Characterization of Two Passive Air Samplers for Per-and

Article pubs.acs.org/est

Characterization of Two Passive Air Samplers for Per- and Polyfluoroalkyl Substances Lutz Ahrens,*,†,‡ Tom Harner,*,† Mahiba Shoeib,† Martina Koblizkova,† and Eric J. Reiner§,⊥ †

Environment Canada, Air Quality Processes Research Section, Toronto, Ontario M3H 5T4, Canada Swedish University of Agricultural Sciences (SLU), Department of Aquatic Sciences and Assessment, Uppsala, Uppland SE-750 07, Sweden § Ontario Ministry of the Environment, 125 Resources Road, Toronto, Ontario M9P 3V6, Canada ⊥ University of Toronto, Department of Chemistry, 80 St. George Street, Toronto, Ontario M5S 3H6, Canada ‡

S Supporting Information *

ABSTRACT: Two passive air sampler (PAS) media were characterized under field conditions for the measurement of per- and polyfluoroalkyl substances (PFASs) in the atmosphere. The PASs, consisting of polyurethane foam (PUF) and sorbent-impregnated PUF (SIP) disks, were deployed for over one year in parallel with high volume active air samplers (HVAAS) and low volume active air samplers (LV-AAS). Samples were analyzed for perfluoroalkyl carboxylic acids (PFCAs), perfluoroalkane sulfonic acids (PFSAs), fluorotelomer alcohols (FTOHs), fluorotelomer methacrylates (FTMACs), fluorotelomer acrylates (FTACs), perfluorooctane sulfonamides (FOSAs), and perfluorooctane sulfonamidoethanols (FOSEs). Sampling rates and the passive sampler medium (PSM)-air partition coefficient (KPSM−A) were calculated for individual PFASs. Sampling rates were similar for PFASs present in the gas phase and particle phase, and the linear sampling rate of 4 m−3 d−1 is recommended for calculating effective air sample volumes in the SIP-PAS and PUF-PAS for PFASs except for the FOSAs and FOSEs in the PUF-PAS. SIP disks showed very good performance for all tested PFASs while PUF disks were suitable only for the PFSAs and their precursors. Experiments evaluating the suitability of different isotopically labeled fluorinated depuration compounds (DCs) revealed that 13C8-perfluorooctanoic acid (PFOA) was suitable for the calculation of site-specific sampling rates. Ambient temperature was the dominant factor influencing the seasonal trend of PFASs.

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INTRODUCTION

High volume active air samplers (HV-AAS) are typically used for measuring PFASs in the atmosphere because of their ability to provide information on the gas- and particle-phase distribution and high temporal resolution. However, HV-AAS depend on power supplies, and sampling artifacts have been reported for PFSAs and PFCAs using conventional HVAAS.11,12 In contrast, passive air samplers (PAS) generate timeintegrated data and are ideal due to their simplicity and low cost, especially for the purpose of spatial and long-term temporal trend studies.13,14 Polyurethane foam (PUF) disks are the most widely used PAS for persistent organic pollutants (POPs).13,15 A new PAS type was developed by Shoeib et al. comprising sorbent-impregnated PUF (SIP) disks to increase the sorptive capacity for more volatile chemicals like FTOHs.16 In general, the uptake of the chemical depends on its diffusivity in air and the passive sampler medium (PSM)-air partition coefficient (KPSM−A), which depends on the PSM and characteristics of the chemical.17 In addition, the chamber

Per- and polyfluoroalkyl substances (PFASs) have received increasing public attention due to their persistence, bioaccumulative potential, and possible adverse effects on humans and wildlife.1 PFASs comprise a diverse group of chemicals including, for example, fluorotelomer alcohols (FTOHs), fluorotelomer acrylates (FTACs), perfluorooctane sulfonamides (FOSAs), perfluorooctane sulfonamidoethanols (FOSEs), perfluoroalkyl carboxylic acids (PFCAs), and perfluoroalkyl sulfonic acids (PFSAs). They have been widely used in a variety of consumer and industrial products such as metal plating, semiconductors, polishing agents, paints, surfactants in textile coatings, paper treatments, and firefighting foams.2,3 Once released into the environment, PFASs can be globally transported by ocean currents and the atmosphere.4,5 However, few data are available for atmospheric PFASs, in particular for the PFSAs and PFCAs due to their unique characteristics (e.g., ionizability) and low concentration levels.6−10 Thus, there is a need for a simple sampling technique to improve our understanding of the temporal trends and spatial distribution of PFASs in a global context. © 2013 American Chemical Society

