Perfluorinated alkyl substances (PFASs) in household dust in Central Europe and North America

Environment International 94 (2016) 315–324

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Perfluorinated alkyl substances (PFASs) in household dust in Central Europe and North America Pavlína Karásková a, Marta Venier b,⁎, Lisa Melymuk a,⁎⁎, Jitka Bečanová a, Šimon Vojta a, Roman Prokeš a, Miriam L. Diamond c, Jana Klánová a a b c

Research Centre for Toxic Compounds in the Environment (RECETOX), Masaryk University, Kamenice 753/3, 625 00 Brno, Czech Republic School of Public and Environmental Affairs, Indiana University, 702 N. Walnut Grove Ave., Bloomington, IN 47405, USA Department of Earth Sciences, University of Toronto, 22 Russell Street, Toronto, Ontario M5S 3B1, Canada

a r t i c l e

i n f o

Article history: Received 22 January 2016 Received in revised form 27 May 2016 Accepted 27 May 2016 Available online xxxx Keywords: Perfluorinated compounds PFASs House dust Indoor environment

a b s t r a c t Concentrations of 20 perfluorinated alkyl substances (PFASs) were measured in dust samples from 41 homes in Canada, the Czech Republic, and United States in the spring-summer of 2013. The most frequently detected compounds were perfluorohexanoic acid (PFHxA) and perfluorooctane sulfonate (PFOS). PFOS and perfluorooctanoic acid (PFOA) had the highest concentrations of PFASs in all countries. PFOS median concentrations for the three countries were between 9.1 and 14.1 ng/g, and PFOA medians ranged between 8.2 and 9.3 ng/g. In general, concentrations in North America were higher than in the Czech Republic, which is consistent with usage patterns. No differences were found for perfluorooctane sulfonamides/sulfonamidoethanols (FOSA/Es) levels due to the low number of detections. Homologue profiles suggest that the shift from longer to shorter chain PFASs is more advanced in North America than in Europe. Significant relationships were found among individual homologues and between PFAS concentrations in dust and type of floor, number of people living in the house, and building age. © 2016 Elsevier Ltd. All rights reserved.

1. Introduction Perfluorinated alkyl substances (PFASs) are industrial chemicals characterized by fluorinated carbon chains with different functional groups. PFASs can be subdivided into several categories, [e.g. perfluoroalkyl carboxylic acids (PFCAs), perfluoroalkane sulfonates (PFSAs), perfluoroalkane sulfonamides (FOSAs), and perfluoroalkane sulfonamidoethanols (FOSEs)]. PFASs have been used for more than 60 years in a variety of applications: surfactants, lubricants, paper and textile coatings, polishes, food packaging, and fire-fighting foams. They are added to consumer products to make them resistant to water, oil, stains, and even fire (Kissa, 2001). PFAS degrade in the environment; for example, volatile sulfonamides such as N-methyl perfluorooctane sulfonamidoethanol (N-MeFOSE) and N-ethyl perfluorooctane sulfonamidoethanol (N-EtFOSE), which are found in surface protection products, are precursors of PFOS, a compound most commonly detected in environmental matrices (Buck et al., 2011; Lehmler, 2005; Renner, 2004). Although there are only a few companies that manufacture ⁎ Correspondence to: M. Venier, School of Public and Environmental Affairs, Indiana University, 702 N. Walnut Grove Ave., Bloomington, IN 47405, USA. ⁎⁎ Correspondence to: L. Melymuk, Research Centre for Toxic Compounds in the Environment (RECETOX), Masaryk University, Kamenice 753/3, 625 00 Brno, Czech Republic. E-mail addresses: [email protected] (M. Venier), [email protected] (L. Melymuk).

http://dx.doi.org/10.1016/j.envint.2016.05.031 0160-4120/© 2016 Elsevier Ltd. All rights reserved.

PFASs (e.g. Arkema, Asahi, Atofina, Ciba, Clariant, Daikin, DuPont, 3M and Solvay Solexis) (Environment Canada, 2012), it has been estimated that more than 100,000 metric tons were produced between 1970 and 2002 (Paul et al., 2009), and because of their stability and transport potential they have become ubiquitous environment contaminants (Lau et al., 2007). Wang et al. (2014) estimated environmental emissions of 2610–21,400 metric tons of C4–C14 PFCAs between 1951 and 2015. PFASs have been detected globally in the abiotic environment (Gomez et al., 2011; Langer et al., 2010; Shoeib et al., 2010; Sun et al., 2011; Wang et al., 2011; Yang et al., 2011), in biota (Giesy and Kannan, 2001; Kannan et al., 2002; Kannan et al., 2001a; Kannan et al., 2001b), in humans (De Silva and Mabury, 2006; Kärrman et al., 2006a; Kärrman et al., 2006b; Lau et al., 2007; Olsen et al., 2003) and in the indoor environment as a result of consumer products (e.g. food packaging, cleaning agents, textiles) containing PFASs (Gewurtz et al., 2009; Herzke et al., 2012; Liu et al., 2014; U.S.EPA, 2009; Liu et al., 2014 #25). Toxicological data are mostly available for PFOA and PFOS, with research continuing to find adverse effects. In general, toxic effects and environmental fate mainly depend on the fluorinated chain length and the type of functional group (Lau et al., 2007). For example, C8 PFASs (PFOS and PFOA) accumulate primarily in blood serum, kidneys and liver, and no metabolism is expected (Calafat et al., 2006; U.S.EPA, 2006b). Their half-life in humans ranges between 2 and 9 years (Kärrman et al., 2006b; Olsen et al., 2007; Wong et al., 2014). Adverse effects on sperm quality (Joensen et al., 2009), reduced body weight

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(Nelson et al., 2010), changes in thyroid hormone levels (Dallaire et al., 2009) and deregulation of lipid and lipoprotein metabolism have been recently reported (U.S. EPA, 2012). PFOA and PFOS in umbilical cord blood and breast milk have been correlated with concentrations in maternal blood indicating transplacental transfer to the fetus (Beesoon et al., 2011; Midasch et al., 2007) and lactational exposure for infants (Tao et al., 2008). In May 2006 PFOA was classified as “likely carcinogenic to humans” (U.S.EPA, 2006c). Prior to 2003, PFAS production consisted mainly of C8 compounds. After 2003, a shift to C4 and C6 PFASs occurred, as these shorter chain compounds have lower toxicity and lower bioaccumulation potential when compared to the legacy C8 compounds (U.S. EPA, 2012; Lim et al., 2011). Shorter chain PFASs have similar functional properties to the longer chain compounds, although information about production volumes, uses, properties, and environmental fate are still limited (U.S. EPA, 2012; Ritter, 2010; Scheringer et al., 2014). In 2009 PFOS, its salts and perfluorooctane sulfonyl fluoride (PFOSF) were listed under Annex B of the Stockholm Convention on Persistent Organic Pollutants (POPs), which restricted their manufacturing and use to a few specific applications (Pollutants, 2012). In addition to this global convention, other jurisdictions have specific regulations. For example, beginning in 2006 Canada established that the sale of longer chain PFAS including PFOA had to be phased out by 2016, and it made the import and sale of long chain PFAS illegal (Environment Canada, 2015). Also, in 2006 U.S. EPA created the 2010/2015 PFOA Stewardship Program inviting the major fluoropolymer and telomer manufacturers to eliminate emissions of PFOA, precursors chemicals and related higher homologue chemicals, and reduce product content levels of these chemicals (U.S.EPA, 2006a). Additionally, the U.S. Environmental Protection Agency (EPA) introduced two Significant New Use Rules (SNURs) in 2002 covering 88 PFOS-related substances and requiring companies to notify the U.S. EPA before manufacture or import. The SNURs were amended in 2007 to include 183 additional PFOS-related substances and to eliminate the use of long-chain PFCs by 2015 (EPA, 2014). Although human exposure to PFAS, particularly PFOA and PFOS, is dominated by dietary (Kärrman et al., 2009; Trudel et al., 2008) and drinking water uptake (Emmett et al., 2006; Hölzer et al., 2008), dust, indoor air, and consumer products may be significant exposure routes (Egeghy and Lorber, 2011; Fraser et al., 2012; Vestergren and Cousins, 2009). Indoor air may contain 20 times higher concentrations of some PFAS than outdoor air (Shoeib et al., 2011), and dust, due to its fine particle size and complex composition of organic and inorganic matrices, is capable of sorbing and concentrating some PFASs. Trace levels of PFAS have been found in consumer products including waterproofing agents, paints, coated fabrics, non-stick-ware and electronics (Washburn et al., 2005) (Herzke et al., 2012; Kotthoff et al., 2015; Vestergren et al., 2015; Washburn et al., 2005). The combination of exposure to indoor air, dust and consumer products treated with PFAS can lead to exposure via dermal contact, ingestion of dust and inhalation of particles and gases (Washburn et al., 2005). For toddlers and children, dust ingestion may be as significant a route as dietary intake due to hand-to-mouth behavior and time spent on the floor (Egeghy and Lorber, 2011; Klepeis et al., 2001; U.S.EPA, 1997). Median total PFASs daily intakes via dust ingestion were 92 ng for children and 46 ng for adults in a study in the USA (Strynar and Lindstrom, 2008). Several studies from different countries have reported concentrations of PFASs in indoor dust, and found that levels of PFASs in dust samples differ according to sampling period, country, type of room and sampling technique (Ericson Jogsten et al., 2012; Eriksson and Kärrman, 2015; Fraser et al., 2013; Huber et al., 2011; Knobeloch et al., 2012; Lankova et al., 2015; Liu et al., 2011; Shoeib et al., 2011; Shoeib et al., 2016; Xu et al., 2013). These studies as a whole suggest that human exposure to PFAS is highly dependent on geography, season and activity patterns.

