw: 4.40 (estimated with regression equation) [75] logKoa = 6.37 (estimated by COSMOtherm) 207-026-3 423-50-7 Perfluorohexanesulphony PHxSF C6 l fluoride Pre-registered 700-323-3 908020-52-0 Ammonium EEA difluoro[1,1,2,2- tetrafluoro-2- (pentafluoroethoxy)ethoxy ]acetate C6 (5 fluorinated) Registered [75] logKaw: 1.83 (estimated by COSMOtherm) => similar Kaw as PFOA [75] logKow (dry): 4.60 (estimated by COSMOtherm) => lower Kow than PFOA [75] logKlipw: 4.76 (estimated with regression equation) [75] logKoa = 6.43 (estimated by COSMOtherm) / / 691-551-1 423-46-1 436-710-6 756-13-8 206-798-9 375-85-9 4:2 fluorotelomer alcohol 4:2 FTOH C6 (4 / fluorinated) 2,2,3,3,4,4,5,5,6,6,6undecafluorohexan-1-ol 5:1 FTOH C6 (5 fluorinated) Not registered 1,1,1,2,2,4,5,5,5nonafluoro-4(trifluoromethyl )-3pentanone Perfluoroheptanoic acid Novec PFHpA C6 (5 fluorinated) Registered C7 Pre-registered [79] log Kaw: 1.26 (calculated from equations) [65] 4:2 FTOH: logKow = 4.05 [75] logKaw: 1.24 (estimated by COSMOtherm) => lower Kaw than 8:2 FTOH [79] log Kow: 0.41 (calculated from equations) [75] logKow (dry): 2.98 (estimated by COSMOtherm) => lower Kow than [75] logKaw: 3.92 (estimated by COSMOtherm) => higher Kaw than PFOS/PFOA [79] log Kaw: 2.00 (calculated from equations) [75] logKow (dry): 4.74 (estimated by COSMOtherm) => higher Kow than [41] PFHpA ChemBioOffice: logKow = 5.71 [75] logKlipw: 3.12 (estimated with regression equation) [75] logKlipw: 4.90 (estimated with regression equation) [65] 4:2 FTOH: logAqS (aqueous solubility): - Y-Exp: 2.99 mg/L - Y-Pred: 1.94 mg/L - at 24 C: 2.99 mg/L - EPI Suite Pred: 1.68 mg/L [65] 4:2 FTOH: logpL (subcooled vapor pressure): - Y-Exp: 0.21 mm Hg - Y-Pred: 0.32 mm Hg - EPI Suite Pred: 1.00 mm Hg - Y Exp: 2.33 Pa - Y-Pred: 2.44 Pa - Exp: 2.33 Pa [79] log Koa: 4.73 (calculated from equations) [79] Water solubility (SL): log (SL/mol L^-1): -1.99 (calculated from equations) [79] Vapor pressure (PL): log (PL/Pa): 2.02 (calculated from equations) [75] logKoa = 4.22 (estimated by COSMOtherm) [65] PFHpA: - logCMC Exp: -1.54 - logCMC Pred: -1.63 - 3: 9.62 [65] PFHpA: logAqS (aqueous solubility): - Y-Pred: 0.82 mg/L - at 24 C: 5.07 mg/L - EPI Suite Pred: -0.45 mg/L [65] PFHpA: logpL (subcooled vapor pressure): - Y-Exp: -0.80 mm Hg - Y-Pred: -0.53 mm Hg - EPI Suite Pred: 0.05 mm Hg - Y Exp: 1.32 Pa - Y-Pred: 1.59 Pa - Pred: 1.66 Pa - ref40: 2.11 Pa [75] logKoa = 0.82 (estimated by COSMOtherm) [79] log Koa: 5.19 (calculated from equations) [65] PFHpA: EPI pred logKow: 5.33 [79] Water solubility (SL): log (SL/mol L^-1): -3.01 (calculated from equations) [79] Vapor pressure (PL): log (PL/Pa): 1.74 (calculated from equations) [79] log Kow: 1.19 (calculated from equations) [94] PFHpA: logKow = 5.33 [80] Study designed to determine if PFAAs in mixtures exhibit additivity in activating PPAR: - PFAAs were tested individually and in binary combinations of PFOA + either PFNA, PFHxA, PFOS or PFHxS in an 8x8 factorial grid design and the responses of mouse PPAR in transiently transfected COS-1 cells were evaluated - Binary combinations of PFAAs (PFOA + PFNA, PFOA + PFOS, PFOA + PFHxA or PFOA + PFHxS) appear to behave additively in the lower PFOA concentration range of 1-32 M => at higher PFOA concentrations with PFAAs, the observed responses exceeded the predicted responses using either RA or CA (model equations) [81] - no sign. cytotoxicity was observed for the shortest chain length PFCs (PFBA, PFHxA, PFBS and PFHxS) after 24h incubation with the human placental choriocarcinoma cell line JEG-3 - differential bioavailability of PFCs in the in-vitro system was assessed by measuring the fraction retained in the cells after exposure: the concentration of PFBA, PFBS and PFHxS was below dection limit under our assay conditions - inhibition of P450 aromatase (CYP19) activity in JEG-3 cells: 84% inhibition by PFHxS when tested at 500 M, IC50: 298 +/- 29 M => sulfonates (PFBS, PFHxS) were stronger inhibitors of aromatase activity than the corresponding acidic compounds (PFBA, PFHxA) - PFBS and PFHxS sign. inhibited CYP19 aromatase activity despite the fact that the measured uptake of the compounds by cells was below detection limit => Therefore, both PFBS and PFHxS may exert the inhibitory effect on CYP19 aromatase activity at rather low endogenous cellular concentrations [82] Study of inhibition of cellular viability in human liver cell line (HL-7702): * PFHxS single: - IC-20 (20% stimulatory effect concentration): 1.65 x 10^(-5) mol/L - IC0 (0% inhibitory concentration): 1.62 x 10^(-3) mol/L - IC10 (10% inhibitory concentration): 1.71 x 10^(-3) mol/L - IC50 (50% inhibitory concentration): 2.14 x 10^(-3) mol/L * PFHxA-PFHxS mixture: - all IC0, IC10 and IC50 values in the binary mixture were less than that of individual PFSA or PFCA values - a synergistic action took place under the effective concentrations of IC0, IC10 and IC50 with various concentration ratios - PFHxA/PFHxS showed the strongest synergistic effect with MTI (mixture toxic index) values from 2.30 to 8.08 for five rays - for the effective concentration of IC-20, only the higher proportion of PFSA in the mixtures (PFHxS > 48.9%) showed a synergistic effect for IC-20 - the mixture of PFHxS and PFHxA showed the highest synergistic effect, especially at the high proportion of PFHxS (70%) and indicated that PFHxS had the highest contribution to the synergistic effect [82] - Mixtures of eleven PFAAs (Mix0, Mix10 and Mix50, PFHxS included) presented a partial addition effect - Mixtures of nine PFAAs (Mix0, Mix10, Mix50, PFHxS included) with J-shaped curves only presented a synergistic effect with MTI values ranging from 1.16 to 1.25, however Mix-20 showed partial addition (MTI = 0.86) - A mixture of seven PFAAs (Mix-20, PFHxS included) showed a synergistic effect (MTI = 1.26) [83] * Estrogen (ER) transactivation (PFHxS tested alone): - LOEC: 2 x 10^(-5) M, MOEC: 9 x 10^(-5) M, EC50: 5.8 x 10^(-5) M - Cytotoxicity: 1 x 10^(-4) M => upon co-exposure with 25 pM E2 (17-estradiol) PFHxS sign. further enhanced the E2-induced ER response => PFHxS has the ability to act as an ER agonist * Androgenic (AR) transactivation (PFHxS tested alone): - LOEC: 1 x 10^(-4) M, MOEC: 1 x 10^(-4) M, no cytotoxicity => upon co-treatment with 25 pM DHT PFHxS elicited sign. concentration-dependent antagonistic effects on DHT-induced AR transactivity => PFHxS has in vitro potency to antagonize AR transactivity in a concentration-dependent manner * Aromatase activity (PFHxS tested alone): - an incipient toxicity of PFHxS was observed at 1 x 10^(-4) M [83] * Estrogen (ER) transactivation (PFHxS tested in mixture): - LOEC: 3.5 x 10^(-5) M, MOEC: 3.5 x 10^(-5) M, EC50: ND - Cytotoxicity: 1.4 x 10^(-4) M => PFAA mixture sign. induced ER transactivity to an effect level of 125% relative to the solvent control => upon co-exposure with 25 pM E2 (17-estradiol) the mixture sign. enhanced the E2-induced ER response * Androgenic (AR) transactivation (PFHxS tested in mixture): - Cytotoxicity: 2.1 x 10^(-4) M => the PFAA mixture alone did not elicit any effect on AR transactivity, however upon co-exposure with 25 pM DHT the mixture sign. antagon-ized the DHT-induced AR transactivity with an IC50 of 6.8 x 10^(-6) M => at the 90 and 85% inhibition levels (IC90 and IC85, respectively), the predicted effect concentrations were found outside the 95% confidence band for the observed effects, and were higher than observed effect concentrations, indicating an effect more than additive (synergistic) of the mixture [84] PFCs can compete with thyroxine (T4, the transport form of thyroid hormone), for binding to the human thyroid hormone transport protein transthyretin (TTR) => such competitive capacity may lead to decreased thyroid hormone levels as previously reported for animals exposed to PFCs => PFCs do not affect the regulatory functions of the thyroid hormone system itself, but it is the competitive binding to transport proteins that alters the free thyroxine (T4) levels in blood => TTR is the main T4 carrier in cerebrospinal fluid, and also important in serum of most mammalian species and birds => Binding potency is clearly associated with the degree of fluorination of the alkyl chain => PFAAs with a carbon chain length longer than eight have low TTR binding potencies => TTR binding potencies were significantly higher for compounds containing a sulfonate functional group than for those containing a carboxylic acid functional group => TTR is both in humans and in rodents the most important carrier protein for thyroid hormone to the developing fetus and the brain [84] PFHxS: - Molecular weight: 400.0 g/mol - T4-TTR binding at maximum concentration (10M): 3% - IC50 (concentration at 50% inhibition): 717 nM - Slope of dose-response curve: -1.40 - T4-REP (relative potency compared to T4) factor: 0.085 - HPLC retention time: 32.3 min - Purity: 98 [85] Mean fold change standard error of mRNA expression after exposure to 100 M of PFHxS (in rat H4IIE hepatoma cells): * Thyroid related genes: - Hex: 25.17 0.22 - PAX 8: 11.98 3.74 => upregulation of thyroid related genes * Cholesterol related genes: - Per-3-Keto-: 24.23 6.50 - SQSYN: 14.81 2.37 => upregulation of cholesterol related genes * Lipoprotein related genes: - Mito-3-Keto-: 0.80 0.08 => downregulation of Mito-3-Keto- - Mito-3-Keto-: 3.05 0.34 => upregulation of Mito-3-Keto- - ApoA4: 0.67 0.13 [86] - weak binding of PFHxS with the protein TR-LBD (TR = thyroid hormone receptor, LBD = ligand-binding domain) * Binding of PFHxS with TR-LBD: - Length: 10.11 A - IC50: 193 62 M - RP (relative potency (IC50 T3 / IC50 chemical)): 0.0015 - Hydrogen bonding: ARG 228 * all tested PFCs fit into the T3-binding pocket of TR-LBD, with the acid or hydroxyl end group residing toward the inner part and the hydrophobic chain toward the entrance of the binding pocket => all PFCs with an acid end group formed a hydrogen bond with [87] * Binding potency of PFHxS to TTR (TTR competitive binding assay): - IC50: 594 63 nM - Kd: 95 10 nM - RP (relative potency (IC50 T4 / IC50 chemical)): 0.053 => perfluoroalkyl acids: Kd values decreased as carbon chain length increased from C4 to C8 => no further increase, but a downward trend was observed for perfluoroalkyl acids with longer chain lengths (C9-C14) * Binding potency of PFHxS to TTRmutK15G: - IC50: ND - Kd: ND - Hydrogen bonding: Lys15 => PFASs exhibited much weaker binding affinities to TTRmutK15G compared with wild-type TTR * Docking study: - Similar to T4, PFOA and PFOS could nearly fill the TTR ligandbinding pocket - Perfluoroalkyl acids with carbon chain length less than C8 did not adequately fill the T4 binding pocket [87] * Binding potency of PFHxS to TBG (TBG competitive binding assay): - IC50: ND - Kd: ND - RP (relative potency (IC50 T4 / IC50 chemical)): ND => only PFTA and PFTdA bound to TBG * Binding potency of PFHxS to TBGmutR378G: - IC50: ND - Kd: ND * Binding potency of PFHxS to TBGmutR381G: - IC50: ND - Kd: ND - Hydrogen bonding: Arg381 => both TBGmut378G and TBGmutR381G exhibited much weaker binding potencies to PFTA and PFTdA when compared with that of wild-type TBG * Docking study: - Perfluoroalkyl acids with carbon chain length less than C12 did not adequately fill the T4 binding pocket - Longer fluorinated carbon chain structures could nearly fill the TBG [98] Beijing prenatal exposure study (human): Correlations between maternal PFHxS and maternal thyroid hormones (adjusted for influential covariates): Not significant for PFHxS [109] * Occupants of all ages can be exposed to indoor contaminants, but preschool-aged children have the greatest risk due to hand to mouth activity and the amount of time they spend playing on or near the floor [111] PFAAs were investigated in home produced eggs and commercially produced eggs surrounding a fluorochemical industrial park in China: PFHxS concentrations: * Home produced eggs (n=4, sites 1-4): Egg yolks: - 2 km distance: <0.02 ng/g - 5 km distance: <0.02 ng/g - 10 km distance: <0.02 ng/g - 20 km distance: <0.02 ng/g Egg whites: - 2 km distance: <0.02 ng/g - 5 km distance: <0.02 ng/g - 10 km distance: <0.02 ng/g - 20 km distance: <0.02 ng/g Whole eggs: - 2 km distance: <0.02 ng/g - 5 km distance: <0.02 ng/g - 10 km distance: <0.02 ng/g - 20 km distance: <0.02 ng/g [111] * Commercially produced eggs (n=12, sites 5-16): Egg yolks: - Min: <0.02 ng/g - Max: <0.02 ng/g - Mean: <0.02 ng/g - Median: <0.02 ng/g - n > LOD (%): 0 (0) Egg whites: - Min: <0.02 ng/g - Max: <0.02 ng/g - Mean: <0.02 ng/g - Median: <0.02 ng/g - n > LOD (%): 0 (0) Whole eggs: - Min: <0.02 ng/g - Max: <0.02 ng/g - Mean: <0.02 ng/g - Median: <0.02 ng/g - n > LOD (%): 0 (0) [116] In May 2009, indoor dust samples (n=17) were collected from fifteen homes located in nine Korean cities for house dust (n=15), and one shopping mall and one library in Seoul for non-residential indoor dust (n=2). Nine cities for house dust samples were chosen to represent densely populated cities, industrialized cities, and rural area. * Concentrations (ng/g) of PFHxS measured in indoor dust samples from Korea (n=17): - House dust (n=15): Range: <LOD - 1.9 Mean ( SD): 0.2 ( 0.5) [116] Median: 0.0 % of < LOD: 73 - Non-residential indoor dust (n=2): Shopping mall: 0.9 Library: 0.0 * There were no significant correlations between PFCAs (or PFSAs) and their precursors in house dust. * In Korean house dust samples, significant correlations (p<0.05) were only found among all the FTOHs and among almost all of the PFCAs. * PFHxS was rarely detected in house dust samples (73% of samples were < LOD). [75] pKa = 0.06 (estimated by SPARC) [75] logKoc: 1.05 [10] - volatile PFASs (FTOHs, FASAs, FASEs) enter the atmosphere, where they can degrade, form intermediates during atmospheric oxidation, or transform into more persistent PFASs, such as PFSAs and PFCAs, which may finally end up in the aquatic environment - PFASs have the potential to affect cell membranes of algae => can alter nutritious quality of biofilms => potential effects on physiological fitness of [69] PFPAs: resistant to heat, oxidants, bases and aerobic biodegradation in surface water => likely highly P [69] PFPAs: - substances bind to proteins and blood cells: elimination half-lives increase with increasing perfluoroalkyl chain length - certain homologs (C8) have long elimination half-life from rainbow trout and rat blood, similar to PFOA and PFOS => some homologs may be B [69] PFPAs: - mixtures of PFPAs and PFPiAs both have modes-ofaction that are similar to, or different from, those of PFCAs and PFSAs - current information is not conclusive => T undetermined [14] no hydrolysis and biodegradibility of GenX was observed in tests according to OECD 111 & 301B [23] Resistant to photolysis, hydroxyl (OH)-radical-mediated reactions, hydrolysis, and biodegradation => known PFECAs and PFESAs are likely to be highly persistent in the environment and not easily metabolized in biota [23] - OECD 111: Hydrolysis: 0% after 5 days - OECD 301B: Biodegradation: 0% after 28 days [75] - half-life in air with AOPWIN: 740.54 hour - half-life in water with BIOWIN3: 17280 hour - half-life in soil with BIOWIN3: 17280 hour [63] Perfluoroalkyl and perfluoroether moieties are highly persistent under natural conditions => Even though some PFASs may partially degrade in the environment and biota, they will all ultimately transform into highly stable end products, wich are usually perfluoroalkyl or perfluoroalkyl(poly)ether acids such [75] overall persistence (Pov): 1038.74 days => same Pov as PFOA [3] protein binding affinity appears to increase with chainlength => short chain: faster elimination and lower distribution to liver [23] - toxicological profiles of GenX and ADONA: acute toxicity, eye and skin irritation, dermal sensitization, genotoxicity, repeated-dose toxicity, and developmental toxicity - ADONA and GenX were shown to cause liver damage in (at least) male rats or mice under repeated low exposure (<= 10 mg/kg/day) - producer suggested to classify GenX as STOT Rep. Exp. 2 H373 => would fulfill toxicity criteria in Annex XIII of REACH [13] - a laboratory study determined the BCF of GenX in common carp and estimated it to be <30, however, dietary accumulation was not studied - animal studies have shown a lower bioaccumulation potential of GenX compared to PFOA and GenX is not metabolized [76] binding affinity of HFPO-DA to hL-FABP (human liver fatty acid binding protein): - IC50: 4.76 +/- 0.31 M - Kd: 15.36 +/- 1.09 M - R^2: 0.997 => Kd value of HFPO-DA higher than that of PFOA and PFOS [23] only serum elimination half-lives of two PFECAs in mammals (in rats and mice for GenX) have been reported => they might be as bioaccumulative as the predecessors due to similar physicochemical properties, but this is uncertain [76] Ebinding values and number of hydrogen bonds (Nhydrogen bond) between hL-FABP and HFPO-DA: - Ebinding: -35.2625 kJ/Mol - Nhydrogen bond: 4 - Hydrogen bond: O13-X:R122:HE/O14-X:R122:HH21/H22X:S124:OG/H23-X:S39:OG => except for 6:2 FTS, all PFASs fit well at the binding pocket, indicating direct interactions between these chemicals and hL-FABP [63] the high solubility and proteinbinding characteristics of ionic PFAAs challenge the conventional assessment of bioaccumulation potential that is through either bioconcentration factor in aquatic species (usually fish) or models based on octanol-water partition coefficients (Kow) => even when PFAAs are not characterized as bioaccumulative under the current regulatory frameworks, the accumulation of these PFAAs in the environment will lead to increasing external exposure [23] Known PFECAs and PFESAs: - P - maybe B - likely T - LRTP [23] on-going (increasing) emissions and high persistence => short-chain PFAAs, PFECAs and PFESAs accumulate in the environment => increasing exposure to these substances over time => even if emissions cease in the future it will take decades to centuries to reverse global contamination of short-chain PFAAs, PFECAs and PFESAs, due to high persistence and environmental mobility [63] the very high persistence of PFAAs leads to poorly reversible exposure to these substances in the global environment and some local/regional environments including groundwater [75] pKa = 0.40 (estimated by SPARC) [75] logKoc: 1.41 [79] - Molar mass (MM): 264.1 g/mol - Molar volume (VM): 174.3 cm^3/mol (by ACD/Labs' ACD/PhysChem Suite) - Total surface area (TSA): 281.0 A^2 (by ChemAxon) - Tm (experimental melting point): 230 K (data obtained from [23] - Photolysis: 0% with UV-B/ visible light - OH* reaction: 1.90-4.27% - Hydrolysis: (almost) 0% [75] - half-life in air with AOPWIN: 740.54 hour - half-life in water with BIOWIN3: 5760 hour - half-life in soil with BIOWIN3: 5760 hour [23] Resistant to photolysis, hydroxyl (OH)-radical-mediated reactions, hydrolysis, and biodegradation => known PFECAs and PFESAs are likely to be highly persistent in the environment and not easily metabolized in biota [63] Perfluoroalkyl and perfluoroether moieties are highly persistent under natural conditions => Even though some PFASs may partially degrade in the environment and biota, they will all ultimately transform into highly stable end products, wich are usually perfluoroalkyl or perfluoroalkyl(poly)ether acids such as PFCAs, PFSAs, PFECAs and PFESAs [75] overall persistence (Pov): 346.25 days => lower Pov than PFOA [75] pKa = 11.86 (estimated by SPARC) [75] pKa = neutral (estimated by SPARC) [22] - increase in the mobility of PFASs with increasing pH => result of repulsive electrostatic interactions with the anionic functional group head and sediment - significant relationships with pH (p<0.05) [118] PFAS species with high Kd or Koc values were strongly adsorbed by sediments and organic matter, and compounds with very low Kd or Koc values were highly mobile with respect to the sediments [118] (i.e. weak adsorption). [79] - Molar mass (MM): 364.1 g/mol - Molar volume (VM): 209.8 cm^3/mol (by ACD/Labs' ACD/PhysChem Suite) - Total surface area (TSA): 331.9 A^2 (by ChemAxon) - Tm (experimental melting point): 303 K (data obtained from ChemSpider) [17] generally logKd increased with chain length, as true for C>6 [22] - PFASs are known to sorb to particulate matter => sorption increases as carbon chain length increases - significant relationships with TOC (p<0.05) => unsurprising given hydrophobicity of perfluorinated chain and potential for hydrophobic partitioning with organic matter [75] - half-life in air with AOPWIN: 2026.75 hour - half-life in water with BIOWIN3: 5760 hour - half-life in soil with BIOWIN3: [75] overall persistence (Pov): 273.13 days => lower Pov than 8:2 FTOH [75] - half-life in water with BIOWIN3: 17280 hour - half-life in soil with BIOWIN3: 17280 hour [75] overall persistence (Pov): 135.59 days => lower Pov than PFOA/PFOS [10] - volatile PFASs (FTOHs, FASAs, FASEs) enter the atmosphere, where they can degrade, form intermediates during atmospheric oxidation, or transform into more persistent PFASs, such as PFSAs and PFCAs, which may finally end up in the aquatic environment - PFASs have the potential to affect cell membranes of algae => can alter nutritious quality of biofilms => potential effects on physiological fitness of invertebrates [21] PFCAs are not degradable and will only be slowly removed from the oceans by vertical mixing into the deeper oceans => Arctic Ocean will have a very slow response time to changing emissions due to its long water residence time [63] Perfluoroalkyl and perfluoroether moieties are highly persistent under natural conditions => Even though some PFASs may partially degrade in the environment and biota, they will all ultimately transform into highly stable end products, wich are usually perfluoroalkyl or perfluoroalkyl(poly)ether acids such as PFCAs, PFSAs, PFECAs and PFESAs [17] - the fate of PFASs sorbed to sludge is generally to be further spread in the environment through application of sludge or sludge-containing products - in 2014: 24% of sludge was applied on agricultural soil, 24% was used to cover mines and dump sites, and 29% was used to manufacture soil intended for less sensitive land use [21] - the atmosphere could have faster responses to changing emissions (precursor concentrations are relatively quickly removed by oxidation processes) - relatively fast response times to changing emissions can also be expected for PFCAs in water bodies with short residence times (rivers and coastal waters) [118] As a result, PFHxA, PFHpA, and LPFBS (with lower Kd and Koc values) could be readily transported from sediments to water in comparison with other PFAS compounds with relatively high log Kd and Koc values. [28] Linear regression showed significant positive correlations of a.o. PFHpA to DOC (p<0.05) => PFASs may bind to and be co-transported by DOC in water => PFCAs and PFSAs are negatively charged in natural waters by proton dissociation of the acid => Short-chains (more hydrophilic) could readily bind to positively charged ions that are complex bound to DOC (such as Ca2+), while longerchained PFASs rather partition to even more hydrophobic phases in the water, such as the organic carbon fraction of suspended particulate matter [20] discharge of municipal wastewaters is one of the principal routes of entry of PFASs in the aquatic environment => often PFAS concentrations increase in WWTPs as a result of biodegradation of precursors during the activated sludge process [22] - at the end of their lifetimes, products containing PFASs may be disposed of to landfill - as a result of decomposition processes within landfills, PFASs associated with waste can become mobile and leach into water that enters a landfill to produce potentially contaminated leachate - small and very small sites may pose a local point source of PFAS contamination into surrounding groundwater (new and large landfills typically use cell liners to prevent leachate migration into groundwater) [23] only serum elimination half-lives of two PFECAs in mammals have been reported => they might be as bioaccumulative as the predecessors due to similar physicochemical properties, but this is uncertain [23] three PFECAs have been shown to cause liver damage in rats in repeated- dose tests; one PFECA is suggested by its manufacturer to be classified as T under the REACH regulation => other known PFECAs may have the same mode-of-action and may fulfill the toxicity criteria under REACH [63] the high solubility and proteinbinding characteristics of ionic PFAAs challenge the conventional assessment of bioaccumulation potential that is through either bioconcentration factor in aquatic species (usually fish) or models based on octanol-water partition coefficients (Kow) => even when PFAAs are not characterized as bioaccumulative under the current regulatory frameworks, the accumulation of these PFAAs in the environment will lead to increasing external exposure [50] human ovarian adenocarcinoma cell line: 4:2 FTOH led to increased estrogen receptor (ER) activity at 10 M (p<0.01) and 100 M (p<0.01) => short-chain PFCAs generally activate both PPAR and PPAR with similar potency and efficacy as long-chain PFCAs [23] Known PFECAs and PFESAs: - P - maybe B - likely T - LRTP [23] on-going (increasing) emissions and high persistence => short-chain PFAAs, PFECAs and PFESAs accumulate in the environment => increasing exposure to these substances over time => even if emissions cease in the future it will take decades to centuries to reverse global contamination of short-chain PFAAs, PFECAs and PFESAs, due to high persistence and environmental mobility [63] the very high persistence of PFAAs leads to poorly reversible exposure to these substances in the global environment and some local/regional environments including groundwater [63] the high solubility and proteinbinding characteristics of ionic PFAAs challenge the conventional assessment of bioaccumulation potential that is through either bioconcentration factor in aquatic species (usually fish) or models based on octanol-water partition coefficients (Kow) => even when PFAAs are not characterized as bioaccumulative under the current regulatory frameworks, the accumulation of these PFAAs in the environment will lead to increasing external exposure [38] PFHpA: Repeated study on aging individuals (SE): associations between change in plasma PFAS concentrations and markers of liver function (95% CI) - Biluribin: -0.93 (-1.14,-0.73) mol/L (p=7.50E-20) - ALT (Alanine aminotransferase): 0.02 (0.01, 0.02) kat/L (p=4.10E04) - ALP (Alkaline phosphatase): 0.06 (0.04, 0.09) kat/L (p=2.10E-07) => Changes in plasma concentrations of PFHpA were positively associated with the changes in ALT and ALP activity and inversely associated with changes in circulatory bilirubin [37] * Frog embryo teratogenicity assay-Xenopus (FETAX): LC50 (concentration at which 50% of the embryos were killed) = 942.4 M, EC50 (concentration at which 50% of the embryos were malformed) = 754.9 M, MCIG (minimum concentration that inhibits growth) = 297.4 M, TI (teratogenic index) = 1.25 => TI was >= 1.2 => compound considered toxic * Exposure to PFHxA or PFHpA greatly increased the mortality and general malformation rates in the developing embryos * PFC exposure induced certain embryonic abnormalities: pericardial edema, dorsal fin blisters, a curved body axis, lack of eye pigment, disturbed facial cells => PFHpA induced severe and diverse malformations [38] => Increased ALT activity is clinically used as a marker of liver dysfunction => Elevation of ALP can, in combination with other markers, be indicative of liver dysfunction => Decreased levels, or highly increased, bilirubin has been associated with adverse health effects, including liver dysfunction, cardiovascular disease, diabetes and the metabolic syndrome => These findings suggest a relationship between low-dose background PFAS exposure and altered liver function in the general population [37] * Whole body length was reduced by 24% in tadpoles treated with 1000 M of PFHpA (p<0.05) * Livers of embryos treated with PFHpA developed abnormally and were unusually large * Hearts from embryos treated with PFHxA and PFHpA showed enlarged atria and a loss of the atrial septum * Hearts from embryos exposed to either PFHxA or PFHpA were smaller and contained no trabeculae + pericardiac cavity was expanded and atrial and ventricular walls were thinner than those observed in the control => Effect more prominant in embryos exposed to PFHpA which reduced trabeculation of the ventricule as well as the thickness of the atrial and ventricular walls [39] Relative response of mouse and human PPAR to PFAAs in transiently transfected COS-1 cells, measured by C20max: PFHpA: - Mouse: 11 M (ranked most active to least active: 5/13 PFASs) - Human: 15 M (ranked most active to least active: 3/13 PFASs) => A lower C20 max value indicates a stronger induction of PPAR alpha activity by that PFAA => PPAR is activated by PFAAs of 4-12 carbons in chain length => Increasing activity of PPAR with increasing chain length of the PFAA up to C9, and lower activity with longer chain PFAAs (C>9) with both mouse and human PPAR => PPAR has been identified as a key player in the mode of action for PFAA-induced toxicity [37] * mRNA expression of xPTB (liver-specific marker) was reduced in PFHpA-treated embryos as compared to the control (p<0.05) * mRNA expression of NKX2.5 (heart-specific marker) was reduced in PFHpA-treated embryos as compared to the control (p<0.05) * PFHpA strongly induced ERK and JNK phosphorylation while PPAR gamma expression remained the same => Activated signaling that caused liver toxicity, known as hepatocarcinogenesis => Observed developmental defects in the liver were associated with the PFHpA-mediated activation of ERK/JNK pathways => PFHpA may cause developmental toxicity and teratogenicity, inducing hepatic and cardiac disorders [41] Progesterone content in mLTC-1 (mouse Leydig tumor cells) decreased with the increase in PFHpA treatment concentration => This relationship was illustrated by a standardized S-shape curve (dose-response curve) => Predicted IC50 (half-maximal inhibitory effect concentration): 47.75 mol/L => A similar dose-response curve was seen for cell viability (decreasing cell viability with higher PFHpA treatment concentration), but progesterone content was more sensitive than cell viability to PFHpA exposure [103] Analysis of 369 pairs of maternal and umbilical cord serum in a prospective birth cohort in Shandong, China: PFHpA: * Maternal serum: - Detection rate: 85.1 % - Geometric mean: 0.06 - Range: <LOD - 0.25 * Cord serum: - Detection rate: 97.3 % - Geometric mean: 0.09 - Range: <LOD - 1.17 * Correlation between maternal and cord levels of PFHpA (stat. sign.): 0.644 (p=0.000) [41] Sign. difference in mitochondrial membrane potential (MMP) between: 50 mol/L PFHpA and control (p<0.05) + 100 mol/L PFHpA and control (p<0.05) + 200 mol/L PFHpA and control (p<0.05) => Inhibition effect of PFHpA on progesterone production might be due, in part, to decrease in MMP in mLTC-1 [103] * Transplacental transfer efficiency: = Individual PFAS concentration in cord serum (ng/mL) / corresponding PFAS concentration in maternal serum (ng/mL): Range: 0.32 - 18.56 => Short-chain PFASs were transferred more efficiently from maternal serum to cord serum * Strong correlations between maternal and cord levels suggest that PFASs could readily pass through the placenta [44] All tested PFCs induced hPXR reporter activites in a dosedependent manner => In general, longer carbon chain length seems to be favorable for the hPXR activity of PFAAs with chain length shorter than 10, as chain length is longer than 10, increasing the chain length of PFAA seems to render the chemicals weaker potency to induce hPXR activity => PFHpA: EC50 = 18.81 M, maximum induction = 30.9 % ==> PXR induction is postulated to be associated with lipid homeostasis, atherosclerosis, carcinogenesis and endocrinedisrupting effects [45] Study on 141 pregnant woman serum samples: - Platelet: -0.167 (p<0.05) (negatively correlated with PFHpA) - Total protein: 0.226 (p<0.01) (positively correlated with PFHpA) - Albumin: 0.324 (p<0.01) (positively correlated with PFHpA) => PFASs may bind to serum albumin which might be related to anti-inflammatory effects of proliferator-activated receptor alpha (PPAR) activation - Total bilirubin: 0.257 (p<0.01) (positively correlated with PFHpA) => PFCAs could be related to adverse liver or bile effects in pregnant women => Results indicate that exposure to PFAAs at high levels may result in health risks to pregnant women or even to the fetus [46] PFASs can compete with thyroxine (T4) for binding to the human thyroid hormone transport protein transthyretin (TTR), which may lead to reduced thyroid hormone levels leading to endocrine disrupting adverse effects => PFHpA: T4-TTR Binding (%): 7 => Experimental Classification: Active (Y) => Predicted Classification: Active (Y) => Experimental pIC50 (conc. with 50% inhibition of T4): 2.81 mM => Predicted pIC50: 2.62 mM => Binding energy: -2.83 kcal/mol => PFASs can compete with thyroxine (T4) for binding to the human thyroid hormone transport protein transthyretin (TTR), which may lead to reduced thyroid hormone levels leading to endocrine disrupting adverse effects [46] Discriminating properties between active (Y) and inactive (N) classes: - Hydrophobicity of the H atom attached to heteroatom - Contributions from the H-050 descriptor PFASs containing acid functional groups: - If the carbon chain length is between 6 and 10, PFASs will be highly toxic or active ones - Toxicity will be lower or non-existent for PFASs containing a C chain length greater than 10 or below 6 => PFHpA IS TOXIC/ACTIVE [104] Breast milk samples taken from 264 Korean lactating women: * Level of PFHpA in breast milk: - n > LOD: 178 => detected in 67.4% of all breast milk samples - Median (Interquartile range): 0.028 (0.014-0.048) - Geometric mean: 0.030 * PFHpA was significantly correlated with PFOA (0.77458, p<0.001) * PFHxA, PFHpA and PFOA showed high correlation coefficients, implying that they may share similar sources of exposure [104] * PFHpA was sign. associated with the use of cosmetics (foundation, powder, etc.) (0.01075, p=0.027) and the use of non-stick coated frying pan (0.00821, p<0.001) => Accumulating body of evidence that PFASs are present in skin care products => The positive association between PFHpA and the use of nonstick coated frying pan may reflect current replacement of PFOArelated compounds with short-carbon chain PFCAs including PFHpA [105] Analysis of breast milk samples (n=293) collected from 127 mothers in Korea. Pairs of pregnant women and fetuses that were recruited from four South Korean cities. Detection rates and concentrations of PFHpA in Korean breast milk samples: * No. detected: 77 => detected in 26% of all breast milk samples * Range (min-max): <10 - 117 ng/L * Average SD: 6.25 12.9 ng/L => Study showed low detection rates of contaminant short-chain PFCAs [106] Study investigated the relationship between serum concentrations of perfluoroalkyl compounds and growth parameters in 2-year-old Korean children. Study included 361 children aged 2 years (192 boys and 169 girls; 22-27 months). PFHpA concentrations in serum from children at 2 years of age: - LOD: 0.157 ng/mL - Detection frequency: 268 => detected in 74.2% of all serum samples - Range (min-max): 0.111 - 5.800 ng/mL [2] human exposure primarily via dietary intake => contribution of drinking water [31] PFHpA serum concentrations in exposed (living in area under impact) and not exposed (subjected to background exposure) subjects => significant positive correlation between PFHpA and other PFAS-analytes (in exposed (all) + not exposed (all, except for