Received: May 22, 2013 Accepted: November 12, 2013 Published: November 12, 2013 14024

dx.doi.org/10.1021/es4048945 | Environ. Sci. Technol. 2013, 47, 14024−14033

Environmental Science & Technology

Article

area 365 cm2 , mass 4.40 g, volume 207 cm 3, Tisch Environmental, Cleves, OH, USA) were impregnated with finely ground XAD-4 resin (Supelco, Bellefonte, PA) (∼0.5 g per PUF disk) (for details, see elsewhere16). During field deployment, the SIP and PUF disks were individually housed inside precleaned stainless steel chambers (“original chamber”, Model TE-200-PAS, Tisch Environmental) and deployed ∼2 m above the ground. To compare different chamber configurations, four different chambers were used with different gaps between the two stainless steel dome housings (i.e., “original chamber” with 1 cm overlap, “flush chamber”, “1 cm gap chamber”, “2 cm gap chamber”) (see Figure S1 in the SI). The PUF and SIP disks were deployed in the different chambers for 28 days over 5 sampling periods (Table S5 and S6 in the SI). The comparison of the different chamber configurations will provide information about the influence of the chamber design on the collection of particles on the PUF and SIP disks and the potential wind speed effects that we expect to be dampened by the more protective (less open) chamber configurations. The PUF-PAS and SIP-PAS (including blanks) were spiked with 20 ng absolute of the DCs 13C8−PFOS, 13C8−PFOA, and 7:2 sFTOH prior to field deployment. The loss of the DCs (i.e., volatilization from the PAS) during the deployment period of the PAS provides information of the site-specific sampling rates that account for the wind and temperature effects.18,20 The HVAAS were not spiked with DCs. DCs were not detected in the HV-AAS samples indicating that no migration of the DCs occurred to other samplers via air transport. For active sampling, high volume air samples (∼330 m3 over 24 h periods, one to two times a week) were collected from March, 2010 to April, 2011 using glass-fiber filters (GFFs) (Type A/E Glass, 102 mm diameter, Pall Corporation) for collecting the particle phase (n = 70) followed by a PUF/ XAD−2 cartridge for trapping the gas-phase compounds (n = 70). In addition, low volume air samples (∼46 m3) were collected from March to October, 2010 using LV-AAS (n = 14) to provide time-integrated concentrations (integrated over 14 days) (for details, see SI). Field blanks for PUFs, SIPs, GFFs, and PUF/XAD−2 cartridges were collected by exposing them for 1 min at the sampling site and then treating them like real samples. To check the efficiency of the collection of PFASs in the gas phase, breakthrough experiments were conducted by operating the active air samplers in series (i.e., two sets of collection PUF/ XAD−2 media) and analyzing the first and second set separately. These tests were conducted for both HV-AAS (n = 3) and LV-AAS (n = 3). All samples were stored at −20 °C until extraction within four weeks. Details of the sampling, dates, air volume, and meteorological data are presented in Tables S3−6 in the SI. Sample Extraction and Instrumental Analysis. The extraction and instrumental analysis is based on the methods described elsewhere.12 Prior to extraction, the PUF/XAD-2 sandwiches, GFFs, SIPs, and PUFs were spiked with 25 ng (mass-labeled FTOHs, FOSAs and FOSEs) and 5 ng (masslabeled PFSAs and PFCAs) (absolute amount) of an IS mixture containing 16 mass-labeled PFASs (Table S2 in the SI). The high volume PUF/XAD-2 sandwiches were Soxhlet extracted with petroleum ether/acetone (85/15, v/v) for ∼6 h, followed by a ∼16 h extraction with methanol. The GFFs were extracted by sonication, three times with dichloromethane and then three times with methanol. The low volume PUF/XAD-2

design has an influence on the amount of particles sampled by the PAS.18 Ultimately, the sampling rate of the chemical is also influenced by meteorological conditions like wind speed and temperature.17−19 To compensate varying meteorological conditions, depuration compounds (DCs) can be used to calculate the site-specific sampling rate by assessing their loss during the deployment period.20 However, it has to be considered that uptake rates of the chemical of interest are not necessarily equal to the loss of the DC due to inhomogeneities in diffusivities within the PSM.21 Furthermore, there is some uncertainty regarding the ability of PAS to capture PFASs (in particular, PFSAs and PFCAs) in air and how to derive air concentrations for PFASs with a high particle associated fraction (e.g., FOSEs, longer chained PFSAs and PFCAs).22 The addition of perfluorooctane sulfonic acid (PFOS) (and its salts and precursors) to the Stockholm Convention on POPs in 2009 means that air monitoring networks reporting to the global monitoring plan (GMP) will be required to measure these compounds in air.23 The results of this study are very relevant therefore and provide guidance on the use of PUF-disks or SIP-disks for monitoring PFOS and precursors (i.e., FOSE and FOSAs) in air. The specific objectives of this study include (i) to assess the sampling rates and KPSM−A values for PFASs for PUF and SIP disks based on field calibration against HV-AAS and low-volume active air samplers (LV-AAS), (ii) to evaluate the suitability of three fluorinated DCs, and (iii) to assess the comparability of four different air sampling techniques for measuring PFASs.