This work is part of a larger study within the Czech-American Scientific Cooperation Program aimed at identifying levels of contamination in residential indoor environments from three countries (Czech Republic, Canada, and USA). Results for flame retardants and legacy POPs are presented elsewhere (Venier et al., 2016). The aim of the present study was to investigate and compare indoor concentrations of 20 PFASs in house dust samples collected in living rooms and bedrooms between April and August 2013 from more than 40 homes in these three countries. Spatial differences in concentrations and congener patterns are evaluated, as well as possible sources within the houses. Finally, the impact on human exposure from dust ingestion was calculated. 2. Materials and methods 2.1. Samples Fifty-six dust samples were collected using polyester vacuum socks between April and August 2013 from living rooms and bedrooms from 41 houses (12 in Czech Republic, 15 in Canada and 14 in USA) – see Table S1 for details. The dimensions of the room, sampled area and type of flooring were recorded at the time of sample collection. At the same time, participants completed a questionnaire regarding room equipment, ventilation, cleaning and other living habits. Details on sampling sites are given in Table S1 in the Supplementary information. The same sampling protocol was used in the three sampling campaigns (Czech Republic, Canada and USA) to ensure minimal differences and sampling artifacts. Also, all samples were analyzed in the same laboratory (RECETOX). Before sampling, polyester vacuum socks were pre-cleaned in a Soxhlet extractor (8 h in acetone, then 8 h in toluene) and stored in clean aluminum foil. For sample collection, socks were inserted into the hose of a household vacuum cleaner, and from 1 to 16 m2 were vacuumed (see Table S1). All samples were packed in clean aluminum foil, sealed, labelled and subsequently stored at −20 °C until analysis. 2.2. Sample preparation and extraction Dust samples were sieved using a 500 μm sieve and weighed. Approximately 100 mg of sieved dust was transferred into polypropylene (PP) Falcon tubes that were pre-cleaned with methanol. Prior to extraction, dust samples were spiked with internal isotopically labelled standards (M8PFOA, M8PFOS). All samples were extracted in methanol with 5 mM ammonium acetate using an ultrasonic bath (15 min, 3 cycles). After each extraction cycle, the supernatant was decanted to pre-cleaned PP Falcon tubes. Extract volumes were reduced under a gentle stream of nitrogen to the last drop and re-diluted into the mobile phase using a solution of ammonium acetate in water (concentration 5 mM) and methanol up to the final volume (50/50, ammonium acetate in water/ammonium acetate in methanol, v/v). The concentrated extracts were cleaned using a syringe filter (nylon membrane, 13 mm diameter and 0.45 μm pore size). Aliquots of 100 μL were removed for HPLC-MS/MS analysis. 2.3. Chemicals Twenty analytical standards (10 PFCAs, 5 PFSAs, FOSA, N-MeFOSA, N-EtFOSA, N-MeFOSE and N-EtFOSE) and 12 mass-labelled standards (MPFHxA, MPFOA, M8PFOA, MPFNA, MPFDA, MPFUnDA, MPFDoDA, MPFHxS, MPFOS, M8PFOS, dMeFOSA, dMeFOSE) were purchased from the Wellington Laboratories Inc. (Guelph, Ontario, Canada). M8PFOA and M8PFOS were used as internal standards and were added before extraction. The other mass-labelled standards were used as recovery standards and were added before analyses to both calibration standards and samples. The solvents used were methanol (LC/MS grade, Biosolve b.v., Valkenswaard, Netherlands) and water (HPLC Gradient Grade, Fisher Scientific, Loughborough, UK). Ammonium acetate was used as an

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addition to the extraction solvent and LC mobile phase [p.a. grade (≥98.0%), Fluka Chemie GmbH, Buchs, Germany]. All chemicals and solvents were used without further purification. 2.4. HPLC-MS/MS analysis The separation, identification, and quantification of all target PFASs were performed with high performance liquid chromatography (HPLC) using an Agilent 1290 instrument (Agilent Technologies, Palo, Alto, California, USA) coupled to a QTRAP 5500 mass spectrometer (AB Sciex, Foster City, California, USA). Chromatographic separation was performed at 20 °C on a SYNERGI 4μ Fusion RP 80Ä 50 mm × 2 mm column with a 4 × 2.00 mm corresponding precolumn (Phenomenex, USA). The flow rate was set to 200 μL/min. Gradient separation was used, with mobile phase A of methanol/5 mM ammonium acetate in water 55/45 (v/v) and mobile phase B of methanol (see Table S2 in Supplementary information). Aliquots of 10 μL were injected onto the column. The mass spectrometer was operated in the electrospray negative ionization mode (ESI-) using two multiple reaction monitoring (MRM) transitions for each compound except N-MeFOSE and N-EtFOSE at 450 °C and ion voltage 4500 V (for details see Tables S3–4 in Supporting information). 2.5. Quality assurance/quality control The efficiency of the extraction method was tested in a spiking experiment using a house dust Standard Reference Material (detailed in SI Table S5). Recoveries of M8PFOA ranged from 72.3% to 103.7%, with an average of 85.7 ± 8.2%, and of M8PFOS from 72.6% to 114.2% with an average of 92.9 ± 8.4%. The method accuracy ranged between 73% and 99% with the mean 82% for all analytes. The method precision was high, with relative standard deviations below 10%, except for PFUnDA where RSD reached 10.7% (for details see SI Table S5). Quantification of data was done by internal standard calculation using eight-point quadratic calibration curves (r2 N 0.99) ranging from 0.004 to 10 ng/mL. Isotopically labelled internal standards (MPFHxA, MPFOA, MPFNA, MPFDA, MPFUnDA, MPFDoDA, MPFHxS, MPFOS, dMeFOSA, dMeFOSE) were added to all samples and standards prior to analysis. Method quantification limits (MQLs) in ng/sample were defined as the mean concentration of procedural blanks plus ten times the standard deviation of blank response. MQL in ng/gram was calculated by adjusting the MQL in ng/sample according to the sample mass (see Table S6 in Supporting information). Concentrations below method quantification limits were replaced by √ 2/2*MQL for the statistical data analysis (Antweiler, 2015). Quality control was also performed by analysis of one procedural blank (consisting of one pre cleaned polyester sock) in each batch of 12 samples. The average blank concentration was compared with samples and if the average of the blanks was N 35% of the sample mass, samples were reported as non-detect (ND). Samples were not blank corrected if average blank was b10% of the sample mass, but they were blank corrected if the average of blanks was 10–35% of the sample mass. 2.6. Human exposure PFASs concentrations in dust samples from this study were used to estimate PFAS exposure via dust ingestion of toddlers and adults. Human exposure to PFOA, PFOS and sum of PFSAs and PFCAs was estimated using two exposure scenarios representing mean and high dust intake rates and three concentration levels of PFASs (5th percentile, median, and 95th percentile). The dust ingestion rate was calculated using the following expression: Eingest ¼ Cdust  Q dust  Fuptake

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where Cdust is the concentration of individual compounds or total PFASs in house dust in ng/g; Qdust is the dust ingestion rate in mg/day (100 and 4.16 for mean scenario, and 200 and 55 for the high scenario for toddlers and adults, respectively (U.S.EPA, 2011)). Fuptake is the uptake fraction of 0.8 for PFSAs as was estimated in previous study on gastrointestinal absorption (Beesoon et al., 2011; Wong et al., 2014).