PFHxS and PFOS)) => may be due to the presence of a dominant common source of exposure in the area [34] significant correlation with other PFAS-analytes detected in groundwater (p<0.05) and surface water (p<0.01), indicating a common or similar source of exposure [63] the very high persistence of PFAAs leads to poorly reversible exposure to these substances in the global environment and some local/regional environments including groundwater [101] - For food items of non-animal origin, PFOA, PFHpA and PFHxA were the most frequently detected homologues - Plant uptake of PFCAs and PFSAs is thought to occur primarily through the roots - It is known that sorption to natural organic matter increases with perfluoroalkyl chain length and is stronger for the sulfonates than the carboxylates => These observations explain why the more water soluble and less sorptive PFCAs (PFOA, PFHpA and PFHxA) are more prevalent in vegetable food samples [101] - Although the bioaccumulation potential of PFSAs and PFCAs increases with chain length in hydroponically grown plants, the uptake of the long-chain PFSAs and PFCAs is limited by the low fraction of freely dissolved ("available") chemical in the soil pore water [101] - Dietary intake of PFOS, PFNA, PFDA, PFUnDA, PFDoDA, PFTrDA and PFTeDA was largely dominated by the consumption of fish products (60-100% of the total dietary intake) - In contrast, dietary intake of PFHxA, PFHpA, PFHxS and PFOA originated from a number of food categories with less than 20% of the total dietary intake coming from consumption of fish [101] - While drinking water intake contributes to a major part of the total exposure (36-53%) for shorter chain PFAAs (PFHxA, PFHpA and PFHxS), dust ingestion is estimated to make a significant contribution (27-49%) to the total exposure of PFHxA, PFHpA, PFNA, PFDoDA and PFTeDA - Nevertheless, dietary intake contributed more than 50% of the total exposure for all PFAAs with more than 8 perfluorinated carbon atoms, as dietary exposure to these longer chain PFAAs is primarily due to fish consumption [109] Study on residential exposure: Vacuum cleaner contents from 39 homes were tested for perfluoroalkyl chemicals: * PFHpA: - Detection frequency: 100 %, Median: 17 (2.4140) ng/g - Correlation between PFHpA and PFBS (0.32, p<0.05) - Correlation between PFHpA and PFHxS (0.37, p<0.05) - Correlation between PFHpA and PFOS (0.58, p<0.05) - Correlation between PFHpA and PFPeA (0.91, p<0.05) - Correlation between PFHpA and PFOA (0.84, p<0.05) - Correlation between PFHpA and PFNA (0.72, p<0.05) [88] 50 paired human maternal and cord serum samples were analyzed => 5 PFAA precursors were detected (a.o. 6:2 FTS) => However, PFOS, PFHxS and longer-chain PFCAs were still the major fluorinated contaminants *Maternal serum: - Percentage > LOD: 100% - Mean SD: 0.075 0.052 ng/mL - Range: 0.015 - 0.28 ng/mL *Cord serum: - Percentage > LOD: 96% - Mean SD: 0.047 0.043 ng/mL - Range: <0.012 - 0.25 ng/mL => Sign. correlation between maternal and cord serum PFHxS [98] Beijing prenatal exposure study (human): PFHxS: 157 paired maternal and cord serum samples around delivery: * Maternal serum (ng/mL): - Percentage > LOD: 100% - Mean SD: 0.63 0.47 - Median: 0.50 - Geometric mean: 0.53 - Range: 0.12-4.22 * Cord serum (ng/mL): - Percentage > LOD: 100% - Mean SD: 0.26 0.30 - Median: 0.18 - Geometric mean: 0.19 - Range: 0.014-3.17 - Maternal-fetal correlation: r=0.684, p<0.001 * Maternal and cord serum concentration ratio: - Mean CS SD:MS: 0.43 0.34:1 [100] * PFHxS concentrations (detected in >50% of samples) in 2011, by gender and age group: - 0-4 y: FEMALE (n=2): 2.2 ng/mL, MALE (n=2): 1.9 ng/mL, ALL (n=4): 2.1 ng/mL - 5-15 y: FEMALE (n=2): 3.1 ng/mL, MALE (n=2): 3.0 ng/mL, ALL (n=4): 3.1 ng/mL - 16-30 y: FEMALE (n=2): 2.4 ng/mL, MALE (n=2): 3.9 ng/mL, ALL (n=4): 3.1 ng/mL - 31-45 y: FEMALE (n=2): 1.7 ng/mL, MALE (n=2): 3.6 ng/mL, ALL (n=4): 2.7 ng/mL - 46-60 y: FEMALE (n=2): 1.7 ng/mL, MALE (n=2): 3.3 ng/mL, ALL (n=4): 2.5 ng/mL - >60 y: FEMALE (n=2): 2.5 ng/mL, MALE (n=2): 3.1 ng/mL, ALL (n=4): 2.8 ng/mL [100] * PFHxS concentrations (detected in >50% of samples) by year: - FEMALE: 2002 (n=2): 8.5 ng/mL, 2004 (n=2): 4.9 ng/mL, 2006 (n=4): 6.2 ng/mL, 2008 (n=4): 3.9 ng/mL, 2011 (n=4): 2.7 ng/mL, 2013 (n=1): 2.0 ng/mL - MALE: 2002 (n=2): 9.9 ng/mL, 2004 (n=4): 6.5 ng/mL, 2006 (n=4): 5.0 ng/mL, 2008 (n=4): 4.0 ng/mL, 2011 (n=4): 2.5 ng/mL, 2013 (n=3): 1.4 ng/mL - ALL: 2002 (n=4): 9.2 ng/mL, 2004 (n=6): 6.0 ng/mL, 2006 (n=8): 5.6 ng/mL, 2008 (n=8): 4.0 ng/mL, 2011 (n=8): 2.6 ng/mL, 2013 (n=4): 1.2 ng/mL => PFHxS followed the same pattern as PFOS with decreasing levels [117] Surface water and groundwater in a large area of the Veneto Region in northeastern Italy was contaminated with PFAS that were also found in drinking water samples. The health surveillance program is a free-of-charge population-based screening program. Residents who decided to participate in the program completed a structured interview administered by a trained public health nurse, followed by blood pressure measurement, and blood and urine sampling. A total of 18,345 residents were included in the present analysis. Only 3 of the 12 PFAS were quantifiable in at least 80% of serum samples: PFOA (99.9%), PFOS (99.8%) and PFHxS (98.1%) [117] Serum concentrations (ng/mL) of PFHxS and percentage of samples above the LOQ in the study population (18,345 subjects, 14-39 years of age): PFHxS: - Min: <0.5 - 5th percentile: 0.7 - 25th percentile: 1.9 - Median: 3.9 - 75th percentile: 7.4 - 95th percentile: 18.1 - Max: 127.0 % samples LOQ: 98.1% [117] PFBA and PFBS were found in high concentrations in drinking water but were detected only in a minority of serum samples at relatively low concentration, whereas PFOS and PFHxS, which were scarcely represented in drinking water, were detected in almost 100% of serum samples. This discrepancy may be explained by the exposure to PFOS and PFHxS from other sources, as demonstrated for the general population, and by the longer human half-lives of PFOS and PFHxS in comparison with PFBA and PFBS. Moreover, the exposure period to PFBA and PFBS was shorter in comparison with that of PFOA and PFOS. [14] - rats: oral, 30 ppm: female: < 12 h, male: < 12 h - mice: oral, 3 ppm: female: > 12 h < 7 d, male: > 12 h, < 7 d [3] liver AUCSS: < 1,9 liver weight ratio: > 1.25 [3] humans: model liver: LOEL = 10 mg/kg/day liver weight ratio: 1.2-1.3 [72] * Direct injection analysis: - Surface water (river Lek, NL): 70 +/- 4 ng/L - Riverbank filtrate (river Oude Rijn, NL): N.D. * Solid-phase extraction analysis: - Riverbank filtrate (river Oude Rijn, NL): N.D. - RO permeate: N.D. [6] tap water MDL = 0.17 ng/L MLQ = 0.52 ng/L not detected [10] total discharge of PFASs into the aquatic environment ranged between 10 g d-1 and 10 000 g d-1, depending on the water usage in the community connected to the sewage treatment plant [13] - river water NL: detectable concentrations: min. 1.7 ng/L - max. 812 ng/L => such high concentrations could lead to elevated exposure of aquatic species downstream from the production plant compared to other locations upstream of the production plant - drinking water: detectable concentrations: min. 0.25 ng/L - max. 11 ng/L [30] levels of PFHpA in serum from participating mothers (SE): 1996-1999 (n=147): - MDL: 0.04 ng/g - Median (range): 0.060 (<MDL-0.400) ng/g - Mean (SE): 0.070 (0.004) ng/g - % of levels <MDL: 28 2008-2011 (n=134): - MDL: 0.04 ng/g - Median (range): 0.045 (<MDL-0.296) ng/g - Mean (SE): 0.059 (0.004) ng/g - % of levels <MDL: 43 => no significant difference between time periods [47] PFHpA in human semen samples: - Detection frequency: 100% - Mean +/- SD: 0.50 +/- 0.45 ng/mL - 5th percentile: 0.091 - Median: 0.36 ng/mL - 95th percentile: 1.5 ng/mL => Much higher concentrations were found in semen, suggesting that blood-testis barrier may not work for these chemicals => positively (0.208) associated with age (p<0.05) [40] Human: * Liver: - Mean: 33.3 ng/g wet weight - Median: 1.5 ng/g wet weight - Range: 638-BDL ng/g wet weight - MLOD: 3.00 ng/g wet weight - % of detection: 5 * Bone: - Mean: 77.1 ng/g wet weight - Median: 2.4 ng/g wet weight - Range: 309-BDL ng/g wet weight - MLOD: 2.89 ng/g wet weight - % of detection: 45 [31] PFHpA serum concentrations in exposed (living in area under impact) and not exposed (subjected to background exposure) subjects EXPOSED (n=257) - Minimum: <LOQ - Median: <LOQ - Maximum: 0.42 ng/g - % levels <LOQ: 49 NOT EXPOSED (n=250) - Minimum: <LOQ - Median: <LOQ - Maximum: 0.26 ng/g - % levels <LOQ: 34 => PFHpA concentration sign. larger in exposed than in not exposed subjects (p<0.001) [78] * PFHpA concentrations in human urine samples (n=86): - Mean: 1.9 ng/L - Median: 0.82 ng/L - Min: Below detection limit - Max: 19 ng/L * PFHpA concentrations in human urine samples (n=86): - Mean: 3.4 ng/g creatinine - Median: 1.0 ng/g creatinine - Min: Below detection limit - Max: 73 ng/g creatinine - Detection frequency: 67 % => a very good correlation was observed between the blood and urine concentrations (p=0.034) [31] - males sign. higher serum concentrations than females in exposed group (p=0.027) - higher serum concentrations with age in exposed group (p=0.002) - higher serum concentrations with BMI in exposed group (p<0.001) - raising own livestock sign. higher serum concentrations in exposed group (p=0.017) - sign. and direct correlation with specific food groups in exposed (cereals and derivatives, wild fish and game, vegetable oils) and not exposed group (liver, wild fish and game) - lower serum concentrations with tap water consumption in not exposed group (p=0.016) => due to different tap water use in residences [40] Human: * Brain: - Mean: BDL - MLOD: 2.70 ng/g wet weight - % of detection: 0 * Lung: - Mean: 17.4 ng/g wet weight - Median: 1.5 ng/g wet weight - Range: 245-BDL ng/g wet weight - MLOD: 3.00 ng/g wet weight - % of detection: 37 [40] Human: * Kidney: - Mean: 7.8 ng/g wet weight - Median: 2.6 ng/g wet weight - Range: 94.3-BDL ng/g wet weight - MLOD: 5.26 ng/g wet weight - % of detection: 95 [32] Concentration of PFHpA in human plasma A samples (14 datasets): - Assigned value: 0.18 - Average: 0.32 ng/ml - Median: 0.19 ng/ml - Min.: 0.14 ng/ml - Max.: 1.70 ng/ml - SD: 0.41 - % relative SD: 127 [40] - Smokers showed smaller accumulation of PFASs - Older people (more than 60 years) showed higher concentrations of PFASs => Clear indication that these compounds accumulate after a long-term exposure [32] Concentration of PFHpA in human plasma B samples (14 datasets): - Assigned value: 0.17 ng/ml - Average: 0.26 ng/ml - Median: 0.18 ng/ml - Min.: 0.10 ng/ml - Max.: 1.10 ng/ml - SD: 0.25 - % relative SD: 98 [71] Shorter chain PFASs have been shown to absorb into the liver more readily than those with longer chains which are concentrated in the blood proteins [38] PFHpA: Repeated study on aging individuals SE: Median (Interquartile range): - 2001-2004 (Age 70): 0.05 (0.03, 0.09) - 2006-2009 (Age 75): 0.07 (0.04, 0.11) - 2011-2014 (Age 80): 0.03 (0.01, 0.06) Decreasing trend (p<0.00001) [45] PFHpA in 141 pregnant woman serum samples: - Mean: 0.20 ng/mL - Max: 0.54 ng/mL - Min: 0.06 ng/mL - Geomean: 0.18 ng/mL - Median: 0.18 ng/mL [47] PFHpA in human blood samples: - Detection frequency: 100% - Mean +/- SD: 0.52 +/- 0.37 ng/mL - 5th percentile: 0.15 ng/mL - Median: 0.41 ng/mL - 95th percentile: 1.3 ng/mL [78] PFHpA concentrations in human blood samples (n=86): - Mean: 0.085 ng/mL - Median: 0.058 ng/mL - Min: Below detection limit - Max: 0.37 ng/mL - Detection frequency: 70% [4] Inhalation exposure rats to 6:2 FTOH: 0.5 ppm PFHpA => Male: 15.4 hrs => Female: 2.1 hrs Inhalation exposure rats to 6:2 FTOH: 5.0 ppm PFHpA => Male: 23.3 hrs => Female: 1.2 hrs [78] PFHpA biological elimination half-lives: * Young females: - Mean: 1.5 years - SE: 0.3 years - GM: 1.0 years - Median: 1.6 years - Min: 0.11 years - Max: 3.3 years - N: 12 * All males and older females: - Mean: 1.2 years - SE: 0.2 years - GM: 0.82 years - Median: 0.79 years - Min: 0.12 years - Max: 5.1 years [78] Owing to the possibility that excretion pathways other than urine may be significant, the estimated elimination half-lives reported in the current study can be viewed as upper limits [100] PFHpA population halving time (time period 20062013, based on human serum samples): 7.8 years [4] Inhalation exposure rats to 6:2 FTOH: 0.5 ppm PFHpA => Male: 0.045 h-1 => Female: 0.330 h-1 Inhalation exposure rats to 6:2 FTOH: 5.0 ppm PFHpA => Male: 0.029 h-1 => Female: 0.577 h-1 [4] Inhalation exposure rats to 6:2 FTOH: 0.5 ppm PFHpA => Male: 0.009 lkg-1h-1 => Female: 0.066 lkg-1h-1 Inhalation exposure rats to 6:2 FTOH: 5.0 ppm PFHpA=> Male: 0.006 lkg-1h-1 => Female: 0.116 lkg-1h-1 Scaled: average estimated rat clearance: 0.5 ppm 6:2 FTOH PFHpA => 0.037 lkg-1h-1 Scaled: average estimated rat clearance: 5 ppm 6:2 FTOH PFHpA => 0.061 lkg-1h-1 Scaled: predicted human clearance: 0.5 ppm 6:2 FTOH PFHpA => 0.808 lkg-1h-1 Scaled: predicted human clearance: 5 ppm 6:2 FTOH PFHpA => 1.315 lkg-1h-1 [4] AUC0- Inhalation exposure rats to 6:2 FTOH: 0.5 ppm PFHpA => Male: 3858.2 nanomoles/l.hr => Female: 247.1 nanomoles/l.hr AUC0- Inhalation exposure rats to 6:2 FTOH: 5.0 ppm PFHpA => Male: 11010.6 nanomoles/l.hr => Female: 306.9 nanomoles/l.hr Occupational exposure to 6:2 FTOH humans: PFHpA => 130.9 ng/ml.d [39] NOEC and LOECs for PFAAs for transactivation of mouse and human PPAR in transiently transfected COS-1 Cells: PFHpA: Mouse: - NOEC: 3 M - LOEC: 5 M - LOEC: 1.82 g/mL (p<0.0001) [78] PFHpA renal clearance: * Young female group: - Mean: 0.61 mL/day/kg - 95% CI: 0.022-1.2 mL/day/kg - GM: 0.27 mL/day/kg - Median: 0.17 mL/day/kg - N: 12 * Male and older female group: - Mean: 0.61 mL/day/kg - 95% CI: 0.38-0.83 mL/day/kg - GM: 0.39 mL/day/kg - Median: 0.41 mL/day/kg - N: 31 [78] - Renal clearance was assumed to be the major pathway for human elimination of the PFCAs and PFSAs, thus CLtotal was set equal to CLrenal - However, considering that menstrual clearance is an important clearance pathway for PFAAs in young females, menstrual clearance was estimated and added to renal clearance for calculation of CLtotal in young females [39] NOEC and LOECs for PFAAs for transactivation of mouse and human PPAR in transiently transfected COS-1 Cells: PFHpA: Human: - NOEC: <0.5 M - LOEC: 0.5 M - LOEC: 0.18 g/mL (p<0.01) [118] Spatial and seasonal distributions of PFAS in air, water, sediment, soil, and fish samples in the Asan Lake area of South Korea. Estimation of bioaccumulation potential of PFAS using sediment-water partition coefficients and bioaccumulation factors. PFHpA concentration in air samples: - Air (unit: pg/m3, n=4): Gaseous: * Range: nd ~ 1.57 * Mean: 1.01 * Median: 1.38 * DF (%): 75.0 [118] Particulate: * Range: nd ~ 0.67 * Mean: 0.17 * Median: 0.67 * DF (%): 25.0 Total: * Range: nd ~ 2.24 * Mean: 1.18 * Median: 1.24 * G (%): 85.8 [101] Concentrations of PFHpA in food basket samples (SE): - Dairy products: 1999-2010: <MDL - Meat products: 1999: 3.2 pg/g, 2005: <MDL, 2010: 2.3 pg/g - Fats: 1999-2010: <MDL - Pastries: 1999: 2.1 pg/g, 2005: <MDL, 2010: <MDL - Fish products: 1999: 27.5 pg/g, 2005: 16.0 pg/g, 2010: 3.1 pg/g - Egg: 1999-2010: <MDL - Cereal products: 1999: 3.0 pg/g, 2005: 3.0 pg/g, 2010: <MDL - Vegetables, including root vegetables: 1999: 2.2 pg/g, 2005: 1.7 pg/g, 2010: 1.8 pg/g [101] - Fruit: 1999-2010: <MDL - Potatoes: 1999: <MDL, 2005: 2.0 pg/g, 2010: 1.5 pg/g - Sugar and sweets: 1999-2010: <MDL - Soft drinks, lemonade: 1999-2010: <MDL => Short-chain PFCAs (PFHxA and PFHpA) were detected at concentrations close to MDL in cereals, vegetables and potatoes => Slightly elevated concentrations of PFHpA were observed in the fish homogenates from 1999 and 2005 (27.5 and 16.1 pg/g) [2] observation well NL max. value: 320 (214318) ng/L [5] LOD = 0.4 ng/L Freq = 29.9 % max = 21 ng/L Average = 1 ng/L med = 0 ng/L Per90 = 1 ng/L [12] SE Surface water: PFCAs ranging from 54% for firefighting sites to 85% for skiing and urban areas (of total PFASs) [28] Rivers SE: - Detection frequency: 48% - Concentration range: 0.36-1.7 ng/L - Average: 1.0 ng/L [16] GR near sea shore: - beached plastic pellets: present in 1 out of 5 samples - sediment: present in 1 out of 8 samples [2] estimated concentration of PFAA in public supply well field NL: 13 ng/L [6] tap water MDL = 0.035 ng/L MLQ = 0.12 ng/L max. value NL: 1.91 ng/L [10] total discharge of PFASs into the aquatic environment ranged between 10 g d-1 and 10 000 g d-1, depending on the water usage in the community connected to the sewage treatment plant [8] concentrations in water CN: - Dissolved phase: mean: 25.7 ng/L contribution to total PFASs: 2.12% - Separate suspended particulate matter: mean: 1.25 ng/L contribution to total PFASs: 2.85% - Soil: mean: 0.59 ng/L contribution to total PFASs: 8.23% - Leaves: mean: 2.97 ng/L contribution to total PFASs: 3.49% - Bark: mean: 1.01 ng/L contribution to total PFASs: 2.66% [34] - Sewage Treatment Plant (STP) effluent: PFHpA: 15% of total PFASs - Unspecific industry: PFHpA: 14% of total PFASs [32] Concentration of PFHpA in canal water samples (17 datasets): - Assigned value: 3.48 ng/L - Average: 3.72 ng/L - Median: 3.66 ng/L - Min.: 1.99 ng/L - Max.: 7.30 ng/L - SD: 3.72 - % relative SD: 36 [21] PFHxA and PFHpA have been reported in fresh and coastal waters at levels comparable to, or much higher than PFOA [12] SE Drinking water: PFCAs 45% of total PFASs [102] Tap water NL (n=4): Average SD (Range): 1.2 0.22 (0.9-1.4) ng/L => Levels of PFAAs observed in four tap water samples closely resembled each other as demonstrated by the relatively low standard deviation of the mean values [17] SE - Other PFCAs detected in filtered effluent water were PFHpA, ... - Other PFCAs detected in sludge were PFHpA, ... - Major contributions to the daily discharge Dd connected to the WWTP were from PFHpA, ... [13] - river water NL: detectable concentrations: min. 1.5 ng/L - max. 2.2 ng/L - drinking water NL: detectable concentrations: min. 1.1 ng/L - max. 2.4 ng/L [102] * Comparison tap water and hot water NL: - Drinking water treatment plant A: Tap water: 0.54 ng/L & Hot water: 0.60 ng/L - Drinking water treatment plant B: Tap water: 1.7 ng/L & Hot water: 1.2 ng/L => Concentrations of PFAAs detected in hot water were similar to those in the corresponding tap water * Post-mixed cola NL (n=6): Average SD (Range): 0.37 0.26 (<0.070.71) ng/L => Ratio tap water/post-mixed cola (Ratio of averages): 3.9 => PFAS concentrations in tap water were higher than in cola => A combination of the dilution step and the purification step inside the cola [12] SE Groundwater: PFCAs ranging from 20% for skiing to 43% for industrial areas (of total PFASs) [102] * Coffee NL: - Brewed coffee from coffee machines (n=12): Average SD (Range): 1.4 0.73 (<0.11-2.4) ng/L - Manually brewed coffee (n=4): Average SD (Range): 1.0 0.22 (0.86-1.3) ng/L - Procedure blank brewed coffee (n=2): Average SD (Range): 0.96 0.13 (0.850.96) ng/L => Coffee collected from coffee machines contained sign. higher concentrations of PFHpA, PFOA and PFOS than the [30] observation wells SE: median water conc. PFHpA: 24 ng/L [102] => Also sign. Increased concentrations of PFOA, PFDA and PFOS were found in manually brewed coffee => The increase of concentrations of PFOA and PFOS in the manually and machine brewed coffee, compared to the tap water used, has been shown to originate from the coffee beans => Fluoropolymer tubes showed leaching of PFAAs at high (80 C) temperature, but its relevance for contamination of beverages in practice is small [34] - Skiing areas: PFHpA: 8% of total PFASs [113] 397 food samples collected from a market in Busan (Korea) in 2011 (n=227) and 2012 (n=170). Food samples of 66 different food types, classified into 7 food categories. 34 tap water samples were also collected from 16 districts in Busan. PFHpA concentrations in each food group (ng/g, beverage unit: ng/L): [113] * Fish and shellfish (n=99): DF (%) = 4.0, Mean = 0.011, Max = 0.410 * Meat and its products (n=39): DF (%) = 0.0, Mean -, Max * Vegetables and fruit (n=78): DF (%) = 1.3, Mean = 0.002, Max = 0.134 * Processed products (n=90): DF (%) = 0.0, Mean -, Max * Dairy (n=37): DF (%) = 2.7, Mean = 0.015, Max = 0.563 [113] * Beverage (n=21): DF (%) = 4.8, Mean = 0.022, Max = 0.462 * Others (n=33): DF (%) = 3.0, Mean = 0.011, Max = 0.359 PFHpA concentrations in bottled water and tap water samples (ng/L): * Bottled water (n=8): DF (%) = 0.0, Mean -, Max - [113] PFHpA was detected in all of the tap water samples. This indicates that Korean tap water is contaminated with PFAAs, and this is likely to be caused by PFAAs contamination in the original sources of the drinking water. 90% of the drinking water in Korea comes from flowing surface water. [34] Mean concentration of 2.9 ng/L in rivers (SE) [28] Recipient sea: - Detection frequency: 30% - Concentration range: 0.61-1.0 ng/L - Average: 0.81 ng/L [18] river water and 3 connected water wells for drinking water: - before oxidation: PFHxA, PFHpA and PFOA ranged between 10-17 ng/L - after oxidation: a rise of +270830% or (+46-127 ng/L) => of this increase 18-53% could be explained by known precursors, in particular 6:2 FTS [113] * Tap water (n=34): DF (%) = 100.0, Mean = 4.85, Max = 6.91 [20] EU WWTP effluents: PFHpA: - LOQ: 1 ng/L - Freq.: detected in 94 % of samples - Max. concentration: 2962 ng/L - Average concentration: 82.9 ng/L - Median concentration: 5.1 ng/L - 90th percentile (Per90): 20.3 ng/L [21] PFHpA near air force base: 149 g/L [118] Spatial and seasonal distributions of PFAS in air, water, sediment, soil, and fish samples in the Asan Lake area of South Korea. Estimation of bioaccumulation potential of PFAS using sediment-water partition coefficients and bioaccumulation factors. [118] PFHpA concentration in water and sediment samples: - Water (unit: ng/L, n=47): * Range: 1.8 ~ 66.9 * Mean: 23.4 * Median: 14.0 * 75%: 38.2 * 95%: 62.4 * DF (%): 100.0 [118] - Sediments (unit: ng/g dw, n=47): * Range: nd ~ 0.24 * Mean: 0.05 * Median: 0.04 * 75%: 0.06 * 95%: 0.18 * DF (%): 80.9 [118] The predominant PFAS in water was PFPeA, followed by PFHxA. Shortchain PFAS such as PFPeA, PFHxA and LPFBS were detected more frequently in water than in the other environmental matrices. Short-chain PFAS and PFOA were detected during all four seasons. The seasonally dominant PFCAs were PFHxA and PFOA during summer and PFPeA and PFHxA during the other seasons. [66] PFHpA levels (ng/L) in outgoing drinking water from two waterworks in Ronneby (SE): - Site 1 (Brantafors): 32 - Site 2 (Krragarden): 1.4 [117] In spring 2013, groundwater of a vast area of the Veneto Region (northeastern Italy) was found to be contaminated by perfluoralkyl substances (PFAS) from a PFAS manufacturing plant active since the late 1960s. [117] Residents were exposed to high concentrations of PFAS, particularly perfluorooctanoic acid (PFOA), through drinking water until autumn 2013. [117] Surface water and groundwater in a large area of the Veneto Region in northeastern Italy was contaminated with PFAS that were also found in drinking water samples. A manufacturing plant located in the town of Trissino that produced PFAS since the late 1960s was identified as the only likely source of water contamination. [117] Measurements of 152 drinking water samples collected in July and August 2013 indicated that the main contaminants were PFOA, PFBA and PFBS, followed by PFPeA, PFHxA, PFOS, PFHpA and PFHxS. The longer-chain PFAS congeners (PFNA, PFDA, PFUnA, PFDoA) were detected only in a minority of samples and at lower concentrations. [117] PFHpA concentrations (ng/L) in 152 samples of drinking water taken during July and August 2013, before the full implementation of granular activated carbon filters: - Min: <10.0 - 25th percentile: <10.0 - Median: 14.0 - 75th percentile: 23.0 [21] PFCAs were found in nearly equal concentrations in municipal (14-459 ng/L) and industrial (3-664 ng/L) wastewaters => origin: PFOA in fluorinated polymeric products + degradation of precursors used in household products [24] river water CN, maximum concentrations in * yearly monitoring: - 2011: 15.9 ng/L - 2012: 45.5 ng/L - 2013: 103 ng/L - 2014: 55.0 ng/L * seasonal monitoring: summer (119 ng/L) > autumn (103 ng/L) > spring (36.8 ng/L) > winter (29.4 ng/L) => peak river contamination periods: summer & autumn [22] Mass of PFHpA annually exported to WWTPs from landfills (AU): - Min: 0.46 g - Mean: 4.6 g - Max: 10 g => PFASs are only partially removed from wastewater + potential for additional formation during treatment from degradation of precursors => WWTPs are considered major point sources of PFASs release into the aquatic environment [119] 3 replicate aqueous and solid samples were collected from each of 19 Australian WWTPs throughout 2017. Samples were taken from various stages within the treatment train from a range of WWTPs to determine the trends in the mass flux and partitioning of PFAS within the sampled WWTPs. [119] Summary statistics for pooled aqueous (n=201, triplicates from 67 individual locations within 19 WWTPs) and pooled solid (n=51, triplicates from 5 primary and secondary sludge locations, 6 lagoon sludges and a lagoon dredge pile) samples. [119] * PFHpA concentration in aqueous samples (ng/L): - Median: 5.0 - Mean: 6.1 - s.d.: 5.1 - min: <LOD - max: 34 - Detect (%): 100% [119] * PFHpA concentration in solid samples (ng/g dw): - Median: <LOQ - Mean: 0.30 - s.d.: 0.66 - min: <LOD - max: 4.1 - Detect (%): 54% 216-283-0 1546-95-8 2,2,3,3,4,4,5,5,6,6,7,7- 7H-PFHpA C7 dodecafluoroheptanoic acid Pre-registered 206-800-8 375-92-8 1,1,2,2,3,3,4,4,5,5,6,6,7 PFHpS C7 ,7,7- pentadecafluoroheptane-1- sulphonic acid Pre-registered [17] generally logKd increased with chain length, as true for C>6 [17] - the fate of PFASs sorbed to sludge is generally to be further spread in the environment through application of sludge or sludge-containing products - in 2014: 24% of sludge was applied on agricultural soil, 24% was used to cover mines and dump sites, and 29% was used to manufacture soil intended for less sensitive land use [48] Change in thyroid hormone TH-responsive genes in neuronal cells of avian species: *Gallus domesticus: D2 mRNA was upregulated following exposure to PFHpA: 10 M: Mean fold change +/- SD: 3.34 +/- 0.08 (p<0.05) TTR was upregulated following exposure to PFHpA: 10 M: Mean fold change +/- SD: 5.10 +/- 1.73 (p<0.05) => short-chained PFCs altered the expression of TH-responsive genes in chicken embryonic neuronal cells to a greater extent than the longchained PFCs => due to bioavailability: could have entered neuronal cells more readily due to their lower binding affinities to extracellular proteins *Larus argentatus: RC3 mRNA was upregulated following exposure to PFHpA: 10 M: Mean fold change +/- SD: 5.74 +/- 0.70 (p<0.05) [50] NIH-3T3 cells: PFHpA led to increased PPAR (at 100 M, p<0.001) and PPAR activity (at 100 M, p<0.001) => as PPAR is involved in adipocyte differentiation, PFAS affecting this receptor may be involved in development of obesity => short-chain PFCAs generally activate both PPAR and PPAR with similar potency and efficacy as long-chain PFCAs [51] - IC50 value and dissociation constant (Kd) with hPPAR-LBD: PFHpA: IC50: 192.4 17.2 M, Kd: 1330.4 119.0 M - Docking interactions of PFHpA with hPPAR-LBD: Hydrogen bonds: Ser-289, His-449, Tyr-473 [106] - Mean SD: 0.454 0.480 ng/mL * Height at 2 years of age was inversely related to PFHpA concentrations: (95% CI): -0.22 (-0.55, 0.11) (p=0.197) * 2-year height gain decreased with increased PFHpA concentrations: (95% CI): -0.09 (-0.47, 0.29) (p=0.628) * Weight at 2 years of age was inversely related to PFHpA concentrations: (95% CI): -0.09 (-0.25, 0.08) (p=0.314) * 2-year weight gain decreased with increased PFHpA concentrations: (95% CI): -0.02 (-0.19, 0.14) (p=0.768) * Results indicated a positive correlation (all p<0.001) between perfluorinated sulfonates (PFHxS and PFOS) and carboxylates (PFHpA, PFOA, PFNA, PFDA and PFUnDA), suggesting common exposure for such perfluoroalkyl compounds * Duration of breastfeeding was positively associated with the serum concentrations of ln PFHpA (p<0.001) [107] Analysis of 369 families with matched parental and cord serum samples from a birth cohort in Shandong, one of the regions seriously polluted by PFASs in China * PFHpA concentrations in 369 matched serum samples: - Paternal serum: Detection frequency: 96.7%, Geometric mean: 0.11 ng/mL - Maternal serum: Detection frequency: 85.1%, Geometric mean: 0.06 ng/mL - Cord serum: Detection frequency: 97.3%, Geometric mean: 0.09 ng/mL * Positive correlation between maternal & cord serum: 0.64 (p<0.01) * Positive correlation between paternal & maternal serum: 0.05 (p>0.05) * Positive correlation between paternal & cord serum: 0.03 (p>0.05) [107] * Positive correlation between PFHpA concentrations and maternal BMI: SE: 0.009 0.004 (p=0.018) * Positive correlation between PFHpA concentrations and maternal smoking history: SE: 0.067 0.027 (p=0.014) * Positive correlation between PFHpA concentrations and gestational age: SE: 0.021 0.010 (p=0.035) [52] human embryonal kidney cell line HEK293, activation of human PPAR, PPAR and PPAR - PFHpA triggered activation of PPAR: c20% value: 5.3 M - Most PFCAs also activated PPAR although higher PFCA concentrations were required for PPAR activation => PFHpA: c20% value: 43.2 M - In case of PPAR only high concentrations of PFHpA and PFOA activated this nuclear receptor => PFHpA: c20% value: 51.0 M => PPAR seems to be the primary molecular target of PFCA => Compounds with a mid chain length such as PFHpA and PFOA had the highest potential for PPAR activation, whereas PFCA with shorter and with longer carbon chain length showed a lower potential for PPAR activation [108] Analysis of 424 mother-fetus pairs from the Maoming Birth Cohort, China. * Concentrations of PFHpA in maternal and cord serum: - Cord serum: Detection rate: 54.95%, median: 0.006 ng/mL, mean SD: 0.02 0.13 ng/mL - Maternal serum: Detection rate: 65.57%, median: 0.007 ng/mL, mean SD: 0.01 0.01 ng/mL * PFAS in cord serum was positively correlated with PFAS in maternal serum for all PFAS (p<0.05) [84] PFCs can compete with thyroxine (T4, the transport form of thyroid hormone), for binding to the human thyroid hormone transport protein transthyretin (TTR) => such competitive capacity may lead to decreased thyroid hormone levels as previously reported for animals exposed to PFCs => PFCs do not affect the regulatory functions of the thyroid hormone system itself, but it is the competitive binding to transport proteins that alters the free thyroxine (T4) levels in blood => TTR is the main T4 carrier in cerebrospinal fluid, and also important in serum of most mammalian species and birds => Binding potency is clearly associated with the degree of fluorination of the alkyl chain, with a maximum potency at a chain length of eight carbons (PFOA) for PFCAs => PFCAs with a carbon chain length longer than eight have low TTR binding potencies => TTR is both in humans and in rodents the most important carrier protein for thyroid hormone to the developing fetus and the brain [108] * Transplacental transfer efficiency (TPT) of PFHpA = concentration cord serum (ng/mL)/ concentration maternal serum (ng/mL): - TPT (n=194): Median: 1.24, mean SD: 10.23 116.96 - U-shaped pattern for TPT in PFCAs: TPT decreased from PFBA to PFDA and then increased to PFTrDA * Higher TPT in PFAS alternatives than PFASs => PFAS alternatives may be more easily transported from mother to infant than conventional PFASs in uterus [84] PFHpA: - Molecular weight: 364.1 g/mol - T4-TTR binding at maximum concentration (10M): 7% - IC50 (concentration at 50% inhibition): 1565 nM - Slope of dose-response curve: -1.26 - T4-REP (relative potency compared to T4) factor: 0.039 - HPLC retention time: 31.9 min - Purity: 96 [86] - weak binding of PFHpA with the protein TR-LBD (TR = thyroid hormone receptor, LBD = ligand-binding domain) * Binding of PFHpA with TR-LBD: - Length: 8.96 A - IC50: 226 M - RP (relative potency (IC50 T3 / IC50 chemical)): 0.0013 - Hydrogen bonding: ARG 228 * all tested PFCs fit into the T3-binding pocket of TR-LBD, with the acid or hydroxyl end group residing toward the inner part and the hydrophobic chain toward the entrance of the binding pocket => all PFCs with an acid end group formed a hydrogen bond with [87] * Binding potency of PFHpA to TTR (TTR competitive binding assay): - IC50: 1128 139 nM - Kd: 180 22 nM - RP (relative potency (IC50 T4 / IC50 chemical)): 0.028 => perfluoroalkyl acids: Kd values decreased as carbon chain length increased from C4 to C8 => no further increase, but a downward trend was observed for perfluoroalkyl acids with longer chain lengths (C9-C14) * Binding potency of PFHpA to TTRmutK15G: - IC50: ND - Kd: ND - Hydrogen bonding: Lys15 => PFASs exhibited much weaker binding affinities to TTRmutK15G compared with wild-type TTR * Docking study: - Similar to T4, PFOA and PFOS could nearly fill the TTR ligandbinding pocket - Perfluoroalkyl acids with carbon chain length less than C8 did not adequately fill the T4 binding pocket [87] * Binding potency of PFHpA to TBG (TBG competitive binding assay): - IC50: ND - Kd: ND - RP (relative potency (IC50 T4 / IC50 chemical)): ND => only PFTA and PFTdA bound to TBG * Binding potency of PFHpA to TBGmutR378G: - IC50: ND - Kd: ND * Binding potency of PFHpA to TBGmutR381G: - IC50: ND - Kd: ND - Hydrogen bonding: Arg381 => both TBGmut378G and TBGmutR381G exhibited much weaker binding potencies to PFTA and PFTdA when compared with that of wild-type TBG * Docking study: - Perfluoroalkyl acids with carbon chain length less than C12 did not adequately fill the T4 binding pocket - Longer fluorinated carbon chain structures could nearly fill the TBG [46] PFASs can compete with thyroxine (T4) for binding to the human thyroid hormone transport protein transthyretin (TTR), which may lead to reduced thyroid hormone levels leading to endocrine disrupting adverse effects => 7H-PFHpA: T4-TTR Binding (%): 45 => Experimental Classification: Active (Y) => Predicted Classification: Active (Y) => Experimental pIC50 (conc. with 50% inhibition of T4): 2.06 mM => Predicted pIC50: 2.56 mM => Binding energy: -3.81 kcal/mol => PFASs can compete with thyroxine (T4) for binding to the human thyroid hormone transport protein transthyretin (TTR), which may lead to reduced thyroid hormone levels leading to endocrine disrupting adverse effects [46] Discriminating properties between active (Y) and inactive (N) classes: - Hydrophobicity of the H atom attached to heteroatom - Contributions from the H-050 descriptor 7H-PFHpA fitted exactly in the identical active sites making interactions with Leu17, Ala108, Ala109, Leu110, Thr118, Thr119 and Lys15, but not with Val121 => A hydrogen bond formed with the positively charged Lys15 and negatively charged oxygen atom of sulfonate or acid functional groups PFASs containing acid functional groups: - If the carbon chain length is between 6 and 10, PFASs will be highly toxic or active ones - Toxicity will be lower or non-existent for PFASs containing a C chain length greater than 10 or below 6 => 7H-PFHpA IS TOXIC/ACTIVE [84] PFCs can compete with thyroxine (T4, the transport form of thyroid hormone), for binding to the human thyroid hormone transport protein transthyretin (TTR) => such competitive capacity may lead to decreased thyroid hormone levels as previously reported for animals exposed to PFCs => PFCs do not affect the regulatory functions of the thyroid hormone system itself, but it is the competitive binding to transport proteins that alters the free thyroxine (T4) levels in blood => TTR is the main T4 carrier in cerebrospinal fluid, and also important in serum of most mammalian species and birds => TTR is both in humans and in rodents the most important carrier protein for thyroid hormone to the developing fetus and the brain [84] 7H-PFHpA: - Molecular weight: 346.1 g/mol - T4-TTR binding at maximum concentration (10M): 45% - IC50 (concentration at 50% inhibition): 8637 nM - Slope of dose-response curve: -1.19 - T4-REP (relative potency compared to T4) factor: 0.007 - HPLC retention time: N.A. - Purity: 98 [108] Analysis of 424 mother-fetus pairs from the Maoming Birth Cohort, China. * Concentrations of PFHpS in maternal and cord serum: - Cord serum: Detection rate: 58.49%, median: 0.01 ng/mL, mean SD: 0.03 0.14 ng/mL - Maternal serum: Detection rate: 92.22%, median: 0.05 ng/mL, mean SD: 0.07 0.11 ng/mL * PFAS in cord serum was positively correlated with PFAS in maternal serum for all PFAS (p<0.05) * Transplacental transfer efficiency (TPT) of PFHpS = concentration cord serum (ng/mL)/ concentration maternal serum (ng/mL): - TPT (n=244): Median: 0.44, mean SD: 0.84 3.55 - U-shaped pattern for TPT in PFCAs: TPT decreased from PFBA to PFDA and then increased to PFTrDA [108] * Higher TPT in PFAS alternatives than PFASs => PFAS alternatives may be more easily transported from mother to infant than conventional PFASs in uterus [109] - Correlation between PFHpA and PFDA (0.56, p<0.05) - Correlation between PFHpA and PFUA (0.64, p<0.05) - Correlation between PFHpA and PFDoA (0.56, p<0.05) - Correlation between PFHpA and PFTrA (0.63, p<0.05) - Correlation between PFHpA and PFTetraA (0.47, p<0.05) * Among 11 homes built after 1990, home age was positively correlated with PFPeA, PFHpA and PFOA levels at p<0.05, and with PFOS at p<0.10 * PFOS, PFOA, PFHpA and PFHxS were the most prominent PFCs found in these dust samples with median concentrations ranging from 16 to 47 ng/g [109] * Occupants of all ages can be exposed to indoor contaminants, but preschool-aged children have the greatest risk due to hand to mouth activity and the amount of time they spend playing on or near the floor [110] Study on chicken/duck poultry products and eggs in China => 2 or 3 samples were collected from henneries or abattoirs at different locations => Chicken (n=53), duck (n=47), chicken eggs (n=56, each sample being a combination of five eggs), and duck eggs (n=56, each sample being a combination of five eggs) => 3 tissues from chicken and ducks: breast meat, liver and subcutaneous fat => Final sample number: 300 samples from chicken (53*3=159) and duck (47*3=141): * Eggs: In chicken and duck eggs, PFHpA at 1.41 g/kg and 0.58 g/kg, were the highest concentrations * This study evaluated 56 chicken egg samples and showed that 23 had higher PFHpA levels [110] * The mean PFHpA concentration of the 23 samples is nearly 10 times the concentration of the remaining samples * In the 23 samples, almost all of the chickens are given feed, and 65% chickens are from Jiangsu Province * Meat: Concentrations of PFHpA were highest in chicken meat (0.23 g/kg) * PFHpA was highest in subcutaneous fat tissues in both chicken and duck (0.99-0.75 g/kg) => The current PFC concentrations in chicken, chicken eggs, duck and duck eggs from the Yangtze River Delta and Pearl River Delta are not of concern to adults => Chicken and chicken eggs could be dangerous to children! [110] => PFHpA was the predominant PFC in eggs of both species and was stored in subcutaneous fat [111] PFAAs were investigated in home produced eggs and commercially produced eggs surrounding a fluorochemical industrial park in China: PFHpA concentrations: * Home produced eggs (n=4, sites 1-4): Egg yolks: - 2 km distance: 0.29 ng/g - 5 km distance: <0.03 ng/g - 10 km distance: <0.03 ng/g - 20 km distance: <0.03 ng/g [111] Egg whites: - 2 km distance: <0.03 ng/g - 5 km distance: <0.03 ng/g - 10 km distance: <0.03 ng/g - 20 km distance: <0.03 ng/g Whole eggs: - 2 km distance: <0.03 ng/g - 5 km distance: <0.03 ng/g - 10 km distance: <0.03 ng/g - 20 km distance: <0.03 ng/g [111] * Commercially produced eggs (n=12, sites 5-16): Egg yolks: - Min: <0.03 ng/g - Max: <0.03 ng/g - Mean: <0.03 ng/g - Median: <0.03 ng/g - n > LOD (%): 0 (0) Egg whites: - Min: <0.03 ng/g - Max: <0.03 ng/g - Mean: <0.03 ng/g - Median: <0.03 ng/g - n > LOD (%): 0 (0) Whole eggs: - Min: <0.03 ng/g - Max: <0.03 ng/g - Mean: <0.03 ng/g - Median: <0.03 ng/g - n > LOD (%): 0 (0) => Unlike in the environmental media, egg yolks contained much lower proportions of PFPeA, PFHxA and PFHpA than of PFBA [116] In May 2009, indoor dust samples (n=17) were collected from fifteen homes located in nine Korean cities for house dust (n=15), and one shopping mall and one library in Seoul for non-residential indoor dust (n=2). Nine cities for house dust samples were chosen to represent densely populated cities, industrialized cities, and rural area. * Concentrations (ng/g) of PFHpA measured in indoor dust samples from Korea (n=17): - House dust (n=15): Range: <LOD - 2.5 Mean ( SD): 0.9 ( 0.9) Median: 0.7 % of < LOD: 40 - Non-residential indoor dust (n=2): Shopping mall: 2.8 Library: 7.1 [116] * There were no significant correlations between PFCAs (or PFSAs) and their precursors in house dust. * In Korean house dust samples, significant correlations (p<0.05) were only found among all the FTOHs and among almost all of the PFCAs. [100] 54 pooled human serum samples from 4920 individual samples (AU): * PFHpA concentrations in 2011, by age group: - 0-4 y (n=4): FOD: 100%, Min: 0.13 ng/mL, Max: 0.19 ng/mL, Mean: 0.16 ng/mL - 5-15 y (n=4): FOD: 100%, Min: 0.08 ng/mL, Max: 0.09 ng/mL, Mean: 0.08 ng/mL - 16-30 y (n=4): FOD: 100%, Min: 0.04 ng/mL, Max: 0.05 ng/mL, Mean: 0.04 ng/mL - 31-45 y (n=4): FOD: 100%, Min: 0.02 ng/mL, Max: 0.03 ng/mL, Mean: 0.03 ng/mL - 46-60 y (n=4): FOD: 100%, Min: 0.02 ng/mL, Max: 0.05 ng/mL, Mean: 0.04 ng/mL - >60 y (n=4): FOD: 100%, Min: 0.04 ng/mL, Max: 0.04 ng/mL, Mean: 0.04 ng/mL [100] * PFHpA concentrations (detected in >50% of samples) in 2011, by gender and age group: - 0-4 y: FEMALE (n=2): 0.16 ng/mL, MALE (n=2): 0.16 ng/mL, ALL (n=4): 0.16 ng/mL - 5-15 y: FEMALE (n=2): 0.08 ng/mL, MALE (n=2): 0.09 ng/mL, ALL (n=4): 0.08 ng/mL - 16-30 y: FEMALE (n=2): 0.05 ng/mL, MALE (n=2): 0.04 ng/mL, ALL (n=4): 0.04 ng/mL - 31-45 y: FEMALE (n=2): 0.02 ng/mL, MALE (n=2): 0.03 ng/mL, ALL (n=4): 0.03 ng/mL - 46-60 y: FEMALE (n=2): 0.03 ng/mL, MALE (n=2): 0.04 ng/mL, ALL (n=4): 0.04 ng/mL - >60 y: FEMALE (n=2): 0.04 ng/mL, MALE (n=2): 0.04 ng/mL, ALL (n=4): 0.04 ng/mL => A reduction with age could be observed for PFHpA => No gender differences could be [100] * PFHpA concentrations (detected in >50% of samples) by year: - FEMALE: 2002 (n=2): 0.17 ng/mL, 2004 (n=2): 0.21 ng/mL, 2006 (n=4): 0.22 ng/mL, 2008 (n=4): 0.15 ng/mL, 2011 (n=4): 0.12 ng/mL, 2013 (n=1): 0.10 ng/mL - MALE: 2002 (n=2): 0.22 ng/mL, 2004 (n=4): 0.23 ng/mL, 2006 (n=4): 0.21 ng/mL, 2008 (n=4): 0.16 ng/mL, 2011 (n=4): 0.12 ng/mL, 2013 (n=3): 0.12 ng/mL - ALL: 2002 (n=4): 0.20 ng/mL, 2004 (n=6): 0.22 ng/mL, 2006 (n=8): 0.21 ng/mL, 2008 (n=8): 0.16 ng/mL, 2011 (n=8): 0.12 ng/mL, 2013 (n=4): 0.11 ng/mL => PFHpA started to decrease after 2006 and was significantly lower in 2011 compared to 2006 [117] Surface water and groundwater in a large area of the Veneto Region in northeastern Italy was contaminated with PFAS that were also found in drinking water samples. The health surveillance program is a free-of-charge population-based screening program. Residents who decided to participate in the program completed a structured interview administered by a trained public health nurse, followed by blood pressure measurement, and blood and urine sampling. A total of 18,345 residents were included in the present analysis. [117] Only 3 of the 12 PFAS were quantifiable in at least 80% of serum samples: PFOA (99.9%), PFOS (99.8%) and PFHxS (98.1%). Serum concentrations (ng/mL) of PFHpA and percentage of samples above the LOQ in the study population (18,345 subjects, 14-39 years of age): PFHpA: - Min: <0.5 - 5th percentile: <0.5 - 25th percentile: <0.5 - Median: <0.5 - 75th percentile: <0.5 - 95th percentile: <0.5 - Max: 15.1 % samples LOQ: 1.4% [100] 54 pooled human serum samples from 4920 individual samples (AU): * PFHpS concentrations in 2011, by age group: - 0-4 y (n=4): FOD: 100%, Min: 0.13 ng/mL, Max: 0.18 ng/mL, Mean: 0.15 ng/mL - 5-15 y (n=4): FOD: 100%, Min: 0.15 ng/mL, Max: 0.20 ng/mL, Mean: 0.17 ng/mL - 16-30 y (n=4): FOD: 100%, Min: 0.16 ng/mL, Max: 0.36 ng/mL, Mean: 0.23 ng/mL - 31-45 y (n=4): FOD: 100%, Min: 0.15 ng/mL, Max: 0.37 ng/mL, Mean: 0.25 ng/mL - 46-60 y (n=4): FOD: 100%, Min: 0.22 ng/mL, Max: 0.46 ng/mL, Mean: 0.34 ng/mL - >60 y (n=4): FOD: 100%, Min: 0.35 ng/mL, Max: [100] * PFHpS concentrations (detected in >50% of samples) in 2011, by gender and age group: - 0-4 y: FEMALE (n=2): 0.15 ng/mL, MALE (n=2): 0.16 ng/mL, ALL (n=4): 0.15 ng/mL - 5-15 y: FEMALE (n=2): 0.16 ng/mL, MALE (n=2): 0.18 ng/mL, ALL (n=4): 0.17 ng/mL - 16-30 y: FEMALE (n=2): 0.17 ng/mL, MALE (n=2): 0.30 ng/mL, ALL (n=4): 0.23 ng/mL - 31-45 y: FEMALE (n=2): 0.15 ng/mL, MALE (n=2): 0.35 ng/mL, ALL (n=4): 0.25 ng/mL - 46-60 y: FEMALE (n=2): 0.24 ng/mL, MALE (n=2): 0.43 ng/mL, ALL (n=4): 0.34 ng/mL - >60 y: FEMALE (n=2): 0.35 ng/mL, MALE (n=2): 0.45 ng/mL, ALL (n=4): 0.40 ng/mL => Both PFHpS and PFOS increased with age from 0 to 4 to >60 years, while PFHxS remained at approximately the same level => Gender-specific patterns were also seen for PFHpS and PFOS, where levels seem to be higher for males [100] * PFHpS concentrations (detected in >50% of samples) by year: - FEMALE: 2002 (n=2): 0.51 ng/mL, 2004 (n=2): 0.36 ng/mL, 2006 (n=4): 0.37 ng/mL, 2008 (n=4): 0.22 ng/mL, 2011 (n=4): 0.15 ng/mL, 2013 (n=1): 0.13 ng/mL - MALE: 2002 (n=2): 0.52 ng/mL, 2004 (n=4): 0.39 ng/mL, 2006 (n=4): 0.38 ng/mL, 2008 (n=4): 0.26 ng/mL, 2011 (n=4): 0.17 ng/mL, 2013 (n=3): 0.12 ng/mL - ALL: 2002 (n=4): 0.51 ng/mL, 2004 (n=6): 0.38 ng/mL, 2006 (n=8): 0.37 ng/mL, 2008 (n=8): 0.24 ng/mL, 2011 (n=8): 0.16 ng/mL, 2013 (n=4): 0.12 ng/mL => PFHpS followed the same pattern as PFOS with decreasing levels throughout the entire study period [100] PFHpS population halving time (time period 20022013, based on human serum samples): 5.3 years [115] In September 2011, foods were purchased in 12 representative cities of Catalonia, all with more than 20,000 inhabitants. Food samples were obtained at each locality in 4 shops/stores of different size. Foods selected for PFAS analysis were among the most consumed in Catalonia. * Mean concentrations (in pg/g fw) of PFHpA in the groups of analyzed foodstuffs: - Meat and meat products: 280 - Fish and seafood: 270 - Vegetables: 360 - Tubers: <340 - Fruits: <350 - Eggs: 200 [115] - Milk: 240 - Dairy products: 170 - Cereals: <120 - Pulses: <250 - Oils: <140 - Industrial bakery: <100 * Fish and shellfish was the group in which more PFASs were detected and where the highest PFAS concentrations were found * Dietary intake of PFHpA by the population of Catalonia (in ng/kg body weight/day): - Children: ND = 0: 0.34; ND = 1/2 LOD: 12.8 - Boys adolescents: ND = 0: 0.07; ND = 1/2 LOD: 4.01 - Girls adolescents: ND = 0: 0.08; ND = 1/2 [115] - Male adults: ND = 0: 0.06; ND = 1/2 LOD: 3.21 - Female adults: ND = 0: 0.08; ND = 1/2 LOD: 3.76 - Male seniors: ND = 0: 0.06; ND = 1/2 LOD: 3.73 - Female seniors: ND = 0: 0.07; ND = 1/2 LOD: 3.24 * Concerning human health risks from dietary exposure to PFASs in Catalonia, it is important to note that for any of the age/gender groups of population estimated, the TDIs recommended by the EFSA were not exceeded. [115] * PFOS was the compound found in the highest number of samples (33 out of 80). PFOA, PFHpA, PFHxS, PFDA and PFDS were the compounds that, concurrently with PFOS, were detected in the greatest number of food samples. * PFHpA could be detected in samples of swordfish, salmon, red mullet, sole and clams, as well as in lettuce, eggs, whole and semi-skimmed milk and yogurt. [117] - 95th percentile: 56.0 - Max: 151.0 - % samples LOQ: 67.1% [119] In aqueous samples: between influent and final effluent, the compounds PFPeA, PFHxA, PFHpA, PFOA, PFNA, and PFDA (all of which are PFCAs) increased significantly. A number of transformation pathways with stable PFCA endproducts are known, this may explain some of the increase in PFCAs from influent to final effluent. Estimated yearly mass discharged: PFHpA: 22 kg annually. [118] Spatial and seasonal distributions of PFAS in air, water, sediment, soil, and fish samples in the Asan Lake area of South Korea. Estimation of bioaccumulation potential of PFAS using sediment-water partition coefficients and bioaccumulation factors. L-PFHpS concentration in air samples: - Air (unit: pg/m3, n=4): Gaseous: * Range: na * Mean: na * Median: na * DF (%): na [113] 397 food samples collected from a market in Busan (Korea) in 2011 (n=227) and 2012 (n=170). Food samples of 66 different food types, classified into 7 food categories. 34 tap water samples were also collected from 16 districts in Busan. PFHpS concentrations in each food group (ng/g, beverage unit: ng/L): * Fish and shellfish (n=99): DF (%) = 2.0, Mean = 0.018, Max = 1.40 * Meat and its products (n=39): DF (%) = 0.0, Mean -, Max * Vegetables and fruit (n=78): DF (%) = 0.0, Mean -, Max - [118] Particulate: * Range: na * Mean: na * Median: na * DF (%): na Total: * Range: na * Mean: na * Median: na * G (%): na [113] * Processed products (n=90): DF (%) = 1.1, Mean = 0.008, Max = 0.750 * Dairy (n=37): DF (%) = 0.0, Mean -, Max * Beverage (n=21): DF (%) = 0.0, Mean -, Max * Others (n=33): DF (%) = 0.0, Mean -, Max PFHpS concentrations in bottled water and tap water samples (ng/L): * Bottled water (n=8): DF (%) = 0.0, Mean -, Max - [118] Spatial and seasonal distributions of PFAS in air, water, sediment, soil, and fish samples in the Asan Lake area of South Korea. Estimation of bioaccumulation potential of PFAS using sediment-water partition coefficients and bioaccumulation factors. [118] L-PFHpS concentration in water and sediment samples: - Water (unit: ng/L, n=47): * Range: na * Mean: na * Median: na * 75%: na * 95%: na * DF (%): na - Sediments (unit: ng/g dw, n=47): * Range: na * Mean: na * Median: na * 75%: na * 95%: na * DF (%): na [66] PFHpS levels (ng/L) in outgoing drinking water from two waterworks in Ronneby (SE): - Site 1 (Brantafors): 60 - Site 2 (Krragarden): <1 [113] * Tap water (n=34): DF (%) = 0.0, Mean -, Max - [17] SE - Other PFSAs detected in filtered effluent water were PFHpS (mean 0.1 ng/L) - Other PFSAs detected in sludge were PFHpS, ... [13] - river water NL: detectable concentrations: min. 0.095 ng/L - max. 0.20 ng/L - drinking water NL: detectable concentrations: min. 0.021 ng/L - max. 0.030 ng/L [119] 3 replicate aqueous and solid samples were collected from each of 19 Australian WWTPs throughout 2017. Samples were taken from various stages within the treatment train from a range of WWTPs to determine the trends in the mass flux and partitioning of PFAS within the sampled WWTPs. Summary statistics for pooled aqueous (n=201, triplicates from 67 individual locations within 19 WWTPs) and pooled solid (n=51, triplicates from 5 primary and secondary sludge locations, 6 lagoon sludges and a lagoon dredge pile) samples. [119] * PFHpS concentration in aqueous samples (ng/L): - Median: <LOQ - Mean: 0.86 - s.d.: 1.7 - min: <LOD - max: 11 - Detect (%): 76% * PFHpS concentration in solid samples (ng/g dw): - Median: <LOD - Mean: 0.29 - s.d.: 0.67 - min: <LOD - max: 3.3 - Detect (%): 26% 480-310-4 / ammonium 2,2,3 trifluor3-(1,1,2,2,3,3-hexafluoro3trifluormethoxypropoxy), propionate ADO N A C7 (6 fluorinated) Registered [75] logKaw: 2.26 (estimated by COSMOtherm) => similar Kaw as PFOA [75] logKow (dry): 4.97 (estimated by COSMOtherm) => similar Kow as PFOA [75] logKlipw: 5.13 (estimated with regression equation) [75] logKoa = 7.23 (estimated by COSMOtherm) [75] pKa = 0.51 (estimated by SPARC) [75] logKoc: 1.78 / [19] 82765-77-3 Perfluoroheptanesulpho FHpSA C7 / namide 206-398-4 335-71-7 Pentadecafluoroheptane-1- PHpSF C7 sulphonyl fluoride 671-215-0 151772-59-7 Perfluoro-3,6,9- / C7 trioxadecanoic acid Pre-registered Not registered / / 5:3 fluorotelomer carboxylic acid 5:3 FTCA C8 (5 / fluorinated) / / Perfluorohexyl ethanoic acid FHEA C8 (6 / fluorinated) / 53826-12-3 6:2 fluorotelomer 6:2 FTCA C8 (6 / carboxylic acid fluorinated) [75] logKaw: 2.44 (estimated by COSMOtherm) => higher Kaw as PFOA [75] logKow (dry): 3.94 (estimated by COSMOtherm) => lower Kow than PFOA [75] logKlipw: 4.09 (estimated with regression equation) [75] logKoa = 6.38 (estimated by COSMOtherm) [75] pKa = 2.82 (estimated by SPARC) [75] logKoc: 0.75 [14] ADONA is not readily biodegradable, but starts to decompose thermally at 125C with completion at 175C leading to formation of volatile substances [23] Resistant to photolysis, hydroxyl (OH)-radical-mediated reactions, hydrolysis, and biodegradation => known PFECAs and PFESAs are likely to be highly persistent in the environment and not easily metabolized in biota [23] - OECD 111: Hydrolysis: 0% after 5 days - OECD 301B: Not readily biodegradable [63] Perfluoroalkyl and perfluoroether moieties are highly persistent under natural conditions => Even though some PFASs may partially degrade in the environment and biota, they will all ultimately transform into highly stable end products, wich are usually perfluoroalkyl or perfluoroalkyl(poly)ether acids such as PFCAs, PFSAs, PFECAs and PFESAs [75] - half-life in air with AOPWIN: 739.55 hour - half-life in water with BIOWIN3: 5760 hour - half-life in soil with BIOWIN3: 5760 hour [75] overall persistence (Pov): 346.25 days => lower Pov than PFOA [23] only serum elimination half-lives of two PFECAs in mammals (in rats and humans for ADONA) have been reported => they might be as bioaccumulative as the predecessors due to similar physicochemical properties, but this is uncertain [23] - toxicological profiles of GenX and ADONA: acute toxicity, eye and skin irritation, dermal sensitization, genotoxicity, repeated-dose toxicity, and developmental toxicity - ADONA and GenX were shown to cause liver damage in (at least) male rats or mice under repeated low exposure (<= 10 mg/kg/day) - ADONA was shown to be a PPAR alpha agonist in male rats [63] the high solubility and proteinbinding characteristics of ionic PFAAs challenge the conventional assessment of bioaccumulation potential that is through either bioconcentration factor in aquatic species (usually fish) or models based on octanol-water partition coefficients (Kow) => even when PFAAs are not characterized as bioaccumulative under the current regulatory frameworks, the accumulation of these PFAAs in the environment will lead to increasing [23] Known PFECAs and PFESAs: - P - maybe B - likely T - LRTP [23] on-going (increasing) emissions and high persistence => short-chain PFAAs, PFECAs and PFESAs accumulate in the environment => increasing exposure to these substances over time => even if emissions cease in the future it will take decades to centuries to reverse global contamination of short-chain PFAAs, PFECAs and PFESAs, due to high persistence and environmental mobility [63] the very high persistence of PFAAs leads to poorly reversible exposure to these substances in the global environment and some local/regional environments including groundwater [14] - rats: 5 x oral: male: 44 h - humans: male (n = 3): 23 +/- 11 d [23] Hydrolysis: (almost) 0% [23] Resistant to photolysis, hydroxyl (OH)-radical-mediated reactions, hydrolysis, and biodegradation => known PFECAs and PFESAs are likely to be highly persistent in the environment and not easily metabolized in biota [23] only serum elimination half-lives of two PFECAs in mammals have been reported => they might be as bioaccumulative as the predecessors due to similar physicochemical properties, but this is uncertain [23] three PFECAs have been shown to cause liver damage in rats in repeated- dose tests; one PFECA is suggested by its manufacturer to be classified as T under the REACH regulation => other known PFECAs may have the same mode-of-action and may fulfill the toxicity criteria under REACH [75] - half-life in air with AOPWIN: 726.57 hour - half-life in water with BIOWIN3: 17280 hour - half-life in soil with BIOWIN3: 17280 hour [75] overall persistence (Pov): 1038.67 days => same Pov as PFOA [93] Rapid elimination rate and nonbioaccumulation for 6:2 FTCA [76] inhibitory effect of 6:2 FTCA on human liver HL-7702 cells after 24h of exposure by MTT assay: - Model: BiPhasic - ^2: 162.0423 - R^2: 0.9542 - IC50: 1.04 x 10^(-3) => much higher IC50 value for 6:2 FTCA implies a much weaker cytotoxicity compared with that of PFOA and PFOS [76] binding affinity of 6:2 FTCA to hL-FABP (human liver fatty acid binding protein): - IC50: 116.88 +/- 0.52 M - Kd: 436.55 +/- 1.92 M - R^2: 0.999 => Kd values of 6:2 FTCA and 6:2 FTS were two orders of magnitude higher than those of PFOA and PFOS, indicating weak binding capacity for 6:2 FTCA and 6:2 FTS compared with that of PFOA and [76] Ebinding values and number of hydrogen bonds (Nhydrogen bond) between hL-FABP and 6:2 FTCA: - Ebinding: -26.3259 kJ/Mol - Nhydrogen bond: 3 - Hydrogen bond: R122:HE-O15/R122:HH21-O15/H26-S124:OG => except for 6:2 FTS, all PFASs fit well at the binding pocket, indicating direct interactions between these chemicals and hL-FABP [23] Known PFECAs and PFESAs: - P - maybe B - likely T - LRTP [23] on-going (increasing) emissions and high persistence => short-chain PFAAs, PFECAs and PFESAs accumulate in the environment => increasing exposure to these substances over time => even if emissions cease in the future it will take decades to centuries to reverse global contamination of short-chain PFAAs, PFECAs and PFESAs, due to high persistence and environmental mobility [93] 6:2 FTCA not detected in either serum or liver => Rapid elimination rate and nonbioaccumulation for 6:2 FTCA [40] Human: * Liver: - Mean: 92.6 ng/g wet weight - Median: 16.7 ng/g wet weight - Range: 289-BDL ng/g wet weight - MLOD: 4.40 ng/g wet weight - % of detection: 45 => FHEA was PFAS-compound with third highest median concentration in liver * Bone: - Mean: 42.5 ng/g wet weight - Median: 2.0 ng/g wet weight - Range: 494-BDL ng/g wet weight - MLOD: 3.52 ng/g wet weight [40] Human: * Brain: - Mean: 18.6 ng/g wet weight - Median: 2.0 ng/g wet weight - Range: 93.1-BDL ng/g wet weight - MLOD: 4.00 ng/g wet weight - % of detection: 25 * Lung: - Mean: 2.4 ng/g wet weight - Median: 3.9 ng/g wet weight - Range: 3.9-BDL ng/g wet weight - MLOD: 5.54 ng/g wet weight [40] Human: * Kidney: - Mean: 23.7 ng/g wet weight - Median: 1.4 ng/g wet weight - Range: 153-BDL ng/g wet weight - MLOD: 2.85 ng/g wet weight - % of detection: 5 [40] - Smokers showed smaller accumulation of PFASs - Older people (more than 60 years) showed higher concentrations of PFASs => Clear indication that these compounds accumulate after a long-term exposure [93] 6:2 FTCA not detected in either serum or liver => Rapid elimination rate and nonbioaccumulation for 6:2 FTCA [4] Inhalation exposure rats to 6:2 FTOH: 0.5 ppm 5:3 FTCA => Male: 5.2 hrs => Female: 5.2 hrs Inhalation exposure rats tp 6:2 FTOH: 5.0 ppm 5:3 FTCA => Male: 11.7 hrs => Female: 14.7 hrs [4] Inhalation exposure rats to 6:2 FTOH: 0.5 ppm 5:3 FTCA => Male: 0.133 h-1 => Female: 0.133 h-1 Inhalation exposure rats to 6:2 FTOH: 5.0 ppm 5:3 FTCA => Male: 0.059 h-1 => Female: 0.047 h-1 [4] Inhalation exposure rats to 6:2 FTOH: 0.5 ppm 5:3 FTCA => Male: 0.027 lkg-1h-1 => Female: 0.027 lkg-1h-1 Inhalation exposure rats to 6:2 FTOH: 5.0 ppm 5:3 FTCA => Male: 0.012 lkg-1h-1 => Female: 0.009 lkg-1h-1 Scaled: average estimated rat clearance: 0.5 ppm 6:2 FTOH 5:3 FTCA => 0.027 lkg-1h-1 Scaled: average estimated rat clearance: 5 ppm 6:2 FTOH 5:3 FTCA => 0.010 lkg-1h-1 Scaled: predicted human clearance: 0.5 ppm 6:2 FTOH 5:3 FTCA => 0.582 lkg-1h-1 Scaled: predicted human clearance: 5 ppm 6:2 FTOH [4] AUC0- Inhalation exposure rats to 6:2 FTOH: 0.5 ppm 5:3 FTCA => Male: 5395.5 nanomoles/l.hr => Female: 5665.9 nanomoles/l.hr AUC0- Inhalation exposure rats to 6:2 FTOH: 5.0 ppm 5:3 FTCA => Male: 56209.5 nanomoles/l.hr => Female: 58604.9 nanomoles/l.hr Occupational exposure to 6:2 FTOH humans: 5:3 FTCA => 354.0 ng/ml.d [14] - ADONA detected in all samples DE (range 0.32-6.2 g/l) in downstream of wastewater effluent discharges - Average deposition rate of ADONA to surface soil near plant DE between dec '09 and may '10 = 684 ng/(m2d) 211-477-1 647-42-7 3,3,4,4,5,5,6,6,7,7,8,8,8- 6:2 FTOH tridecafluorooctan-1-ol C8 (6 fluorinated) Registered [79] log Kaw: 2.40 (calculated from equations) [65] 6:2 FTOH: logKow = 4.45 [79] log Kow: 1.62 (calculated from equations) [65] 6:2 FTOH: logAqS (aqueous solubility): - Y-Exp: 1.27 mg/L - Y-Pred: 0.87 mg/L - at 24 C: 1.08 mg/L - EPI Suite Pred: -0.56 mg/L [65] 6:2 FTOH: logpL (subcooled vapor pressure): - Y-Exp: -0.46 mm Hg - Y-Pred: -0.78 mm Hg - EPI Suite Pred: -0.02 mm Hg - Y Exp: 1.66 Pa - Y-Pred: 1.34 Pa - Exp: 1.26 Pa - ref40: 2.85 Pa [79] log Koa: 5.44 (calculated from equations) [79] Water solubility (SL): log (SL/mol L^-1): -3.57 (calculated from equations) [79] Vapor pressure (PL): log (PL/Pa): 1.58 (calculated from equations) [18] 6:2 FTOH: logKoc of 2.5 [79] - Molar mass (MM): 364.1 g/mol - Molar volume (VM): 229.2 cm^3/mol (by ACD/Labs' ACD/PhysChem Suite) - Total surface area (TSA): 366.8 A^2 (by ChemAxon) - Tm (experimental melting point): 240 K (data obtained from ChemSpider) 248-580-6 2 7 6 1 9 -9 7 -2 3,3,4,4,5,5,6,6,7,7,8,8,8- 6:2 FTS tridecafluorooctanesulph onic acid C8 (6 fluorinated) Registered [18] 6:2 FTS: - logKoc sediment: 2.2-4.4 - logKoc: ~2.9 - logKoc: 2.6 261-818-3 59587-38-1 Potassium 3,3,4,4,5,5,6,6,7,7,8,8,8tridecafluorooctanesulph onate K-6:2 FTS C8 (6 fluorinated) Registered 700-183-3 52299-25-9 bis(nonafluorobutyl)phos C4/C4 phinic acid PFPiA C8 (2 times 4 Registered fluorinated) 279-481-6 80475-32-7 N-[3(dimethylamino)propyl]3,3,4,4,5,5,6,6,7,7,8,8,8tridecafluorooctanesulph onamide N-oxide Forafac C13 (6 fluorinated) Registered [75] logKaw: 10.57 (estimated by COSMOtherm) => lower Kaw than PFOA [75] logKow (dry): 0.92 (estimated by COSMOtherm) [75] logKlipw: 1.04 (estimated with regression equation) 252-046-8 3 4 4 5 5 -2 9 -3 Carb o xym eth yl d i m eth yl -3 [[(3,3,4,4,5,5,6,6,7,7,8,8 ,8tridecafluorooctyl)sulpho nyl]amino]propylammoni um hydroxide 6:2 FTAB C15 (6 / fluorinated) [75] logKoa = 11.49 (estimated by COSMOtherm) [75] pKa = 13.27 (estimated by SPARC) [75] logKoc: 1.78 / / 6:2 dipolyfluorinated phosphate ester 6:2 diPAP C28 (2 times / 6 fluorinated) [17] of the total amount of PFHxA released from WWTPs, 1.0-14% can be related to degradation from diPAPs (6:2 diPAP degrades to 6:2 FTOH) and 6:2 FTS during waste water treatment [92] - No conclusive evidence for biodegradation of PFAS in sludge under aerobic conditions over a period of up to either 9 or 15 weeks - Although decreases in PFHxA, 6:2 FTOH, and 8:2 FTOH concentrations were observed, their concentrations in the control bottles also decreased, and it is therefore not possible to confirm that they were indeed due to biodegradation - The concentration decrease could also be due to a non-biological degradation process, losses not due to degradation or even to incomplete sterilization of the control bottles [53] Molar yield comparison of stable transformation products of 6:2 FTOH: * aerobic sludge (28-60d): - 5:3 Acid: 14 - PFBA: ND - PFPeA: 4.4 - PFHxA: 11 - PFHpA: ND * aerobic sediment (90d): - 5:3 Acid: 22 - PFBA: 1.5 - PFPeA: 10 - PFHxA: 8.4 [53] Molar yield comparison of stable transformation products of 6:2 FTOH: * aerobic soil (180d): - 5:3 Acid: 15 - PFBA: 1.8 - PFPeA: 30 - PFHxA: 8.1 - PFHpA: ND * anaerobic sludge (90-176d): - 5:3 Acid: 21 - PFBA: ND - PFPeA: ND - PFHxA: <0.4 [53] Molar yield comparison of stable transformation products of 6:2 FTOH: * anaerobic sediment (100d): - 5:3 Acid: 12 - PFBA: ND - PFPeA: ND - PFHxA: 0.6 - PFHpA: ND [92] - No evidence for degradation of any of the PFAS was observed under anaerobic conditions - This result is consistent with the lack of anaerobic biodegradation of PFAS reported elsewhere in the literature => The PFAS tested in these experiments are non-biodegradable under the experimental conditions used in this study, despite using municipal sewage sludge, which presumably has a history of exposure to PFAS [49] - Activated carbon filtration and an anthracite/ sand combination led to an increase in the concentrations of certain PFCAs and a decrease in 6:2 FTS concentration, suggesting microbial metabolization of FTs into PFCAs during this treatment - after ozonation, a significant increase of the total PFAS concentration was observed, which was mainly linked to an increase of the 6:2 FTS concentration - only nanofiltration was able to remove efficiently all the analyzed PFASs from raw water [17] of the total amount of PFHxA released from WWTPs, 1.0-14% can be related to degradation from diPAPs and 6:2 FTS during waste water treatment [49] 6:2 FTSA was also detected in drinking water sources, suggesting that it has some mobility and sufficient stability in the alluvial aquifer to reach some wells [71] - Historically, longer chain PFASs and PFAS precursors were used as the main components in AFFF - Today, longer chain PFASs have been replaced by 6:2 fluorotelomers which can - when released into the environment - transform into shorter chain PFAAs [53] Molar yield comparison of stable transformation products of 6:2 FTS: * aerobic sediment (90d): - 5:3 Acid: 16 - PFBA: ND - PFPeA: 21 - PFHxA: 20 - PFHpA: 0.55 * anaerobic sediment (100d): - 5:3 Acid: ND - PFBA: ND - PFPeA: ND - PFHxA: ND - PFHpA: ND [54] Molar yield comparison of stable transformation products of 6:2 FTS: * aerobic sludge (90d): - 5:3 Acid: 0.12 - PFBA: 0.14 - PFPeA: 1.5 - PFHxA: 1.1 - PFHpA: ND + degradation scheme in article: - major pathway leading to PFPeA and PFHxA - minor pathway leading to 5:3 Acid [92] - No conclusive evidence for biodegradation of PFAS in sludge under aerobic conditions over a period of up to either 9 or 15 weeks - Although decreases in PFHxA, 6:2 FTOH, and 8:2 FTOH concentrations were observed, their concentrations in the control bottles also decreased, and it is therefore not possible to confirm that they were indeed due to biodegradation - The concentration decrease could also be due to a non-biological degradation process, losses not due to degradation or even to incomplete [92] - No evidence for degradation of any of the PFAS was observed under anaerobic conditions - This result is consistent with the lack of anaerobic biodegradation of PFAS reported elsewhere in the literature => The PFAS tested in these experiments are non-biodegradable under the experimental conditions used in this study, despite using municipal sewage sludge, which presumably has a history of exposure to PFAS [59] Electrochemical oxidation of 6:2 FTS: degradation scheme in article: - 1 pathway leading to C5F11 + CO2 - 1 pathway leading to C5F11COO- (and decomposition into shorter-chain perfluorocarboxylic acids) [60] Heat-activated persulfate oxidation of 6:2 FTS: Both the ethyl linkage and CF2-CH2 bond of 6:2 FTS oxidize simultaneously, resulting in ~25% PFHpA and 75% PFHxA [91] Zebrafish embryos were treated with different concentrations of 6:2 FTCA (0, 4, 8 and 12 mg/L) from 6 to 120 hpf (hours post fertilization), with three replicates of the control and each treatment group Results: - 6:2 FTCA exposure influenced the development of zebrafish embryos, including a reduction in the hatching and survival percentage, a decrease in heart rate, and an increase in the occurrence of cardiac malformations - 6:2 FTCA exposure decreased the number of erythrocytes in a concentration-dependent manner - 6:2 FTCA exposure affected a number of erythrocyte-related genes => 6:2 FTCA exposure might affect heme biosynthesis and degradation during early development of zebrafish embryos [91] - The median LC50 of 6:2 FTCA on zebrafish embryos was 25.1 1.5 mg/L at 72 hpf and 7.33 0.50 mg/L at 120 hpf - The 120h LC50-value was lower than that for PFOA with the same carbon chain length (>500 mg/L), but higher than that for PFOS (2.20 mg/L) => 6:2 FTCA is more toxic than PFOA, but less so than PFOS [93] - Treatment of male CD1 mice with 5 mg/kg/day 6:2 FTCA for 28 days consecutively => killed at the end of treatment - No significant changes observed in neither liver weight or relative liver weight - No statistical differences in the major biochemical indices were observed, indicating no liver damage occurrence - Only a slight increase in liver T-CHO [93] - No significant difference in cellular structure of hepatic tissue could be observed - No changes in cell numbers were observed - 6:2 FTCA only induced the upregulation of 16 genes and downregulation of 23 genes, with no significant change in metabolism-related genes observed - Only TNF in liver increased significantly - Inflammation-related protein levels were increased significantly [44] All tested PFCs induced hPXR reporter activites in a dosedependent manner => In general, longer carbon chain length seems to be favorable for the hPXR activity of PFAAs with chain length shorter than 10, as chain length is longer than 10, increasing the chain length of PFAA seems to render the chemicals weaker potency to induce hPXR activity => 6:2 FTOH: EC50 = 9.87 M, maximum induction = 21.2 % => Activation of hPXR: All docked compounds showed a similar pose, in which the polar terminal group (carbonyl group, hydroxyl group or sulfo group) of the PFCs interacted with Ser247 through a hydrogen bonding, and fluorinated carbon chain inserted itself into a hydrophobic domain comprising Met243, Met246, Phe288, Trp299, Tyr306 and Val211 => Agonistic activity depends on interactions of PFCs with hydrophobic residues in the binding pocket => 6:2 FTOH shortens the distance between the edge of its main chain with face of Phe288, Trp299 and Tyr306, increasing its agonistic activity towards hPXR ==> PXR induction is postulated to be associated with lipid homeostasis, atherosclerosis, carcinogenesis and endocrine- [46] PFASs can compete with thyroxine (T4) for binding to the human thyroid hormone transport protein transthyretin (TTR), which may lead to reduced thyroid hormone levels leading to endocrine disrupting adverse effects => 6:2 FTOH: T4-TTR Binding (%): 117 => Experimental Classification: Inactive (N) => Predicted Classification: Inactive (N) => Binding energy: -1.10 kcal/mol => PFASs can compete with thyroxine (T4) for binding to the human thyroid hormone transport protein transthyretin (TTR), which may lead to reduced thyroid hormone levels leading to endocrine disrupting adverse effects [46] Discriminating properties between active (Y) and inactive (N) classes: - Hydrophobicity of the H atom attached to heteroatom - Contributions from the H-050 descriptor Fluorotelomer alcohols: - Fluorotelomer alcohols are inactive, irrespetive of their carbon chain length => 6:2 FTOH is not toxic/inactive [50] - NCI-H295R human adrenal corticocarcinoma cell line: exposure to 6:2 FTOH caused a significant increase in 17-estradiol levels => 6:2 FTOH exhibited effects at 6.3 M (p<0.01), 12.5 M (p<0.001), 25 M (p<0.001) and 50 M (p<0.01) - human ovarian adenocarcinoma cell line: 6:2 FTOH led to increased estrogen receptor (ER) activity (a.o. at 100 M, p<0.001) [51] - IC50 value and dissociation constant (Kd) with hPPAR-LBD: 6:2 FTOH: IC50: ND, Kd: ND - 6:2 FTOH had no hPPAR activation potency even at a high concentration of 200 M [84] PFCs can compete with thyroxine (T4, the transport form of thyroid hormone), for binding to the human thyroid hormone