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EXPERIMENTAL SECTION Chemicals. The target analytes included C4, C6, C8, C10 (PFBS, PFHxS, PFOS, PFDS) PFSAs (CnF2n+1SO3H), C4−C12, C14 (PFBA, PFPeA, PFHxA, PFHpA, perfluorooctanoic acid (PFOA), PFNA, PFDA, PFUnDA, PFDoDA, PFTeDA) PFCAs (C n F 2 n + 1 COOH), 6:2, 8:2, 10:2 FTOHs (C n F 2n+1 CH 2 CH 2 OH), 6:2 fluorotelomer methacrylate (FTMAC, C6F13CH2CH2OC(O)C(CH3)=CH2), 8:2, 10:2 FTACs (C n F 2 n + 1 CH 2 CH 2 OC(O)CHCH 2 ), FOSA (C8F17SO2NH2), methyl and ethyl FOSAs (C8F17SO2N(CnH2n+1)H), and methyl and ethyl FOSEs (C8F17SO2N(CnH2n+1)CH2CH2OH). In addition, 17 mass-labeled internal standards (IS), three injection standards (InjS) (i.e., N,Ndimethyl perfluorooctane sulfonamide (Me 2 FOSA, C8F17SO2N(CH3)(CH3)), 13C8−PFOS, and 13C8−PFOA), and three DCs (i.e., perfluoroheptylethanol (7:2 sFTOH, C7F15CH(OH)CH3), 13C8−PFOS, and 13C8−PFOA) were used. Details are provided in Tables S1 and S2 of the Supporting Information (SI). Sampling. The calibration of the PAS was conducted from March 30 to October 13, 2010 at a semiurban meteorological station in Toronto (Environment Canada field site, 43°46′ N, 79°28′ W). Two different PAS media (i.e., PUF and SIP disks) were evaluated against parallel samples collected using LV-AAS and HV-AAS. After the completion of the calibration component (October 2010), sampling for determining ambient air concentrations of PFASs continued until the end of April 2011, using both PAS media and HV-AAS.21 Under the calibration study component, the PUF-PAS and SIP-PAS were deployed for 7, 21, 28, 42, 56, 84, 112, 140, 168, and 197 days. Duplicate PUF-PAS were collected on days 28, 84, and 197 to verify reproducibility. SIP-PAS were prepared according to the protocol from Shoeib et al. 16 Briefly, precleaned PUF disks (14 cm diameter ×1.35 cm thick; surface 14025



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Figure 1. Uptake profiles of PFASs for (A) PUF-PAS and (B) SIP-PAS.

blanks + 3× standard deviation (σ)) are given in Tables S9 and S10 in the SI. Average recoveries were 78%, 96%, 67%, 81%, and 93% for the LV-AAS, SIP-PAS, PUF-PAS, and gas phase and particle using HV-AAS, respectively (Table S11 in the SI). Breakthrough experiments were conducted to check the efficiency of the PUF/XAD sandwich for trapping the gasphase compounds using HV-AAS (n = 3, air volume ∼330 m3) and LV-AAS (n = 3, air volume ∼46 m3) (for details, see Figure S3 in the SI). Theory of PAS. The uptake of chemicals by PUF disks and other PSMs has been shown to be controlled by the air-side mass-transfer coefficient (kA), when the sampling medium has a high sorption capacity for the chemical.17 However, the uptake of chemicals may be also influenced by sampler-side resistance within the PSM for some compounds.24 The uptake profile can be described by the following equation:

sandwiches, SIPs, and PUFs were extracted by using a pressurized liquid extraction (PLE) system (ASE 350, Accelerated Solvent Extraction System from Dionex Corporation, Sunnyvale, CA, USA). The extraction was carried out using petroleum ether/acetone (83/17, v/v; 2 cycles) and thereafter acetonitrile (2 cycles) using the same ASE conditions in either case as follows: 100 °C, 5 min static cycle with a 100% flush and 240 s purge. The petroleum ether/acetone and dichloromethane extracts contain the more volatile PFASs (i.e., 6:2 FTMAC, FTACs, FTOHs, FOSAs, and FOSEs), and the methanol and acetonitrile extracts contain the PFCAs and PFSAs. The petroleum ether/acetone and dichloromethane extracts, and methanol extracts were concentrated by rotary evaporation followed by gentle nitrogen blow-down to 0.5 using iso-octane as a keeper solvent and 1 mL using methanol as a keeper solvent, respectively. Prior to injection, 10 ng absolute of Me2FOSA was added to the iso-octane extracts, respectively, and 4 ng absolute of 13C8−PFOS and 13C8−PFOA were added to the methanol extract in a polypropylene (PP) vial (Canadian Life Science, Peterborough, ON, Canada). The methanol extracts were further filtered using Mini-Uniprep PP filters (0.2 μm pore size, Whatman, Piscataway, NJ, USA) and finally transferred to PP vials (for details, see Figure S2 in the SI). The separation and detection of the 6:2 FTMAC, FTACs, FTOHs, FOSAs, and FOSEs were performed using gas chromatography−mass spectrometry (Agilent 5975C; Agilent Technologies, Palo Alto, CA, USA) (GC/MS) in selective ion monitoring (SIM) mode using positive chemical ionization (PCI). Aliquots of 2 μL were injected on a DB-WAX column (30 m, 0.25 mm inner diameter, 0.25 μm film, J&W Scientific, Folsom, CA, USA). Analyses of PFCAs, PFSAs, and FOSA were performed by liquid chromatography (Agilent 1100; Agilent Technologies, Palo Alto, CA, USA) using a triple quadrupole mass spectrometer interfaced with an electrospray ionization source in negative-ion mode (LC−(−)ESI−MS/MS; API 4000, Applied Biosystems/MDS SCIEX, Foster City, CA, USA). Aliquots of 25 μL were injected on a Luna C8(2) 100A column (50 × 2 mm, 3 μm particle size; Phenomenex, Torrance, CA) using a gradient of 250 μL min−1 methanol and water (both with 10 mM aqueous ammonium acetate solution (NH4OAc)) (for details, see Tables S7 and S8 in the SI).12 The isotope dilution method was used for quantification, which is based on the ratio of the peak-areas of the target analyte to the IS (Tables S7 and S8 in the SI). The blank concentrations and limits of detection (LODs) (average of