3. Results 3.1. Concentrations of PFASs in dust samples Table 1 shows summary results for PFAS concentrations, and individual concentrations are presented in Tables S7–9 in the SI. Box plots with ANOVA results, calculated using log-transformed data, are shown in Fig. 1. In the following discussion ∑PFCAs = sum of PFPA, PFHxA, PFOA, PFNA, PFDA, PFUnDA, PFTrDA, and PFTeDA; ∑PFSAs = sum of PFBS, PFHxS, PFHpS, PFOS and PFDS; ∑ FOSA/Es = sum of FOSA, NMeFOSA, N-EtFOSA, N-MeFOSE and N-EtFOSE. Nineteen of the 20 target PFASs were detected in dust samples, although the detection frequency was quite variable depending on compound and geographical location (see Fig. 2). Among the PFCAs and PFSAs, PFHxA was detected in all samples regardless of the country, and PFOS was detected in 100% and 94% of samples from North America and the Czech Republic, respectively. PFOA was detected in 89% of all samples. Five other compounds were detected in more than 50% of all samples: PFHpA, PFNA, PFDoDA, PFHxS and PFDS. PFNA was among the most abundant compound in samples from North America with an incidence of 98%, and N-MeFOSE was detected in 70% of samples from USA only. N-MeFOSA and N-EtFOSA were detected only in one or two cases, and FOSA was not detected in any sample. Although the most frequently detected compound was PFHxA, PFOS and PFOA were generally measured at the highest concentrations (see Fig. 2). PFOS median concentrations were between 9.1 and 14.1 ng/g for the three countries; PFOA medians were between 2.0 and 9.0 ng/g, with a maximum concentration of 318 ng/g measured in one sample from USA. Medians for ∑ PFCAs and ∑ PFSAs ranged between 10.2 and 49.4 ng/g and between 16.1 and 33.8 ng/ g, respectively. ∑ FOSA/Es concentrations were much lower than the other two groups with medians ranging between not detected and 1.3 ng/g. In general, concentrations in North America were higher than in Czech Republic (see Fig. 1). ∑ PFCAs were highest in samples from North America (USA ~ CAN N CZ) and ∑ PFSAs followed the spatial trend USA N CAN with CZ not distinguishable. ∑FOSA/Es showed no statistically significant difference between countries, probably due to the relatively low detection frequencies. Fewer differences between countries were found for individual compounds. We observed significant differences for nine compounds: PFPA, PFHpA, PFOA, PFNA, PFDA, PFHxS, PFOS, PFDS and N-MeFOSE. PFPA, PFHpA, PFOA, PFNA, PFDA and PFDS concentrations were significantly higher in North America than in the Czech Republic (p b 0.05). For PFOS, the concentrations measured in the USA were significantly higher than those measured in Canada, with the Czech samples not significantly different from those from North America. For PFHxS, the trend was USA N CAN ~ CZ. We suggest that these differences may be explained by differences in the market, import history and usage of these substances, potentially compounded by differences in product turn-over rates between CZ, CAN and USA. Decreasing concentrations of PFOS and related compounds are expected in all countries due to national and international controls, although the rate of change will be dampened by the time over which products containing these products come out of service. In general, these spatial differences should be interpreted cautiously because large variability could exist within countries.

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Table 1 Summary results for dust samples (ng/g). Czech Republic Mean ±

Canada

Medianb Range

SEa

n N MQL (%)

Mean ±

USA Medianb Range

SEa

PFPA PFHxA PFHpA PFOA PFNA PFDA PFUnDA PFDoDA PFTrDA PFTeDA PFBS PFHxS PFHpS PFOS PFDS

2.4 ± 0.5 12.8 ± 5.0 3.1 ± 0.5 8.9 ± 2.5 3.0 ± 0.9 5.2 ± 1.7 4.3 2.5 ± 1.0 3.5 3.6 ± 1.2 3.6 ± 1.4 2.8 ± 0.6 1.4 ± 0.5 20.7 ± 7.3 1.6 ± 0.4

bMQL 3.8 1.8 2.0 bMQL bMQL bIQL 0.5 bIQL bMQL bMQL 2.0 bMQL 10.3 0.8

ND–6.3 1.4–69.1 bMQL–7.8 bMQL–26.7 ND–11.0 ND–17.1 ND–4.3 ND–13.1 ND–3.5 ND–14.8 ND–14.4 bMQL–9.3 ND–4.6 4.8–118 (262) ND–6.6

6 (37.5) 16 (100) 11 (68.8) 11 (68.8) 8 (50.0) 5 (31.3) 1 (6.3) 9 (56.3) 1 (6.3) 7 (43.8) 6 (37.5) 15 (93.8) 5 (31.3) 15 (94) 12 (75.0)

b a b b b b a a a a a b a ab b

4.5 ± 1.0 14.5 ± 7.0 7.1 ± 2.4 17.7 ± 5.4 19.4 ± 10.2 8.5 ± 4.3 8.7 ± 3.4 6.3 ± 3.5 8.2 ± 1.8 4.8 ± 2.0 1.6 ± 0.3 3.1 ± 0.7 2.4 ± 1.0 10.8 ± 1.7 3.7 ± 0.7

2.0 5.6 3.4 8.2 4.4 2.4 1.1 1.1 bMQL 1.4 bMQL 1.9 bIQL 9.1 3.2

bMQL–17.0 1.7–146 bMQL–47.4 2.1–92.7 bMQL–195 0.9–86.2 ND–49.6 ND–61.1 ND–19.4 bMQL–33.6 ND–5.8 ND–11.5 ND–12.7 3.3–31.8 0.5–14.5

FOSA N-MeFOSA N-EtFOSA N-MeFOSE N-EtFOSE ∑PFCAs ∑PFSAs ∑FOSA/Es

– – – 8.9 ± 2.5 1.7 ± 0.6 28.5 ± 1.4 34.0 ± 2.4 1.7 ± 1.9

– – – bIQL bIQL 10.2 16.1 ND

– – – ND–16.0 ND–6.4 2.5–120 2.9–266 ND–22.4

0 (0) 0 (0) 0 (0) 2 (12.5) 6 (37.5) 75 (49) 54 (68) 8 (10)

a a a b a b ab a

– 0.7 – 12.2 ± 5.1 0.5 ± 0.05 83.3 ± 2.1 19.1 ± 0.7 2.6 ± 4.6

– bIQL – bIQL bIQL 38.7 13.6 ND

– 0.7 – ND–46.3 ND–0.8 5.2–428 6.1–36.0 ND–47.1

n N MQL (%)

Mean ±

Medianb Range

SEa

15 (75.0) 20 (100) 17 (85.0) 20 (100) 19 (95.0) 20 (100) 12 (60.0) 15 (75.0) 4 (20.0) 13 (65.0) 11 (55.0) 18 (90.0) 7 (35.0) 20 (100) 20 (100)

a a ab a a a a a a a a b a b a

5.4 ± 1.4 20.9 ± 9.5 11.8 ± 4.8 38.6 ± 17.1 10.9 ± 3.8 6.9 ± 3.1 3.6 ± 0.8 2.0 ± 0.5 1.8 ± 0.1 1.4 ± 0.14 1.4 ± 0.13 13.8 ± 4.0 1.3 ± 0.2 42.4 ± 12.5 2.9 ± 0.5