transport protein transthyretin (TTR) => such competitive capacity may lead to decreased thyroid hormone levels as previously reported for animals exposed to PFCs => PFCs do not affect the regulatory functions of the thyroid hormone system itself, but it is the competitive binding to transport proteins that alters the free thyroxine (T4) levels in blood => TTR is the main T4 carrier in cerebrospinal fluid, and also important in serum of most mammalian species and birds => TTR is both in humans and in rodents the most important carrier protein for thyroid hormone to the developing fetus and the brain [56] Data indicate that K-6:2 FTS (6:2 FTS) is not bioaccumulative in aquatic organisms [84] 6:2 FTOH: - Molecular weight: 364.1 g/mol - T4-TTR binding at maximum concentration (10M): 117% - IC50 (concentration at 50% inhibition): N.D. - Slope of dose-response curve: / - T4-REP (relative potency compared to T4) factor: N.D. - HPLC retention time: 34.4 - Purity: 98 [86] - weak binding of 6:2 FTOH with the protein TR-LBD (TR = thyroid hormone receptor, LBD = ligand-binding domain) * Binding of 6:2 FTOH with TR-LBD: - Length: 12.27 A - IC50: ND - RP (relative potency (IC50 T3 / IC50 chemical)): ND - Hydrogen bonding: / * all tested PFCs fit into the T3-binding pocket of TR-LBD, with the acid or hydroxyl end group residing toward the inner part and the [87] * Binding potency of 6:2 FTOH to TTR (TTR competitive binding assay): - IC50: ND - Kd: ND - RP (relative potency (IC50 T4 / IC50 chemical)): ND => perfluoroalkyl acids: Kd values decreased as carbon chain length increased from C4 to C8 => no further increase, but a downward trend was observed for perfluoroalkyl acids with longer chain lengths (C9-C14) * Binding potency of 6:2 FTOH to TTRmutK15G: - IC50: ND - Kd: ND => PFASs exhibited much weaker binding affinities to TTRmutK15G compared with wild-type TTR * Docking study: - Similar to T4, PFOA and PFOS could nearly fill the TTR ligandbinding pocket - Perfluoroalkyl acids with carbon chain length less than C8 did not adequately fill the T4 binding pocket [87] * Binding potency of 6:2 FTOH to TBG (TBG competitive binding assay): - IC50: ND - Kd: ND - RP (relative potency (IC50 T4 / IC50 chemical)): ND => only PFTA and PFTdA bound to TBG * Binding potency of 6:2 FTOH to TBGmutR378G: - IC50: ND - Kd: ND * Binding potency of 6:2 FTOH to TBGmutR381G: - IC50: ND - Kd: ND => both TBGmut378G and TBGmutR381G exhibited much weaker binding potencies to PFTA and PFTdA when compared with that of wild-type TBG * Docking study: - Perfluoroalkyl acids with carbon chain length less than C12 did not adequately fill the T4 binding pocket - Longer fluorinated carbon chain structures could nearly fill the TBG [90] * Hazard classification for human health endpoints of 6:2 FTOH is Category 4 for acute oral toxicity based on an LD50 of 1750 mg/kg Other acute health endpoints, including eye and skin irrita on, skin sensitization, as well as genotoxicity, did not meet the criteria for hazard classification * Based on the results of the 90-day subchronic study, the NOAEL for 6:2 FTOH is 5 mg/kg/day This NOAEL is due to effects observed at 25 mg/kg/day and higher, including changes in clinical signs, hematological, clinical chemistry and urinalysis parameters and various histopathological effects in the liver and kidney [95] - Yeast cells were modified by incorporation of hER (human estrogen receptor) isoforms (hER or hER) - Study of estrogenic effects of FTOHs for the hER and hER using a yeast two-hybrid assay - Treatments with 6:2 FTOH, 8:2 FTOH and NFDH dose-dependently induced hER-mediated transcriptional activity with interaction between the hER or hER ligand binding domain and TIF2, whereas no activation of hERs was observed when treated with PFOS and PFOA - 6:2 FTOH - hER: ECx10 = 2.3 M, relative activity = 3.7x10^(-3) - 6:2 FTOH - hER: ECx10 = 4.1 M, relative activity = 2.5x10^(-3) [95] - The relative ranks of tested chemicals on the estrogenic effects for hERs descended in the order of estradiol-17 >>> 6:2 FTOH > NFDH > 8:2 FTOH - These results suggest that FTOHs including 6:2 FTOH, 8:2 FTOH and NFDH interact with hER isoforms and in vitro, and that certain FTOHs may be associated with potential biological effects via the ER signaling pathway in humans - Overall ECx10 values of FTOHs on hER activation were lower than those for hER, indicating a differential responsiveness of hERs to [96] - Exposure of MCF-7 human Caucasian breast adenocarcinoma cells to E2 or test compounds - 6:2 FTOH and 8:2 FTOH behave like xenoestrogens in vitro - These compounds clearly induce cell proliferation at 10 M, the concentration at which many other xenoestrogens are also active - Although growth-arrested MCF-7 cells are predominantly in the G0/G1-phase of the cell cycle, addition of (xeno)estrogens makes cells proliferate again, shown by the marked increase in the percentage of cells in the S-phase after 24 hr of exposure - The increase of cell numbers in the S-phase was 31% for 6:2 FTOH [96] - Altered expression of estrogen-responsive genes: * Upregulation of TFF1 mRNA by 6:2 FTOH (6.2x) * Upregulation of PGR mRNA by 6:2 FTOH (10.4x) * Upregulation of ESR1 mRNA by 6:2 FTOH (2.2x) * Upregulation of PDZK1 mRNA by 6:2 FTOH (5.4x) * Downregulation of ERBB2 by 6:2 FTOH (2.4x) - 6:2 FTOH was characterized as a stronger xenoestrogen than 8:2 FTOH [56] Acute and chronic aquatic hazard endpoints indicate 6:2 FTS is not classified for aquatic hazard according to GHS or European CLP legislation [98] Beijing prenatal exposure study (human): Correlations between fetal 6:2 FTS and maternal thyroid hormones (adjusted for influential covariates): Not significant for 6:2 FTS [93] 6:2 FTSA showed a tendency to accumulate more easily in the liver, with a concentration 10 times greater than that in serum [76] inhibitory effect of 6:2 FTS on human liver HL-7702 cells: - Model: BiPhasic - ^2: 8.6137 - R^2: 0.9985 - IC50 : 3.54 x 10^(-4) => 6:2 FTS showed a lower IC50 value compared with that of PFOA and PFOS, suggesting stronger effects on cell viability [97] 6:2 FTS was also found in fish liver (from a lake receiving water from a fire-fighting training area) suggesting a bioaccumulation, however the concentration could not be calculated due to signal enhancement and lack of labelled internal standard [76] binding affinity of 6:2 FTS to hL-FABP (human liver fatty acid binding protein): - IC50: 78.97 +/- 0.56 M - Kd: 345.54 +/- 2.18 M - R^2: 0.996 => Kd values of 6:2 FTCA and 6:2 FTS were two orders of magnitude higher than those of PFOA and PFOS, indicating weak binding capacity for 6:2 FTCA and 6:2 FTS compared with that of PFOA and PFOS [76] Ebinding values and number of hydrogen bonds (Nhydrogen bond) between hL-FABP and 6:2 FTS: - Ebinding: No binding - Nhydrogen bond: / - Hydrogen bond: No binding => except for 6:2 FTS, all PFASs fit well at the binding pocket, indicating direct interactions between these chemicals and hL-FABP [93] - Treatment of male CD1 mice with 5 mg/kg/day 6:2 FTSA for 28 days consecutively => killed at the end of treatment - Absolute liver weights were significantly increased by 19% and relative liver weights increased by 22% in animals exposed to 6:2 FTSA - AST and ALB levels significantly increased compared to the control, while levels of ALT and ALP showed nonsignificant increases, indicating that 6:2 FTSA induced early liver damage in mice [61] Degradation scheme in article of 6:2 FTS-K in UV/H2O2 system: - Substitution of hydrogen atom by hydroxyl radicals - Desulfonation - Carboxylation (formation of PFHxA and PFHpA) - Sequential flake-off of CF2 unit until F- + CO2 + H2O remains [69] PFPiAs: - hydrolyze to yield PFPAs when heated or alkalized - biotransform to yield PFPAs in rainbow trout and possibly also rats => likely P (high persistence of its transformation products: PFPAs and possibly PFCAs) [75] - half-life in air with AOPWIN: 11.23 hour - half-life in water with BIOWIN3: 17280 hour - half-life in soil with BIOWIN3: 17280 hour [75] overall persistence (Pov): 1038.74 days => same Pov as PFOA/PFOS [49] - for the fourth sampling campaign, 6:2 FTAB was no longer detectable after ozonation => disappearance is due to a reaction between ozone and 6:2 FTAB, leading to the formation of 6:2 FTS - only nanofiltration was able to remove efficiently all the analyzed PFASs from raw water [36] * The main fluorosurfactant in Forafac 1157 is a 6:2 fluorotelomer sulfonamide alkylbetaine (6:2 FTAB) => Contains a 6:2 fluorotelomer sulfonamide moiety [F3C(CF2)5CH2CH2SO2NH-] and a betaine moiety [-N(R,R')+CH2COO-], separated by an alkyl spacer * Observed degradation products: - A tertiary N,N-dimethylamine - A secondary amine - A primary amine - An N-demethylated species of the native 6:2 FTAB - 6:2 fluorotelomer sulfonamide [57] Electrochemical degradation: 6:2 FTAB was partially degraded into 6:2 FTS which in these conditions would be simultaneously generated and broken => 6:2 FTS degradation would start with the attack of hydroxyl radical at the positions of the two unfluorinated carbons => Further reaction with hydroxyl radical caused the desulfonation, cleaving the bond between the end group of the sulfonate and the polyfluorinated tail => After that, the polyfluorinated tail could be carboxylized at the end to form PFHpA and PFHxA [57] => These PFCAs were broken into shorter-chain perfluorocarboxylates through a step by step mechanism that involved the loss of one electron to the anode to form a perfluoroalkyl carboxyl radical, which was later decarboxylated and defluorinated by hydroxyl radical mediated reaction => In each step, the PFCA molecule loses a CF2 unit, and generates CO2 and fluoride ions [58] Boron doped diamond electrooxidation: - Fluorotelomer sulfonamides 6:2 FTAB and M4 would be degraded into 6:2 FTS, which conversely would give rise to PFHpA and preferentially PFHxA - The latter PFCAs were transformed into shorter-chain PFCAs and eventually into CO2 and fluoride (Elaborate degradation scheme from 6:2 FTAB and M4 to CO2 + F- drawn [67] Degradation and defluorination of 6:2 FTAB by Gordonia sp. strain NB4-1Y under sulfur-limiting conditions (elaborate degradation scheme in article): NB4-1Y appears to utilize two degradation pathways: - Major pathway: Conversion of 6:2 FTUA to 5:2 ketone => end products PFHxA and PFPeA - Minor pathway: Conversion of 6:2 FTUA to 5:3 FTCA => end products [17] of the total amount of PFHxA released from WWTPs, 1.0-14% can be related to degradation from diPAPs (6:2 diPAP degrades to 6:2 FTOH) and 6:2 FTS during waste water treatment [93] - In comparison with the normal cellular structure of hepatic tissue in the control group, early liver damage (necrosis) was observed in the 6:2 FTSA group - Nucleus counting results showed that 6:2 FTSA caused a significant decrease in cell numbers in homolographic visions, indicating obvious hepatocellular hypertrophy - 217 genes were upregulated and 194 genes were downregulated, among which many inflammation- or necrosis-related genes were changed significantly - PPAR was significantly upregulated in the livers - Significant increases in TNF, IL-10, and IL-6 levels in the livers => 6:2 FTSA induced an inflammatory reaction [98] Beijing prenatal exposure study (human): Correlations between maternal 6:2 FTS and maternal thyroid hormones (adjusted for influential covariates): - Maternal FT3: 0.160 (p<0.05) - Maternal FT4: 0.164 (p<0.05) - Maternal T3: 0.205 (p<0.05) - Maternal T4: 0.172 (p<0.05) - Maternal TSH: Not significant => 6:2 FTS is positively correlated with maternal FT3, FT4, T3 and T4 [56] Bioconcentration and dietary bioaccumulation test in Oncorhynchus mykiss : - All bioconcentration factors (BCF) were <40 - Dietary assimilation effiency = 0.435d - Growth-corrected half-life = 23.1d - Dietary biomagnification factor (BMF) = 0.295 => These data indicate that K-6:2 FTS (6:2 FTS) is not bioaccumulative in aquatic organisms [69] PFPiAs: - bind to proteins - elimination half-lives increase with increasing perfluoroalkyl chain length - certain homologs (C6/C6) have long elimination half-life from rainbow trout and rat blood, similar to PFOA and PFOS => some homologs may be B [56] Acute and chronic aquatic hazard endpoints indicate K-6:2 FTS is not classified for aquatic hazard according to GHS or European CLP legislation [69] PFPiAs: - C4/C4 PFPiA shows acute toxicity to rats that is similar to that of PFOS and much higher than for PFOA and PFBS - mixtures of PFPAs and PFPiAs both have modes-of-action that are similar to, or different from, those of PFCAs and PFSAs - current information is not conclusive => T undetermined [55] Developmental toxicity in zebrafish embryos: - 72 hpf: LC50 = 67.12 +/- 3.82 mg/L 96 hpf: LC50 = 64.39 +/- 4.23 mg/L 120 hpf: LC50 = 43.73 +/- 3.24 mg/L - Main morphological defect occurred in the tail region, with the formation of rough-edged skin/fins in the 20 and 40 mg/L 6:2 FTAB-treated groups + 10% and 23.3% of embryos exhibited the malformation type of uninflated swim bladders in 20 and 40 mg/L 6:2 FTAB treated groups at 120 hpf respectively Embryos developed normally in the control group - A large number of apoptotic cells were observed in the tail area in the 40 mg/L 6:2 FTAB treated embryos At 96 hpf no obvious apoptotic cells were detected in the control group => Cell apoptosis might explain the observed tail malformations of embryos after 6:2 FTAB exposure - Transcriptional levels of p53 and apaf1 (related to apoptosis) were significantly up-regulated in the 40 mg/L 6:2 FTAB-treated [55] - Transciptional level of bax was upregulated in 20 and 40 mg/L treated groups - Transcriptional level of mdm2 increased in the 10, 20 and 40 mg/L 6:2 FTAB treated groups - Caspase enzyme activities increased with 10, 20 and 40 mg/L 6:2 FTAB exposure => 6:2 FTAB-induced cell apoptosis was likely through p53 and tnf activation, leading to caspase-dependent cell apoptosis, and thus embryonic malformation in the tail region - ROS and MDA levels were sign. higher in the 20 and 40 mg/L esposure groups => Embryos sustained serious oxidative stress - Transcriptional level of ccl1 was downregulated in a concentration-dependent manner - Transcriptional levels of il-1, il-8 and tnf were sign. upregulated after exposure to 10, 20 and 40 mg/L of 6:2 FTAB - Compared with the control, sign. increases in cxcl-c1c and ifn mRNA expression were observed after exposure to 40 mg/L of 6:2 FTAB [116] In May 2009, indoor dust samples (n=17) were collected from fifteen homes located in nine Korean cities for house dust (n=15), and one shopping mall and one library in Seoul for non-residential indoor dust (n=2). Nine cities for house dust samples were chosen to represent densely populated cities, industrialized cities, and rural area. * Concentrations (ng/g) of 6:2 FTOH measured in indoor dust samples from Korea (n=17): - House dust (n=15): Range: <LOD - 12.7 Mean ( SD): 3.7 ( 4.1) Median: 3.2 % of < LOD: 47 - Non-residential indoor dust (n=2): Shopping mall: < LOD Library: 65.5 [4] 6:2 FTOH metabolized to form transient intermediates: 6:2 FTAL, 6:2 FTCA and 6:2 FTUCA (phase I metabolism) => 2 possible ways: - Parent 6:2 FTOH and intermediates can undergo phase II competitive conjugation reactions to form glutathione, glucoronide and sulfate conjugates => can undergo phase III metabolism to be eliminated mainly in the urine and to some extent in the bile - Transient intermediates may follow another metabolic pathway to form terminal products that include PFCAs (such as PFBA, PFPeA, PFHxA, PFHpA, 5:3 FTCA and 4:3 FTCA, also representative of phase I metabolism) => hypothesized that terminal products lead to a long-term internal exposure that contributes to systemic toxicity of 6:2 FTOH [116] * There were no significant correlations between PFCAs (or PFSAs) and their precursors in house dust. * In Korean house dust samples, significant correlations (p<0.05) were only found among all the FTOHs and among almost all of the PFCAs. [34] significant correlation with other PFAS-analytes detected in surface water (p<0.01), indicating a common or similar source of exposure [88] 50 paired human maternal and cord serum samples were analyzed for 15 PFAA precursors => 5 PFAA precursors were detected (a.o. 6:2 FTS) => 6:2 FTS had the highest levels and was detected in a similar number of samples in both maternal and cord serum *Maternal serum: - Percentage > LOD: 64% - Mean SD: 13.39 12.78 pg/mL - Range: <3.00 - 48.24 pg/mL *Cord serum: - Percentage > LOD: 68% - Mean SD: 20.99 22.67 pg/mL - Range: <3.00 - 90.21 pg/mL => No sign. differences and no correlation between [88] => lack of correlation: maternal 6:2 FTS might not have been the only source for cord serum 6:2 FTS (placenta-enzyme mediated biodegradation of fluorotelomer surfactants or other 6:2 FTS precursors as source) *Placental transfer ratio: - Mean CS/MS SD: 1.85 1.59:1 => the 5 detected PFAA precursors were found to be transferred from mother into fetus through the placental barrier with different placental transfer ratios => among the detected PFAA precursors, only cord serum 6:2 FTS differed across the 3 maternal age groups, with the highest concentrations in the youngest age group [88] concentration of 6:2 FTS in cord sera decreased as mothers' age increased (might be related to decreased bioactivities of placental enzymes/ transport systems as women become older) [93] 6:2 FTSA (6:2 fluorotelomer sulfonic acid) concentration serum: 18.52 g/mL [98] Beijing prenatal exposure study (human): 6:2 FTS: 157 paired maternal and cord serum samples around delivery: * Maternal serum (pg/mL): - Percentage > LOD: 82% - Mean SD: 48.38 40.94 - Median: 41.44 - Geometric mean: 26.04 - Range: <LOD-203.00 * Cord serum (pg/mL): - Percentage > LOD: 90% - Mean SD: 33.54 38.52 - Median: 24.68 - Geometric mean: 19.98 - Range: <LOD-252.12 - Maternal-fetal correlation: r=0.336, p<0.001 * Maternal and cord serum concentration ratio: - Mean CS SD:MS: 3.08 4.34:1 [100] 54 pooled human serum samples from 4920 individual samples (AU): * 6:2 FTSA concentrations in 2011, by age group: - 0-4 y (n=4): FOD: 50%, Min: <LOD, Max: 0.02 ng/mL, Mean: 0.01 ng/mL - 5-15 y (n=4): FOD: 0%, Min: <LOD, Max: /, Mean: / - 16-30 y (n=4): FOD: 25%, Min: <LOD, Max: 0.05 ng/mL, Mean: 0.01 ng/mL - 31-45 y (n=4): FOD: 25%, Min: <LOD, Max: 0.03 ng/mL, Mean: 0.01 ng/mL - 46-60 y (n=4): FOD: 0%, Min: <LOD, Max: / , [100] 54 pooled human serum samples from 4920 individual samples (AU): * 6:2 diPAP concentrations in 2011, by age group: - 0-4 y (n=4): FOD: 75%, Min: <LOD, Max: 0.21 ng/mL, Mean: 0.15 ng/mL - 5-15 y (n=4): FOD: 50%, Min: <LOD, Max: 0.14 ng/mL, Mean: 0.05 ng/mL - 16-30 y (n=4): FOD: 75%, Min: <LOD, Max: 0.22 ng/mL, Mean: 0.10 ng/mL - 31-45 y (n=4): FOD: 50%, Min: <LOD, Max: 0.09 ng/mL, Mean: 0.04 ng/mL - 46-60 y (n=4): FOD: 50%, Min: <LOD, Max: 0.12 ng/mL, Mean: 0.04 ng/mL - >60 y (n=4): FOD: 100%, Min: 0.10 ng/mL, Max: [100] * 6:2 diPAP concentrations (detected in >50% of samples) in 2011, by gender and age group: - 0-4 y: FEMALE (n=2): 0.19 ng/mL, MALE (n=2): 0.10 ng/mL, ALL (n=4): 0.15 ng/mL - 5-15 y: FEMALE (n=2): <0.15 ng/mL, MALE (n=2): 0.11 ng/mL, ALL (n=4): 0.05 ng/mL - 16-30 y: FEMALE (n=2): <0.14 ng/mL, MALE (n=2): 0.06 ng/mL, ALL (n=4): 0.10 ng/mL - 31-45 y: FEMALE (n=2): 0.03 ng/mL, MALE (n=2): 0.04 ng/mL, ALL (n=4): 0.04 ng/mL - 46-60 y: FEMALE (n=2): 0.06 ng/mL, MALE (n=2): 0.02 ng/mL, ALL (n=4): 0.04 ng/mL [71] Shorter chain PFASs have been shown to absorb into the liver more readily than those with longer chains which are concentrated in the blood proteins [93] 6:2 FTSA concentration liver: 194.44 g/g => 6:2 FTSA showed a tendency to accumulate more easily in the liver, with a concentration 10 times greater than that in serum [18] the influent to an industrial WWTP plant was dominated by 6:2 FTOH (mean ~500 000 ng/L) => in the effluent no FTOHs were observed => volatilization and biodegradation were put forward as explanations for disappearance [12] Groundwater SE: 6:2 FTS detected at areas near firefighting sites, unspecific industry and landfill/ waste disposal sites but not at areas influenced by skiing and urban sources [18] In highly contaminated groundwater 6:2 FTS showed an average concentration of 25 000 ng/L [12] Surface water SE: 6:2 FTS detected at areas influenced by firefighting (9.1% of total PFASs), unspecific industry (6.0% of total PFASs), STP (9.3% of total PFASs) and landfill/ waste disposal (3.9% of total PFASs), but not at skiing and urban areas [49] - 6:2 FTS concentrations ranged from 22 ng/L in October 2013 to 195 ng/L in December 2013 - at sampling point 4 there was a decrease in concentration => likely due to the fact that WWTP effluent is not thoroughly mixed with the river water until sampling point 4 - within sampling points 4 and 11, observed 6:2 FTS concentrations were relatively constant => no major environmental sink (volatilization or adsorption onto sediment) nor transformation that could have affected 6:2 FTS between [49] - opposite pattern for PFCA mass flows: systematically higher in the river at sampling point 4 than in WWTP effluent => 6:2 FTS and 6:2 FTAB partially adsorbed onto sediments and trandsformed into PFCA in the river [12] Drinking water SE: 6:2 FTS 4.5 % of total PFASs [17] - 6:2 FTS was found in filtered effluent water of all three WWTPs (mean 3.1 ng/L) - levels of 6:2 FTS were similar in influent compared to effluent water in all three WWTPs [18] 6:2 FTS measured downstream of a WWTP of a PFAS production facility => concentrations decreased from 200 to 100 ng/L over a distance of 62 km [49] - although 6:2 FTS concentrations were consistently higher in the river, PFCA concentrations were consistently higher in alluviate well F6 + shortchain PFASs were only detected in well F6 => suggests transformation of PFCA precursors, such as 6:2 FTS and 6:2 FTAB, in the sediments of the alluvial groundwater, leading to endstage metabolites (PFCAs) in well F6 [18] SE: 6:2 FTS was found to be the major PFAS in outgoing water from WWTPs (n=20) with a median concentration of around 10 ng/L [71] among the PFAS precursor compounds analyzed in AFFF used in Sweden, 6:2 FTS is frequently detected in high concentrations and AFFF contaminated areas [119] 3 replicate aqueous and solid samples were collected from each of 19 Australian WWTPs throughout 2017. Samples were taken from various stages within the treatment train from a range of WWTPs to determine the trends in the mass flux and partitioning of PFAS within the sampled WWTPs. Summary statistics for pooled aqueous (n=201, triplicates from 67 individual locations within 19 WWTPs) and pooled solid (n=51, triplicates from 5 primary and secondary sludge locations, 6 lagoon sludges and a lagoon dredge pile) samples. [97] - soil at the fire-fighting training ground (NO): 612 ng/g DW 6:2 FTS - soil 10-20 m from the training gound (NO): 2101 ng/g DW 6:2 FTS => Levels thereafter decreased with increasing distance => Movement of the chemicals not only vertically, but also horizontally - seepage water (water that has seeped through an oil separator): 5110-6693 ng/L 6:2 FTS => 6:2 FTS was quantified at up to 25% lower concentrations in the diluted samples - sediment from three locations in the lake receiving water from the training area (NO): on average 7 ng/g DW 6:2 FTS [119] * 6:2 FTS concentration in aqueous samples (ng/L): - Median: 2.4 - Mean: 7.3 - s.d.: 12 - min: <LOD - max: 61 - Detect (%): 99% * 6:2 FTS concentration in solid samples (ng/g dw): - Median: <LOD - Mean: 0.26 - s.d.: 0.69 - min: <LOD - max: 2.7 - Detect (%): 26% [49] - although monitored only in December 2013, 6:2 FTAB was detected in the river from sampling point 3 at a high concentration (968 ng/L) - in December 2013 6:2 FTAB concentration pattern along the river followed a trend similar to that of 6:2 FTS, but at higher concentration levels - at 62 km from the industrial WWTP (sampling point 11), the 6:2 FTAB concentration was still 248 ng/L - opposite pattern for PFCA mass flows: systematically higher in the river at sampling point 4 than in WWTP effluent => 6:2 FTS and 6:2 FTAB partially adsorbed onto sediments and [49] - although 6:2 FTS concentrations were consistently higher in the river, PFCA concentrations were consistently higher in alluviate well F6 + shortchain PFASs were only detected in well F6 => suggests transformation of PFCA precursors, such as 6:2 FTS and 6:2 FTAB, in the sediments of the alluvial groundwater, leading to endstage metabolites (PFCAs) in well F6 [18] 6:2 FTAB measured downstream of a WWTP of a PFAS production facility => concentrations decreased from nearly 1000 to 250 ng/L over a distance of 62 km => ozonation led to removal of 6:2 FTAB by oxidation into 6:2 FTS [17] - 6:2 diPAP could not be detected above LOD in any effluent water - concentration of 6:2 diPAP in influent water = 7.6 ng/L - mean 6:2 diPAP concentration in sludge = 2.0 ng/g / / 6:2 chlorinated 6:2 Cl- C8 / polyfluorinated ether PFESA sulfonate Blue = substance not specifically mentioned, but mentioned as part of a group of substances [108] Analysis of 424 mother-fetus pairs from the Maoming Birth Cohort, China. * Concentrations of 6:2 Cl-PFESA in maternal and cord serum: - Cord serum: Detection rate: 99.53%, median: 0.32 ng/mL, mean SD: 0.38 0.33 ng/mL - Maternal serum: Detection rate: 99.76%, median: 0.63 ng/mL, mean SD: 0.80 0.57 ng/mL * PFAS in cord serum was positively correlated with PFAS in maternal serum for all PFAS (p<0.05) * Transplacental transfer efficiency (TPT) of 6:2 Cl-PFESA = concentration cord serum (ng/mL)/ concentration maternal serum (ng/mL): - TPT (n=421): Median: 0.46, mean SD: 0.49 0.26 - U-shaped pattern for TPT in PFCAs: TPT decreased from PFBA to PFDA and then increased to PFTrDA * Higher TPT in PFAS alternatives than PFASs => Cl-PFESAs with an additional oxygen atom in the carbon chain and the attachment of a chlorine atom may be transferred to the fetus more efficiently than PFOS and PFOA => PFAS alternatives may be more easily transported from mother to infant than conventional PFASs in uterus * PFBA and 6:2 Cl-PFESA are the second and fourth highest PFAS [100] * 6:2 diPAP concentrations (detected in >50% of samples) in 2011, by year: - FEMALE: 2002 (n=2): 0.02 ng/mL, 2004 (n=2): <0.03 ng/mL, 2006 (n=4): 0.02 ng/mL, 2008 (n=4): <0.32 ng/mL, 2011 (n=4): 0.09 ng/mL, 2013 (n=1): <0.32 ng/mL - MALE: 2002 (n=2): 0.07 ng/mL, 2004 (n=4): 0.06 ng/mL, 2006 (n=4): 0.04 ng/mL, 2008 (n=4): 0.02 ng/mL, 2011 (n=4): 0.11 ng/mL, 2013 (n=3): 0.12 ng/mL - ALL: 2002 (n=4): 0.05 ng/mL, 2004 (n=6): 0.04 ng/mL, 2006 (n=8): 0.03 ng/mL, 2008 (n=8): 0.01 ng/mL, 2011 (n=8): 0.10 ng/mL, 2013 (n=4): 0.09 ng/mL [119] 3 replicate aqueous and solid samples were collected from each of 19 Australian WWTPs throughout 2017. Samples were taken from various stages within the treatment train from a range of WWTPs to determine the trends in the mass flux and partitioning of PFAS within the sampled WWTPs. Summary statistics for pooled aqueous (n=201, triplicates from 67 individual locations within 19 WWTPs) and pooled solid (n=51, triplicates from 5 primary and secondary sludge locations, 6 lagoon sludges and a lagoon dredge pile) samples. * 6:2 Cl-PFESA concentration in aqueous samples (ng/L): - Median: <LOD - Mean: <LOQ - s.d.: / - min: <LOD - max: <LOQ - Detect (%): 4% [119] * 6:2 Cl-PFESA concentration in solid samples (ng/g dw): - Median: <LOD - Mean: <LOQ - s.d.: / - min: <LOD - max: <LOQ - Detect (%): 16% Ecotoxicological information Near landfill Sludge Soil Vertebrates (fish) Invertebrates Algae [70] * General observations and geographical trends of PFASs in leachate: - PFCAs are generally found to be the dominant PFASs - C4-C7 chain length PFAAs are more abundant than their longer-chain (C8) homologues - Short-chain PFAAs are prone to preferential release and leaching from municipal solid waste (MSW), consistent with their higher aqueous solubilities and lower organic carbonwater partition coefficients relative to longer-chain PFAAs [70] - Dominance of C4-C7 PFAAs could be related to the shift towards production of shorter-chain perfluorinated compounds since the early 2000's * Concentrations and trends in China: In contrast to other studies, PFOA (mean contribution 29%) was found to be the most abundant PFAA, followed by PFBS (26%) and PFPrA (16%) [70] General observations and geographical trends of PFASs in leachate: - PFCAs are generally found to be the dominant PFASs - C4-C7 chain length PFAAs are more abundant than their longer-chain (C8) homologues [32] Concentration of PFBA in sludge samples (3 datasets): - Assigned value: NA - Average: 16.84 ng/L - Median: 22.90 ng/L - Min.: 0.62 ng/L - Max.: 27.00 ng/L - SD: 14.20 - % relative SD: 84 [118] Spatial and seasonal distributions of PFAS in air, water, sediment, soil, and fish samples in the Asan Lake area of South Korea. Estimation of bioaccumulation potential of PFAS using sedimentwater partition coefficients and bioaccumulation factors. [29] FRESHWATER Danio rerio (acute aquatic toxicity): - 144h post fertilization LC50: >3000 mg/L based on: mortality (test without pH adjustment) - 144h post fertilization EC50: >2200 (120022,000) mg/L based on: mortality and malformation effects (test without pH adjustment) - 96h post fertilization LC50: >3000 mg/L based on: mortality and malformation effects (test with pH adjustment) - 120h post fertilization LC50: > 3000 mg/L based on: mortality and malformation effects (test with pH adjustment) [10] effects on the next generation can already be observed if a species is exposed to PFASs for only 1 generation => more than 1 generation may suffer from PFAS exposure [29] FRESHWATER Pseudokirchneriella subcapitata (acute aquatic toxicity): - 4.5h EC50: 260.96 (213.03-319.55) mg/L based on: photosynthesis (test without pH adjustment) (Cyano)bacteria and sewage microorganisms [29] MARINE WATER Photobacterium phosphoreum (acute aquatic toxicity): - 15min EC50: 14.07 mg/L based on: luminescence inhibition Other organisms [70] - Short-chain PFAAs are prone to preferential release and leaching from municipal solid waste (MSW), consistent with their higher aqueous solubilities and lower organic carbonwater partition coefficients relative to longer-chain PFAAs - Dominance of C4-C7 PFAAs could be related to the shift towards production of shorter-chain perfluorinated compounds since the early 2000's [118] PFBA concentration in soil samples: - Soils (unit: ng/g dw, n=48): * Range: na * Mean: na * Median: na * 75%: na * 95%: na * DF (%): na [32] Concentration of PFBA in fish samples (1 dataset): - Assigned value: NA - Average: 0.16 ng/g ww - Median: 0.16 ng/g ww - Min.: NA - Max.: NA - SD: NA - % relative SD: NA [29] FRESHWATER Daphnia magna (acute aquatic toxicity): - 24h EC50: 184.27 (182.78-185.55) mg/L based on: immobilization (test without pH adjustment) - 24h EC50: >4260.6 mg/L based on: immobilization (test with pH adjustment) - 48h EC50: 180.65 (179.15-182.35) mg/L based on: immobilization (test without pH adjustment) - 48h EC50: >4260.6 mg/L based on: immobilization (test with pH adjustment) [99] PFBA: leachate from 6 landfills and a laboratory bioreactor (US) (ng/L): - Site A: 1700 63 - Site B: 170 6 - Site C: 1400 25 - Site D2: 430 34 - Site D3: 250 29 - Site D6: 540 48 - Laboratory bioreactor: 63 22 [99] - Perfluoroalkyl carboxylates accounted for the majority (67 4%) of the fluorochemicals quantified in leachates - Of the 14 individual (C4 to C14) carboxylate forms measured in the present study, the most abundant were C4-C10 with only infrequent detection of C11-C14 homologs above quantification limits [120] 17 PFAS were analysed in groundwater surrounding legacy landfills in a major Australian urban re-development precinct (Fishermans Bend, Melbourne). * PFBA concentrations in Fishermans Bend groundwater (concentrations in ng/L; average of duplicate sample analysis): - LOD: 0.2 - LOQ: 1.7 - Detection frequency (%): 85 - Minimum: <0.2 - Maximum: 49 - Median: 11 [118] Spatial and seasonal distributions of PFAS in air, water, sediment, soil, and fish samples in the Asan Lake area of South Korea. Estimation of bioaccumulation potential of PFAS using sedimentwater partition coefficients and bioaccumulation factors. PFBA concentration in fish samples: - Fish (unit: ng/g ww, n=42): * Range: nd ~ 3.42 * Mean: 0.12 * Median: 0.00 * 75%: 0.00 [118] * 95%: 0.28 * DF (%): 21.4 Longer-chain PFAS, especially PFOS, can accumulate more in the muscles and lipids of fish than can short-chain PFAS. [29] FRESHWATER Chydorus sphaericus (acute aquatic toxicity): - 24h EC50: 534.49 (513.19-555.37) mg/L based on: immobilization (test without pH adjustment) - 24h EC50: >4.26 mg/L based on: immobilization (test with pH adjustment) - 48h EC50: 460.14 (440.97-479.10) mg/L based on: immobilization (test without pH adjustment) - 48h EC50: > 4260.6 mg/L based on: immobilization (test with pH adjustment) [42] FRESHWATER Brachionus calyciflorus (acute aquatic toxicity): - 24h LC50: 110 mg/L based on: mortality (test without pH adjustment) Quality standards (QS) Plants Presence in consumer products Sources Production/import volume ranges Global distribution Long-range transport potential Suggested alternatives [71] ultra short-chain (<C4) or very long-chain (>C14) PFAAs may also form following oxidation of some PFAS precursors which are not measured with current analytical methods [63] past and ongoing production and use will lead to the accumulation of PFAAs in the global environment, with very slow mixing/sedimentation to the deep oceans and sedimentation/burial in deep sediments as the only known global environmental sinks [63] due to their high persistence and water solubility, PFAAs will be transported to remote locations from sources through water currents and aerosols [63] the most common industrial practice of phasing out one PFAS is to replace it with another structurally similar PFAS => will not solve issues in relation to PFASs as a whole group => let us start the dialogue in defining "essential" and "nonessential" uses of PFASs, while simultaneously developing safe alternative substances and processes for those essential uses [29] FRESHWATER - MAC-EQS (maximum allowable concentration - environmental quality standard): Assessment factor: 100 on LC50: 110 mg/L => QS water = 1100 g/L - AA-QS (annual average - quality standard): Assessment factor: 1000 on LC50: 110 mg/L => QS water = 110 g/L - AA-EQS water = 7 g/L [71] Shorter chains have been reported to have a higher uptake in lettuce leaves, while longer chain PFASs are found primarily in the roots [120] The most common applications include textile protection (ScotchgardTM), surface coating for cooking implements (TeflonTM), food contact paper, and Aqueous Film Forming Foams (AFFFs). [1], [2], [8], [14], [27], [62] [27] - 1986: 4.54 - 227 t/yr - 1990: not reported - 1994: not reported - 1998: not reported - 2002: not reported - 2006: not reported [26] Estimated global emissions of PFBA: 1951-2002: lower: 5 tonnes => 72% from direct sources, higher: 402 tonnes => 50% from direct sources 2003-2015: lower: 5 tonnes => 58% from direct sources, higher: 220 tonnes => 14% from direct sources 2016-2030: lower: 6 tonnes => 17% from direct sources higher: 293 tonnes => 3% from direct sources [23] in comparison to long-chain homologues, short-chain PFAAs are more mobile in soil and sediment => higher solubility in water and lower sorption to solids => higher mobility in the environment [23] Replacements in the future might still use the same concept by slightly modifying the currently used fluorinated alternatives => non-fluorinated parts of these structures may undergo (bio)degradation but transformation products are still likely to