⎛ ⎡⎛ A kA ⎞ c PSM = KPSM−A × cA × ⎜⎜1 − exp −⎢⎜ PSM × ⎟ KPSM−A ⎠ ⎣⎢⎝ VPSM ⎝ ⎤⎞ × t ⎥⎟⎟ ⎥⎦⎠

(1)

where cPSM is the concentration of chemical in the PSM (pg m−3), KPSM−A is the PSM-air partition coefficient, cA is the total concentrations of the target analyte in air measured by HV-AAS (pg m−3), APSM is the planar area of the passive sampler in cm2 (i.e., 370 cm2), VPSM is the volume of the PSM in cm3 (i.e., 210 cm3), kA is the air-side mass-transfer coefficient (cm d−1), and t is the exposure time in days. The sample air volume is chemical specific and based on KPSM−A for each chemical. The uptake profile of the chemical to the PSM can be divided into three sections. Initially, the uptake is linear because the amount in the PSM is small. As cPSM increases over time, the term cPSM/KPSM−A becomes more important and the uptake is reduced and becomes curvilinear, and finally, cPSM reaches an equilibrium plateau (equal fugacity). In addition to depending on the properties of the PSM, the uptake profile also depends on the chamber housing design and meteorological factors such as temperature and wind speed.18 Further details for calibration of PAS are described elsewhere.17,25 14026



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Table 1. Calibration Results for PUF-PASa cA (pg m−3) PFBS PFHxS PFOS PFDS MeFOSA EtFOSA MeFOSE EtFOSE

0.24 0.13 0.96 0.09 1.24 0.89 3.37 1.77

± ± ± ± ± ± ± ±

0.18 0.11 0.46 0.08 0.81 0.58 2.07 1.16

cPUF (pg disk−1)

cPUF (pg m−3 disk)

62 21 287 41 86 47 134 208

× × × × × × × ×

2.94 9.84 1.37 1.98 4.12 2.15 6.38 9.90

5

10 104 106 105 105 105 105 105

KPUF‑A and QPUF‑Ab

log KPUF‑A and log QPUF‑Ab

6b

b

>1.22 × 10 >7.39 × 105 b >1.43 × 106 b >2.23 × 106 b 3.31 × 105 2.53 × 105 1.90 × 105 5.58 × 105

>6.09 >5.87b >6.16b >6.35b 5.52 5.40 5.28 5.75

kA (m d−1)

R (m−3 d−1)

48 ncc 55 70 ncc ncc ncc ncc

1.8 ncc 2.0 2.6 ncc ncc ncc ncc

a

PFCAs, FTOHs, 6:2 FTMAC, FTACs, and FOSA were not detected in the PUF-PAS and are therefore not shown. bAverage temperature at 18 °C between 20/07/2010 and 13/10/2010. For some analytes, the PUF-PAS did not reach equilibrium by the end of the 197 day uptake study (168 days for PFOS) so the lower limits of the partition coefficients were calculated as QPUF‑A. cnc = not calculable, because of insufficient data points.