1.7 6.5 3.6 9.0 3.9 1.8 1.2 0.6 bMQL 0.8 0.9 8.7 bIQL 14.1 2.8

0 (0) 1 (5.0) 0 (0) 4 (20.0) 5 (25.0) 155 (78) 76 (76) 10 (10)

a a a b a a b a

– 0.6 5.9 2.2 ± 0.5 32.0 ± 12.0 94.9 ± 3.0 60.0 ± 3.9 6.7 ± 4.9

– bIQL bIQL 1.0 bMQL 49.4 33.8 1.3

ND–24.8 2.5–190 0.9–86.7 2.9–318 1.1–62.9 0.4–64.0 ND–13.1 ND–9.0 ND–2.1 bMQL–3.0 ND–2.6 1.4–84.4 ND–2.9 5.7–239 0.5–9.8 (764) – 0.6 5.9 ND–9.9 ND–93.9 8.0–573 12.2–256 ND–102.3

n N MQL (%) 15 (75.0) 20 (100) 19 (95.0) 19 (95.0) 20 (100) 20 (100) 12 (60.0) 12 (60.0) 3 (15.0) 10 (50.0) 12 (60.0) 20 (100) 6 (30.0) 20 (100) 18 (90)

a a a a a a a a a a a a a a a

0 (0) 1 (5.0) 1 (5.0) 14 (70.0) 3 (15.0) 150 (75) 78 (78) 19 (19)

a a a a a a a a

ND = not detected. bIQL = below instrumental quantification limit. bMQL = below quantification limit. Italics in brackets = outliers. a Mean was calculated only from values N MQL. b Median was calculated for all samples using substitution of values b MQL.

3.2. Comparison with published data and temporal trends A comparison of median concentrations for PFASs in house dust in this study with values available in the literature for Europe and North America is presented in Tables 2 and 3, respectively. Concentrations and congener patterns of PFASs in Czech and North American dust in this study are generally consistent with data from previous studies, with a few exceptions, especially for samples collected recently. In the only other study that measured dust from Czech Republic, Lankova et al. (2015) reported much lower concentrations for PFOS than in the present study (median of 1.5 ng/g in Lankova et al. vs. 10.3 ng/g in this study), but similar concentrations of PFOA (median of 2.4 ng/g in Lankova et al. vs. 2.0 ng/g in this study). The differences in PFOS levels may be due to differences in timing of sample collection and surfaces sampled; Lankova et al. sampled raised surfaces in addition to floor dust. Eriksson and Kärrman (2015) collected samples in Ottawa, Canada, in the same year as the present study, and the results are comparable. Shoeib et al. (2011) reported median concentrations of 30 ng/g for PFOA and 71 ng/g for PFOS in Vancouver, Canada, which are significantly higher than our study. The wide range of dust sampling techniques employed (e.g., differences in vacuum cleaners, targeted surfaces, sieving of samples) could affect comparability between studies. However, we suspect that some of the larger differences between this study and some of the published data are due to the fact that samples were collected several years previous to our study. Some of the samples included in Tables 2 and 3 were collected as early as 2008 and 5 years can be significant in a rapidly changing market such as that of PFASs. The results from similar studies may be used to deduce apparent temporal trends of PFASs measured in dust in indoor environments. We focused on PFOA, PFOS and PFHxA since they were reported consistently in the literature; we looked at Europe and North America separately. Plots of ranges and medians dust concentrations versus sampling year are shown in Fig. S1. Results from this analysis should

be interpreted cautiously because limited data availability precluded more sophisticated approach such as whole distribution comparisons; also, possible differences in sampling and instrumental analysis as well as within-country variability could influence these results. However, despite all the limitations described above, this comparison provides useful preliminary insights on the temporal trends of these compounds. Keeping in mind the caveats above, the analysis revealed that no temporal trends were observed in Europe while PFHxA, PFOA and PFOS are generally decreasing with time in North America. The decrease was significant for median PFOA (p = 0.045) and PFOS (p = 0.004) (for detailed information on calculations and results see Table S14). The presence of a temporal trend in North America but not in Europe is probably related to higher initial concentrations and data starting earlier (1998 for North America vs. 2006 for Europe). PFHxA data are more limited, and thus temporal trends are harder to deduce, particularly for Europe; however, recent measurements in North America shows a significant decline (p = 0.0083 based on median values) while for Europe there is no observable change. The trends for PFOA and PFOS suggest that both regulations and voluntary industry actions are effective in reducing the use of PFASs, and thus levels in indoor environments. A certain time lag in the decrease in concentrations should be expected based on the in-use lifetime of products containing these compounds (e.g. Abbasi et al. (2015). 3.3. PFAS profiles in dust samples Fig. 3 shows the percentage pattern of PFASs in samples from the three countries calculated using median concentrations. The dominant compound in all three countries was PFOS, representing ~ 30–50% of total PFASs measured. The profiles for the remaining compounds were different between CZ and North America. In the USA and Canada, PFOA was the second most abundant compound, followed by PFHxA. In CZ, the order was reversed, with PFHxA as second most abundant

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Fig. 1. Box and whisker plots of concentration (ng/g) showing the distribution of PFASs in dust samples from Czech Republic, Canada and USA. The lower and upper ends of the box are the 25th and 75th percentiles of data. The horizontal line within the box is the median value. The whiskers define the 5th and 95th percentiles and symbol ● illustrates outliers. Boxes that share the same letter are not significantly different at a 5% level in ANOVA analysis using Tukey's test.

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Fig. 2. Plot of median concentrations (ng/g) versus detection frequency (%) for each target compound in the three countries.

compound and PFOA as the third. PFNA was much more abundant in North America (13–18%) than in the Czech Republic (6%). These differences between the Czech Republic and North America may suggest a faster shift from long chain PFASs to their shorter chain homologues in Europe than in North America. Europe implemented a directive restricting the production and use of PFASs and related chemicals as early as 2006 (2006/122/EC) and the EPA followed suit only in 2010 (US EPA 2010/2015 Stewardship Program). The US EPA 2010/2015

Stewardship Program included major European producers such as Clariant.

3.4. Correlations among PFAS concentrations Relationships between different PFAS levels in all dust samples were investigated with Pearson correlation coefficients based on log-

Table 2 Summary of PFAS medians and ranges (ng/g) in European dust samples reported in the present study and published literature. Location

Greece, Athens

Spain, Catalonia

Sweden, Örebro, Växjö, Nyköping

Norway, Oslo

Norway, Tromsø

Belgium, Flanders

Czech Republic, Prague

Czech Republic, Brno

Type of room (N) Year of sampling PFPA

houses (7)

houses (10)

houses (10)

homes (18)

houses (16)

2009

2013–2014

living rooms (7) 2007/2008

houses (45)

2013–2014

bookshelves and window sills (41) 2008

2008

2013

2013

b0.36 (b0.36–8.7) 3.9 (b3.21–26.2) na (b1.23–30.1) 12.8 (b7.01–129) 3.6 (b0.96–17.7) 5.4 (b1.85–20.6) 2.9 (0.72–7.88) 3.6 (1.09–8.01) 1.1 (0.6–3.02) 1.3 (0.55–6.29) 1.2 (b0.39–2.74) 0.4 (0.26–11.3) –

0.37 (b0.36–2.11) 3.4 (b3.21–5.45) 0.9 (b1.23–5.86) 8.8 (b7.01–39.7) 1.7 (b0.96–42.6) 3.7 (2.12–80.6) 1.1 (b0.42–33.4) 2.0 (1.04–59.7) 0.7 (0.36–12.6) 1.1 (b0.02–14.6) 1.4 (b0.39–3.89) 0.4 (b0.09–1.04) –

na (b0.36–5.26)

3 (1.50–29)

6 (2.2–21.1)

–

(b1–4.73)

7.1 (b3.21–39.6)

28 (4.30–96)

–

(b1–9.73)

1.5 (b1.23–8.8)

9.4 (4.50–28)

10.1 (1.6–26.7) 9.2 (2.8–18.5)

bMQL (ND–6.3) 3.8 (1.4–69.1)

–

(b1–13.3)

14.4 (b7.01–49.5)

18 (6.20–56)

0.7 (b0.05–109)

2.4 (b1–9.12)

2.0 (b0.96–8.54)

23 (3.90–92)

38.8 (10.2–80.1) 7 (3.3–26.7)

–

(b1–8.88)

3.7 (b1.85–35.1)

4.1 (1.1–12.0)

7.5 (2–10.5)

–

(b1–12.4)

0.9 (b0.42–6.08)

–

–

(b1–3.26)

3.4 (0.58–17.4)

19 (1.40–78)