be environmentally stable PFECAs or short-chain PFCAs => a non-fluorinated replacement strategy should be pursued [29] MARINE WATER - MAC-EQS: Assessment factor: 1000 on LC50: 110 mg/L => QS water = 110 g/L - AA-QS: Assessment factor: 10000 on LC50: 110 mg/L => QS water = 11 g/L [9] waste water treatment plants have been suggested as one of the major point sources of PFASs to surface waters and the atmosphere [10] - point sources: landfills, manufacturing plants, application of PFAS-containing products at a concentrated area, industrial and municipal sewage treatment plants - nonpoint sources: wet and dry atmospheric deposition [71] - point sources: wastewater treatment plants (WWTPs), industrial emissions from PFAS production sites, commercial and military airfields (use of AFFF-containing PFASs and their precursors), landfills that contain PFASs-contaminated waste [71] - diffuse sources: contamination caused from a range of dispersed urban and rural land use activities, such as atmospheric deposition and upstream water input [28] - Northern rivers had lower levels of PFASs than rivers on south-east and south-west coast (SE) - Northern sub-basins of Baltic Sea had lower levels of PFASs than in southern sub-basins => higher riverine input of PFASs in the south with its higher population density compared to the north [28] Long-chain PFASs were predominant in sea, while short-chain PFASs predominated in rivers (SE) => long-chain PFASs are highly present in atmospheric deposition, while in the terrestrial environment long-chain substances are retained and do not readily reach the rivers => in addition short-chain PFASs have higher water solubility and reach water bodies more easily than longer chained PFASs [63] the most common industrial practice of phasing out one PFAS is to replace it with another structurally similar PFAS => will not solve issues in relation to PFASs as a whole group => let us start the dialogue in defining "essential" and "nonessential" uses of PFASs, while simultaneously developing safe alternative substances and processes for those essential uses [28] - Shorter chained PFCAs were shown to correlate with low latitude (south) (high population density) - Long chained PFCAs were shown to correlate with high latitude (north) => higher relative abundance of longer chained PFCAs in the north due to precursor degradation and aerosol associated stabilization of PFCAs and their precursors in the atmosphere [28] Northern rivers had higher fractions of longchained PFCAs than rivers in the south and west where higher fractions of short-chain PFCAs were observed => stronger relative influence of atmospheric deposition in the north due to long-range transport of long-chained PFCAs and their precursors => stronger relative influence of contemporary point sources in the south due to the replacement of C8based PFASs, by shorter-chained PFAss [63] past and ongoing production and use will lead to the accumulation of PFAAs in the global environment, with very slow mixing/sedimentation to the deep oceans and sedimentation/burial in deep sediments as the only known global environmental sinks [63] due to their high persistence and water solubility, PFAAs will be transported to remote locations from sources through water currents and aerosols [74] transport of PFBA in a water-saturated sediment column => recoveries of PFBA were in the same range as the tracer indicating complete breakthrough => breakthrough curves, recoveries and partition coefficients showed that short-chain PFCAs and PFSAs (i.e. with up to six C-atoms) are only slightly retarded in the water-saturated sediment column => if contaminated surface waters are used as a resource for drinking water production via sediment passage, short-chain PFCAs and PFSAs will not be subject to attenuation [77] - one of the major challenges associated with PFASs is the relatively high mobility and persistence of these compounds in the subsurface => this creates the potential for large plumes in transmissive hydrogeological settings - the persistence of PFAAs coupled with their high solubility, low/moderate sorption to soils and lack of volatility make many PFAAs highly mobile, resulting in long (i.e., potentially multiple mile) groundwater plumes - shorter chain PFAAs generally have lower organic carbon partitioning coefficients than the longer chain compounds => therefore, they are expected to be more mobile in aquifer systems [70] * Concentration and trends in the USA: - PFSA concentrations in leachate have varied from 50 to 3200 ng/L in the USA, with median concentrations of a few hundred ng/L for PFBS, PFHxS and PFOS - While PFOS was detected in all leachate samples, its concentration was generally lower than that of PFBS and PFHxS - This dominance of shorter-chain PFSAs over historically used PFOS could be indicative of the transition towards C4-based chemistry (even prior to 2002) [32] Concentration of lin-PFBS in sludge samples (7 datasets): - Assigned value: NA - Average: 2.28 ng/L - Median: 1.00 ng/L - Min.: 0.00 ng/L - Max.: 8.50 ng/L - SD: 2.93 - % relative SD: 128 [70] * Concentration and trends in Canada: - Landfill gas condensate was reported to contain C4-C8 PFAAs, with PFBS being the dominant compound at a concentration of 1000 ng/L * Concentrations and trends in China: In contrast to other studies, PFOA mean contribution 29%) was found to be the most abundant PFAA, followed by PFBS (26%) and PFPrA (16%) [118] Spatial and seasonal distributions of PFAS in air, water, sediment, soil, and fish samples in the Asan Lake area of South Korea. Estimation of bioaccumulation potential of PFAS using sedimentwater partition coefficients and bioaccumulation factors. [118] L-PFBS concentration in soil samples: - Soils (unit: ng/g dw, n=48): * Range: 0.05 ~ 0.16 * Mean: 0.02 * Median: 0.00 * 75%: 0.00 * 95%: 0.12 * DF (%): 22.9 [29] FRESHWATER Danio rerio (acute aquatic toxicity): - 144h post fertilization LC50: 1500 (1100-1900) mg/L based on: mortality and malformation effects - 144h post fertilization EC50: 450 (350-600) mg/L based on: mortality and malformation effects - 96h post fertilization LC50: >3000 mg/L based on: mortality and malformation effects - 96h post fertilization EC50: 1900.78 (1728.76- 2089.92) mg/L based on: mortality and malformation effects - 120h post fertilization LC50: >3000 mg/L based on: mortality and malformation effects - 120h post fertilization EC50: 1592.32 (1316.19- [10] effects on the next generation can already be observed if a species is exposed to PFASs for only 1 generation => more than 1 generation may suffer from PFAS exposure [29] FRESHWATER Pseudokirchneriella subcapitata (acute aquatic toxicity): - 72h EC50: > 20250 mg/L (37% growth inhibition at 20,250 mg/L) based on: growth inhibition (biomass) - 96h EC50: 5733 (5659-5817) mg/L based on: growth inhibition - 96h EC50: 2347 (2018-2707) mg/L based on: biomass [29] FRESHWATER Pimephales promelas (acute aquatic toxicity): - 96h LC50: 1938 (888-3341) mg/L based on: mortality Lepomis macrochirus (acute aquatic toxicity): - 96h LC50: 6452 mg/L based on: mortality [29] FRESHWATER Daphnia magna (acute aquatic toxicity): - 48h EC50: 2183 (1707-3767) mg/L based on: immobilization MARINE WATER Mysidopsis bahia (acute aquatic toxicity): - 96h EC50: 372 (314-440) mg/L based on: mortality and abnormal behaviour [29] FRESHWATER Pseudokirchneriella subcapitata (chronic aquatic toxicity): - 96h NOEC: 1077 mg/L based on: growth inhibition - 96h NOEC: 1077 mg/L based on: biomass [29] FRESHWATER Anabaena CPB4337 (acute aquatic toxicity): - 30min EC50: 8386 (7752-8693) mg/L based on: bioluminescence inhibition Sewage microorganisms (acute aquatic toxicity): - 3h EC50: >1000 mg/L based on: respiration inhibition MARINE WATER Vibrio fischeri (acute aquatic toxicity): - 15min EC50: 17,520 (16,850-18,200) mg/L based on: bioluminescence inhibition [99] PFBS: leachate from 6 landfills and a laboratory bioreactor (US) (ng/L): - Site A: 750 50 - Site B: 280 13 - Site C: 810 36 - Site D2: 280 12 - Site D3: 390 6.3 - Site D6: 890 100 - Laboratory bioreactor: 2300 130 [32] Concentration of lin-PFBS in fish samples (2 datasets): - Assigned value: NA - Average: 0.10 ng/g ww - Median: 0.10 ng/g ww - Min.: 0.00 ng/g ww - Max.: 0.00 ng/g ww - SD: 0.14 - % relative SD: 141 [29] FRESHWATER Daphnia magna (chronic aquatic toxicity): - 21d NOEC: 502 mg/L based on: reproduction/length [99] Perfluoroalkyl sulfonates were the second most abundant class of fluorochemicals at 22 2% in landfill leachates [120] 17 PFAS were analysed in groundwater surrounding legacy landfills in a major Australian urban re-development precinct (Fishermans Bend, Melbourne). * PFBS concentrations in Fishermans Bend groundwater (concentrations in ng/L; average of duplicate sample analysis): - LOD: 0.2 - LOQ: 0.2 [120] - Detection frequency (%): 100 - Minimum: 2.0 - Maximum: 31 - Median: 9.0 [118] Spatial and seasonal distributions of PFAS in air, water, sediment, soil, and fish samples in the Asan Lake area of South Korea. Estimation of bioaccumulation potential of PFAS using sedimentwater partition coefficients and bioaccumulation factors. L-PFBS concentration in fish samples: - Fish (unit: ng/g ww, n=42): * Range: nd ~ 0.46 * Mean: 0.01 * Median: 0.00 * 75%: 0.00 * 95%: 0.00 * DF (%): 2.4 Longer-chain PFAS, especially PFOS, can accumulate more in the muscles and lipids of fish than can short-chain PFAS. [29] FRESHWATER - MAC-EQS (maximum allowable concentration - environmental quality standard): Assessment factor: 100 on EC50: 372 mg/L => QS water = 3720 g/L - AA-QS (annual average - quality standard): Assessment factor: 1000 on EC50: 372 mg/L => QS water = 372 g/L - AA-EQS water = 3 g/L [35] - control plants (wheat, soybean, pumpkin): PFBS (0.0057-0.0067 nmol/g) in shoots - exposured plants with solution of NEtFOSA (wheat, soybean, pumpkin): PFBS (0.0234-0.1220 nmol/g) in roots PFBS (0.0074-0.0130 nmol/g) in shoots (sign. higher than control p<0.05) - translocation ratio (TR) wheat: PFBS (0.210) - translocation ratio (TR) soybean: PFBS (0.013) - translocation ratio (TR) pumpkin: PFBS (0.060) [114] Analysis of perfluoroalkyl substances in 126 individual samples of building materials, consumer products, car interior materials and wastes which potentially affect indoor environments where people spend most of their time. PFBS concentrations in individual samples: * Textile (1A): - New curtain: 0.398 g/kg - Pillow fill: 1.06 g/kg - Curtain I: 0.159 g/kg - Curtain II: 0.635 g/kg - Foam: 2.53 g/kg - Foam: 3.39 g/kg - Bed cover: 6.14 g/kg [1], [2], [6], [7], [8], [14], [27], [62] [29] MARINE WATER - MAC-EQS: Assessment factor: 1000 on EC50: 372 mg/L => QS water = 372 g/L - AA-QS: Assessment factor: 10000 on EC50: 372 mg/L => QS water = 37 g/L [71] Shorter chains have been reported to have a higher uptake in lettuce leaves, while longer chain PFASs are found primarily in the roots [114] - Hanging: 5.82 g/kg - Teddy bear (filling): 0.855 g/kg - Teddy bear (cover): 0.311 g/kg - Foam: 2.92 g/kg - Upholstery material: 0.583 g/kg - Coir (coconut fibre): 2.9 g/kg - Foam: 1.43 g/kg - Blanket: 0.286 g/kg * Floor covering (1B): - Carpet - grey: 0.966 g/kg - Carpet - green: 0.348 g/kg - Carpet - brown: 0.488 g/kg - Persian carpet: 0.874 g/kg * EEE (1C): - Switch: 0.028 g/kg - Vacuum cleaner II: 0.09 g/kg [9] waste water treatment plants have been suggested as one of the major point sources of PFASs to surface waters and the atmosphere [112] Study investigated plant uptake of PFAAs by a native wetland species in the US, Juncus effusus * PFBS concentration in shoots: - Timepoint: p<0.001 - Concentration: p<0.001 - Timepoint x Concentration: p<0.001 * PFBS concentration in roots: - Timepoint: p=0.006 - Concentration: p<0.001 - Timepoint x Concentration: p=0.045 * J. effusus plants could translocate all 7 PFAAs during the 21-day experimental [114] - Keyboard I: 0.163 g/kg - Screen I: 0.305 g/kg - Printed wiring: 0.065 g/kg - Vacuum cleaner II: 0.102 g/kg - Keyboard II: 0.2 g/kg - Screen II: 11.4 g/kg - TV: 0.051 g/kg - fridge - door insulation foam: 0.073 g/kg * Plastics (1D): DVD cover: 0.384 g/kg * OSB and wood (2A): - Wooden board: 0.201 g/kg [10] - point sources: landfills, manufacturing plants, application of PFAS-containing products at a concentrated area, industrial and municipal sewage treatment plants - nonpoint sources: wet and dry atmospheric deposition [112] * Plants exposed to PFAAs at 10x had higher PFAAs concentrations than the plants exposed to those at 1x * Uptake of PFAAs increased with increasing exposure time * PFAAs, except PFOS, had higher concentrations in the plant shoots than in the plant roots * PFAAs concentration in the growth medium was a significant factor on the uptake of all 7 PFAAs for both shoots and roots * Exposure time also had significant impact on the plant uptake of PFAAs, except PFPeA and PFHxS by roots * Significant interactions between exposure time and PFAAs concentration regarding their impact on the uptake of PFHxA, PFHpA and PFOS by shoots and PFBS and [114] * Insulation materials (2B): - KNAUF insulation Ecose technology: 0.414 g/kg - Insulation glass fibre: 3.87 g/kg - Phenolic foam insulation ISOVER WEBER TERRANOVA: 1.53 g/kg - Insulation FOAMGLAS PERINSUL: 0.101 g/kg - Sound insulation WOLF Professional: 0.086 g/kg - Cotton insulation: 1.47 g/kg * Mounting and sealing foam (2C): - Asphalt: 0.423 g/kg * Facade materials (2D): - Water-resisting paint: 0.536 g/kg * Car interior materials (3): - Hyundai - plastic material: 0.254 g/kg - Hyundai - textile material: 2.18 g/kg [71] - point sources: wastewater treatment plants (WWTPs), industrial emissions from PFAS production sites, commercial and military airfields (use of AFFF-containing PFASs and their precursors), landfills that contain PFASs-contaminated waste [112] * Translocation factor (TF) = ratio of pollutant concentration in shoots to that in roots => TF values of PFAAs decreased with increasing carbon chain length and PFSAs had lower TF than PFCAs with similar carbon chain length => PFAAs with shorter carbon chain length can be taken up by J. effusus roots and tended to translocate to plant shoots * TF for PFPeA, PFBS, PFHxA and PFHpA were always greater than 2 at different timepoints and reached the highest values on Day-21 [114] - Hyundai - textile material: 1.84 g/kg - Skoda - textile material: 0.197 g/kg - Skoda - plastic material: 0.274 g/kg - Skoda - plastic material: 0.196 g/kg - Skoda - textile material: 0.068 g/kg * WEEE (4): - Spaghetti insulation I: 0.308 g/kg - Spaghetti insulation II: 0.074 g/kg - Mixed waste - small electronical devices: 0.327 g/kg [71] - diffuse sources: contamination caused from a range of dispersed urban and rural land use activities, such as atmospheric deposition and upstream water input [112] * Removal efficiencies of PFAAs by shoots and roots increased with increasing exposure time * Regarding each PFAA, J. effusus roots had lower removal efficiency of PFAAs (except PFOS) than shoots * On Day-21, the whole J. effusus had relatively higher removal efficiencies of PFHxA, PFOA and PFOS * Plants exposed to PFAAs at 1x generally had higher removal efficiencies than at 10x * Both concentrations of PFAAs in the growth medium and PFAAs exposure time can affect the PFAAs removal efficiency by J. effusus [114] - Mixed waste: 0.753 g/kg - Rubber: 0.3 g/kg - Impactor undersized particles: 1.38 g/kg [120] The most common applications include textile protection (ScotchgardTM), surface coating for cooking implements (TeflonTM), food contact paper, and Aqueous Film Forming Foams (AFFFs). [19] Calculated global emissions of PFBS (based on PFOS emissions): 1958-2002: lower: 13 tonnes, higher: 132 tonnes 2003-2015: lower: 1 tonnes, higher: 26 tonnes 2016-2030: lower: 0 tonnes, higher: 5 tonnes [23] in comparison to long-chain homologues, short-chain PFAAs are more mobile in soil and sediment => higher solubility in water and lower sorption to solids => higher mobility in the environment [23] Replacements in the future might still use the same concept by slightly modifying the currently used fluorinated alternatives => non-fluorinated parts of these structures may undergo (bio)degradation but transformation products are still likely to be environmentally stable PFECAs or short-chain PFCAs => a non-fluorinated replacement strategy should be pursued [28] - Northern rivers had lower levels of PFASs than rivers on south-east and south-west coast (SE) - Northern sub-basins of Baltic Sea had lower levels of PFASs than in southern sub-basins => higher riverine input of PFASs in the south with its higher population density compared to the north [28] Long-chain PFASs were predominant in sea, while short-chain PFASs predominated in rivers (SE) => long-chain PFASs are highly present in atmospheric deposition, while in the terrestrial environment long-chain substances are retained and do not readily reach the rivers => in addition short-chain PFASs have higher water solubility and reach water bodies more easily than longer chained PFASs [63] the most common industrial practice of phasing out one PFAS is to replace it with another structurally similar PFAS => will not solve issues in relation to PFASs as a whole group => let us start the dialogue in defining "essential" and "nonessential" uses of PFASs, while simultaneously developing safe alternative substances and processes for those essential uses [28] - Shorter chained PFCAs were shown to correlate with low latitude (south) (high population density) - Long chained PFCAs were shown to correlate with high latitude (north) => higher relative abundance of longer chained PFCAs in the north due to precursor degradation and aerosol associated stabilization of PFCAs and their precursors in the atmosphere [28] Northern rivers had higher fractions of longchained PFCAs than rivers in the south and west where higher fractions of short-chain PFCAs were observed => stronger relative influence of atmospheric deposition in the north due to long-range transport of long-chained PFCAs and their precursors => stronger relative influence of contemporary point sources in the south due to the replacement of C8based PFASs, by shorter-chained PFAss [34] Surface water data SE: slightly higher PFAS concentrations in southern Sweden (higher population density) compared to northern Sweden [63] due to their high persistence and water solubility, PFAAs will be transported to remote locations from sources through water currents and aerosols [63] past and ongoing production and use will lead to the accumulation of PFAAs in the global environment, with very slow mixing/sedimentation to the deep oceans and sedimentation/burial in deep sediments as the only known global environmental sinks [74] transport of PFBS in a water-saturated sediment column => recoveries of PFBS were in the same range as the tracer indicating complete breakthrough => breakthrough curves, recoveries and partition coefficients showed that short-chain PFCAs and PFSAs (i.e. with up to six C-atoms) are only slightly retarded in the water-saturated sediment column => if contaminated surface waters are used as a resource for drinking water production via sediment passage, short-chain PFCAs and PFSAs will not be subject to attenuation [77] - one of the major challenges associated with PFASs is the relatively high mobility and persistence of these compounds in the subsurface => this creates the potential for large plumes in transmissive hydrogeological settings - the persistence of PFAAs coupled with their high solubility, low/moderate sorption to soils and lack of volatility make many PFAAs highly mobile, resulting in long (i.e., potentially multiple mile) groundwater plumes - shorter chain PFAAs generally have lower organic carbon partitioning coefficients than the longer chain compounds => therefore, they are expected to be more mobile in aquifer systems [70] General observations and geographical trends of PFASs in leachate: - PFCAs are generally found to be the dominant PFASs - C4-C7 chain length PFAAs are more abundant than their longer-chain (C8) homologues [32] Concentration of PFPeA in sludge samples (7 datasets): - Assigned value: NA - Average: 3.06 ng/L - Median: 3.12 ng/L - Min.: 0.77 ng/L - Max.: 5.00 ng/L - SD: 1.59 - % relative SD: 52 [118] Spatial and seasonal distributions of PFAS in air, water, sediment, soil, and fish samples in the Asan Lake area of South Korea. Estimation of bioaccumulation potential of PFAS using sedimentwater partition coefficients and bioaccumulation factors. [29] FRESHWATER Pimephales promelas (acute aquatic toxicity): - 96h LC50: 31.8 (10.3-98.3) mg/L based on: mortality [10] effects on the next generation can already be observed if a species is exposed to PFASs for only 1 generation => more than 1 generation may suffer from PFAS exposure [29] FRESHWATER Pseudokirchneriella subcapitata (acute aquatic toxicity): - 72h EC50: 81.7 (76.7-87.5) mg/L based on: biomass [29] MARINE WATER Photobacterium phosphoreum (acute aquatic toxicity): - 15 min EC50: 16.22 mg/L based on: luminescence inhibition [70] - Short-chain PFAAs are prone to preferential release and leaching from municipal solid waste (MSW), consistent with their higher aqueous solubilities and lower organic carbonwater partition coefficients relative to longer-chain PFAAs - Dominance of C4-C7 PFAAs could be related to the shift towards production of shorter-chain perfluorinated compounds since the early 2000's [99] PFPeA: leachate from 6 landfills and a laboratory bioreactor (US) (ng/L): - Site A: 1100 170 - Site B: 120 13 - Site C: 1500 36 - Site D2: 730 36 - Site D3: 500 29 - Site D6: 470 34 - Laboratory bioreactor: 460 23 [99] - Perfluoroalkyl carboxylates accounted for the majority (67 4%) of the fluorochemicals quantified in leachates - Of the 14 individual (C4 to C14) carboxylate forms measured in the present study, the most abundant were C4-C10 with only infrequent detection of C11-C14 homologs above quantification limits [120] 17 PFAS were analysed in groundwater surrounding legacy landfills in a major Australian urban re-development precinct (Fishermans Bend, Melbourne). * PFPeA concentrations in Fishermans Bend groundwater (concentrations in ng/L; average of duplicate sample analysis): - LOD: 0.2 - LOQ: 0.2 - Detection frequency (%): 46 - Minimum: <0.2 - Maximum: 15 - Median: <0.2 [118] PFPeA concentration in soil samples: - Soils (unit: ng/g dw, n=48): * Range: nd ~ 0.37 * Mean: 0.04 * Median: 0.00 * 75%: 0.00 * 95%: 0.23 * DF (%): 22.9 [32] Concentration of PFPeA in fish samples (1 dataset): - Assigned value: NA - Average: 0.24 ng/g ww - Median: 0.24 ng/g ww - Min.: NA - Max.: NA - SD: NA - % relative SD: NA [29] FRESHWATER Daphnia magna (acute aquatic toxicity): - 48h EC50: >112 mg/L based on: immobilization [118] Spatial and seasonal distributions of PFAS in air, water, sediment, soil, and fish samples in the Asan Lake area of South Korea. Estimation of bioaccumulation potential of PFAS using sedimentwater partition coefficients and bioaccumulation factors. PFPeA concentration in fish samples: - Fish (unit: ng/g ww, n=42): * Range: nd ~ 0.58 * Mean: 0.01 * Median: 0.00 * 75%: 0.00 * 95%: 0.00 [42] FRESHWATER Brachionus calyciflorus (acute aquatic toxicity): - 24h LC50: 130 mg/L based on: mortality [118] * DF (%): 2.4 Longer-chain PFAS, especially PFOS, can accumulate more in the muscles and lipids of fish than can short-chain PFAS. [27] - 1986: 4.54 - 227 t/yr - 1990: 4.54 - 227 t/yr - 1994: 4.54 - 227 t/yr - 1998: 4.54 - 227 t/yr - 2002: 4.54 - 227 t/yr - 2006: <227 t/yr [19] Calculated global emissions of FBSA/Es (based on PFOS emissions): 1958-2002: lower: 1 tonnes, higher: 472 tonnes 2003-2015: lower: 1 tonnes, higher: 58 tonnes 2016-2030: lower: 0 tonnes, higher: 42 tonnes [27] [27] - 1986: 4.54 - 227 t/yr [19] Calculated global emissions of PBSF - 1990: 4.54 - 227 t/yr (based on PFOS emissions): - 1994: 4.54 - 227 t/yr 1958-2002: lower: 8 tonnes, higher: 25 tonnes - 1998: 4.54 - 227 t/yr 2003-2015: lower: 0 tonnes, higher: 1 tonnes - 2002: 4.54 - 227 t/yr 2016-2030: lower: 0 tonnes, higher: 0 tonnes - 2006: 227 - 454 t/yr [75] - critical travel distance (CTD): 41179.08 km - travel efficiency (TE): 163.729 % => higher CTD and TE than 8:2 FTOH [29] FRESHWATER - MAC-EQS (maximum allowable concentration - environmental quality standard): Assessment factor: 10 on LC50: 31.8 mg/L => QS water = 3180 g/L - AA-QS (annual average - quality standard): Assessment factor: 1000 on LC50: 31.8 mg/L => QS water = 32 g/L - AA-EQS water = 3 g/L [71] Shorter chains have been reported to [114] Analysis of perfluoroalkyl [6] have a higher uptake in lettuce leaves, substances in 126 individual samples of while longer chain PFASs are found building materials, consumer products, primarily in the roots car interior materials and wastes which potentially affect indoor environments where people spend most of their time. PFPeA concentrations in individual samples: * Textile (1A): - Curtain I: 0.122 g/kg - Foam: 3.87 g/kg - Hanging: 0.381 g/kg - Teddy bear (cover): 0.221 g/kg - Foam: 2.9 g/kg - Upholstery material: 2.58 g/kg - Coir (coconut fibre): 3.68 g/kg * Floor covering (1B): - Carpet - grey: 0.719 g/kg [29] MARINE WATER - MAC-EQS: Assessment factor: 100 on LC50: 31.8 mg/L => QS water = 318 g/L - AA-QS: Assessment factor: 10000 on LC50: 31.8 mg/L => QS water = 3.2 g/L [112] Study investigated plant uptake of PFAAs by a native wetland species in the US, Juncus effusus * PFPeA concentration in shoots: - Timepoint: p<0.001 - Concentration: p<0.001 - Timepoint x Concentration: p=0.113 * PFPeA concentration in roots: - Timepoint: p=0.562 - Concentration: p<0.001 - Timepoint x Concentration: p=0.666 * J. effusus plants could translocate all 7 PFAAs during the 21-day experimental period [112] * Plants exposed to PFAAs at 10x had higher PFAAs concentrations than the plants exposed to those at 1x * Uptake of PFAAs increased with increasing exposure time * PFAAs, except PFOS, had higher concentrations in the plant shoots than in the plant roots * PFAAs concentration in the growth medium was a significant factor on the uptake of all 7 PFAAs for both shoots and roots * Exposure time also had significant impact on the plant uptake of PFAAs, except PFPeA and PFHxS by roots * Significant interactions between exposure time and PFAAs concentration regarding their impact on the uptake of PFHxA, PFHpA and PFOS by shoots and PFBS and [114] - Laminated plastic floor covering: 2.14 g/kg - Persian carpet: 0.853 g/kg * EEE (1C): - Switch: 0.043 g/kg - Keyboard I: 0.251 g/kg - Screen I: 0.25 g/kg - Printed wiring: 0.091 g/kg - Vacuum cleaner II: 0.534 g/kg - Keyboard II: 0.06 g/kg - Keyboard: 0.326 g/kg * OSB and wood (2A): - Formica: 1.89 g/kg - Oriented Strand Board: 2.85 g/kg - Oriented Strand Board - white: 0.385 g/kg - Wooden board: 1.79 g/kg [114] - Chipboard - brown: 2.21 g/kg - Oriented Strand Board - white: 1.75 g/kg - Chipboard - pale: 3.33 g/kg - Chipboard: 10 g/kg - Chipboard: 2.79 g/kg - Oriented Strand Board: 0.756 g/kg - Oriented Strand Board: 3.53 g/kg - Chipboard: 1.54 g/kg - Chipboard: 1.35 g/kg * Insulation materials (2B): - Insulation glass fibre: 1.55 g/kg [9] waste water treatment plants have been suggested as one of the major point sources of PFASs to surface waters and the atmosphere [10] - point sources: landfills, manufacturing plants, application of PFAS-containing products at a concentrated area, industrial and municipal sewage treatment plants - nonpoint sources: wet and dry atmospheric deposition [112] * Translocation factor (TF) = ratio of pollutant concentration in shoots to that in roots => TF values of PFAAs decreased with increasing carbon chain length and PFSAs had lower TF than PFCAs with similar carbon chain length => PFAAs with shorter carbon chain length can be taken up by J. effusus roots and tended to translocate to plant shoots * TF for PFPeA, PFBS, PFHxA and PFHpA were always greater than 2 at different timepoints and reached the highest values on Day-21 [114] - Insulation hemp rope: 2.94 g/kg - Wooden fibre insulation: 5 g/kg - Pipe insulation aeroflex: 0.232 g/kg - Phenolic foam insulation ISOVER WEBER TERRANOVA: 2.01 g/kg - Insulation CANABEST PANEL: 2.87 g/kg - TETRAPAK FLEXIBUILD: 1.62 g/kg - Wooden fibre insulation: 6.49 g/kg - Insulation FOAMGLAS PERINSUL: 0.089 g/kg - Sound insulation WOLF PROFESSIONAL: 0.387 g/kg * Mounting and sealing foam (2C): - Asphalt: 0.483 g/kg [71] - point sources: wastewater treatment plants (WWTPs), industrial emissions from PFAS production sites, commercial and military airfields (use of AFFF-containing PFASs and their precursors), landfills that contain PFASs-contaminated waste [112] * Removal efficiencies of PFAAs by shoots and roots increased with increasing exposure time * Regarding each PFAA, J. effusus roots had lower removal efficiency of PFAAs (except PFOS) than shoots * On Day-21, the whole J. effusus had relatively higher removal efficiencies of PFHxA, PFOA and PFOS * Plants exposed to PFAAs at 1x generally had higher removal efficiencies than at 10x * Both concentrations of PFAAs in the growth medium and PFAAs exposure time can affect the PFAAs removal efficiency by J. effusus [114] * Facade materials (2D): - Window corner bead: 1.94 g/kg * Air conditioning (2F): - Heat exchanger: 0.058 g/kg - AC - inside foil: 0.194 g/kg - AC - alu foil: 0.094 g/kg * Car interior materials (3): - Hyundai - plastic material: 0.086 g/kg - Hyundai - textile material: 0.159 g/kg - Hyundai - textile material: 0.302 g/kg - Skoda - textile material: 0.079 g/kg - Skoda - plastic material: 0.208 g/kg - Skoda - plastic material: 0.417 g/kg - Skoda - textile material: 0.056 g/kg [71] - diffuse sources: contamination caused from a range of dispersed urban and rural land use activities, such as atmospheric deposition and upstream water input [114] * WEEE (4): - Spaghetti insulation II: 0.057 g/kg - Mixed waste - small electronical devices: 0.04 g/kg - Mixed waste: 0.044 g/kg - Impactor undersized particles: 0.043 g/kg [26] Estimated global emissions of PFPeA: 1951-2002: lower: 14 tonnes => 39% from direct sources, higher: 690 tonnes => 20% from direct sources 2003-2015: lower: 5 tonnes => 37% from direct sources, higher: 305 tonnes => 7% from direct sources 2016-2030: lower: 7 tonnes => 8% from direct sources higher: 382 tonnes => 2% from direct sources [63] due to their high persistence and water solubility, PFAAs will be transported to remote locations from sources through water currents and aerosols [63] the most common industrial practice of phasing out one PFAS is to replace it with another structurally similar PFAS => will not solve issues in relation to PFASs as a whole group => let us start the dialogue in defining "essential" and "nonessential" uses of PFASs, while simultaneously developing safe alternative substances and processes for those essential uses [34] Surface water data SE: slightly higher PFAS concentrations in southern Sweden (higher population density) compared to northern Sweden [74] transport of PFPeA in a water-saturated sediment column => recoveries of PFPeA were in the same range as the tracer indicating complete breakthrough => breakthrough curves, recoveries and partition coefficients showed that short-chain PFCAs and PFSAs (i.e. with up to six C-atoms) are only slightly retarded in the water-saturated sediment column => if contaminated surface waters are used as a resource for drinking water production via sediment passage, short-chain PFCAs and PFSAs will not be subject to attenuation [63] past and ongoing production and use will lead to the accumulation of PFAAs in the global environment, with very slow mixing/sedimentation to the deep oceans and sedimentation/burial in deep sediments as the only known global environmental sinks [77] - one of the major challenges associated with PFASs is the relatively high mobility and persistence of these compounds in the subsurface => this creates the potential for large plumes in transmissive hydrogeological settings - the persistence of PFAAs coupled with their high solubility, low/moderate sorption to soils and lack of volatility make many PFAAs highly mobile, resulting in long (i.e., potentially multiple mile) groundwater plumes - shorter chain PFAAs generally have lower organic carbon partitioning coefficients than the longer chain compounds => therefore, they are expected to be more mobile in aquifer systems [120] The most common applications include textile protection (ScotchgardTM), surface coating for cooking implements (TeflonTM), food contact paper, and Aqueous Film Forming Foams (AFFFs). [120] 17 PFAS were analysed in groundwater surrounding legacy landfills in a major Australian urban re-development precinct (Fishermans Bend, Melbourne). * PFPeS concentrations in Fishermans Bend groundwater (concentrations in ng/L; average of duplicate sample analysis): - LOD: 0.2 - LOQ: 0.2 [120] - Detection frequency (%): 77 - Minimum: <0.2 - Maximum: 16 - Median: 6.3 [120] The most common applications include textile protection (ScotchgardTM), surface coating for cooking implements (TeflonTM), food contact paper, and Aqueous Film Forming Foams (AFFFs). [100] - exposure of PFASs to humans occurs mainly through food, water, and the indoor environment - in addition to direct exposure, humans are also indirectly exposed to PFAS through precursor compounds that degrade to PFCA and PFSA [19] Calculated global emissions of PFPeS (based on PFOS emissions): 1958-2002: lower: 10 tonnes, higher: 164 tonnes 2003-2015: lower: 1 tonnes, higher: 32 tonnes 2016-2030: lower: 0 tonnes, higher: 6 tonnes [22] * PFHxA in leachate landfill (AU): - concentrations: mean: 1700 ng/L, range: 73-25000 ng/L - contribution to total PFASs: mean: 37%, range: 14-72% * newer landfills have higher PFAS concentrations => current influx of PFAS-containing products into the waste stream, despite regulatory action more than a decade ago [32] Concentration of PFHxA in sludge samples (10 datasets): - Assigned value: 0.84 - Average: 7.33 ng/L - Median: 0.99 ng/L - Min.: 0.37 ng/L - Max.: 61.31 ng/L - SD: 19.02 - % relative SD: 259 [118] Spatial and seasonal distributions of PFAS in air, water, sediment, soil, and fish samples in the Asan Lake area of South Korea. Estimation of bioaccumulation potential of PFAS using sedimentwater partition coefficients and bioaccumulation factors. [29] FRESHWATER Oncorhynchus mykiss (acute aquatic toxicity): - 96h LC50: >99.2 mg/L based on: mortality [10] effects on the next generation can already be observed if a species is exposed to PFASs for only 1 generation => more than 1 generation may suffer from PFAS exposure [29] FRESHWATER Pseudokirchneriella subcapitata (acute aquatic toxicity): - 72h LC50: >100 mg/L based on: biomass MARINE WATER Chlorella vulgaris (acute aquatic toxicity): - 72h EC50: 4019.51 +/- 200.35 mg/L based on: biomass Skeletonema marinoi (acute aquatic toxicity): - 72h EC50: 1477.58 +/- 72.00 mg/L based on: biomass [29] MARINE WATER Vibrio fischeri (acute aquatic toxicity): - 30 min EC50: 1335.39 +/- 123.18 mg/L based on: luminescence inhibition Photobacterium phosphoreum (acute aquatic toxicity): - 15 min EC50: 17.20 mg/L based on: luminescence inhibition Geitlerinema amphibium (acute aquatic toxicity): - 72 h EC50: 995.49 +/- 50.09 mg/L based on: biomass [70] * General observations and geographical trends of PFASs in leachate: - PFCAs are generally found to be the dominant PFASs - C4-C7 chain length PFAAs are more abundant than their longer-chain (C8) homologues [118] PFHxA concentration in soil samples: - Soils (unit: ng/g dw, n=48): * Range: nd ~ 0.33 * Mean: 0.10 * Median: 0.09 * 75%: 0.14 * 95%: 0.26 * DF (%): 81.3 [32] Concentration of PFHxA in fish samples (4 datasets): - Assigned