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RESULTS AND DISCUSSION Uptake Profile for PUF-PAS and SIP-PAS for PFASs. The equivalent air volume for a passive air sampler is a measure of the amount of air that it has sampled after a given exposure period. It can also be regarded as equivalent to the sampling rate of the passive sampler (R, m3 d−1) × the number of days of exposure assuming a linear uptake during the deployment period. It can be calculated by dividing the amount of chemical in the PSM (cPSM, pg sample−1) by the total concentrations of the target analyte in ambient air using the HV-AAS (cA, pg m−3).17 The uptake profiles of PFASs for PUF-PAS and SIPPAS over the deployment time are given in Figure 1 and Figures S4 and S5 in the SI. The HV-AAS concentrations (average concentration over one month) were used to calculate the uptake profiles for the PAS because of lower detection frequency of PFASs in the LV-AAS (see Table S12 in the SI). For the PUF-PAS, Me- and EtFOSA, and Me- and EtFOSE had short linear uptake curves (56 days) in comparison to the PUF-PAS ( 0.05, Pearson Correlation). Sampling Rates for PUF-PAS and SIP-PAS for PFASs. The sampling rate (R, m3 d−1) was derived from the linear uptake phase of the uptake profiles, by taking the slope of the plot of cPSM/cA versus time. As a general rule for estimating Rvalues, ideally there should be at least 3 data points within the linear region, which we define here as the time up to 25% of equilibrium (t25). Inclusion of data points in the curvilinear region, defined here as the time in the range of 25%−90% of equilibrium (t25− t90), will result in underestimates of R. The KPSM−A at equilibrium can be described as the volume of ambient air (VAIR) that contains an equivalent amount of chemical contained in a PSM having a volume (VPSM). It is also the ratio of the concentration of the chemical in the PSM (cPSM) divided by the concentration of the target analyte in air (cA) when the system is at equilibrium.17 KPSM−A =

c VAIR = PSM VPSM cA

(2)

The KPSM−A for a chemical in the PUF-PAS and SIP-PAS (i.e., KPUF−A and KSIP−A, respectively) can be determined from its concentration in the PSM when it has reached equilibrium relative to air. This is reflected by a flattening of the uptake profile with time. However, the PFSAs in the PUF-PAS and SIP-PAS and the C12−16 PFCAs, FTOHs, FOSAs, and FOSEs in the SIP-PAS did not equilibrate, and therefore, a minimum 14027



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Table 2. Calibration Results for SIP-PASa cA (pg m−3) PFBS PFHxS PFOS PFDS PFBA PFPeA PFHxA PFHpA PFOA PFNA PFDA PFUnDA PFDoDA PFTrDA PFTeDA PFPeDA PFHxDA PFODA 8:2 FTOH 10:2 FTOH MeFOSA EtFOSA MeFOSE EtFOSE

0.24 0.13 0.96 0.09 4.80 1.28 0.45 nd 1.71 0.50 0.31 0.54 0.09 0.06 0.03 0.02 0.03 nd 53.5 21.7 1.24 0.89 3.37 1.77

± ± ± ± ± ± ±

0.18 0.11 0.46 0.08 3.09 0.93 0.33

± ± ± ± ± ± ± ± ±

1.15 0.35 0.27 0.85 0.06 0.04 0.02 0.02 0.02

± ± ± ± ± ±

23.8 11.9 0.81 0.58 2.07 1.16

cSIP (pg disk−1)

cSIP (pg m−3 disk)

76 35 511 20 3268 448 446 291 572 345 215 244 77 102 45 37 29 20 27024 7838 825 642 1493 840

× × × × × × × × × × × × × × × × × × × × × × × ×

3.60 1.69 2.44 9.49 1.56 2.14 2.13 1.39 2.73 1.65 1.02 1.17 3.69 4.86 2.15 1.75 1.38 9.50 1.29 3.74 3.94 3.06 7.13 4.01

KSIP‑A and QPUF‑Ab 6b

5

10 105 106 104 107 106 106 106 106 106 106 106 105 105 105 105 105 104 108 107 106 106 106 106

log KSIP‑A and log QPUF‑Ab b

kA (m d−1)

R (m−3 d−1)

>1.50 × 10 >1.26 × 106 b >2.55 × 106 b >1.07 × 106 b 3.25 × 106 1.67 × 106 4.68 × 106

>6.18 >6.10b >6.41b >6.03b 6.51 6.22 6.67

53 55 68 50 110 91 135

2.0 2.0 2.5 1.8 4.1 3.4 4.2

1.60 × 106 3.32 × 106 3.28 × 106 2.17 × 106 >3.94 × 106 b >8.16 × 106 b >6.27 × 106 b >7.12 × 106 b >5.33 × 106 b

6.20 6.52 6.52 6.36 >6.60b >6.91b >6.80b >6.85b >6.73b

104 96 105 100 110 149 142 125 120

3.8 3.6 3.9 3.7 4.1 5.5 5.3 4.6 4.4

106 b 106 b 106 b 106 b 106 b 106 b

>6.38b >6.24b >6.50b >6.54b >6.33b >6.35b

93 89 118 113 79 77

3.4 3.3 4.4 4.2 2.9 2.9

>2.41 >1.72 >3.17 >3.45 >2.12 >2.26

× × × × × ×

a

nd = not detected. 6:2 FTMAC, FTACs, and FOSA were not detected in the SIP-PAS and are therefore not shown. bAverage temperature at 18 °C between 20/07/2010 and 13/10/2010. For some analytes, the SIP-PAS did not reach equilibrium by the end of the 197 day uptake study so the lower limits of the partition coefficients were calculated as QSIP‑A.