96.8 (0.9–322) 0.8 (0.2–3)

1.8 (bMQL–7.8) 2.0 (bMQL–26.7) bMQL (ND–11.0) bMQL (ND–17.1) bIQL (ND–4.3)

–

(b1–32.2)

0.5 (ND–13.1)

0.3 (b0.02–3.45)

6.8 (1.1–46.0)

–

–

(b1–5.31)

bIQL (ND–3.5)

0.3 (b0.02–12.7)

3.3 (1.1–35.0)

–

–

(b1–34.7)

1.2 (b0.39–2.47)

0.4 (0.17–9.8)

7.6 (3.8–38.7)

–

–

0.2 (b0.09–3.62)

0.6 (0.21–142)

1.4 (0.8–3.1)

–

–

–

0.2 (0.10–2.1)

–

–

–

2.8 (b0.68–9.67)

3.1 (1.2–94.0)

9.1 (4.4–23.7)

0.5 (b0.1–211)

1.5 (b0.5–3.38)

0.01 (b0.01–0.44)

1.1 (0.15–42)

1.5 (1–13.7)

(b0.5–222)

0.8 (ND–6.6)

FOSA N-MeFOSA N-EtFOSA N-MeFOSE

7.2 (2.77–81) 5.3 (2.12–7.16) na na (b0.01–0.43) (b0.01–0.02) – – – – – – – –

bMQL (ND–14.8) bMQL (ND–14.4) 2.0 (bMQL–9.3) bMQL (ND–4.6) 10.3 (4.8–118)

– – – –

0.5 (0.22–41) 0.5 (0.27–1.1) 0.6 (0.26–33) 9 (3.5–90.0)

– – – –

– – – –

– – – –

N-EtFOSE References

– – – Eriksson and Kärrman (2015)

– Haug et al. (2011)

– Huber et al. (2011)

– D'Hollander et al. (2010)

– Lankova et al. (2015)

– – – bIQL (ND–16.0) bIQL (ND–6.4) Present study

PFHxA PFHpA PFOA PFNA PFDA PFUnDA PFDoDA PFTrDA PFTeDA PFBS PFHxS PFHpS PFOS PFDS

P. Karásková et al. / Environment International 94 (2016) 315–324

321

Table 3 Summary of PFASs median and ranges (ng/g) in North America dust samples reported in the present study and published literature. Location

Canada, Ottawa

Canada, Vancouver

Canada, Ottawa

Canada, Toronto

USA, Ohio, North Carolina

USA, Boston, MA

USA, Wisconsin

USA, Bloomington

Type of room (N) Year of sampling PFPA

Homes (67)

Homes (132)

Houses (10)

Houses (20)

Homes (30)

2007–2008

2013–2014

2013

2009

Residential dust (39) 2008

Houses (20)

2002/2003

Homes (102), daycare centers (10) 2000–2001

–

–

1.62 (b0.36–36.3)

–

(5.39–249)

5.4 (ND–32)

PFHxA

–

n.a.

7.44 (2.67–97.1)

54.2 (1250)

–

69 (BDL-1561)

4.75 (b1.23–53.3)

PFOA

19.7 (1.15–1234) 30 (1.9–1390)

21 (b7.01–310)

PFNA

–

(BDL-680)

13.5 (2.15–419)

PFDA

–

(BDL-251)

7.07 (2.32–20.2)

6.65 (267)

8.65a (4.85–1380) 12a (4.93–586) 23.7a (5.71–894) 10.9a (6.21–1420) (6.97–26.8)

0 (ND–180)

PFHpA

5.7 (ND–60)

PFUnDA

–

(BDL-370)

7.83 (1.01–140)

7.57 (588)

(10.8–39.4)

3.1 (ND–48)

PFDoDA

–

(BDL-301)

4.19 (1.13–11.6)

7.78 (520)

(5.09–13.3)

5 (ND–41)

PFTrDA

–

–

1.12 (b0.02–7.02)

–

(10.3–10.3)

2.1 (ND–11)

PFTeDA

–

(BDL-478)

0.72 (b0.02–8.81)

–

(11.2–11.2)

3.7 (ND–24)

PFBS

–

–

3.44 (b0.39–10.9)

2.83 (bMQL-17.0) 5.61 (1.7–146) 3.91 (bMQL–47.4) 8.18 (2.1–92.7) 4.54 (bMQL–195) 2.40 (0.9–86.2) 1.75 (ND–49.6) 1.22 (ND–61.1) 6.10 (ND–19.4) 1.65 (bMQL–33.6) 1.09 (ND–5.8)

9.11 (1150)

(4.98–4.98)

1.8 (ND–31)

PFHxS

23.1 (2.28–4305) –

3.75 (0.26–472)

45.5 (35,700)

(6.05–430)

16 (2.1–1000)

PFHpS

–

–

–

–

9.9 (ND–37)

PFOS

37.8 (2.28–5065) 71 (1.5–4661)

7.29 (1.69–699)

201 (12,100)

–

–

0.42 (0.23–2.69)

–

26.9a (14.1–280) –

47 (8.7–1100)

PFDS

5.4 (ND–120)

FOSA N-MeFOSA N-EtFOSA N-MeFOSE

– – – –

– 1.5 (0.9–14) 0.14 (BDL-73) 38 (12–1676)

– – – –

– – – –

– – – (18–488)

– – – –

N-EtFOSE

–

–

–

(12.2–3280)

–

References

Kubwabo et al. (2005)

7.1 (BDL-1590) Shoeib et al. (2011)

2.11 (ND–11.5) 0.59 (ND–12.7) 9.10 (3.3–31.8) 3.15 (0.5–14.5) – 0.7 – 0.94 (ND–46.3) 0.44 (ND–0.8)

Strynar and Lindstrom (2008)

Fraser et al. (2013)

Knobeloch et al. (2012)

–

Eriksson and Kärrman (2015)

Present study

Fig. 3. Percentage of total PFASs in dust samples from Czech Republic, Canada and USA for eight major contaminants detected in ≥50% of samples.

50.2 (1150) 142 (1960) 7.99 (263)

17 (2.4–140) 44 (6.5–420) 12 (1.3–280)

2013 3.19 (ND–24.8) 6.52 (2.5–190) 3.99 (0.9–86.7) 9.29 (2.9–318) 3.88 (1.1–62.9) 1.85 (0.4–64) 2.23 (ND–13.1) 0.91 (ND–9.0) 1.69 (ND–2.1) 1.21 (bMQL–3.0) 1.25 (ND–2.6) 8.74 (1.4–84.4) 0.98 (ND–2.9) 14.12 (5.7–239) 2.6 (0.5–9.8) – 0.6 5.9 1.58 (ND–9.9) 2.04 (ND–93.9) Present study

transformed PFASs concentrations. Results of the most significant correlations are showed in Fig. S2. Statistically significant correlations (p b 0.05) were found between PFASs concentrations in dust samples when all data were considered (see Table S10 in the SI), especially for PFCAs. Long-chain PFCA concentrations generally correlated with one another (p b 0.05), as did short chain PFCA concentrations (p b 0.001). Similar relationships for these two groups of homologues were reported in a previous study from the US (Fraser et al., 2013). In particular, PFOA concentrations were highly correlated (p b 0.05) with those of both shorter-chain homologues (i.e. PFPA, PFHxA and PFHpA), and longer-chain homologues (i.e. PFNA, PFDA and PFDoDA) as is shown in Fig. S2, top and middle panels. The obtained results may be explained by common occurrence of PFOA with long-chain PFDA and PFDoDA in fluorotelomer-based products, where long-chain PFCAs are exclusively found as impurities. Moreover, a statistically significant correlation (p b 0.05) between PFOA and PFOS concentrations was found (see Fig. S2, bottom panel), consistent with previous studies on house dust (Björklund et al., 2009; D'Hollander et al., 2010; Fraser et al., 2013; Haug et al., 2011; Moriwaki et al., 2003; Shoeib et al., 2011). The significant correlation between PFOS and PFOA suggests that these PFASs may be from a common source in the different microenvironments or may originate from the same precursor compound in relation to the use of POSF-based products (Prevedouros et al., 2006).

322

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Fig. 4. Box and whiskers plots of log-transformed concentration (ng/g) showing the results of ANOVA analysis comparing PFAS concentrations in dust samples according to (a) floor type; (b) number of people living in the house; (c) age of building.