value: NA - Average: 2.07 ng/g ww - Median: 2.15 ng/g ww - Min.: 0.09 ng/g ww - Max.: 3.90 ng/g ww - SD: 1.56 - % relative SD: 75 [29] FRESHWATER Daphnia magna (acute aquatic toxicity): - 48h LC50: >96.5 mg/L based on: mortality [56] FRESHWATER Pseudokirchneriella subcapitata (chronic aquatic toxicity): 72h NOEC > 100 mg/L [70] - Short-chain PFAAs are prone to preferential release and leaching from municipal solid waste (MSW), consistent with their higher aqueous solubilities and lower organic carbonwater partition coefficients relative to longer-chain PFAAs - Dominance of C4-C7 PFAAs could be related to the shift towards production of shorter-chain perfluorinated compounds since the early 2000's * Concentration and trends in Australia: Median concentrations were <550 ng/L for all PFAAs, expect PFHxA (970 ng/L), similar to those reported in European Countries [56] FRESHWATER Oncorhynchus mykiss (chronic aquatic toxicity): 90d NOEC = 9.96 mg/L (as PFHx anion) [42] FRESHWATER Brachionus calyciflorus (acute aquatic toxicity): - 24h LC50: 140 mg/L based on: mortality [94] Algae isolated from Baltic Sea coastal waters and maintained as monoalgal cultures Chlorella vulgaris (acute aquatic toxicity): - 72h EC50: 12.84 0.64 mM, logEC50: 1.11 Skeletonema marinoi (acute aquatic toxicity): - 72h EC50: 4.72 0.23 mM, logEC50: 0.67 Geitlerinema amphibium (acute aquatic toxicity): - 72h EC50: 3.18 0.16 mM, logEC50: 0.50 => Results show that these algae are highly sensitive to PFCAs, although the cyanobacteria and diatoms are far more sensitive to these compounds than the green algae => A linear relationship was also found between lipophilicity expressed as alkyl chain length or partition coefficient and toxicity expressed as logEC50 [29] FRESHWATER - AA-EQS water = 1 g/L [27] - 1986: 4.54 - 227 t/yr - 1990: 4.54 - 227 t/yr - 1994: 4.54 - 227 t/yr - 1998: 4.54 - 227 t/yr - 2002: not reported - 2006: not reported [71] Shorter chains have been reported to have a higher uptake in lettuce leaves, while longer chain PFASs are found primarily in the roots [114] Analysis of perfluoroalkyl substances in 126 individual samples of building materials, consumer products, car interior materials and wastes which potentially affect indoor environments where people spend most of their time. PFHxA concentrations in individual samples: * Textile (1A): - Foam: 33.0 g/kg - Upholstery material: 0.759 g/kg - Coir (coconut fibre): 0.444 g/kg * Floor covering (1B): - Carpet - grey: 2.38 g/kg - Carpet - grey: 1.19 g/kg - Laminated plastic floor covering: 2.57 g/kg [1], [2], [6], [8], [14], [22], [27] [112] Study investigated plant uptake of PFAAs by a native wetland species in the US, Juncus effusus * PFHxA concentration in shoots: - Timepoint: p<0.001 - Concentration: p<0.001 - Timepoint x Concentration: p<0.001 * PFHxA concentration in roots: - Timepoint: p=0.013 - Concentration: p<0.001 - Timepoint x Concentration: p=0.066 * J. effusus plants could translocate all 7 PFAAs during the 21-day experimental period [114] * EEE (1C): - Switch: 0.09 g/kg - Screen I: 0.269 g/kg - Printed wiring: 0.15 g/kg - Keyboard II: 0.071 g/kg - Screen II: 0.216 g/kg * OSB and wood (2A): - Oriented Strand Board: 1.4 g/kg - Wooden board: 3.14 g/kg - Chipboard - brown: 0.665 g/kg - Oriented Strand Board - white: 2.03 g/kg - Chipboard - pale: 1.94 g/kg - Chipboard: 1.52 g/kg - Chipboard: 3.07 g/kg - Oriented Strand Board: 1.32 g/kg - Oriented Strand Board: 0.596 g/kg [112] * Plants exposed to PFAAs at 10x had higher PFAAs concentrations than the plants exposed to those at 1x * Uptake of PFAAs increased with increasing exposure time * PFAAs, except PFOS, had higher concentrations in the plant shoots than in the plant roots * PFAAs concentration in the growth medium was a significant factor on the uptake of all 7 PFAAs for both shoots and roots * Exposure time also had significant impact on the plant uptake of PFAAs, except PFPeA and PFHxS by roots * Significant interactions between exposure time and PFAAs concentration regarding their impact on the uptake of PFHxA, PFHpA and PFOS by shoots and PFBS and [114] - Wooden board: 2.35 g/kg - Chipboard: 1.47 g/kg - Chipboard: 0.72 g/kg * Insulation materials (2B): - Insulation hemp rope: 0.452 g/kg - Blow cellulose insulation: 0.417 g/kg - Wooden fibre insulation: 0.798 g/kg - Insulation CANABEST PANEL: 0.599 g/kg - TETRAPAK FLEXIBUILD: 0.164 g/kg - Wooden fibre insulation: 5.79 g/kg * Facade materials (2D): - Window corner bead: 0.971 g/kg [10] - point sources: landfills, manufacturing plants, application of PFAS-containing products at a concentrated area, industrial and municipal sewage treatment plants - nonpoint sources: wet and dry atmospheric deposition [21] emissions of PFCAs divided in two source categories: - direct emission sources = emissions that come from the manufacture, intentional use and disposal of PFCAs throughout their product life-cycle - indirect emission sources = emissions of a given PFCA that is present as an impurity in a product, or formed by degradation (in environment, wildlife, humans) of a precursor substance (PFCA formed from the atmospheric degradation of perfluorooctane sulfonamido ethanol or from the biotransformation of a fluorotelomer alcohol) [19] Calculated global emissions of FPeSA/Es (based on PFOS emissions): 1958-2002: lower: 0 tonnes, higher: 236 tonnes 2003-2015: lower: 0 tonnes, higher: 29 tonnes 2016-2030: lower: 0 tonnes, higher: 21 tonnes [19] Calculated global emissions of PPeSF (based on PFOS emissions): 1958-2002: lower: 7 tonnes, higher: 32 tonnes 2003-2015: lower: 0 tonnes, higher: 1 tonnes 2016-2030: lower: 0 tonnes, higher: 0 tonnes [23] PFECAs and PFESAs are expected to be highly persistent and have physicochemical properties (high water solubility and low pKa) similar to those of PFCAs and PFSAs => likely to have similarly high long-range transport potential in water [23] Replacements in the future might still use the same concept by slightly modifying the currently used fluorinated alternatives => non-fluorinated parts of these structures may undergo (bio)degradation but transformation products are still likely to be environmentally stable PFECAs or short-chain PFCAs => a non-fluorinated replacement strategy should be pursued [21] all manufacturing of PFCAs occurs in northern hemisphere => PFCAs can only be detected in oceans north of the equator + concentrations of volatile precursors in remote environments originating from consumer products are lower in southern hemisphere => BUT Australians have similar PFCA-serum levels to humans in industrialized countries in northern hemisphere => localized environmental or consumer exposure from product use are similar in both hemispheres [23] in comparison to long-chain homologues, short-chain PFAAs are more mobile in soil and sediment => higher solubility in water and lower sorption to solids => higher mobility in the environment [23] Replacements in the future might still use the same concept by slightly modifying the currently used fluorinated alternatives => non-fluorinated parts of these structures may undergo (bio)degradation but transformation products are still likely to be environmentally stable PFECAs or short-chain PFCAs => a non-fluorinated replacement strategy should be pursued [21] levels of volatile precursors of PFCAs and levels of PFCAs in precipitation are highest near urban areas => urban areas are sources of PFCAs to the atmosphere [25] PFHxA is more water soluble and less sorptive than PFOS => will be even more mobile in the subsurface environment [25] all PFASs which contain persistent perfluoroalkyl moieties will ultimately degrade to form non-degradable P chemicals in the environment => poorly reversible PFAS exposures will arise in (parts of) the environment => if poorly reversible exposure had been used as a criterion, then PFAS would have been regulated at an earlier stage => usual risk-based regulatory approach has been applied instead of precautionary approach => eventual replacement of persistent PFASs in fire-fighting foams with degradable alternatives that achieve acceptable functionality is a strategy that is in harmony with a future sustainable society [21] significant (p<0.01) correlation between concentrations of PFHxA and PFOA in ocean water samples indicating that these compounds have a common source [28] Long-chain PFASs were predominant in sea, while short-chain PFASs predominated in rivers (SE) => long-chain PFASs are highly present in atmospheric deposition, while in the terrestrial environment long-chain substances are retained and do not readily reach the rivers => in addition short-chain PFASs have higher water solubility and reach water bodies more easily than longer chained PFASs [63] the most common industrial practice of phasing out one PFAS is to replace it with another structurally similar PFAS => will not solve issues in relation to PFASs as a whole group => let us start the dialogue in defining "essential" and "nonessential" uses of PFASs, while simultaneously developing safe alternative substances and processes for those essential uses [99] PFHxA: leachate from 6 landfills and a laboratory bioreactor (US) (ng/L): - Site A: 790 50 - Site B: 270 17 - Site C: 620 14 - Site D2: 360 12 - Site D3: 350 21 - Site D6: 430 19 - Laboratory bioreactor: 2200 140 [99] - Perfluoroalkyl carboxylates accounted for the majority (67 4%) of the fluorochemicals quantified in leachates - Of the 14 individual (C4 to C14) carboxylate forms measured in the present study, the most abundant were C4-C10 with only infrequent detection of C11-C14 homologs above quantification limits [120] 17 PFAS were analysed in groundwater surrounding legacy landfills in a major Australian urban re-development precinct (Fishermans Bend, Melbourne). * PFHxA concentrations in Fishermans Bend groundwater (concentrations in ng/L; average of duplicate sample analysis): - LOD: 0.2 - LOQ: 0.7 - Detection frequency (%): 85 - Minimum: <0.2 - Maximum: 46 - Median: 19 [120] * PFHxS had the highest median concentration (34 ng/L; range: 2.6 280 ng/L) followed by PFOS (26 ng/L; range: 1.3 - 4,800 ng/L), PFHxA (19 ng/L; range: <LOQ - 46 ng/L) and PFOA (12 ng/L; range: 1.7 - 74 ng/L). [118] Spatial and seasonal distributions of PFAS in air, water, sediment, soil, and fish samples in the Asan Lake area of South Korea. Estimation of bioaccumulation potential of PFAS using sedimentwater partition coefficients and bioaccumulation factors. PFHxA concentration in fish samples: - Fish (unit: ng/g ww, n=42): * Range: nd * Mean: * Median: * 75%: 0.00 * 95%: 0.00 * DF (%): 0.0 [118] Longer-chain PFAS, especially PFOS, can accumulate more in the muscles and lipids of fish than can short-chain PFAS. [112] * Translocation factor (TF) = ratio of pollutant concentration in shoots to that in roots => TF values of PFAAs decreased with increasing carbon chain length and PFSAs had lower TF than PFCAs with similar carbon chain length => PFAAs with shorter carbon chain length can be taken up by J. effusus roots and tended to translocate to plant shoots * TF for PFPeA, PFBS, PFHxA and PFHpA were always greater than 2 at different timepoints and reached the highest values on Day-21 [114] * Car interior materials (3): - Hyundai - textile material: 0.135 g/kg - Hyundai - textile material: 0.271 g/kg - Skoda - textile material: 0.138 g/kg - Skoda - plastic material: 0.433 g/kg * WEEE (4): - Spaghetti insulation II: 0.07 g/kg - Mixed waste: 0.172 g/kg - Rubber: 0.162 g/kg [21] high PFHxA and PFHpA levels often found at locations with point sources of PFCAs (such as air force bases where AFFF is used or wastewater treatment plants) [112] * Removal efficiencies of PFAAs by shoots and roots increased with increasing exposure time * Regarding each PFAA, J. effusus roots had lower removal efficiency of PFAAs (except PFOS) than shoots * On Day-21, the whole J. effusus had relatively higher removal efficiencies of PFHxA, PFOA and PFOS * Plants exposed to PFAAs at 1x generally had higher removal efficiencies than at 10x * Both concentrations of PFAAs in the growth medium and PFAAs exposure time can affect the PFAAs removal efficiency by J. effusus [120] The most common applications include textile protection (ScotchgardTM), surface coating for cooking implements (TeflonTM), food contact paper, and Aqueous Film Forming Foams (AFFFs). [25] PFHxA = Impurity and ultimate degradation product of 6:2 fluorotelomer-based AFFFs [71] - point sources: wastewater treatment plants (WWTPs), industrial emissions from PFAS production sites, commercial and military airfields (use of AFFF-containing PFASs and their precursors), landfills that contain PFASs-contaminated waste [71] - diffuse sources: contamination caused from a range of dispersed urban and rural land use activities, such as atmospheric deposition and upstream water input [26] Estimated global emissions of PFHxA: 1951-2002: lower: 16 tonnes => 26% from direct sources, higher: 1061 tonnes => 26% from direct sources 2003-2015: lower: 17 tonnes => 80% from direct sources, higher: 513 tonnes => 16% from direct sources 2016-2030: lower: 5 tonnes => 98% from direct sources higher: 117 tonnes => 48% from direct sources [28] Northern rivers had higher fractions of longchained PFCAs than rivers in the south and west where higher fractions of short-chain PFCAs were observed => stronger relative influence of atmospheric deposition in the north due to long-range transport of long-chained PFCAs and their precursors => stronger relative influence of contemporary point sources in the south due to the replacement of C8based PFASs, by shorter-chained PFAss [28] - Northern rivers had lower levels of PFASs than rivers on south-east and south-west coast (SE) - Northern sub-basins of Baltic Sea had lower levels of PFASs than in southern sub-basins => higher riverine input of PFASs in the south with its higher population density compared to the north [63] due to their high persistence and water solubility, PFAAs will be transported to remote locations from sources through water currents and aerosols [28] - Shorter chained PFCAs were shown to correlate with low latitude (south) (high population density) - Long chained PFCAs were shown to correlate with high latitude (north) => higher relative abundance of longer chained PFCAs in the north due to precursor degradation and aerosol associated stabilization of PFCAs and their precursors in the atmosphere [74] transport of PFHxA in a water-saturated sediment column => recoveries of PFHxA were in the same range as the tracer indicating complete breakthrough => breakthrough curves, recoveries and partition coefficients showed that short-chain PFCAs and PFSAs (i.e. with up to six C-atoms) are only slightly retarded in the water-saturated sediment column => if contaminated surface waters are used as a resource for drinking water production via sediment passage, short-chain PFCAs and PFSAs will not be subject to attenuation [34] Surface water data SE: slightly higher PFAS concentrations in southern Sweden (higher population density) compared to northern Sweden [77] - one of the major challenges associated with PFASs is the relatively high mobility and persistence of these compounds in the subsurface => this creates the potential for large plumes in transmissive hydrogeological settings - the persistence of PFAAs coupled with their high solubility, low/moderate sorption to soils and lack of volatility make many PFAAs highly mobile, resulting in long (i.e., potentially multiple mile) groundwater plumes - shorter chain PFAAs generally have lower organic carbon partitioning coefficients than the longer chain compounds => therefore, they are expected to be more mobile in aquifer systems [63] past and ongoing production and use will lead to the accumulation of PFAAs in the global environment, with very slow mixing/sedimentation to the deep oceans and sedimentation/burial in deep sediments as the only known global environmental sinks [22] * PFHxS in leachate landfill (AU): - concentrations: mean: 1200 ng/L, range: 56-16000 ng/L - contribution to total PFASs: mean: 24%, range: 1-49% * newer landfills have higher PFAS concentrations => current influx of PFAS-containing products into the waste stream, despite regulatory action more than a decade ago [32] Concentration of lin-PFHxS in sludge samples (12 datasets): - Assigned value: NA - Average: 1.67 ng/L - Median: 1.40 ng/L - Min.: 0.24 ng/L - Max.: 4.46 ng/L - SD: 1.28 - % relative SD: 77 [118] Spatial and seasonal distributions of PFAS in air, water, sediment, soil, and fish samples in the Asan Lake area of South Korea. Estimation of bioaccumulation potential of PFAS using sedimentwater partition coefficients and bioaccumulation factors. [32] Concentration of lin-PFHxS in fish samples (7 datasets): - Assigned value: NA - Average: 0.89 ng/g ww - Median: 0.11 ng/g ww - Min.: 0.00 ng/g ww - Max.: 5.10 ng/g ww - SD: 1.87 - % relative SD: 210 [10] effects on the next generation can already be observed if a species is exposed to PFASs for only 1 generation => more than 1 generation may suffer from PFAS exposure [70] * Concentration and trends in the USA: - PFSA concentrations in leachate have varied from 50 to 3200 ng/L in the USA, with median concentrations of a few hundred ng/L for PFBS, PFHxS and PFOS - While PFOS was detected in all leachate samples, its concentration was generally lower than that of PFBS and PFHxS - This dominance of shorter-chain PFSAs over historically used PFOS could be indicative of the transition towards C4-based chemistry (even prior to 2002) * Concentration and trends in Canada: - PFSAs in leachates collected from Canadian landfills: Median PFHxS concentration of 200 ng/L higher than for PFOS and PFBS in a cross-Canada study [99] PFHxS: leachate from 6 landfills and a laboratory bioreactor (US) (ng/L): - Site A: 700 19 - Site B: 160 8.2 - Site C: 430 13 - Site D2: 170 7 - Site D3: 200 24 - Site D6: 360 110 - Laboratory bioreactor: 120 14 [99] Perfluoroalkyl sulfonates were the second most abundant class of fluorochemicals at 22 2% in landfill leachates [118] L-PFHxS concentration in soil samples: - Soils (unit: ng/g dw, n=48): * Range: nd * Mean: * Median: * 75%: 0.00 * 95%: 0 * DF (%): 0.0 [118] Spatial and seasonal distributions of PFAS in air, water, sediment, soil, and fish samples in the Asan Lake area of South Korea. Estimation of bioaccumulation potential of PFAS using sedimentwater partition coefficients and bioaccumulation factors. L-PFHxS concentration in fish samples: - Fish (unit: ng/g ww, n=42): * Range: nd ~ 6.09 * Mean: 0.45 * Median: 0.17 * 75%: 0.66 * 95%: 1.06 * DF (%): 59.5 Longer-chain PFAS, especially PFOS, can accumulate more in the muscles and lipids of fish than can short-chain PFAS. [120] 17 PFAS were analysed in groundwater surrounding legacy landfills in a major Australian urban re-development precinct (Fishermans Bend, Melbourne). * PFHxS concentrations in Fishermans Bend groundwater (concentrations in ng/L; average of duplicate sample analysis): - LOD: 0.2 - LOQ: 0.2 [120] - Detection frequency (%): 100 - Minimum: 2.6 - Maximum: 280 - Median: 34 * PFHxS had the highest median concentration (34 ng/L; range: 2.6 280 ng/L) followed by PFOS (26 ng/L; range: 1.3 - 4,800 ng/L), PFHxA (19 ng/L; range: <LOQ - 46 ng/L) and PFOA (12 ng/L; range: 1.7 - 74 ng/L). [35] control plants (wheat, soybean, pumpkin): PFHxS (0.0009-0.00091 nmol/g) in shoots - exposured plants with solution of N-EtFOSA (wheat, soybean, pumpkin): PFHxS (0.0237-0.0674 nmol/g) in roots PFHxS (0.0013-0.0292 nmol/g) in shoots (sign. higher than control p<0.05) - translocation ratio (TR) wheat: PFHxS (0.739) - translocation ratio (TR) soybean: PFHxS (0.017) - translocation ratio (TR) pumpkin: PFHxS (0.062) [114] Analysis of perfluoroalkyl substances in 126 individual samples of building materials, consumer products, car interior materials and wastes which potentially affect indoor environments where people spend most of their time. PFHxS concentrations in individual samples: * Textile (1A): - New curtain: 2.18 g/kg - Foam: 0.586 g/kg - Foam: 0.409 g/kg - Bed cover: 0.161 g/kg - Hanging: 0.512 g/kg - Teddy bear (cover): 0.439 g/kg [2], [6], [7], [8], [14], [22], [62] [71] Shorter chains have been reported to have a higher uptake in lettuce leaves, while longer chain PFASs are found primarily in the roots [114] - Tablecloth: 0.437 g/kg - Upholstery material: 2.93 g/kg - Foam: 1.085 g/kg - Upholstery material: 0.835 g/kg - Coir (coconut fibre): 0.68 g/kg - Foam: 1.09 g/kg - Blanket: 0.219 g/kg * Floor covering (1B): - New carpet - red: 0.256 g/kg - Carpet - grey: 5.16 g/kg - Foam: 1.43 g/kg - Blanket: 0.286 g/kg - Carpet - grey: 0.966 g/kg - Carpet - green: 0.448 g/kg - Carpet - brown: 0.274 g/kg - Persian carpet: 3.24 g/kg - Carpet - green: 1.25 g/kg * EEE (1C): - Switch: 0.054 g/kg - Keyboard II: 0.059 g/kg - TV: 0.115 g/kg [9] waste water treatment plants have been suggested as one of the major point sources of PFASs to surface waters and the atmosphere [112] Study investigated plant uptake of PFAAs by a native wetland species in the US, Juncus effusus * PFHxS concentration in shoots: - Timepoint: p=0.009 - Concentration: p<0.001 - Timepoint x Concentration: p=0.074 * PFHxS concentration in roots: - Timepoint: p=0.191 - Concentration: p<0.001 - Timepoint x Concentration: p=0.102 * J. effusus plants could translocate all 7 PFAAs during the 21-day experimental period [112] * Plants exposed to PFAAs at 10x had higher PFAAs concentrations than the plants exposed to those at 1x * Uptake of PFAAs increased with increasing exposure time * PFAAs, except PFOS, had higher concentrations in the plant shoots than in the plant roots * PFAAs concentration in the growth medium was a significant factor on the uptake of all 7 PFAAs for both shoots and roots * Exposure time also had significant impact on the plant uptake of PFAAs, except PFPeA and PFHxS by roots * Significant interactions between exposure time and PFAAs concentration regarding their impact on the uptake of PFHxA, PFHpA and PFOS by shoots and PFBS and [114] - Fridge - rubber insulation: 0.09 g/kg * OSB and wood (2A): - Chipboard: 0.207 g/kg * Insulation materials (2B): - Blow cellulose insulation: 0.642 g/kg - Pipe insulation aeroflex: 1.43 g/kg - TETRAPAK FLEXIBUILD: 0.073 g/kg - Insulation FOAMGLAS PERINSUL: 0.078 g/kg - Sound insulation WOLF Professional: 0.036 g/kg - Paper insulation: 0.372 g/kg [114] * Mounting and sealing foam (2C): - Mounting foam - green: 0.662 g/kg * Facade materials (2D): - Window finishing bead: 0.252 g/kg - Window corner bead: 1.48 g/kg - Drywall: 24.5 g/kg - Plaster: 0.182 g/kg * Car interior materials (3): - Hyundai - textile material: 0.147 g/kg - Hyundai - textile material: 0.479 g/kg - Skoda - textile material: 0.027 g/kg * WEEE (4): - Impactor undersized particles: 0.018 g/kg [112] * Translocation factor (TF) = ratio of pollutant concentration in shoots to that in roots => TF values of PFAAs decreased with increasing carbon chain length and PFSAs had lower TF than PFCAs with similar carbon chain length => PFAAs with shorter carbon chain length can be taken up by J. effusus roots and tended to translocate to plant shoots * TF for PFOA and PFHxS were approximately 1 on Day-7 and increased to [120] The most common applications include textile protection (ScotchgardTM), surface coating for cooking implements (TeflonTM), food contact paper, and Aqueous Film Forming Foams (AFFFs). [112] * Removal efficiencies of PFAAs by shoots and roots increased with increasing exposure time * Regarding each PFAA, J. effusus roots had lower removal efficiency of PFAAs (except PFOS) than shoots * On Day-21, the whole J. effusus had relatively higher removal efficiencies of PFHxA, PFOA and PFOS * Plants exposed to PFAAs at 1x generally had higher removal efficiencies than at 10x * Both concentrations of PFAAs in the growth medium and PFAAs exposure time can affect the PFAAs removal efficiency by J. effusus [10] - point sources: landfills, manufacturing plants, application of PFAS-containing products at a concentrated area, industrial and municipal sewage treatment plants - nonpoint sources: wet and dry atmospheric deposition [71] - point sources: wastewater treatment plants (WWTPs), industrial emissions from PFAS production sites, commercial and military airfields (use of AFFF-containing PFASs and their precursors), landfills that contain PFASs-contaminated waste [71] - diffuse sources: contamination caused from a range of dispersed urban and rural land use activities, such as atmospheric deposition and upstream water input [100] - exposure of PFASs to humans occurs mainly through food, water, and the indoor environment - in addition to direct exposure, humans are also indirectly exposed to PFAS through precursor compounds that degrade to PFCA and PFSA [19] Calculated global emissions of FHxSA/Es (based on PFOS emissions): 1958-2002: lower: 5 tonnes, higher: 551 tonnes 2003-2015: lower: 2 tonnes, higher: 68 tonnes 2016-2030: lower: 0 tonnes, higher: 49 tonnes [19] Calculated global emissions of PFHxS (based on PFOS emissions): 1958-2002: lower: 21 tonnes, higher: 208 tonnes 2003-2015: lower: 2 tonnes, higher: 40 tonnes 2016-2030: lower: 0 tonnes, higher: 8 tonnes [23] in comparison to long-chain homologues, short-chain PFAAs are more mobile in soil and sediment => higher solubility in water and lower sorption to solids => higher mobility in the environment [23] Replacements in the future might still use the same concept by slightly modifying the currently used fluorinated alternatives => non-fluorinated parts of these structures may undergo (bio)degradation but transformation products are still likely to be environmentally stable PFECAs or short-chain PFCAs => a non-fluorinated replacement strategy should be pursued [28] - Northern rivers had lower levels of PFASs than rivers on south-east and south-west coast (SE) - Northern sub-basins of Baltic Sea had lower levels of PFASs than in southern sub-basins => higher riverine input of PFASs in the south with its higher population density compared to the north [28] Long-chain PFASs were predominant in sea, while short-chain PFASs predominated in rivers (SE) => long-chain PFASs are highly present in atmospheric deposition, while in the terrestrial environment long-chain substances are retained and do not readily reach the rivers => in addition short-chain PFASs have higher water solubility and reach water bodies more easily than longer chained PFASs [25] all PFASs which contain persistent perfluoroalkyl moieties will ultimately degrade to form non-degradable P chemicals in the environment => poorly reversible PFAS exposures will arise in (parts of) the environment => if poorly reversible exposure had been used as a criterion, then PFAS would have been regulated at an earlier stage => usual risk-based regulatory approach has been applied instead of precautionary approach => eventual replacement of persistent PFASs in fire-fighting foams with degradable alternatives that achieve acceptable functionality is a strategy that is in harmony with a future sustainable society [28] - Shorter chained PFCAs were shown to correlate with low latitude (south) (high population density) - Long chained PFCAs were shown to correlate with high latitude (north) => higher relative abundance of longer chained PFCAs in the north due to precursor degradation and aerosol associated stabilization of PFCAs and their precursors in the atmosphere [28] Northern rivers had higher fractions of longchained PFCAs than rivers in the south and west where higher fractions of short-chain PFCAs were observed => stronger relative influence of atmospheric deposition in the north due to long-range transport of long-chained PFCAs and their precursors => stronger relative influence of contemporary point sources in the south due to the replacement of C8based PFASs, by shorter-chained PFAss [63] the most common industrial practice of phasing out one PFAS is to replace it with another structurally similar PFAS => will not solve issues in relation to PFASs as a whole group => let us start the dialogue in defining "essential" and "nonessential" uses of PFASs, while simultaneously developing safe alternative substances and processes for those essential uses [34] Surface water data SE: slightly higher PFAS concentrations in southern Sweden (higher population density) compared to northern Sweden [63] due to their high persistence and water solubility, PFAAs will be transported to remote locations from sources through water currents and aerosols [63] past and ongoing production and use will lead to the accumulation of PFAAs in the global environment, with very slow mixing/sedimentation to the deep oceans and sedimentation/burial in deep sediments as the only known global environmental sinks [74] transport of PFHxS in a water-saturated sediment column => recoveries of PFHxS were in the same range as the tracer indicating complete breakthrough => breakthrough curves, recoveries and partition coefficients showed that short-chain PFCAs and PFSAs (i.e. with up to six C-atoms) are only slightly retarded in the water-saturated sediment column => if contaminated surface waters are used as a resource for drinking water production via sediment passage, short-chain PFCAs and PFSAs [10] effects on the next generation can already be observed if a species is exposed to PFASs for only 1 generation => more than 1 generation may suffer from PFAS exposure [22] * PFHpA in leachate landfill (AU): - concentrations: mean: 430 ng/L, range: 18-4400 ng/L - contribution to total PFASs: mean: 11%, range: 5-21% * newer landfills have higher PFAS concentrations => current influx of PFAS-containing products into the waste stream, despite regulatory action more than a decade ago [32] Concentration of PFHpA in sludge samples (12 datasets): - Assigned value: NA - Average: 1.37 ng/L - Median: 0.65 ng/L - Min.: 0.21 ng/L - Max.: 8.00 ng/L - SD: 2.14 - % relative SD: 156 [70] General observations and geographical trends of PFASs in leachate: - PFCAs are generally found to be the dominant PFASs - C4-C7 chain length PFAAs are more abundant than their longer-chain (C8) homologues [70] - Short-chain PFAAs are prone to preferential release and leaching from municipal solid waste (MSW), consistent with their higher aqueous solubilities and lower organic carbonwater partition coefficients relative to longer-chain PFAAs - Dominance of C4-C7 PFAAs could be related to the shift towards production of shorter-chain perfluorinated compounds since the early 2000's [118] Spatial and seasonal distributions of PFAS in air, water, sediment, soil, and fish samples in the Asan Lake area of South Korea. Estimation of bioaccumulation potential of PFAS using sedimentwater partition coefficients and bioaccumulation factors. [118] PFHpA concentration in soil samples: - Soils (unit: ng/g dw, n=48): * Range: nd ~ 0.74 * Mean: 0.16 * Median: 0.12 * 75%: 0.22 * 95%: 0.58 * DF (%): 93.8 [32] Concentration of PFHpA in fish samples (3 datasets): - Assigned value: NA - Average: 1.28 ng/g ww - Median: 1.00 ng/g ww - Min.: 0.80 ng/g ww - Max.: 2.05 ng/g ww - SD: 0.67 - % relative SD: 52 [10] effects on the next generation can already be observed if a species is exposed to PFASs for only 1 generation => more than 1 generation may suffer from PFAS exposure [118] Spatial and seasonal distributions of PFAS in air, water, sediment, soil, and fish samples in the Asan Lake area of South Korea. Estimation of bioaccumulation potential of PFAS using sedimentwater partition coefficients and bioaccumulation factors. PFHpA concentration in fish samples: - Fish (unit: ng/g ww, n=42): * Range: nd ~ 0.33 * Mean: 0.01 * Median: 0.00 * 75%: 0.00 * 95%: 0.00 * DF (%): 2.4 [94] Algae isolated from Baltic Sea coastal waters and maintained as monoalgal cultures Chlorella vulgaris (acute aquatic toxicity): - 72h EC50: 5.21 0.26 mM, logEC50: 0.72 Skeletonema marinoi (acute aquatic toxicity): - 72h EC50: 2.40 0.12 mM, logEC50: 0.38 Geitlerinema amphibium (acute aquatic toxicity): - 72h EC50: 1.42 0.07 mM, logEC50: 0.15 => Results show that these algae are highly sensitive to PFCAs, although the cyanobacteria and diatoms are far more sensitive to these compounds than the green algae => A linear relationship was also found between lipophilicity expressed as alkyl chain length or partition coefficient and toxicity expressed as logEC50 [118] The predominant PFAS in soils was PFOA (24.8% of the total), followed by PFOS (22.5% of the total). PFHpA and PFDA were the most frequently detected species in soils, followed by PFOA. [118] Longer-chain PFAS, especially PFOS, can accumulate more in the muscles and lipids of fish than can short-chain PFAS. [99] PFHpA: leachate from 6 landfills and a laboratory bioreactor (US) (ng/L): - Site A: 328 21 - Site B: 100 14 - Site C: 340 15 - Site D2: 170 4.3 - Site D3: 150 10 - Site D6: 170 3.6 - Laboratory bioreactor: 2800 89 [99] - Perfluoroalkyl carboxylates accounted for the majority (67 4%) of the fluorochemicals quantified in leachates - Of the 14 individual (C4 to C14) carboxylate forms measured in the present study, the most abundant were C4-C10 with only infrequent detection of C11-C14 homologs above quantification limits [120] 17 PFAS were analysed in groundwater surrounding legacy landfills in a major Australian urban re-development precinct (Fishermans Bend, Melbourne). * PFHpA concentrations in Fishermans Bend groundwater (concentrations in ng/L; average of duplicate sample analysis): - LOD: 0.2 [120] - LOQ: 0.2 - Detection frequency (%): 31 - Minimum: <0.2 - Maximum: 22 - Median: <0.2 [27] - 1986: not reported - 1990: 4.54 - 227 t/yr - 1994: not reported - 1998: not reported - 2002: not reported - 2006: not reported [10] - point sources: landfills, manufacturing plants, application of PFAS-containing products at a concentrated area, industrial and municipal sewage treatment plants - nonpoint sources: wet and dry [69] PFPAs: - high acidity - adsorption on organic matter increases with increasing perfluoroalkyl chain length - certain homologs (C6) stay primarily in aqueous phase => likely LRTP [63] past and ongoing production and use will lead to the accumulation of PFAAs in the global environment, with very slow mixing/sedimentation to the deep oceans and sedimentation/burial in deep sediments as the only known global environmental sinks [23] PFECAs and PFESAs are expected to be highly persistent and have physicochemical properties (high water solubility and low pKa) similar to those of PFCAs and PFSAs => likely to have similarly high long-range transport potential in water [23] Replacements in the future might still use the same concept by slightly modifying the currently used fluorinated alternatives => non-fluorinated parts of these structures may undergo (bio)degradation but transformation products are still likely to be environmentally stable PFECAs or short-chain PFCAs => a non-fluorinated replacement strategy should be pursued [63] due to their high persistence and water solubility, PFAAs will be transported to remote locations from sources through water currents and aerosols [63] the most common industrial practice of phasing out one PFAS is to replace it with another structurally similar PFAS => will not solve issues in relation to PFASs as a whole group => let us start the dialogue in defining "essential" and "nonessential" uses of PFASs, while simultaneously developing safe alternative substances and processes for those essential uses [75] - critical travel distance (CTD): 1745.31 km - travel efficiency (TE): 0.0025 % => same CTD as PFOA => lower TE than PFOA [27] - 1986: 4.54 - 227 t/yr - 1990: 4.54 - 227 t/yr - 1994: 4.54 - 227 t/yr - 1998: 4.54 - 227 t/yr - 2002: 227 - 454 