define the uptake profile can be used to make decisions regarding ideal deployment times for particular chemicals or when groups of chemicals are being investigated. The estimated t25 ranged between a few weeks and three months, whereas the estimated t95 ranged between several months and two and a half years for both PAS types (Table S13 in the SI). Generally, the estimated times to equilibrate for the SIP-PAS were a factor of ∼2 higher compared to the PUF-PAS. This is a relatively small difference when compared to other compound classes such as PCBs and organochlorine pesticides (OCPs) where capacity of the SIP-PAS was up to 2 orders of magnitude greater compared to PUF-PAS.14,27,28 This finding suggests that the PFASs partition differently to PUF-PAS and SIP-PAS compared to semivolatile organic compounds like PCBs and OCPs, which sorb much less to surfaces due to their nonpolar hydrophobic characteristics. This is consistent with observations from particle-gas partitioning investigations of the PFASs, that indicate that they do not obey the typical KOA-based and subcooled liquid vapor pressure (poL)based relationships that have been derived for nonpolar hydrophobic chemicals.22 However, more work is required to further investigate the sorption mechanism of PFASs. The correlation of log KPSM−A (or log QPSM‑A) against log KOA for PFASs in PUF-PAS and SIP-PAS was investigated (see SI), and the results are shown in Table S14 and Figure S7 in the SI. Overall, only a weak correlation was observed indicating that the PFASs do not undergo KOA-driven partitioning into PSM. Thus, there might be other factors influencing the sorption of PFASs to PSM, for example, amount of particles/ aqueous aerosols sampled by the PAS and influence of the hydrophobic fluorocarbon chain length and hydrophilic functional groups of PFASs.11,22,29

value for the partition coefficient is estimated for these compounds (in this case, QPSM−A is used instead of KPSM−A, to indicate that it is not a true partition coefficient) (see Tables 1 and 2). It is important to note, that KPSM−A can increase by a factor of 2.5−3.0 with every 10 °C decrease in temperature.17 This means longer linear phases for PAS when operating at colder temperatures. For the PUF-PAS, the sampling rates for the PFSAs (except PFHxS) ranged between 1.8 and 2.6 m3 d−1 (PFCAs were not detected in the PUF-PAS). The sampling rates for the FOSAs, FOSEs, and PFHxS could not be reported due to rapid equilibration of these compounds in the PUF-PAS and lack of sufficient data points to allow for a reliable estimate of the linear sampling rate. For the SIP-PAS, the sampling rates for FTOHs (3.3−4.3 m3 d−1), FOSAs (4.2−4.4 m3 d−1), and FOSEs (2.9 m3 d−1) in this study are in reasonable agreement with previous reported indoor derived sampling rates (i.e., 4.6 m3 d−1 for FTOHs, 2.6 for FOSAs, and 1.4−1.5 m3 d−1for FOSEs).16 Differences can be explained by different environmental conditions (outdoor vs indoor). PFSAs were similar compared to the PUF-PAS, ranging from 1.8 to 2.5 m3 d−1. The sampling rates for PFCAs were, on average, 4.2 m3 d−1 for the SIP-PAS. The uptake profile for PAS can be described by the uptake constant kU (day−1).17 kU =

APSM kA × VPSM KPSM−A

(3)

The kU can be used to calculate the extent of the linear uptake phase as t25 = ln(0.75)/kU (i.e., time when the PSM has accumulated 25% of the equilibrium value) and the time of 95% of equilibrium value as t95 = ln(0.05)/kU. These values that 14028



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Figure 2. Loss of DCs during deployment as fraction of the starting amount. The concentration of the chemical in the PSM during the deployment time (c) is divided by the concentration of the chemical at deployment time t = 0 (c0).