Additionally, PFOS was significantly correlated with PFHxA (p = 0.012) and other short-chain PFCAs, as reported previously (Fraser et al., 2013). PFOS concentrations were also correlated with those of PFHxS (p = 0.004, r = 0.384).

observed lower levels of PFASs in older houses in Canada (Kubwabo et al., 2005). This analysis doesn't include information about the presence of PFAS-containing products in the house (e.g. treated carpet) or recent renovations (furniture or carpet change), all factors that are expected to affect the levels of PFASs.

3.5. PFAS sources 3.6. Human exposure No significant differences in concentrations were found between the bedroom and living room in the same household. This finding is consistent with PFASs being added to a wide variety of consumer products. In an effort to identify potential sources of PFASs in the homes sampled, we evaluated the relationship between their concentrations in dust and various household information obtained from the questionnaires. Total PFASs in dust samples from all three countries were analyzed using ANOVA analysis of log-transformed concentrations. Input data were grouped into categories to reduce fragmentation and increase statistical power. For example, floor types were assigned to one of the following categories: carpet, wood, and others. Results of these comparisons are illustrated in Fig. 4 and Table S15. The parameters that were related to the PFAS levels in dust samples were: type of flooring, number of people living in the house and the age of house. Higher concentrations of total PFASs were found on carpet and wood than other types of floorings (e.g., linoleum) (Fig. 4a), although differences were only weakly significant (p = 0.102). Higher levels of PFASs in carpeted floors have been observed previously (Fraser et al., 2013; Gewurtz et al., 2009; Knobeloch et al., 2012; Kubwabo et al., 2005). PFASs are used in carpets and carpet treatment products to make them resistant to stains. It has been estimated that the levels of PFASs in carpets are almost 30 times higher than those in treated floor waxes and wood sealants (U.S.EPA, 2009). Also, carpets can trap dust much more effectively than smooth surface floors such as wood or laminate and thus absorb PFAS more readily than other surface. The number of residents in the house also was related to PFAS levels (Fig. 4b) (p = 0.012), with higher levels in homes with 1–2 residents. The presence of more people in a household has been found to be associated with more frequent cleaning (Moran et al., 2012), which may result in greater removal of the surface protective layer from consumer products treated with PFASs. In contrast to our study, Fraser et al. reported higher PFASs concentration in dust from homes with more than three residents (Fraser et al., 2013). A significant positive relationship between PFAS concentrations and the age of house (p = 0.026) was observed with three categories for building age: houses in group #1 were constructed before 1950 (before industrial production of PFASs); group #2 were constructed between 1951 and 2003 (during industrial production of PFASs); group #3 were constructed after 2003 (after PFOS production ended). The levels from homes built after 2003 were significantly higher than in older homes (p = 0.047 for b 1950 vs. N 2003, p = 0.008 for 1951–2003 vs. N2003) (Fig. 4c). This result is consistent with Kubwabo et al., who

Estimates of PFAS intake via dust ingestion for toddlers and adults for the different scenarios are summarized in Table S16. For toddlers, who ingest more dust due to hand-to-mouth activity, human exposure under the mean dust ingestion scenario and median dust concentration was between 0.21 ng/kg body weight/day (in CZ) and 0.48 ng/kg body weight/day (in USA). For adults, exposure via dust ingestion was about one order of magnitude lower, 0.002–0.004 ng/kg body weight/ day. This represents exposure for the majority of the population. In a high exposure scenario, with the 95th percentile dust concentration and high dust ingestion rates, the daily intake of ∑PFSAs and PFCAs via dust ingestion for children could reach up to 2.0 ng/kg body weight/day in CZ and 7.9 ng/kg body weight/day in USA. For adults, the same high exposure scenario gives an estimated daily intake of 0.12–0.47 ng/kg body weight/day. While calculated intakes were below the tolerable daily intake values of both main compounds – PFOA (1500 ng/kg bw/day) and PFOS (150 ng/kg bw/day) (EFSA, 2008) dust is only one of the contributors in the total exposure to these compounds. Non- food sources might contribute to about 50% of the overall dietary exposure for PFOA and ~3% for PFOS (EFSA, 2008). It should be noted that the contributions of precursors that could be transformed into PFOA or PFOS are not taken into account in these calculations, despite the fact that they could play an important role. Also, the geographic differences in concentrations and the large range of observed indoor concentrations could possible significantly influence the overall human exposure to PFASs. 4. Conclusion This is the first study to compare directly PFAS concentrations in North America and Europe in homes. In general, concentrations in North America were higher than in Czech Republic, which is consistent with usage patterns. Also, based on a literature comparison, dust concentration of PFOS and PFOA seem to be decreasing with time, suggesting that policy controls and regulations are effective. Conversely, concentrations of replacements products are increasing, although literature data from previous years are quite limited. No differences were found between living rooms and bedrooms suggesting either similar usage patterns or that these compounds are well-mixed indoors. Significant relationships were found between PFASs concentrations in dust and type of floors, number of people living in the house and age of the house. The relatively large range of dust ingestion between different

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countries warrants caution when considering the overall exposure to these compounds. Acknowledgments This project was supported by the Czech-American Scientific Cooperation Program (AMVIS/KONTAKT II, LH12074), National Sustainability Programme of the Czech Ministry of Education, Youth and Sports (LO1214) and the RECETOX Research Infrastructure (LM2015051). We also thank all volunteers from Brno, Czech Republic, Toronto, Canada and Bloomington, IN, USA. We thank Ron Hites for helpful discussions and comments and Kevin Romanak for his contribution to the sampling campaign. Appendix A. Supplementary data Supplementary data to this article can be found online at http://dx. doi.org/10.1016/j.envint.2016.05.031. References Abbasi, G., Buser, A.M., Soehl, A., Murray, M.W., Diamond, M.L., 2015. Stocks and flows of pbdes in products from use to waste in the U.S. and Canada from 1970 to 2020. Environmental Science & Technology 49, 1521–1528. Antweiler, R.C., 2015. Evaluation of statistical treatments of left-censored environmental data using coincident uncensored data sets. II. Group comparisons. Environmental Science & Technology 49, 13439–13446. Beesoon, S., Webster, G.M., Shoeib, M., Harner, T., Benskin, J.P., Martin, J.W., 2011. Isomer profiles of perfluorochemicals in matched maternal, cord, and house dust samples: manufacturing sources and transplacental transfer. Environmental Health Perspectives 119, 1659–1664. Björklund, J.A., Thuresson, K., de Wit, C.A., 2009. Perfluoroalkyl compounds (PFCs) in indoor dust: concentrations, human exposure estimates, and sources. Environmental Science & Technology 43, 2276–2281. Buck, R.C., Franklin, J., Berger, U., Conder, J.M., Cousins, I.T., de Voogt, P., Jensen, A.A., Kannan, K., Mabury, S.A., van Leeuwen, S.P.J., 2011. Perfluoroalkyl and polyfluoroalkyl substances in the environment: terminology, classification, and origins. Integrated Environmental Assessment and Management 7, 513–541. Calafat, A.M., Needham, L.L., Kuklenyik, Z., Reidy, J.A., Tully, J.S., Aguilar-Villalobos, M., Naeher, L.P., 2006. Perfluorinated chemicals in selected residents of the American continent. Chemosphere 63, 490–496. Dallaire, R., Dewailly, É., Pereg, D., Dery, S., Ayotte, P., 2009. Thyroid function and plasma concentrations of polyhalogenated compounds in Inuit adults. Environmental Health Perspectives. 117, 1380–1386. De Silva, A.O., Mabury, S.A., 2006. Isomer distribution of perfluorocarboxylates in human blood: potential correlation to source. Environmental Science & Technology. 40, 2903–2909. D'Hollander, W., Roosens, L., Covaci, A., Cornelis, C., Reynders, H., Campenhout, K.V., Voogt, P.D., Bervoets, L., 2010. Brominated flame retardants and perfluorinated compounds in indoor dust from homes and offices in Flanders, Belgium. Chemosphere 81, 478–487. Egeghy, P.P., Lorber, M., 2011. An assessment of the exposure of Americans to perfluorooctane sulfonate: a comparison of estimated intake with values inferred from Nhanes data. J Expos Sci Environ Epidemiol 21, 150–168. Emmett, E.A., Shofer, F.S., Zhang, H., Freeman, D., Desai, C., Shaw, L.M., 2006. Community exposure to perfluorooctanoate: relationships between serum concentrations and exposure sources. Journal of Occupational and Environmental Medicine 48, 759–770. Environment Canada, 2012. http://www.ec.gc.ca/ese-ees/451C95ED-6236-430C-BE5A22F91B36773F/PFOA%20%26%20PFCAs_RMA_EN.pdf. Environment Canada, 2015. Perfluorooctane Sulfonate (Pfos), its Salts and its Precursors. EPA, E.C.P.S.P.a.P.A.P. Epa, 2014. Emerging Contaminants – Perfluorooctane Sulfonate (PFOS) and Perfluorooctanoic Acid (PFOA). Ericson Jogsten, I., Nadal, M., van Bavel, B., Lindström, G., Domingo, J.L., 2012. Per- and polyfluorinated compounds (PFCs) in house dust and indoor air in Catalonia. Spain: Implications for Human Exposure. Environment International. 39, 172–180. Eriksson, U., Kärrman, A., 2015. World-wide indoor exposure to polyfluoroalkyl phosphate esters (Paps) and other PFASs in household dust. Environmental Science & Technology. European Food Safety Authority (EFSA), 2008. Perfluorooctane sulfonate (PFOS), perfluorooctanoic acid (PFOA) and their salts Scientific Opinion of the Panel on Contaminants in the Food chain, EFSA-Q-2004-163. http://www.efsa.europa.eu/en/ efsajournal/pub/653. Fraser, A.J., Webster, T.F., Watkins, D.J., Nelson, J.W., Stapleton, H.M., Calafat, A.M., Kato, K., Shoeib, M., Vieira, V.M., McClean, M.D., 2012. Polyfluorinated compounds in serum linked to indoor air in office environments. Environmental Science & Technology. 46, 1209–1215. Fraser, A.J., Webster, T.F., Watkins, D.J., Strynar, M.J., Kato, K., Calafat, A.M., Vieira, V.M., McClean, M.D., 2013. Polyfluorinated compounds in dust from homes, offices, and vehicles as predictors of concentrations in office workers' serum. Environment International. 60, 128–136.