t/yr - 2006: not reported [19] Calculated global emissions of PHxSF (based on PFOS emissions): 1958-2002: lower: 14 tonnes, higher: 40 tonnes 2003-2015: lower: 1 tonnes, higher: 2 tonnes 2016-2030: lower: 0 tonnes, higher: 1 tonnes [63] past and ongoing production and use will lead to the accumulation of PFAAs in the global environment, with very slow mixing/sedimentation to the deep oceans and sedimentation/burial in deep sediments as the only known global environmental sinks [23] PFECAs and PFESAs are expected to be highly persistent and have physicochemical properties (high water solubility and low pKa) similar to those of PFCAs and PFSAs => likely to have similarly high long-range transport potential in water [23] Replacements in the future might still use the same concept by slightly modifying the currently used fluorinated alternatives => non-fluorinated parts of these structures may undergo (bio)degradation but transformation products are still likely to be environmentally stable PFECAs or short-chain PFCAs => a non-fluorinated replacement strategy should be pursued [63] due to their high persistence and water solubility, PFAAs will be transported to remote locations from sources through water currents and aerosols [63] the most common industrial practice of phasing out one PFAS is to replace it with another structurally similar PFAS => will not solve issues in relation to PFASs as a whole group => let us start the dialogue in defining "essential" and "nonessential" uses of PFASs, while simultaneously developing safe alternative substances and processes for those essential uses [75] - critical travel distance (CTD): 592.69 km - travel efficiency (TE): 0.0146 % => lower CTD than PFOA => similar TE as PFOA [71] Shorter chains have been reported to have a higher uptake in lettuce leaves, while longer chain PFASs are found primarily in the roots [114] Analysis of perfluoroalkyl substances in 126 individual samples of building materials, consumer products, car interior materials and wastes which potentially affect indoor environments where people spend most of their time. PFHpA concentrations in individual samples: * Textile (1A): - Foam: 12.0 g/kg - Coir (coconut fibre): 0.324 g/kg * Floor covering (1B): - Carpet - grey: 2.87 g/kg - Carpet - grey: 0.548 g/kg [2], [6], [8], [22] [112] Study investigated plant uptake of PFAAs by a native wetland species in the US, Juncus effusus * PFHpA concentration in shoots: - Timepoint: p<0.001 - Concentration: p<0.001 - Timepoint x Concentration: p=0.016 * PFHpA concentration in roots: - Timepoint: p=0.036 - Concentration: p<0.001 - Timepoint x Concentration: p=0.104 * J. effusus plants could translocate all 7 PFAAs during the 21-day experimental period * Plants exposed to PFAAs at 10x had higher PFAAs concentrations than the plants exposed to those at 1x [114] - Laminated plastic floor covering: 1.98 g/kg * OSB and wood (2A): - Formica: 4.62 g/kg - Oriented Strand Board: 0.879 g/kg - Oriented Strand Board - white: 1 g/kg - Wooden board: 13.35 g/kg - Chipboard - brown: 1.38 g/kg - Oriented Strand Board - white: 12.79 g/kg - Chipboard: 0.592 g/kg - Chipboard: 1.81 g/kg - Chipboard: 1.58 g/kg * Insulation materials (2B): - Insulation hemp rope: 0.738 g/kg - Blow cellulose insulation: 0.828 g/kg [112] * Uptake of PFAAs increased with increasing exposure time * PFAAs, except PFOS, had higher concentrations in the plant shoots than in the plant roots * PFAAs concentration in the growth medium was a significant factor on the uptake of all 7 PFAAs for both shoots and roots * Exposure time also had significant impact on the plant uptake of PFAAs, except PFPeA and PFHxS by roots * Significant interactions between exposure time and PFAAs concentration regarding their impact on the uptake of PFHxA, PFHpA and PFOS by shoots and PFBS and PFOA by both shoots and roots [114] - Wooden fibre insulation: 28.38 g/kg - Insulation CANABEST PANEL: 0.64 g/kg - Wooden fibre insulation: 20.59 g/kg * Facade materials (2D): - Window corner bead: 1.72 g/kg * Car interior materials (3): - Hyundai - textile material: 0.249 g/kg - Hyundai - textile material: 0.569 g/kg - Skoda - plastic material: 0.101 g/kg * WEEE (4): - Mixed waste - small electronical devices: 0.03 g/kg - Mixed waste: 0.115 g/kg - Rubber: 0.058 g/kg [10] - point sources: landfills, manufacturing plants, application of PFAS-containing products at a concentrated area, industrial and municipal sewage treatment plants - nonpoint sources: wet and dry atmospheric deposition [21] emissions of PFCAs divided in two source categories: - direct emission sources = emissions that come from the manufacture, intentional use and disposal of PFCAs throughout their product life-cycle - indirect emission sources = emissions of a given PFCA that is present as an impurity in a product, or formed by degradation (in environment, wildlife, humans) of a precursor substance (PFCA formed from the atmospheric degradation of perfluorooctane sulfonamido ethanol or from the biotransformation of a fluorotelomer alcohol) [112] * Translocation factor (TF) = ratio of pollutant concentration in shoots to that in roots => TF values of PFAAs decreased with increasing carbon chain length and PFSAs had lower TF than PFCAs with similar carbon chain length => PFAAs with shorter carbon chain length can be taken up by J. effusus roots and tended to translocate to plant shoots * TF for PFPeA, PFBS, PFHxA and PFHpA were always greater than 2 at different timepoints and reached the highest values on Day-21 [114] - In group 2B, two samples of wood fibre insulation (produced in 2010) contained high amounts of PFHpA (20.6 and 28.4 g/kg) and other 5- to 8-carbon chain PFCAs (12.3 and 5.8 g/kg). - Following a recalculation using insulation area and material density, the PFHpA concentration was found to be 61.5 g/m2 (wall insulation) and 181.8 g/m2 (floor insulation). [21] high PFHxA and PFHpA levels often found at locations with point sources of PFCAs (such as air force bases where AFFF is used or wastewater treatment plants) [112] * Removal efficiencies of PFAAs by shoots and roots increased with increasing exposure time * Regarding each PFAA, J. effusus roots had lower removal efficiency of PFAAs (except PFOS) than shoots * On Day-21, the whole J. effusus had relatively higher removal efficiencies of PFHxA, PFOA and PFOS * Plants exposed to PFAAs at 1x generally had higher removal efficiencies than at 10x * Both concentrations of PFAAs in the growth medium and PFAAs exposure time can affect the PFAAs removal efficiency by J. effusus [114] - The presence of PFHpA, which is predominantly a breakdown product of coatings on carpets, textiles and food packaging suggests breakdown of coatings, potentially beginning rapidly after product manufacture. [71] - point sources: wastewater treatment plants (WWTPs), industrial emissions from PFAS production sites, commercial and military airfields (use of AFFF-containing PFASs and their precursors), landfills that contain PFASs-contaminated waste [120] The most common applications include textile protection (ScotchgardTM), surface coating for cooking implements (TeflonTM), food contact paper, and Aqueous Film Forming Foams (AFFFs). [71] - diffuse sources: contamination caused from a range of dispersed urban and rural land use activities, such as atmospheric deposition and upstream water input [100] - exposure of PFASs to humans occurs mainly through food, water, and the indoor environment - in addition to direct exposure, humans are also indirectly exposed to PFAS through precursor compounds that degrade to PFCA and PFSA [75] - critical travel distance (CTD): 69066.89 km - travel efficiency (TE): 91.819 % => higher CTD and TE than 8:2 FTOH [21] all manufacturing of PFCAs occurs in northern hemisphere => PFCAs can only be detected in oceans north of the equator + concentrations of volatile precursors in remote environments originating from consumer products are lower in southern hemisphere => BUT Australians have similar PFCA-serum levels to humans in industrialized countries in northern hemisphere => localized environmental or consumer exposure from product use are similar in both hemispheres [75] - critical travel distance (CTD): 4864.48 km - travel efficiency (TE): 0.01353 % => higher CTD than PFOA/PFOS => similar TE as PFOA [28] Long-chain PFASs were predominant in sea, while short-chain PFASs predominated in rivers (SE) => long-chain PFASs are highly present in atmospheric deposition, while in the terrestrial environment long-chain substances are retained and do not readily reach the rivers => in addition short-chain PFASs have higher water solubility and reach water bodies more easily than longer chained PFASs [63] the most common industrial practice of phasing out one PFAS is to replace it with another structurally similar PFAS => will not solve issues in relation to PFASs as a whole group => let us start the dialogue in defining "essential" and "nonessential" uses of PFASs, while simultaneously developing safe alternative substances and processes for those essential uses [21] levels of volatile precursors of PFCAs and levels of PFCAs in precipitation are highest near urban areas => urban areas are sources of PFCAs to the atmosphere [28] Northern rivers had higher fractions of longchained PFCAs than rivers in the south and west where higher fractions of short-chain PFCAs were observed => stronger relative influence of atmospheric deposition in the north due to long-range transport of long-chained PFCAs and their precursors => stronger relative influence of contemporary point sources in the south due to the replacement of C8based PFASs, by shorter-chained PFAss [26] Estimated global emissions of PFHpA: 1951-2002: lower: 44 tonnes => 17% from direct sources, higher: 2123 tonnes => 19% from direct sources 2003-2015: lower: 13 tonnes => 51% from direct sources, higher: 774 tonnes => 24% from direct sources 2016-2030: lower: 2 tonnes => 94% from direct sources higher: 358 tonnes => 64% from direct sources [63] due to their high persistence and water solubility, PFAAs will be transported to remote locations from sources through water currents and aerosols [28] - Northern rivers had lower levels of PFASs than rivers on south-east and south-west coast (SE) - Northern sub-basins of Baltic Sea had lower levels of PFASs than in southern sub-basins => higher riverine input of PFASs in the south with its higher population density compared to the north [28] - Shorter chained PFCAs were shown to correlate with low latitude (south) (high population density) - Long chained PFCAs were shown to correlate with high latitude (north) => higher relative abundance of longer chained PFCAs in the north due to precursor degradation and aerosol associated stabilization of PFCAs and their precursors in the atmosphere [34] Surface water data SE: slightly higher PFAS concentrations in southern Sweden (higher population density) compared to northern Sweden [63] past and ongoing production and use will lead to the accumulation of PFAAs in the global environment, with very slow mixing/sedimentation to the deep oceans and sedimentation/burial in deep sediments as the only known global environmental sinks [120] 17 PFAS were analysed in groundwater surrounding legacy landfills in a major Australian urban re-development precinct (Fishermans Bend, Melbourne). * PFHpS concentrations in Fishermans Bend groundwater (concentrations in ng/L; average of duplicate sample analysis): - LOD: 0.2 - LOQ: 0.2 - Detection frequency (%): 39 - Minimum: <0.2 - Maximum: 7.1 [118] Spatial and seasonal distributions of PFAS in air, water, sediment, soil, and fish samples in the Asan Lake area of South Korea. Estimation of bioaccumulation potential of PFAS using sedimentwater partition coefficients and bioaccumulation factors. [118] L-PFHpS concentration in soil samples: - Soils (unit: ng/g dw, n=48): * Range: na * Mean: na * Median: na * 75%: na * 95%: na * DF (%): na [118] Spatial and seasonal distributions of PFAS in air, water, sediment, soil, and fish samples in the Asan Lake area of South Korea. Estimation of bioaccumulation potential of PFAS using sedimentwater partition coefficients and bioaccumulation factors. L-PFHpS concentration in fish samples: - Fish (unit: ng/g ww, n=42): * Range: nd ~ 0.58 * Mean: 0.02 * Median: 0.00 * 75%: 0.00 * 95%: 0.07 * DF (%): 14.3 [118] Longer-chain PFAS, especially PFOS, can accumulate more in the muscles and lipids of fish than can short-chain PFAS. [114] Analysis of perfluoroalkyl substances in 126 individual samples of building materials, consumer products, car interior materials and wastes which potentially affect indoor environments where people spend most of their time. PFHpS concentrations in individual samples: * Textile (1A): - Stain resistant upholstery material: 73.8 g/kg - New curtain: 0.164 g/kg - Foam: 0.109 g/kg - Tablecloth: 0.105 g/kg [114] - Upholstery material: 0.388 g/kg - Foam: 0.227 g/kg - Upholstery material: 0.232 g/kg - Coir (coconut fibre): 0.285 g/kg - Foam: 0.765 g/kg - Blanket: 0.088 g/kg * Floor covering (1B): - Carpet - grey: 0.443 g/kg - Carpet - grey: 0.132 g/kg - Carpet - green: 0.474 g/kg - Persian carpet: 0.727 g/kg - New carpet - red: 0.256 g/kg - Carpet - grey: 5.16 g/kg * EEE (1C): - Screen I: 0.026 g/kg - Keyboard II: 0.025 g/kg * OSB and wood (2A): - Oriented Strand Board: 1.36 g/kg - Wooden board: 0.132 g/kg [100] - exposure of PFASs to humans occurs mainly through food, water, and the indoor environment - in addition to direct exposure, humans are also indirectly exposed to PFAS through precursor compounds that degrade to PFCA and PFSA [114] * Insulation materials (2B): - Phenolic foam insulation ISOVER WEBER TERRANOVA: 1.26 g/kg - TETRAPAK FLEXIBUILD: 0.519 g/kg - Insulation aluminium foil: 0.098 g/kg - Paper insulation: 0.178 g/kg * Mounting and sealing foam (2C): - Mounting foam - green: 0.49 g/kg * Facade materials (2D): - Water-resisting paint: 0.306 g/kg - Glass fibre net: 0.191 g/kg - Plaster: 0.452 g/kg * Car interior materials (3): - Hyundai - textile material: 0.02 g/kg - Hyundai - textile material: 0.01 g/kg - Skoda - textile material: 0.021 g/kg * WEEE (4): - Spaghetti insulation II: 0.019 g/kg [114] - Mixed waste - small electronical devices: 0.019 g/kg - Mixed waste: 0.013 g/kg - Impactor undersized particles: 0.131 g/kg [120] The most common applications include textile protection (ScotchgardTM), surface coating for cooking implements (TeflonTM), food contact paper, and Aqueous Film Forming Foams (AFFFs). [19] Calculated global emissions of PFHpS (based on PFOS emissions): 1958-2002: lower: 10 tonnes, higher: 164 tonnes 2003-2015: lower: 1 tonnes, higher: 32 tonnes 2016-2030: lower: 0 tonnes, higher: 6 tonnes [63] past and ongoing production and use will lead to the accumulation of PFAAs in the global environment, with very slow mixing/sedimentation to the deep oceans and sedimentation/burial in deep sediments as the only known global environmental sinks [23] PFECAs and PFESAs are expected to be highly persistent and have physicochemical properties (high water solubility and low pKa) similar to those of PFCAs and PFSAs => likely to have similarly high long-range transport potential in water [23] Replacements in the future might still use the same concept by slightly modifying the currently used fluorinated alternatives => non-fluorinated parts of these structures may undergo (bio)degradation but transformation products are still likely to be environmentally stable PFECAs or short-chain PFCAs => a non-fluorinated replacement strategy should be pursued [63] due to their high persistence and water solubility, PFAAs will be transported to remote locations from sources through water currents and aerosols [63] the most common industrial practice of phasing out one PFAS is to replace it with another structurally similar PFAS => will not solve issues in relation to PFASs as a whole group => let us start the dialogue in defining "essential" and "nonessential" uses of PFASs, while simultaneously developing safe alternative substances and processes for those essential uses [75] - critical travel distance (CTD): 592.68 km - travel efficiency (TE): 0.0070 % => lower CTD and TE than PFOA [27] - 1986: 4.54 - 227 t/yr - 1990: 4.54 - 227 t/yr - 1994: 4.54 - 227 t/yr - 1998: 4.54 - 227 t/yr - 2002: not reported - 2006: not reported [19] Calculated global emissions of FHpSA/Es (based on PFOS emissions): 1958-2002: lower: 9 tonnes, higher: 472 tonnes 2003-2015: lower: 4 tonnes, higher: 58 tonnes 2016-2030: lower: 0 tonnes, higher: 42 tonnes [19] Calculated global emissions of PHpSF (based on PFOS emissions): 1958-2002: lower: 7 tonnes, higher: 32 tonnes 2003-2015: lower: 0 tonnes, higher: 1 tonnes 2016-2030: lower: 0 tonnes, higher: 0 tonnes [23] PFECAs and PFESAs are expected to be highly persistent and have physicochemical properties (high water solubility and low pKa) similar to those of PFCAs and PFSAs => likely to have similarly high long-range transport potential in water [23] Replacements in the future might still use the same concept by slightly modifying the currently used fluorinated alternatives => non-fluorinated parts of these structures may undergo (bio)degradation but transformation products are still likely to be environmentally stable PFECAs or short-chain PFCAs => a non-fluorinated replacement strategy should be pursued [75] - critical travel distance (CTD): 1745.19 km - travel efficiency (TE): 0.8797 % => same CTD as PFOA => higher TE than PFOA [70] Occurrence of PFASs in landfill ambient air: - Ambient landfill air predominantly contained FTOHs, with concentrations being >90% of total precursor compounds measured - 8:2 FTOH was found to be the highest sole contributor (50-65% of the FTOHs, FOSAs, FOSEs), followed by 6:2 FTOH (15-40%) FTOHs - Higher abundance of 8:2 FTOH compared with 6:2 FTOH has been reported to be typical of urban air => Landfills likely act as emission sources of atmospheric PFASs [18] 31-57 g/kg DM 6:2 FTOH in two biosolids-amended soils [18] 6:2 FTS was a major compound in landfill leachate with a level of ~220 ng/L (n=18) [99] 6:2 FTS: leachate from 6 landfills and a laboratory bioreactor (US) (ng/L): - Site A: 280 11 - Site B: 370 20 - Site C: 280 6.8 - Site D2: 29 2.6 - Site D3: 56 13 - Site D6: 270 67 - Laboratory bioreactor: 260 21 [99] Fluorotelomer sulfonates were the fourth most abundant class of fluorochemicals (2.4 1.3%) in landfill leachates (after perfluoroalkyl sulfonamides) [18] - in soil from fire exercise sites the concentration of 6:2 FTS was 26 g/kg DM - average level of 6:2 FTS was 68-85 g/kg for soil and aquifer solids [56] Oncorhynchus mykiss (acute aquatic toxicity): 96h LC50 >108 mg/L [56] FRESHWATER Daphnia magna (acute aquatic toxicity): 48h EC50 >112 mg/L [56] FRESHWATER - Pseudokirchneriella subcapitata (acute aquatic toxicity): 72h EC50 >125 mg/L - Pseudokirchneriella subcapitata (chronic aquatic toxicity): 72h NOEC >125 mg/L [97] 6:2 FTS was also found in fish liver (from a lake receiving water from a fire-fighting training area) suggesting a bioaccumulation, however the concentration could not be calculated due to signal enhancement and lack of labelled internal standard [64] * PFAS concentrations in chronomids at the end of E1 (started with L2 larvae until the end of the L4 stage and lasted for 9 days) and E2 (started directly with L4 larvae) exposures: - E1: 0.07 +/- 0.03 ng/g ww - E2: 0.08 +/- 0.005 ng/g ww * Biota-to-sediment accumulation factor (BSAFww) (= Corg/Csed,oc = PFAS concentration in the organism at steady state/PFAS concentration in sediment normalized by the organic carbon content) = 0.018 => Contamination pathways appeared to be trophic (sediment particles) and tegumentary (pore water) [71] Shorter chains have been reported to have a higher uptake in lettuce leaves, while longer chain PFASs are found primarily in the roots [71] - point sources: wastewater treatment plants (WWTPs), industrial emissions from PFAS production sites, commercial and military airfields (use of AFFF-containing PFASs and their precursors), landfills that contain PFASs-contaminated waste [71] - diffuse sources: contamination caused from a range of dispersed urban and rural land use activities, such as atmospheric deposition and upstream water input [97] 6:2 FTS or its precursor(s) are a likely component in AFFFs used at the site (a fire-fighting training site in NO), given the high levels found in soil and seepage water [100] - exposure of PFASs to humans occurs mainly through food, water, and the indoor environment - in addition to direct exposure, humans are also indirectly exposed to PFAS through precursor compounds that degrade to PFCA and PFSA [100] - Both 6:2 FTSA and 8:2 FTSA have been identified as ingredients in AFFF formulations, with 6:2 FTSA as the predominating FTSA - Direct exposure to 6:2 FTSA from AFFF is not an expected pathway for the general population - However, indirect exposure is possible through contaminated drinking water, at local point sources from fire fighting practicing areas - Uses of 6:2 FTSA other than in AFFF and chrome plating are as an ingredient in consumer products such as cleaning products and color ink - 6:2 FTSA has been observed in indoor dust and in carpets [34] Surface water data SE: slightly higher PFAS concentrations in southern Sweden (higher population density) compared to northern Sweden [56] - Oncorhynchus mykiss (acute aquatic toxicity): 96h LC50 >107 mg/L - Oncorhynchus mykiss (chronic aquatic toxicity): 90d NOEC = 2.62 mg/L based on mean, measured K-6:2 FTS concentrations and first day of hatching [56] FRESHWATER Daphnia magna (acute aquatic toxicity): 48h EC50 >109 mg/L [56] FRESHWATER - Pseudokirchneriella subcapitata (acute aquatic toxicity): 72h EC50 > 96 mg/L - Pseudokirchneriella subcapitata (chronic aquatic toxicity): 72h NOEC = 47.6 mg/L [56] - Eisenia fetida (acute aquatic toxicity): 14d LC50 = 500 mg/kg - Eisenia fetida (EC10, repro.): 56d EC10 repro. = 247 mg/kg [69] PFPiAs: - high acidity - adsorption on organic matter increases with increasing perfluoroalkyl chain length - certain homologs (C6/C6) adsorb more strongly on organic matter than PFOS => LRTP undetermined, but can contribute to the LRTP of PFPAs [75] - critical travel distance (CTD): 1745.33 km - travel efficiency (TE): 0.000385 % => same CTD as PFOA/PFOS => lower TE than PFOA [100] - exposure of PFASs to humans occurs mainly through food, water, and the indoor environment - in addition to direct exposure, humans are also indirectly exposed to PFAS through precursor compounds that degrade to PFCA and PFSA ,,,.....,,,,,.. Number [1] [2] [3] [4] [5] [6] [7] [8] [9] [10] [11] [12] Main author(s) + year Brendel et al. (2018) Eschauzier et al. (2013) Gomis et al. (2018) Kabadi et al. (2018) (accepted manuscript) Title Short-chain perfluoroalkyl acids: environmental concerns and a regulatory strategy under REACH Perfluorinated alkylated acids in groundwater and drinking water: Identification, origin and mobility Comparing the toxic potency in vivo of long-chain perfluoroalkyl acids and fluorinated alternatives Internal exposure-based pharmacokinetic evaluation of potential for biopersistence of 6:2 fluorotelomer alcohol (FTOH) and its metabolites Loos et al. (2010) Pan-European survey on the occurrence of selected polar organic persistent pollutants in ground water Ullah et al. (2011) Colles et al. (2018) (manuscript) Simultaneous determination of perfluoroalkyl phosphonates, carboxylates, and sulfonates in drinking water Perfluorinated substances in the Flemish population: levels and determinants of exposure Shan et al. (2014) Rahman et al. (2014) Ahrens et al. (2014) Gebbink et al. (2015) Ahrens et al. (2016) Concentration profiles and spatial distribution of perfluoroalkyl substances in an industrial center with condensed fluorochemical facilities Behaviour and fate of perfluoroalkyl and polyfluoroalkyl substances (PFASs) in drinking water treatment: A review Fate and effects of poly- and perfluoroalkyl substances in the aquatic environment: a review Estimating human exposure to PFOS isomers and PFCA homologues: The relative importance of direct and indirect (precursor) exposure Screening of PFASs in groundwater and surface water [13] Gebbink et al. (2017) [14] Wang et al. (2013) [15] Olsen et al. (2009) [16] Llorca et al. (2014) [17] Eriksson et al. (2017) [18] van Hees [19] Wang et al. (2017) [20] Loos et al. (2013) Presence of emerging per- and polyfluoroalkyl substances (PFASs) in river and drinking water near a fluorochemical production plant in the Netherlands Fluorinated alternatives to long-chain perfluoroalkyl carboxylic acids (PFCAs), perfluoroalkane sulfonic acids (PFSAs) and their potential precursors A comparison of the pharmacokinetics of perfluorobutanesulfonate (PFBS) in rats, monkeys, and humans Levels and fate of perfluoroalkyl substances in beached plastic pellets and sediments collected from Greece Contribution of precursor compounds to the release of per- and polyfluoroalkyl substances (PFASs) from waste water treatment plants (WWTPs) Analysis of the unknown pool of PFAS: Total Oxidizable Precursors (TOP), PFOS Precursor (PreFOS) and Telomer Degradation Toward a comprehensive global emission inventory of C4-C10 perfluoroalkanesulfonic acids (PFSAs) and related precursors: focus on the life cycle of C8-based products and ongoing industrial transition EU-wide monitoring survey on emerging polar organic contaminants in wastewater treatment plant effluents [21] Cousins et al. (2011) [22] Gallen et al. (2017) [23] Wang et al. (2015) [24] Wang et al. (2016) [25] Cousins et al. (2016) Reconciling measurement and modelling studies of the sources and fate of perfluorinated carboxylates Australia-wide assessment of perfluoroalkyl substances (PFASs) in landfill leachates Hazard assessment of fluorinated alternatives to long-chain perfluoroalkyl acids (PFAAs) and their precursors: Status quo, ongoing challenges and possible solutions Coupled production and emission of short chain perfluoroalkyl acids from a fast developing fluorochemical industry: Evidence from yearly and seasonal monitoring in Daling River Basin, China The precautionary principle and chemicals management: The example of perfluoroalkyl acids in groundwater [26] Wang et al. (2014) [27] Wang et al. (2014) [28] Nguyen et al. (2017) [29] Valsecchi et al. (2017) Global emission inventories for C4-C14 perfluoroalkyl carboxylic acid (PFCA) homologues from 1951 to 2030, Part I: production and emissions from quantifiable sources Global emission inventories for C4-C14 perfluoroalkyl carboxylic acid (PFCA) homologues from 1951 to 2030, part II: The remaining pieces of the puzzle Spatial distribution and source tracing of per- and polyfluoroalkyl substances (PFASs) in surface water in Northern Europe Deriving environmental quality standards for perfluorooctanoic acid (PFOA) and related short chain perfluorinated alkyl acids [30] Gyllenhammar et al. (2015) Influence of contaminated drinking water on perfluoroalkyl acid levels in human serum - A case study from Uppsala, Sweden [31] Ingelido et al. (2018) Biomonitoring of perfluorinated compounds in adults exposed to contaminated drinking water in the Veneto Region, Italy [32] van Leeuwen et al. (2011) Third interlaboratory study on perfluorinated compounds in environmental and human matrices [33] Qin et al. (2017) Association of perfluoroalkyl substances exposure with impaired lung function in children [34] Gobelius et al. (2018) Per- and polyfluoroalkyl substances (PFASs) in Swedish ground- and surface water - Implications for environmental quality standards and drinking water guidelines [35] Zhao et al. (2018) Uptake, translocation and biotransformation of N-ethyl perfluorooctanesulfonamide (N-EtFOSA) by hydroponically grown plants [36] Moe et al. (2012) The structure of the fire fighting foam surfactant Forafac 1157 and its biological and photolytic transformation products [37] Kim et al. (2015) Perfluoroheptanoic acid affects amphibian embryogenesis by inducing the phosphorylation of ERK and JNK [38] Salihovic et al. (2018) Changes in markers of liver function in relation to changes in perfluoroalkyl substances - A longitudinal study [39] Wolf et al. (2012) [40] Prez et al. (2013) Activation of mouse and human peroxisome proliferator-activated receptor-alpha (PPAR) by perfluoroalkyl acids (PFAAs): Further investigation of C4-C12 compounds Accumulation of perfluoroalkyl substances in human tissues [41] Zhao et al. (2017) Inhibition effects of perfluoroalkyl acids on progesterone production in mLTC-1 [42] Wang et al. (2014) [43] Blanc et al. (2017) [44] Zhang et al. (2017) Toxicity assessment of perfluorinated carboxylic acids (PFCAs) towards the rotifer Brachionus calyciflorus Mixture-specific gene expression in zebrafish (Danio rerio ) embryos exposed to perfluorooctane sulfonic acid (PFOS), perfluorohexanoic acid (PFHxA) and 3,3',4,4',5-pentachlorobiphenyl (PCB126) Poly- and perfluorinated compounds activate human pregnane X receptor [45] Jiang et al. (2014) Serum levels of perfluoroalkyl acids (PFAAs) with isomer analysis and their associations with medical parameters in Chinese pregnant women [46] Kar et al. (2017) Endocrine-disrupting activity of per- and polyfluoroalkyl substances: Exploring combined approaches of ligand and structure based modeling [47] Song et al. (2018) Biomonitoring PFAAs in blood and semen samples: Investigation of a potential link between PFAAs exposure and semen mobility in China [48] Vongphachan et al. (2011) Effects of perfluoroalkyl compounds on mRNA expression levels of thyroid hormone-responsive genes in primary cultures of avian neuronal cells [49] Boiteux et al. (2017) Concentrations and patterns of perfluoroalkyl and polyfluoroalkyl substances in a river and three drinking water treatment plants near and far from a major production source [50] Rosenmai et al. (2016) Fluorinated alkyl substances and technical mixtures used in food paper-packaging exhibit endocrine-related activity in vitro [51] Zhang et al. (2014) Structure-dependent binding and activation of perfluorinated compounds on human peroxisome proliferator-activated receptor [52] Buhrke et al. (2013) [53] Zhang et al. (2016) [54] Wang et al. (2011) In vitro toxicological characterization of perfluorinated carboxylic acids with different carbon chain lengths Biotransformation potential of 6:2 fluorotelomer sulfonate (6:2 FTSA) in aerobic and anaerobic sediment 6:2 Fluorotelomer sulfonate aerobic biotransformation in activated sludge of waste water treatment plants [55] Shi et al. (2018) 6:2 fluorotelomer sulfonamide alkylbetaine (6:2 FTAB), a novel perfluorooctane sulfonate alternative, induced developmental toxicity in zebrafish embryos [56] Hoke et al. (2015) Aquatic hazard, bioaccumulation and screening risk assessment for 6:2 fluorotelomer sulfonate [57] Gomez-Ruiz et al. (2017) Efficient electrochemical degradation of poly- and perfluoroalkyl substances (PFASs) from the effluents of an industrial wastewater treatment plant [58] Gomez-Ruiz et al. (2017) Boron doped diamond electrooxidation of 6:2 fluorotelomers and perfluorocarboxylic acids. Application to industrial wastewaters treatment [59] Zhuo et al. (2014) Electrochemical oxidation of 1H,1H,2H,2H-perfluorooctane sulfonic acid (6:2 FTS) on DSA electrode: Operating parameters and mechanism [60] Park et al. (2016) Heat-activated persulfate oxidation of PFOA, 6:2 fluorotelomer sulfonate, and PFOS under conditions suitable for in-situ groundwater remediation [61] Yang et al. (2014) Stability of 6:2 fluorotelomer sulfonate in advanced oxidation processes: degradation kinetics and pathway [62] Betts (2007) Perfluoroalkyl acids: What is the evidence telling us? [63] Wang et al. (2017) A never-ending story of per- and polyfluoroalkyl substances (PFASs)? [64] Bertin et al. (2014) Bioaccumulation of perfluoroalkyl compounds in midge (Chironomus riparius ) larvae exposed to sediment [65] Bhhatarai & Gramatica (2011) Prediction of aqueous solubility, vapor pressure and critical micelle concentration for aquatic partitioning of perfluorinated chemicals [66] Li et al. (2017) Half-lives of PFOS, PFHxS and PFOA after end of exposure to contaminated drinking water [67] Shaw et al. (2019) Degradation and defluorination of 6:2 fluorotelomer sulfonamidoalkyl betaine and 6:2 fluorotelomer sulfonate by Gordonia sp. strain NB4-1Y under sulfur- limiting conditions [68] Ng & Hungerbhler (2013) Bioconcentration of perfluorinated alkyl acids: how important is specific binding? [69] Wang et al. (2016) [70] Hamid et al. (2018) [71] Banzhaf et al. (2017) Comparative assessment of the environmental hazards of and exposure to perfluoroalkyl phosphonic and phosphinic acids (PFPAs and PFPiAs): Current knowledge, gaps, challenges and research needs Review of the fate and transformation of per- and polyfluoroalkyl substances (PFASs) in landfills A review of contamination of surface-, ground-, and drinking water in Sweden by perfluoroalkyl and polyfluoroalkyl substances (PFASs) [72] Albergamo et al. (2018) Direct injection analysis of polar micropollutants in natural drinking water sources with biphenyl liquid chromatography coupled to high-resolution time-of-flight [73] Zhao et al. (2012) mass spectrometry Distribution and long-range transport of polyfluoroalkyl substances in the Arctic, Atlantic Ocean and Antarctic coast [74] Vierke et al. (2014) [75] Gomis et al. (2015) Transport of perfluoroalkyl acids in a water-saturated sediment column investigated under near-natural conditions A modeling assessment of the physicochemical properties and environmental fate of emerging and novel per- and polyfluoroalkyl substances [76] [77] [78] [79] [80] [81] [82] [83] [84] [85] [86] [87] [88] [89] [90] [91] [92] [93] [94] [95] [96] [97] [98] [99] [100] [101] [102] [103] Sheng et al. (2018) Ross et al. (2018) Cytotoxicity of novel fluorinated alternatives to long-chain perfluoroalkyl substances to human liver cell line and their binding capacity to human liver fatty acid binding protein A review of emerging technologies for remediation of PFASs Zhang et al. (2013) Kim et al. (2015) Biomonitoring of perfluoroalkyl acids in human urine and estimates of biological half-life Selecting reliable physicochemical properties of perfluoroalkyl and polyfluoroalkyl substances (PFASs) based on molecular descriptors Wolf et al. (2014) Gorrochategui et al. (2014) Evaluating the additivity of perfluoroalkyl acids in binary combinations on peroxisome proliferator-activated receptor- activation Perfluorinated chemicals: Differential toxicity, inhibition of aromatase activity and alteration of cellular lipids in human placental cells Hu et al. (2014) Kjeldsen & Bonefeld-Jorgensen et al. (2013) Weiss et al. (2009) Synergistic effects of perfluoroalkyl acids mixtures with J-shaped concentrationresponses on viability of a human liver cell line Perfluorinated compounds affect the function of sex hormone receptors Competitive binding of poly- and perfluorinated compounds to the thyroid hormone transport protein transthyretin Naile et al. (2012) Transcriptional effects of perfluorinated compounds in rat hepatoma cells Ren et al. (2015) Structure-activity relations in binding of perfluoroalkyl compounds to human thyroid hormone T3 receptor Ren et al. (2016) Binding interactions of perfluoroalkyl substances with thyroid hormone transport proteins and potential toxicological implications Yang et al. (2016) Human placental transfer of perfluoroalkyl acid precursors: Levels and profiles in paired maternal and cord serum Stahl et al. (2014) Analysis of selected perfluoroalkyl substances (PFASs) in beer to evaluate the effect of beer consumption on human PFAS exposure: a pilot