Loss of Depuration Compounds. The use of various DCs was explored to cover the range of PFASs. DCs are useful for calculating site-specific R-values under different meteorological conditions. The R-values (m3 d−1) can be calculated by multiplying the kA value, which was derived from the loss of the DCs, by the surface area of the PUF-PAS (for details, see elsewhere15,20). Conditions under which DCs are selected and applied include: (i) DCs should not be present in ambient air. (ii) DCs should belong to the same compound class as the target analytes. (iii) DCs should not degrade in the PSM during the deployment period. Losses should only be due to volatilization to air (i.e., air-side mass transfer). (iii) Target losses for DCs during the deployment period is in the range of >40% to up to 90%. This ensures that losses are large enough to be distinguished from analytical variability. (iv) Several DCs should be used so that an average sampling rate can be determined. This also helps to reduce variability associated with analysis of individual compounds.20 In Figure 2, the loss of the DCs (i.e., 13C8−PFOS, 13C8− PFOA, and 7:2 sFTOH) is shown during the deployment period. For 13C8−PFOS, no substantial loss was observed for the PUF-PAS and SIP-PAS during the deployment period of 197 days. This demonstrates the low volatility and high adsorption and stability of 13C8−PFOS (and therefore also for the native PFOS) to both PAS media. The high adsorption potential of PFOS is also observed by its strong sorption to airborne particles (i.e., 46%).12 Given the negligible loss of 13C8 PFOS from the PAS, this chemical is better suited as a quality control/recovery surrogate for the DCs. In contrast, the 13C8− PFOA concentration decreased linearly for both PAS types during the deployment period. Overall, the loss of 13C8−PFOA was ∼85% for the PUF-PAS and ∼45% for the SIP-PAS after 168 days. For 7:2 sFTOH, a different adsorption behavior was observed between the PUF-PAS and SIP-PAS. For the SIPPAS, 7:2 sFTOH had a linear loss of ∼17% during the deployment period of 197 days showing the high sorption capacity of SIP-PAS for more volatile chemicals like 7:2 sFTOH. In contrast, 7:2 sFTOH was only found in the spiked blank PUF disk samples but was not detected in any deployed PUF-PAS which means that 7:2 sFTOH was completely volatilized within seven days of deployment which represents the time when the first uptake sample was collected. This is in agreement with the uptake results for the FTOHs, which have a

similar structure to the sFTOHs, showing that they have a very limited sorption capacity in PUF-PAS. Overall, of the DCs tested in this study, only 13C8 PFOA was suitable for the calculation of the site specific sampling rates. The R-value of 13C8 PFOA was determined to be 3.7 ± 0.9 m3 d−1 for the SIP-PAS. This is in agreement with the R-value calculated for native compounds in this uptake study (see Table 2). This is in accordance to a previous study showing comparable R-values determined from time integrated active sampling and the DC approach.30 More work is required to understand the influence of the PSM-side kinetic resistance on the loss of the DC and identify additional PFASs that are suitable as DCs. Implications for the Calculation of the Equivalent Sampling Air Volume. The approach for calculating the equivalent sampling air volume (VAIR) for the PAS was described previously17 (for the uptake parameters, see Tables 1 and 2). ⎛ ⎡⎛ A kA ⎞ VAIR = KPSM−A × VPSM × ⎜⎜1 − exp −⎢⎜ PSM × ⎟ ⎢⎣⎝ VPSM KPSM−A ⎠ ⎝ ⎤⎞ × t ⎥⎟⎟ ⎥⎦⎠ (4)

For analytes which are still in the linear phase, VAIR can simply be calculated by multiplying the R-value of the analyte with the days of deployment. VAIR = R × t

(5)

These same expressions are applied to both gas- and particlephase compounds.31−33 In order to investigate the robustness of the passive sampling chamber design, different configurations were tested by varying the gap (opening) between the upper and lower domes. A larger opening between the upper and lower domes could result in increased sampling of gas-phase compounds if the wind effect is important, over the range of wind speeds at the sampling site.34 The larger gaps should also allow for unimpeded movement of particles into the sampler. If particle sampling is somehow reduced by the conventional chamber design, we should see greater sampling of particlebound PFASs in the larger gap configurations.18 The results indicate that the concentration difference for PFSAs, FOSAs, and FOSEs that are associated with particles12 and for other PFASs that are mainly in the gas phase was not significant between the different chambers for SIP-PAS and PUF-PAS (p > 0.05, Student’s t-test) (Figures S8 and S9 in the SI). 14029



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Figure 3. Comparison between total air concentrations (gas and particle phase) using HV-AAS and concentrations derived by PUF-PAS and SIPPAS in pg m−3 using linear regression. Each dot represents average concentration over one month for HV-AAS and integrated concentration over one month for PAS for individual PFAS.

Comparison of Four Different Sampling Techniques. PFAS were measured in air using four different sampling techniques: (i) HV-AAS to measure gas and particle phase separately, (ii) LV-AAS comprising the sum of the gas and particle phase, (iii) SIP-PAS, and (iv) PUF-PAS. In general, the averages agree generally within a factor of 2 and no significant differences were found for the PFAS concentrations measured by the PUF-PAS, SIP-PAS, LV-AAS, and HV-AAS (p > 0.05, Kruskal−Wallis test) (Figure S10 in the SI). The performance of the PUF-PAS and SIP-PAS for measuring FOSAs/FOSEs and PFSAs in the atmosphere was compared using linear regression. Both the FOSA/FOSE and PFSA concentrations were generally within a factor of 2 for the two PAS types (r2 = 0.66 and r2 = 0.98, respectively) (Figure S11 in the SI). The air concentration of PFASs measured by the HV-AAS (representing 14−29% of the time for the monthly average) was compared with the air concentration derived by the SIPPAS and PUF-PAS (which sample 100% of the time) using linear regression (Figure 3). Generally, the SIP-PAS showed a good agreement with the air concentration determined by the HV-AAS for all PFAS classes (r2 > 0.9). The average difference between the two sampling techniques was less than 50% for individual PFASs. For the PUF-PAS, the FOSA/FOSE concentrations showed a higher scattering of the data (r2 = 0.76) but the linear regressions with the HV-AAS measurements were close to unity. The higher scattering of the FOSA and FOSE concentrations can be explained by lower accumulation of these compounds in the PUF-PAS due to a limited uptake capacity (compared to the SIP-PAS). Consequently, these lower concentrations in the PUF-PAS may approach detection limits for these compounds, which results in derived air concentrations that have greater analytical