323

Gewurtz, S.B., Bhavsar, S.P., Crozier, P.W., Diamond, M.L., Helm, P.A., Marvin, C.H., Reiner, E.J., 2009. Perfluoroalkyl contaminants in window film: indoor/outdoor, urban/rural, and winter/summer contamination and assessment of carpet as a possible source. Environmental Science & Technology. 43, 7317–7323. Giesy, J.P., Kannan, K., 2001. Global distribution of perfluorooctane sulfonate in wildlife. Environmental Science & Technology. 35, 1339–1342. Gomez, C., Vicente, J., Echavarri-Erasun, B., Porte, C., Lacorte, S., 2011. Occurrence of perfluorinated compounds in water, sediment and mussels from the Cantabrian Sea (North Spain). Marine Pollution Bulletin. 62, 948–955. Haug, L.S., Huber, S., Schlabach, M., Becher, G., Thomsen, C., 2011. Investigation on perand polyfluorinated compounds in paired samples of house dust and indoor air from Norwegian homes. Environmental Science & Technology. 45, 7991–7998. Herzke, D., Olsson, E., Posner, S., 2012. Perfluoroalkyl and polyfluoroalkyl substances (PFASs) in consumer products in Norway - a pilot study. Chemosphere 88, 980–987. Hölzer, J., Midasch, O., Rauchfuss, K., Kraft, M., Reupert, R., Angerer, J., Kleeschulte, P., Marschall, N., Wilhelm, M., 2008. Biomonitoring of perfluorinated compounds in children and adults exposed to perfluorooctanoate-contaminated drinking water. Environmental Health Perspectives. 116, 651–657. Huber, S., Haug, L.S., Schlabach, M., 2011. Per- and polyfluorinated compounds in house dust and indoor air from northern Norway – a pilot study. Chemosphere 84, 1686–1693. Joensen, U.N., Bossi, R., Leffers, H., Jensen, A.A., Skakkebæk, N.E., Jørgensen, N., 2009. Do perfluoroalkyl compounds impair human semen quality? Environmental Health Perspectives 117, 923–927. Kannan, K., Franson, J.C., Bowerman, W.W., Hansen, K.J., Jones, P.D., Giesy, J.P., 2001a. Perfluorooctane sulfonate in fish-eating water birds including bald eagles and albatrosses. Environmental Science & Technology 35, 3065–3070. Kannan, K., Koistinen, J., Beckmen, K., Evans, T., Gorzelany, J.F., Hansen, K.J., Jones, P.D., Helle, E., Nyman, M., Giesy, J.P., 2001b. Accumulation of perfluorooctane sulfonate in marine mammals. Environmental Science & Technology 35, 1593–1598. Kannan, K., Corsolini, S., Falandysz, J., Oehme, G., Focardi, S., Giesy, J.P., 2002. Perfluorooctanesulfonate and related fluorinated hydrocarbons in marine mammals, fishes, and birds from coasts of the Baltic and the Mediterranean seas. Environmental Science & Technology 36, 3210–3216. Kärrman, A., Mueller, J.F., van Bavel, B., Harden, F., Toms, L.-M.L., Lindstrom, G., 2006a. Levels of 12 perfluorinated chemicals in pooled australian serum, collected 2002– 2003, in relation to age, gender, and region. Environmental Science & Technology 40, 3742–3748. Kärrman, A., van Bavel, B., Järnberg, U., Hardell, L., Lindström, G., 2006b. Perfluorinated chemicals in Relation to other persistent organic pollutants in human blood. Chemosphere 64, 1582–1591. Kärrman, A., Harada, K.H., Inoue, K., Takasuga, T., Ohi, E., Koizumi, A., 2009. Relationship between dietary exposure and serum Perfluorochemical (PFC) levels—a case study. Environment International 35, 712–717. Kissa, E., 2001. Fluorinated Surfactants and Repellents. Marcel Dekker Inc., New York. Klepeis, N.E., Nelson, W.C., Ott, W.R., Robinson, J.P., Tsang, A.M., Switzer, P., Behar, J.V., Hern, S.C., Engelmann, W.H., 2001. The National Human Activity Pattern Survey (NHAPS): a resource for assessing exposure to environmental pollutants. J. Expo. Anal. Environ. Epidemiol. 11, 231–252. Knobeloch, L., Imm, P., Anderson, H., 2012. Perfluoroalkyl chemicals in vacuum cleaner dust from 39 Wisconsin homes. Chemosphere 88, 779–783. Kotthoff, M., Müller, J., Jürling, H., Schlummer, M., Fiedler, D., 2015. Perfluoroalkyl and polyfluoroalkyl substances in consumer products. Environmental Science and Pollution Research 22, 14546–14559. Kubwabo, C., Stewart, B., Zhu, J.P., Marro, L., 2005. Occurrence of perfluorosulfonates and other perfluorochemicals in dust from selected homes in the City of Ottawa, Canada. Journal of Environmental Monitoring 7, 1074–1078. Langer, V., Dreyer, A., Ebinghaus, R., 2010. Polyfluorinated compounds in residential and nonresidential indoor air. Environmental Science & Technology 44, 8075–8081. Lankova, D., Svarcova, A., Kalachova, K., Lacina, O., Pulkrabova, J., Hajslova, J., 2015. Multianalyte method for the analysis of various organohalogen compounds in house dust. Analytica Chimica Acta 854, 61–69. Lau, C., Anitole, K., Hodes, C., Lai, D., Pfahles-Hutchens, A., Seed, J., 2007. Perfluoroalkyl acids: a review of monitoring and toxicological findings. Toxicologal Sciences 99, 366–394. Lehmler, H.-J., 2005. Synthesis of environmentally relevant fluorinated surfactants—a review. Chemosphere 58, 1471–1496. Lim, T.C., Wang, B., Huang, J., Deng, S., Yu, G., 2011. Emission inventory for Pfos in China: review of past methodologies and suggestions. TheScientificWorldJOURNAL 11, 1963–1980. Liu, W., Chen, S., Harada, K.H., Koizumi, A., 2011. Analysis of perfluoroalkyl carboxylates in vacuum cleaner dust samples in Japan. Chemosphere 85, 1734–1741. Liu, X., Guo, Z., Krebs, K.A., Pope, R.H., Roache, N.F., 2014. Concentrations and trends of perfluorinated chemicals in potential indoor sources from 2007 through 2011 in the US. Chemosphere 98, 51–57. Midasch, O., Drexler, H., Hart, N., Beckmann, M.W., Angerer, J., 2007. Transplacental exposure of neonates to perfluorooctanesulfonate and perfluorooctanoate: a pilot study. International Archives of Occupational and Environmental Health. 80, 643–648. Moran, R.E., Bennett, D.H., Tancredi, D.J., Wu, X., Ritz, B., Hertz-Picciotto, I., 2012. Frequency and longitudinal trends of household care product use. Atmospheric Environment 55, 417–424. Moriwaki, H., Takata, Y., Arakawa, R., 2003. Concentrations of perfluorooctane sulfonate (Pfos) and perfluorooctanoic acid (Pfoa) in vacuum cleaner dust collected in Japanese homes. Journal of Environmental Monitoring 5, 753–757. Nelson, J.W., Hatch, E.E., Webster, T.F., 2010. Exposure to polyfluoroalkyl chemicals and cholesterol, body weight, and insulin resistance in the general U.S. population. Environmental Health Perspectives 118, 197–202.