study Serex et al. (2014) Toxicological evaluation of 6:2 fluorotelomer alcohol Shi et al. (2017) Sez et al. (2008) Sheng et al. (2017) Latala et al. (2009) Ishibashi et al. (2007) 6:2 fluorotelomer carboxylic acid (6:2 FTCA) exposure induces developmental toxicity and inhibits the formation of erythrocytes during zebrafish embryogenesis Persistence of perfluoroalkylated substances in closed bottle tests with municipal sewage sludge Comparative hepatotoxicity of 6:2 fluorotelomer carboxylic acid and 6:2 fluorotelomer sulfonic acid, two fluorinated alternatives to long-chain perfluoroalkyl acids, on adult male mice Acute toxicity assessment of perfluorinated carboxylic acids towards the Baltic microalgae Estrogenic effects of fluorotelomer alcohols for human estrogen receptor isoforms and in vitro Maras et al. (2006) Estrogen-like properties of fluorotelomer alcohols as revealed by MCF-7 breast cancer cell proliferation Krrman et al. (2011) Environmental levels and distribution of structural isomers of perfluoroalkyl acids after aqueous fire-fighting foam (AFFF) contamination Yang et al. (2016) Placental transfer of perfluoroalkyl substances and associations with thyroid hormones: Beijing prenatal exposure study Huset et al. (2011) Quantitative determination of fluorochemicals in municipal landfill leachates Eriksson et al. (2017) Vestergren et al. (2012) Eschauzier et al. (2013) Wang et al. (2019) Temporal trends of PFSAs, PFCAs and selected precursors in Australian serum from 2002 to 2013 Dietary exposure to perfluoroalkyl acids for the Swedish population in 1999, 2005 and 2010 Presence and sources of anthropogenic perfluoroalkyl acids in high-consumption tap-water based beverages Efficiency of maternal-fetal transfer of perfluoroalkyl and polyfluoroalkyl substances [104] [105] [106] [107] [108] Kang et al. (2016) Lee et al. (2018) Lee et al. (2018) Han et al. (2018) Cai et al. (2020) [109] [110] [111] [112] [113] [114] [115] [116] [117] Knobeloch et al. (2012) Qi et al. (2019) Su et al. (2017) Zhang et al. (2019) Heo et al. (2014) Becanov et al. (2016) Domingo et al. (2012) Tian et al. (2016) Pitter et al. (2020) [118] [119] [120] Lee et al. (2020) Coggan et al. (2019) Hepburn et al. (2019) Elevated levels of short-carbon chain PFCAs in breast milk among Korean women: Current status and potential challenges Perfluoroalkyl substances (PFASs) in breast milk from Korea: Time-course trends, influencing factors, and infant exposure The serum concentrations of perfluoroalkyl compounds were inversely associated with growth parameters in 2-year old children Perfluoroalkyl and polyfluoroalkyl substances in matched parental and cord serum in Shandong, China High trans-placental transfer of perfluoroalkyl substances alternatives in the matched maternal-cord blood serum: Evidence from a birth cohort study Perfluoroalkyl chemicals in vacuum cleaner dust from 39 Wisconsin homes Perfluorinated compounds in poultry products from the Yangtze River Delta and Pearl River Delta regions in China Home produced eggs: An important pathway of human exposure to perfluorobutanoic acid (PFBA) and perfluorooctanoic acid (PFOA) around a fluorochemical industrial park in China Exposure of Juncus effusus to seven perfluoroalkyl acids: Uptake, accumulation and phytotoxicity Foodstuff analyses show that seafood and water are major perfluoroalkyl acids (PFAAs) sources to humans in Korea Screening for perfluoroalkyl acids in consumer products, building materials and wastes Human dietary exposure to perfluoroalkyl substances in Catalonia, Spain. Temporal trend. Human exposure to per- and polyfluoroalkyl substances (PFASs) via house dust in Korea: Implication to exposure pathway. Serum levels of perfluoroalkyl substances (PFAS) in adolescents and young adults exposed to contaminated drinking water in the Veneto Region, Italy: A crosssectional study based on a health surveillance program. Concentration and distribution of per- and polyfluoroalkyl substances (PFAS) in the Asan Lake area of South Korea. An investigation into per- and polyfluoroalkyl substances (PFAS) in nineteen Australian wastewater treatment plants (WWTPs). Contamination of groundwater with per- and polyfluoroalkyl substances (PFAS) from legacy landfills in an urban re-development precinct. Disclaimer: A number [ ] mentioned above refers to a specific paper and the references therein. Especially information out of literature reviews can be retrieved from other studies. Type Literature review Own research Modelling of pre-existing data Modelling of pre-existing data Own research Own research Own research Own research Literature review Literature review Modelling of pre-existing data Own research Reference Brendel, S., Fetter, E., Staude, C., Vierke, L. & Biegel-Engler, A. (2018). Short-chain perfluoralkyl acids: environmental concerns and a regulatory strategy under REACH. Environmental Sciences Europe, 30: 1-11. Eschauzier, C., Raat, K.J., Stuyfzand, P.J. & De Voogt, P. (2013). Perfluorinated alkylated acids in groundwater and drinking water: Identification, origin and mobility. Science of the Total Environment, 458-460: 477-485. Gomis, M.I., Vestergren, R., Borg, D. & Cousins, I.T. (2018). Comparing the toxic potency in vivo of long-chain perfluoroalkyl acids and fluorinated alternatives. Environment International, 113: 1-9. Kabadi, S.V., Fisher, J., Aungst, J. & Rice, P. (2018). Internal exposure-based pharmacokinetic evaluation of potential for biopersistence of 6:2 fluorotelomer alcohol (FTOH) and its metabolites. Food and Chemical Toxicology, doi: 10.1016/j.fct.2018.01.012. Loos, R., Locoro, G., Comero, S., Contini, S., Schwesig, D., Werres, F., Balsaa, P., Gans, O., Weiss, S., Blaha, L., Bolchi, M. & Gawlik, B.M. (2010). Pan-European survey on the occurrence of selected polar organic persistent pollutants in ground water. Water Research, 44: 4115-4126. Ullah, S., Alsberg, T. & Berger, U. (2011). Simultaneous determination of perfluoroalkyl phosphonates, carboxylates, and sulfonates in drinking water. Journal of Chromatography A, 1218: 6388-6395. Colles, A., Den Hond, E., Bruckers, L., Govarts, E., Morrens, B., Sioen, I., Schettgen, T., Buekers, J., Coertjens, D., Nawrot, T., Loots, I., Nelen, V., De Henauw, S., Schoeters, G., Baeyens, W. & van Larebeke, N. (2018). Perfluorinated substances in the Flemish population: levels and determinants of exposure. Shan, G., Wei, M., Zhu, L., Liu, Z. & Zhang, Y. (2014). Concentration profiles and spatial distribution of perfluoroalkyl substances in an industrial center with condensed fluorochemical facilities. Science of the Total Environment, 490: 351-359. Rahman, M.F., Peldszus, S. & Anderson, W.B. (2014). Behaviour and fate of perfluoroalkyl and polyfluoroalkyl substances (PFASs) in drinking water treatment: A review. Water Research, 50: 318-340. Ahrens, L. & Bundschuh, M. (2014). Fate and effects of poly- and perfluoroalkyl substances in the aquatic environment: a review. Environmental Toxicology and Chemistry, 33: 1921-1929. Gebbink, W.A., Berger, U. & Cousins, I.T. (2015). Estimating human exposure to PFOS isomers and PFCA homologues: The relative importance of direct and indirect (precursor) exposure. Environment International, 74: 160-169. Ahrens, L., Hedlund, J., Drig, W., Trger, R. & Wiberg, K. (2016). Screening of PFASs in groundwater and surface water. Rapport 2016:2, Sveriges lantbruksuniversitet, Institutionen fr vatten och milj, ISBN 978-91-576-9386-0. Own research Literature review Own research Own research Own research Literature review Modelling of pre-existing data Gebbink, W.A., van Asseldonk, L. & van Leeuwen, S.P.J. (2017). Presence of emerging per- and polyfluoroalkyl substances (PFASs) in river and drinking water near a fluorochemical production plant in the Netherlands. Environmental Science & Technology, 51: 11057-11065. Wang, Z., Cousins, I.T., Scheringer, M. & Hungerbhler, K. (2013). Fluorinated alternatives to long-chain perfluoroalkyl carboxylic acids (PFCAs), perfluoroalkane sulfonic acids (PFSAs) and their potential precursors. Environment International, 60: 242-248. Olsen, G.W., Chang, S.-C., Noker, P.E., Gorman, G.S., Ehresman, D.J., Lieder, P.H. & Butenhoff, J.L. (2009). A comparison of the pharmacokinetics of perfluorobutanesulfonate (PFBS) in rats, monkeys, and humans. Toxicology, 256: 65-74. Llorca, M., Farr, M., Karapanagioti, H.K. & Barcel, D. (2014). Levels and fate of perfluoroalkyl substances in beached plastic pellets and sediments collected from Greece. Marine Pollution Bulletin, 87: 286-291. Eriksson, U., Haglund, P. & Krrman, A. (2017). Contribution of precursor compounds to the release of per- and polyfluoroalkyl substances (PFASs) from waste water treatment plants (WWTPs). Journal of Environmental Sciences, 61: 80-90. van Hees. Analysis of the unknown pool of PFAS: Total Oxidizable Precursors (TOP), PFOS Precursor (PreFOS) and Telomer Degradation. Eurofins Environment Testing Sweden AB and Man-Technology-Environment Research Centre, rebro University. Wang, Z., Boucher, J.M., Scheringer, M., Cousins, I.T. & Hungerbhler, K. (2017). Toward a comprehensive global emission inventory of C4-C10 perfluoroalkanesulfonic acids (PFSAs) and related precursors: focus on the life cycle of C8-based products and ongoing industrial transition. Environmental Science & Technology, 51: 4482-4493. Own research Literature review Own research Literature review Own research Loos, R., Carvalho, R., Antnio, D.C., Comero, S., Locoro, G., Tavazzi, S., Paracchini, B., Ghiani, M., Lettieri, T., Blaha, L., Jarosova, B., Voorspoels, S., Servaes, K., Haglund, P., Fick, J., Lindberg, R.H., Schwesig, D. & Gawlik, B.M. (2013). EU-wide monitoring survey on emerging polar organic contaminants in wastewater treatment plant effluents. Water Research, 47: 6475-6487. Cousins, I.T., Kong, D. & Vestergren, R. (2011). Reconciling measurement and modelling studies of the sources and fate of perfluorinated carboxylates. Environmental Chemistry, 8: 339-354. Gallen, C., Drage, D., Eaglesham, G., Grant, S., Bowman, M. & Mueller, J.F. (2017). Australia-wide assessment of perfluoroalkyl substances (PFASs) in landfill leachates. Journal of Hazardous Materials, 331: 132-141. Wang, Z., Cousins, I.T., Scheringer, M. & Hungerbhler, K. (2015). Hazard assessment of fluorinated alternatives to long-chain perfluoroalkyl acids (PFAAs) and their precursors: Status quo, ongoing challenges and possible solutions. Environment International, 75: 172-179. Wang, P., Lu, Y., Wang, T., Zhu, Z., Li, Q., Meng, J., Su, H., Johnson, A.C. & Sweetman, A.J. (2016). Coupled production and emission of short chain perfluoroalkyl acids from a fast developing fluorochemical industry: Evidence from yearly and seasonal monitoring in Daling River Basin, China. Environmental Pollution, 218: 1234-1244. Modelling of pre-existing Cousins, I.T., Vestergren, R., Wang, Z., Scheringer, M. & McLachlan, M.S. (2016). The precautionary principle and data chemicals management: The example of perfluoroalkyl acids in groundwater. Environment International, 94: 331- 340. Modelling of pre-existing data Modelling of pre-existing data Own research Literature review Own research Own research Wang, Z., Cousins, I.T., Scheringer, M., Buck, R.C. & Hungerbhler, K. (2014). Global emission inventories for C4-C14 perfluoroalkyl carboxylic acid (PFCA) homologues from 1951 to 2030, Part I: production and emissions from quantifiable sources. Environment International, 70: 62-75. Wang, Z., Cousins, I.T., Scheringer, M., Buck, R.C. & Hungerbhler, K. (2014). Global emission inventories for C4-C14 perfluoroalkyl carboxylic acid (PFCA) homologues from 1951 to 2030, part II: The remaining pieces of the puzzle. Environment International, 69: 166-176. Nguyen, M.A., Wiberg, K., Ribeli, E., Josefsson, S., Futter, M., Gustavsson, J. & Ahrens, L. (2017). Spatial distribution and source tracing of per- and polyfluoroalkyl substances (PFASs) in surface water in Northern Europe. Environmental Pollution, 220: 1438-1446. Valsecchi, S., Conti, D., Crebelli, R., Polesello, S., Rusconi, M., Mazzoni, M., Preziosi, E., Carere, M., Lucentini, L., Ferretti, E., Balzamo, S., Simeone, M.G. & Aste, F. (2017). Deriving environmental quality standards for perfluorooctanoic acid (PFOA) and related short chain perfluorinated alkyl acids. Journal of Hazardous Materials, 323: 84-98. Gyllenhammar, I., Berger, U., Sundstrm, M., McCleaf, P., Eurn, K., Eriksson, S., Ahlgren, S., Lignell, S., Aune, M., Kotova, N. & Glynn, A. (2015). Influence of contaminated drinking water on perfluoroalkyl acid levels in human serum - A case study from Uppsala, Sweden. Environmental Research, 140: 673-683. Ingelido, A.M., Abballe, A., Gemma, S., Dellatte, E., Iacovella, N., De Angelis, G., Zampaglioni, F., Marra, V., Miniero, R., Valentini, S., Russo, F., Vazzoler, M., Testai, E. & De Felip, E. (2018). Environment International, 110: 149-159. Own research Own research Own research van Leeuwen, S., Strub, M.-P., Cofino, W., Lindstrm, G. & van Bavel, B. (2011). Third interlaboratory study on perfluorinated compounds in environmental and human matrices. Report R-11/04, 27 May 2011, IVM Institute for Environmental Studies. Qin, X.-D., Qian, Z.(M.), Dharmage, S.C., Perret, J., Geiger, S.D., Rigdon, S.E., Howard, S., Zeng, X.-W., Hu, L.-W., Yang, B.-Y., Zhou, Y., Li, M., Xu, S.-L., Bao, W.-W., Zhang, Y.-Z., Yuan, P., Wang, J., Zhang, C., Tian, Y.-P., Nian, M., Xiao, X., Chen, W., Lee, Y.L., Dong, G.-H. (2017). Association of perfluoroalkyl substances exposure with impaired lung function in children. Environmental Research, 155: 15-21. Gobelius, L., Hedlund, J.K., Drig, W., Trger, R., Lilja, K., Wiberg, K. & Ahrens, L. (2018). Per- and polyfluoroalkyl substances (PFASs) in Swedish ground- and surface water - Implications for environmental quality standards and drinking water guidelines. Ecotoxicology and Human Environmental Health, doi: 10.1021/acs.est.7b05718. Own research Own research Own research Own research Own research Own research Own research Own research Own research Own research Own research Zhao, S., Zhou, T., Zhu, L., Wang, B., Li, Z., Yang, L. & Liu, L. (2018). Uptake, translocation and biotransformation of Nethyl perfluorooctanesulfonamide (N-EtFOSA) by hydroponically grown plants. Environmental Pollution, doi: 10.1016/j.envpol.2017.12.053. Moe, M.K., Huber, S., Svenson, J., Hagenaars, A., Pabon, M., Trmper, M., Berger, U., Knapen, D. & Herzke, D. (2012). The structure of the fire fighting foam surfactant Forafac 1157 and its biological and photolytic transformation products. Chemosphere, 89: 869-875. Kim, M., Park, M.S., Son, J., Park, I., Lee, H.-K., Kim, C., Min, B.-H., Ryoo, J., Choi, K.S., Lee, D.-S. & Lee, H.-S. (2015). Perfluoroheptanoic acid affects amphibian embryogenesis by inducing the phosphorylation of ERK and JNK. International Journal of Molecular Medicine, 36: 1693-1700. Salihovic, S., Stubleski, J., Krrman, A., Larsson, A., Fall, T., Lind, L. & Lind, P.M. (2018). Changes in markers of liver function in relation to changes in perfluoroalkyl substances - A longitudinal study. Environment International, 117: 196-203. Wolf, C.J., Schmid, J.E., Lau, C. & Abbott, B.D. (2012). Activation of mouse and human peroxisome proliferatoractivated receptor-alpha (PPAR) by perfluoroalkyl acids (PFAAs): Further investigation of C4-C12 compounds. Reproductive Toxicology, 33: 546-551. Prez, F., Nadal, M., Navarro-Ortega, A., Fbrega, F., Domingo, J.L., Barcel, D. & Farr, M. (2013). Accumulation of perfluoroalkyl substances in human tissues. Environment International, 59: 354-362. Zhao, W., Cui, R., Wang, J. & Dai, J. (2017). Inhibition effects of perfluoroalkyl acids on progesterone production in mLTC-1. Journal of Environmental Sciences, 56: 272-280. Wang, Y., Niu, J., Zhang, L. & Shi, J. (2014). Toxicity assessment of perfluorinated carboxylic acids (PFCAs) towards the rotifer Brachionus calyciflorus . Science of the Total Environment, 491-492: 266-270. Blanc, M., Krrman, A., Kukucka, P., Scherbak, N. & Keiter, S. (2017). Mixture-specific gene expression in zebrafish (Danio rerio ) embryos exposed to perfluorooctane sulfonic acid (PFOS), perfluorohexanoic acid (PFHxA) and 3,3',4,4',5-pentachlorobiphenyl (PCB126). Science of the Total Environment, 590-591: 249-257. Zhang, Y.-M., Dong, X.-Y., Fan, L.-J., Zhang, Z.-L., Wang, Q., Jiang, N. & Yang, X.-S. (2017). Poly- and perfluorinated compounds activate human pregnane X receptor. Toxicology, 380: 23-29. Jiang, W., Zhang, Y., Zhu, L. & Deng, J. (2014). Serum levels of perfluoroalkyl acids (PFAAs) with isomer analysis and their associations with medical parameters in Chinese pregnant women. Environment International, 64: 40-47. Modelling of pre-existing Kar, S., Seplveda, M.S., Roy, K. & Leszczynski, J. (2017). Endocrine-disrupting activity of per- and polyfluoroalkyl data substances: Exploring combined approaches of ligand and structure based modeling. Chemosphere, 184: 514-523. Own research Own research Own research Song, X., Tang, S., Zhu, H., Chen, Z., Zang, Z., Zhang, Y., Niu, X., Wang, X., Yin, H., Zeng, F. & He, C. (2018). Biomonitoring PFAAs in blood and semen samples: Investigation of a potential link between PFAAs exposure and semen mobility in China. Environment International, 113: 50-54. Vongphachan, V., Cassone, C.G., Wu, D., Chiu, S., Crump, D. & Kennedy, S.W. (2011). Effects of perfluoroalkyl compounds on mRNA expression levels of thyroid hormone-responsive genes in primary cultures of avian neuronal cells. Toxicological Sciences, 120: 392-402. Boiteux, V., Dauchy, X., Bach, C., Colin, A., Hemard, J., Sagres, V., Rosin, C. & Munoz, J.-F. (2017). Concentrations and patterns of perfluoroalkyl and polyfluoroalkyl substances in a river and three drinking water treatment plants near and far from a major production source. Science of the Total Environment, 583: 393-400. Own research Own research Own research Own research + literature review Own research Own research Own research + literature review Own research Own research Own research Own research Own research Literature review Literature review Own research Modelling of pre-existing data Own research Rosenmai, A.K., Taxvig, C., Svingen, T., Trier, X., van Vugt-Lussenburg, B.M.A., Pedersen, M., Lesn, L., Jgou, B. & Vinggaard, A.M. (2016). Fluorinated alkyl substances and technical mixtures used in food paper-packaging exhibit endocrine-related activity in vitro. Andrology, 4: 662-672. Zhang, L., Ren, X.-M., Wan, B. & Guo, L.-H. (2014). Structure-dependent binding and activation of perfluorinated compounds on human peroxisome proliferator-activated receptor . Toxicology and Applied Pharmacology, 279: 275-283. Buhrke, T., Kibellus, A. & Lampen, A. (2013). In vitro toxicological characterization of perfluorinated carboxylic acids with different carbon chain lengths. Toxicology Letters, 218: 97-104. Zhang, S., Lu, X., Wang, N. & Buck, R.C. (2016). Biotransformation potential of 6:2 fluorotelomer sulfonate (6:2 FTSA) in aerobic and anaerobic sediment. Chemosphere, 154: 224-230. Wang, N., Liu, J., Buck, R.C., Korzeniowski, S.H., Wolstenholme, B.W., Folsom, P.W. & Sulecki, L.M. (2011). 6:2 Fluorotelomer sulfonate aerobic biotransformation in activated sludge of waste water treatment plants. Chemosphere, 82: 853-858. Shi, G., Xie, Y., Guo, Y. & Dai, J. (2018). 6:2 fluorotelomer sulfonamide alkylbetaine (6:2 FTAB), a novel perfluorooctane sulfonate alternative, induced developmental toxicity in zebrafish embryos. Aquatic Toxicology, 195: 24-32. Hoke, R.A., Ferrell, B.D., Ryan, T., Sloman, T.L., Green, J.W., Nabb, D.L., Mingoia, R., Buck, R.C. & Korzeniowski, S.H. (2015). Aquatic hazard, bioaccumulation and screening risk assessment for 6:2 fluorotelomer sulfonate. Chemosphere, 128: 258-265. Gomez-Ruiz, B., Gmez-Lavn, S., Diban, N., Boiteux, V., Colin, A., Dauchy, X. & Urtiaga, A. (2017). Efficient electrochemical degradation of poly- and perfluoroalkyl substances (PFASs) from the effluents of an industrial wastewater treatment plant. Chemical Engineering Journal, 322: 196-204. Gomez-Ruiz, B., Gmez-Lavn, S., Diban, N., Boiteux, V., Colin, A., Dauchy, X. & Urtiaga, A. (2017). Boron doped diamond electrooxidation of 6:2 fluorotelomers and perfluorocarboxylic acids. Application to industrial wastewaters treatment. Journal of Electroanalytical Chemistry, 798: 51-57. Zhuo, Q., Li, X., Yan, F., Yang, B., Deng, S., Huang, J. & Yu, G. (2014). Electrochemical oxidation of 1H,1H,2H,2Hperfluorooctane sulfonic acid (6:2 FTS) on DSA electrode: Operating parameters and mechanism. Journal of Environmental Sciences, 26: 1733-1739. Park, S., Lee, L.S., Medina, V.F., Zull, A. & Waisner, S. (2016). Heat-activated persulfate oxidation of PFOA, 6:2 fluorotelomer sulfonate, and PFOS under conditions suitable for in-situ groundwater remediation. Chemosphere, 145: 376-383. Yang, X., Huang, J., Zhang, K., Yu, G., Deng, S. & Wang, B. (2014). Stability of 6:2 fluorotelomer sulfonate in advanced oxidation processes: degradation kinetics and pathway. Environmental Science and Pollution Research, 21: 46344642. Betts, K.S. (2007). Perfluoroalkyl acids: What is the evidence telling us? Environmental Health Perspectives, 115: A250-A256. Wang, Z., DeWitt, J.C., Higgins, C.P. & Cousins, I.T. (2017). A never-ending story of per- and polyfluoroalkyl substances (PFASs)? Environmental Science and Technology, 51: 2508-2518. Bertin, D., Ferrari, B.J.D., Labadie, P., Sapin, A., Garric, J., Budzinski, H., Houde, M. & Babut, M. (2014). Bioaccumulation of perfluoroalkyl compounds in midge (Chironomus riparius ) larvae exposed to sediment. Environmental Pollution, 189: 27-34. Bhhatarai, B. & Gramatica, P. (2011). Prediction of aqueous solubility, vapor pressure and critical micelle concentration for aquatic partitioning of perfluorinated chemicals. Environmental Science and Technology, 45: 81208128. Li, Y., Fletcher, T., Mucs, D., Scott, K., Lindh, C.H., Tallving, P. & Jakobsson, K. (2017). Half-lives of PFOS, PFHxS and PFOA after end of exposure to contaminated drinking water. Occupational and Environmental Medicine, 0: 1-6. Own research Modelling of pre-existing data Literature review Literature review Literature review Own research Own research Own research Own research Shaw, D.M.J., Munoz, G., Bottos, E.M., Vo Duy, S., Sauv, S., Liu, J. & Van Hamme, J.D. (2019). Degradation and defluorination of 6:2 fluorotelomer sulfonamidoalkyl betaine and 6:2 fluorotelomer sulfonate by Gordonia sp. strain NB4-1Y under sulfur-limiting conditions. Science of the Total Environment, 647: 690-698. Ng, C.A. & Hungerbhler, K. (2013). Bioconcentration of perfluorinated alkyl acids: how important is specific binding? Environmental Science and Technology, 47: 7214-7223. Wang, Z., Cousins, I.T., Berger, U., Hungerbhler, K. & Scheringer, M. (2016). Comparative assessment of the environmental hazards of and exposure to perfluoroalkyl phosphonic and phosphinic acids (PFPAs and PFPiAs): Current knowledge, gaps, challenges and research needs. Environment International, 89-90: 235-247. Hamid, H., Li, L.Y. & Grace, J.R. (2018). Review of the fate and transformation of per- and polyfluoroalkyl substances (PFASs) in landfills. Environmental Pollution, 235: 74-84. Banzhaf, S., Filipovic, M., Lewis, J., Sparrenbom, C.J. & Barthel, R. (2017). A review of contamination of surface-, ground-, and drinking water in Sweden by perfluoroalkyl and polyfluoroalkyl substances (PFASs). Ambio, 46: 335346. Albergamo, V., Helmus, R. & de Voogt, P. (2018). Direct injection analysis of polar micropollutants in natural drinking water sources with biphenyl liquid chromatography coupled to high-resolution time-of-flight mass spectrometry. Journal of Chromatography A, 1569: 53-61. Zhao, Z., Xie, Z., Mller, A., Sturm, R., Tang, J., Zhang, G. & Ebinghaus, R. (2012). Distribution and long-range transport of polyfluoroalkyl substances in the Arctic, Atlantic Ocean and Antarctic coast. Environmental Pollution, 170: 71-77. Vierke, L., Mller, A. & Klitzke, S. (2014). Transport of perfluoroalkyl acids in a water-saturated sediment column investigated under near-natural conditions. Environmental Pollution, 186: 7-13. Gomis, M.I., Wang, Z., Scheringer, M. & Cousins, I.T. (2015). A modeling assessment of the physicochemical properties and environmental fate of emerging and novel per- and polyfluoroalkyl substances. Science of the Total Environment, 505: 981-991. Own research Literature review Own research Own research Own research Own research Own research Own research Own research Own research Own research Own research Own research Own research Own research Own research Own research Own research Own research Own research Own research Own research Own research Own research Own research Own research Own research Own research Sheng, N., Cui, R., Wang, J., Guo, Y., Wang, J. & Dai, J. (2018). Cytotoxicity of novel fluorinated alternatives to longchain perfluoroalkyl substances to human liver cell line and their binding capacity to human liver fatty acid binding protein. Archives of Toxicology, 92: 359-369. Ross, I., McDonough, J., Miles, J., Storch, P., Kochunarayanan, P.T., Kalve, E., Hurst, J., Dasgupta, S.S. & Burdick, J. (2018). A review of emerging technologies for remediation of PFASs. Remediation, 28: 101-126. Zhang, Y., Beesoon, S., Zhu, L. & Martin, J.W. (2013). Biomonitoring of perfluoroalkyl acids in human urine and estimates of biological half-life. Environmental Science and Technology, 47: 10619-10627. Kim, M., Li, L.Y., Grace, J.R. & Yue, C. (2015). Selecting reliable physicochemical properties of perfluoroalkyl and polyfluoroalkyl substances (PFASs) based on molecular descriptors. Environmental Pollution, 196: 462-472. Wolf, C.J., Rider, C.V., Lau, C. & Abbott, B.D. (2014). Evaluating the additivity of perfluoroalkyl acids in binary combinations on peroxisome proliferator-activated receptor- activation. Toxicology, 316: 43-54. Gorrochategui, E., Prez-Albaladejo, E., Casas, J., Lacorte, S. & Porte, C. (2014). Perfluorinated chemicals: Differential toxicity, inhibition of aromatase activity and alteration of cellular lipids in human placental cells. Toxicology and Applied Pharmacology, 277: 124-130. Hu, J., Li, J., Wang, J., Zhang, A. & Dai, J. (2014). Synergistic effects of perfluoroalkyl acids mixtures with J-shaped concentration-responses on viability of a human liver cell line. Chemosphere, 96: 81-88. Kjeldsen, L.S. & Bonefeld-Jorgensen, E.C. (2013). Perfluorinated compounds affect the function of sex hormone receptors. Environmental Science and Pollution Research, 20: 8031-8044. Weiss, J.M., Andersson, P.L., Lamoree, M.H., Leonards, P.E.G., van Leeuwen, S.P.J. & Hamers, T. (2009). Competitive binding of poly- and perfluorinated compounds to the thyroid hormone transport protein transthyretin. Toxicological Sciences, 109: 206-216. Naile, J.E., Wiseman, S., Bachtold, K., Jones, P.D. & Giesy, J.P. (2012). Transcriptional effects of perfluorinated compounds in rat hepatoma cells. Chemosphere, 86: 270-277. Ren, X.-M., Zhang, Y.-F., Guo, L.-H., Qin, Z.-F., Lv, Q.-Y. & Zhang, L.-Y. (2015). Structure-activity relations in binding of perfluoralkyl compounds to human thyroid hormone T3 receptor. Archives of Toxicology, 89: 233-242. Ren, X.-M., Qin, W.-P., Cao, L.-Y., Zhang, J., Yang, Y., Wan, B. & Guo, L.-H. (2016). Binding interactions of perfluoroalkyl substances with thyroid hormone transport proteins and potential toxicological implications. Toxicology, 366-367: 32-42. Yang, L., Wang, Z., Shi, Y., Li, J., Wang, Y., Zhao, Y., Wu, Y. & Cai, Z. (2016). Human placental transfer of perfluoroalkyl acid precursors: Levels and profiles in paired maternal and cord serum. Chemosphere, 144: 1631-1638. Stahl, T., Hofmann, A., Cllen, M., Falk, S. & Brunn, H. (2014). Analysis of selected perfluoroalkyl substances (PFASs) in beer to evaluate the effect of beer consumption on human PFAS exposure: a pilot study. European Food Research and Technology, 238: 443-449. Serex, T., Anand, S., Munley, S., Donner, E.M., Frame, S.R., Buck, R.C. & Loveless, S.E. (2014). Toxicological evaluation of 6:2 fluorotelomer alcohol. Toxicology, 319: 1-9. Shi, G., Cui, Q., Pan, Y., Sheng, N., Guo, Y. & Dai, J. (2017). 6:2 fluorotelomer carboxylic acid (6:2 FTCA) exposure induces developmental toxicity and inhibits the formation of erythrocytes during zebrafish embryogenesis. Aquatic Toxicology, 190: 53-61. Sez, M., de Voogt, P. & Parsons, J.R. (2008). Persistence of perfluoroalkylated substances in closed bottle tests with municipal sewage sludge. Environmental Science and Pollution Research, 15: 472-477. Sheng, N., Zhou, X., Zheng, F., Pan, Y., Guo, X., Guo, Y., Sun, Y. & Dai, J. (2017). Comparative hepatotoxicity of 6:2 fluorotelomer carboxylic acid and 6:2 fluorotelomer sulfonic acid, two fluorinated alternatives to long-chain perfluoroalkyl acids, on adult male mice. Archives of Toxicology, 91: 2909-2919. Latala, A., Nedzi, M. & Stepnowski, P. (2009). Acute toxicity assessment of perfluorinated carboxylic acids towards the Baltic microalgae. Environmental Toxicology and Pharmacology, 28: 167-171. Ishibashi, H., Ishida, H., Matsuoka, M., Tominaga, N. & Arizono, K. (2007). Estrogenic effects of fluorotelomer alcohols for human estrogen receptor isoforms and in vitro. Biological and Pharmaceutical Bulletin, 30: 13581359. Maras, M., Vanparys, C., Muylle, F., Robbens, J., Berger, U., Barber, J.L., Blust, R. & De Coen, W. (2006). Estrogen-like properties of fluorotelomer alcohols as revealed by MCF-7 breast cancer cell proliferation. Environmental Health Perspectives, 114: 100-105. Krrman, A., Elgh-Dalgren, K., Lafossas, C. & Mskeland, T. (2011). Environmental levels and distribution of structural isomers of perfluoroalkyl acids after aqueous fire-fighting foam (AFFF) contamination. Environmental Chemistry, 8: 372-380. Yang, L., Li, J., Lai, J., Luan, H., Cai, Z., Wang, Y., Zhao, Y. & Wu, Y. (2016). Placental transfer of perfluoroalkyl substances and associations with thyroid hormones: Beijing prenatal exposure study. Scientific Reports, doi:10.1038/srep21699. Huset, C.A., Barlaz, M.A., Barofsky, D.F. & Field, J.A. (2011). Quantitative determination of fluorochemicals in municipal landfill leachates. Chemosphere, 82: 1380-1386. Eriksson, U., Mueller, J.F., Toms, L.-M.L., Hobson, P. & Krrman, A. (2017). Temporal trends of PFSAs, PFCAs and selected precursors in Australian serum from 2002 to 2013. Environmental Pollution, 220: 168-177. Vestergren, R., Berger, U., Glynn, A. & Cousins, I.T. (2012). Dietary exposure to perfluoroalkyl acids for the Swedish population in 1999, 2005 and 2010. Environment International, 49: 120-127. Eschauzier, C., Hoppe, M., Schlummer, M. & de Voogt, P. (2013). Presence and sources of anthropogenic perfluoroalkyl acids in high-consumption tap-water based beverages. Chemosphere, 90: 36-41. Wang, Y., Han, W., Wang, C., Zhou, Y., Shi, R., Bonefeld-Jorgensen, E.C., Yao, Q., Yuan, T., Gao, Y., Zhang, J. & Tian, Y. (2019). Efficiency of maternal-fetal transfer of perfluoroalkyl and polyfluoroalkyl substances. Environmental Science and Pollution Research, 26: 2691-2698. Own research Own research Own research Own research Own research Own research Own research Own research Own research Own research Own research Own research Own research Own research Own research Own research Own research Kang, H., Choi, K., Lee, H.-S., Kim, D.-H., Park, N.-Y., Kim, S. & Kho, Y. (2016). Elevated levels of short carbon-chain PFCAs in breast milk among Korean women: Current status and potential challenges. Environmental Research, 148: 351-359. Lee, S., Kim, S., Park, J., Kim, H.-J., Choi, G., Choi, S., Kim, S., Kim, S.Y., Kim, S., Choi, K. & Moon, H.-B. (2018). Perfluoroalkyl substances (PFASs) in breast milk from Korea: Time-course trends, influencing factors, and infant exposure. Science of the Total Environment, 612: 286-292. Lee, Y.A., Kim, J.H., Jung, H.W., Lim, Y.-H., Bae, S., Kho, Y., Hong, Y.-C., Shin, C.H. & Yang, S.W. (2018). The serum concentrations of perfluoroalkyl compounds were inversely associated with growth parameters in 2-year old children. Science of the Total Environment, 628-629: 226-232. Han, W., Gao, Y., Yao, Q., Yuan, T., Wang, Y., Zhao, S., Shi, R., Bonefeld-Jorgensen, E.C., Shen, X. & Tian, Y. (2018). Perfluoroalkyl and polyfluoroalkyl substances in matched parental and cord serum in Shandong, China. Environment International, 116: 206-213. Cai, D., Li, Q.-Q., Chu, C., Wang, S.-Z., Tang, Y.-T., Appleton, A.A., Qiu, R.-L., Yang, B.-Y., Hu, L.-W., Dong, G.-H. & Zeng, X.-W. (2020). High trans-placental transfer of perfluoroalkyl substances alternatives in the matched maternalcord blood serum: Evidence from a birth cohort study. Science of the Total Environment, 705: 135885. Knobeloch, L., Imm, P. & Anderson, H. (2012). Perfluoroalkyl chemicals in vacuum cleaner dust from 39 Wisconsin homes. Chemosphere, 88: 779-783. Qi, X., Zhou, J., Wang, M., Yang, M.-R., Tang, X.-Y., Mao, X.-F. & Wang, T.-T. (2019). Perfluorinated compounds in poultry products from the Yangtze River Delta and Pearl River Delta regions in China. Science of the Total Environment, 689: 1079-1086. Su, H., Shi, Y., Lu, Y., Wang, P., Zhang, M., Sweetman, A., Jones, K. & Johnson, A. (2017). Home produced eggs: An important pathway of human exposure to perfluorobutanoic acid (PFBA) and perfluorooctanoic acid (PFOA) around a fluorochemical industrial park in China. Environment International, 101: 1-6. Zhang, W., Zhang, D., Zagorevski, D.V. & Liang, Y. (2019). Exposure of Juncus effusus to seven perfluoroalkyl acids: Uptake, accumulation and phytotoxicity. Chemosphere, 233: 300-308. Heo, J.-J., Lee, J.-W., Kim, S.-K. & Oh, J.-E. (2014). Foodstuff analyses show that seafood and water are major perfluoroalkyl acids (PFAAs) sources to humans in Korea. Journal of Hazardous Materials, 279: 402-409. Becanov, J., Melymuk, L., Vojta, S. Komprdov, K. & Klnov, J. (2016). Screening for perfluoroalkyl acids in consumer products, building materials and wastes. Chemosphere, 164: 322-329. Domingo, J.L., Jogsten, I.E., Eriksson, U., Martorell, I., Perell, G., Nadal, M. & van Bavel, B. (2012). Human dietary exposure to perfluoroalkyl substances in Catalonia, Spain. Temporal trend. Food Chemistry, 135: 1575-1582. Tian, Z., Kim, S.-K., Shoeib, M., Oh, J.-E. & Park, J.-E. (2016). Human exposure to per- and polyfluoroalkyl substances (PFASs) via house dust in Korea: Implication to exposure pathway. Science of the Total Environment, 553: 266-275. Pitter, G., Da Re, F., Canova, C., Barbieri, G., Jeddi, M.Z., Dapr, F., Manea, F., Zolin, R., Bettega, A.M., Stopazzolo, G., Vittorii, S., Zambelli, L., Martuzzi, M., Mantoan, D. & Russo, F. (2020). Serum levels of perfluoroalkyl substances (PFAS) in adolescents and young adults exposed to contaminated drinking water in the Veneto Region, Italy: A crosssectional study based on a health surveillance program. Environmental Health Perspectives, 128: 027007-1 - 02700712. Lee, Y.-M., Lee, J.-Y., Kim, M.-K., Yang, H., Lee, J.-E., Son, Y., Kho, Y., Choi, K. & Zoh, K.-D. (2020). Concentration and distribution of per- and polyfluoroalkyl substances (PFAS) in the Asan Lake area of South Korea. Journal of Hazardous Materials, 381: 120909. Coggan, T.L., Moodie, D., Kolobaric, A., Szabo, D., Shimeta, J., Crosbie, N.D., Lee, E., Fernandes, M. & Clarke, B.O. (2019). An investigation into per- and polyfluoroalkyl substances (PFAS) in nineteen Australian wastewater treatment plants (WWTPs). Heliyon, 5: e02316. Hepburn, E., Madden, C., Szabo, D., Coggan, T.L., Clarke, B. & Currell, M. (2019). Contamination of groundwater with per- and polyfluoroalkyl substances (PFAS) from legacy landfills in an urban re-development precinct. Environmental Pollution, 248: 101-113. information out of literature reviews can be retrieved from other studies.