324

P. Karásková et al. / Environment International 94 (2016) 315–324

Olsen, G.W., Burris, J.M., Burlew, M.M.M.S., Mandel, J.H., 2003. Epidemiologic assessment of worker serum perfluorooctanesulfonate (Pfos) and perfluorooctanoate (Pfoa) concentrations and medical surveillance examinations. Journal of Occupational & Environmental Medicine 45, 260–270. Olsen, G.W.; Burris, J.M.; Ehresman, D.J.; Froehlich, J.W.; Seacat, A.M.; Butenhoff, J.L.; Zobel, L.R. Half-life of serum elimination of perfluorooctanesulfonate, perfluorohexanesulfonate, and perfluorooctanoate in retired fluorochemical production workers. Environmental Health Perspectives 115:1298–1305; 2007 Paul, A.G., Jones, K.C., Sweetman, A.J., 2009. A first global production, emission, and environmental inventory for perfluorooctane sulfonate. Environmental Science & Technology. 43, 386–392. Pollutants, S.C.o.P.O., 2012. Guidance for the Inventory of Perfluorooctane Sulfonic Acid (Pfos) and Related Chemicals Listed under the Stockholm Convention on Persistent Organic Pollutants. Prevedouros, K., Cousins, I.T., Buck, R.C., Korzeniowski, S.H., 2006. Sources, fate and transport of perfluorocarboxylates. Environmental Science & Technology 40, 32–44. Renner, R., 2004. Perfluorinated sources outside and inside. Environmental Science & Technology 38, 80A. Ritter, S.K., 2010. Fluorochemicals go short. Chemical & Engineering News Archive 88, 12–17. Scheringer, M., Trier, X., Cousins, I.T., de Voogt, P., Fletcher, T., Wang, Z., Webster, T.F., 2014. Helsingør statement on poly- and perfluorinated alkyl substances (PFASs). Chemosphere 114, 337–339. Shoeib, M., Harner, T., M. Webster, G., Lee, S.C., 2011. Indoor sources of poly- and perfluorinated compounds (PFCs) in Vancouver, Canada: implications for human exposure. Environmental Science & Technology 45, 7999–8005. Shoeib, T., Hassan, Y., Rauert, C., Harner, T., 2016. Poly- and perfluoroalkyl substances (PFASs) in indoor dust and food packaging materials in Egypt: trends in developed and developing countries. Chemosphere 144, 1573–1581. Shoeib, M., Vlahos, P., Harner, T., Peters, A., Graustein, M., Narayan, J., 2010. Survey of polyfluorinated chemicals (PFCs) in the atmosphere over the Northeast Atlantic Ocean. Atmospheric Environment 44, 2887–2893. Strynar, M.J., Lindstrom, A.B., 2008. Perfluorinated compounds in house dust from Ohio and North Carolina. USA. Environmental Science & Technology. 42, 3751–3756. Sun, H., Li, F., Zhang, T., Zhang, X., He, N., Song, Q., Zhao, L., Sun, L., Sun, T., 2011. Perfluorinated compounds in surface waters and Wwtps in Shenyang. China: Mass Flows and Source Analysis. Water Research. 45, 4483–4490. Tao, L., Ma, J., Kunisue, T., Libelo, E.L., Tanabe, S., Kannan, K., 2008. Perfluorinated compounds in human breast milk from several Asian countries, and in infant formula and dairy milk from the United States. Environmental Science & Technology. 42, 8597–8602. Trudel, D., Horowitz, L., Wormuth, M., Scheringer, M., Cousins, I.T., Hungerbühler, K., 2008. Estimating consumer exposure to Pfos and Pfoa. Risk Analysis. 28, 251–269.

U.S. EPA, 2012. Perfluoroalkyl Sulfonated and Long-chain Perfluoroalkyl Carboxylate Chemical Substances; Proposed Significant New Use Rule. U.S.EPA, 1997. Exposure Factors Handbook (1997 Final Report). Agency U.S.E.P., Washington DC. U.S.EPA, 2006a. 2010/15 Pfoa Stewardship Program. Agency U.S.E.P., Washington, DC. U.S.EPA, 2006b. In: Register, F. (Ed.), PFAS-Proposed Significant New Use Rule. U.S.EPA, 2006c. Sab Review of EPA's Draft Risk Assessment of Potential Human Health Effects Associated with Pfoa and Its Salts Washington, DC . U.S.EPA, 2009. Perfluorocarboxylic Acid Content in 116 Articles of Commerce. Development O.o.R.a., Research Triangle Park, Durhem, NC. U.S.EPA, 2011. Exposure Factors Handbook. Agency U.S.E.P., Washington, DC. Venier, M., Audy, O., Vojta, S., Bečanová, J., Romanak, K., Melymuk, L., Krátká, M., Kukučka, P., Okeme, J., Saini, A., Diamond, M.L., Klanova, J., 2016. Brominated flame retardants in indoor environment - comparative study of indoor contamination from three countries. Environment International 94, 150–160. Vestergren, R., Cousins, I.T., 2009. Tracking the pathways of human exposure to Perfluorocarboxylates. Environmental Science & Technology 43, 5565–5575. Vestergren, R., Herzke, D., Wang, T., Cousins, I.T., 2015. Are imported consumer products an important diffuse source of PFASs to the Norwegian environment? Environmental Pollution 198, 223–230. Wang, Z., Cousins, I.T., Scheringer, M., Buck, R.C., Hungerbühler, 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, T., Lu, Y., Chen, C., Naile, J.E., Khim, J.S., Park, J., Luo, W., Jiao, W., Hu, W., Giesy, J.P., 2011. Perfluorinated compounds in estuarine and coastal areas of north Bohai Sea, China. Marine Pollution Bulletin 62, 1905–1914. Washburn, S.T., Bingman, T.S., Braithwaite, S.K., Buck, R.C., Buxton, L.W., Clewell, H.J., Haroun, L.A., Kester, J.E., Rickard, R.W., Shipp, A.M., 2005. Exposure assessment and risk characterization for perfluorooctanoate in selected consumer articles. Environmental Science & Technology 39, 3904–3910. Wong, F., MacLeod, M., Mueller, J.F., Cousins, I.T., 2014. Enhanced elimination of perfluorooctane sulfonic acid by menstruating women: evidence from populationbased pharmacokinetic modeling. Environmental Science & Technology 48, 8807–8814. Xu, Z., Fiedler, S., Pfister, G., Henkelmann, B., Mosch, C., Völkel, W., Fromme, H., Schramm, K.-W., 2013. Human exposure to fluorotelomer alcohols, perfluorooctane sulfonate and perfluorooctanoate via house dust in Bavaria, Germany. Science of The Total Environment. 443, 485–490. Yang, L., Zhu, L., Liu, Z., 2011. Occurrence and partition of perfluorinated compounds in water and sediment from Liao River and Taihu Lake. China. Chemosphere. 83, 806–814.