In this study, all PFAS (except FOSAs and FOSEs in the PUF-PAS) showed a lengthy linear/curvilinear uptake phase with an average linear-phase R-value ranging from 1.8−5.5 m3 d−1 (on average, 3.5 m3 d−1) for PUF-PAS and SIP-PAS and the R-value derived from the DC 13C8 PFOA was determined to be 3.7 m3 d−1 for the SIP-PAS. These sampling rates are very close to the suggested R-value of 4 m3 d−1 reported previously for the classical POPs indicating a similar uptake characteristic for POPs and PFASs.15 Minor variations in derived sampling rates, between compounds classes, are likely due to analytical or experimental variability. Thus, to simplify the assessment of equivalent sample air volumes, eq 5 was applied using a common linear sampling rate of 4 m3 d−1 (with the exception of FOSAs and FOSEs in the PUF-PAS, see below). This results in a VAIR of 112 m3 for a one month deployment period. The sample volumes for FOSAs and FOSEs in the PUF-PAS were calculated using eq 4 to account for their approach to equilibrium and reduced sample air volumes. VAIR ranged from 39 to 72 m3 for the FOSAs and FOSEs in PUF-PAS. Ultimately, the calculated VAIR can be used to calculate the concentration of the analyte in air (cA) by dividing cPSM (pg disk−1) with VAIR (m3).

cA = c PSM /VAIR

(6)

Overall, VAIR can be calculated by applying a R-value of 4 m3 d−1 for all PFASs except for the FOSAs and FOSEs in the PUFPAS for which the full uptake expression of eq 4 has been used. It is interesting to note that there was no significant correlation of the R-value for the individual analytes (see Tables 1 and 2) with the particle associated fraction of the analyte in air12 (p > 0.05, Pearson Correlation). This indicates that the SIP-PAS capture gas-phase and particle-phase PFASs with similar efficiency, consistent with the recent findings by Harner et al.33 14030



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Figure 4. PFOS, PFOA, 8:2 FTOH, and MeFOSE concentrations in air measured by four different sampling techniques over one year: HV-AAS (sum of gas and particle phase; integrated over 24 h (black bars) and average concentration over one month) and LV-AAS (sum of gas and particle phase; integrated over 14 days), and SIP-PAS and PUF-PAS (integrated over one month).

dominant compound (48% of the ΣFTOHs). The total ΣPFCA concentration ranged from 0.7 to 20 pg m−3 with PFBA as the dominant compound (∼54% of the ΣPFCA) and with a tendency of decreasing air concentrations for the longer chain PFCAs. The C4-based PFASs (e.g., PFBA) are the main replacement compound of the voluntary phase-out C8-based products (i.e., PFOA and PFOS) which may explain the elevated concentration of PFBA in air in Toronto.2,3 It is interesting to note that the average ΣFTAC concentrations in this study were higher (8.3 pg m−3) compared to the ΣFOSAs and ΣFOSEs (1.4 and 3.4 pg m−3, respectively). This demonstrates the importance of FTACs as the next most relevant precursor class after the FTOHs. However, a recent study of PFASs in the Asian atmosphere showed that 8:2 fluorotelomer olefin (FTO) is the second most abundant PFAS class after the FTOHs.35 The ΣFOSA and ΣFOSE concentrations were about 5 times lower than previously reported for suburban or urban areas5,36 which might be due to the phase out of perfluorooctyl sulfonyl fluoride (POSF), reduced PFAS emissions by optimization of the production process,2 or a production shift to shorter chain PFASs and new fluorinated chemicals.37,38 Lower total air concentrations were observed for ΣPFSAs (on average 1.0 pg m−3) with PFOS as the predominant compound in this PFAS class (74% of the ΣPFSAs). This is in agreement with recent measurement in the Toronto atmosphere.39,40 The air concentrations of individual PFASs in Toronto were compared over one year, using the four different sampling approaches used in this study (see Figure 4 and Figure S13 in the SI). For the PFASs in LV-AAS and the PFSAs and PFCAs

uncertainty. In contrast, the PFSA concentrations derived from the SIP-PAS and PUF-PAS showed a good linear regression with r2 of 0.91 and 0.90, respectively (p < 0.05, Pearson Correlation). However, the concentrations were lower compared to the total air concentration measured by the HVAAS, which can be due to the high association of PFSAs to the particle phase22 which might be less efficiently collected by the PAS. The PFCA concentrations derived from the SIP-PAS also showed a good linear regression with r2 of 0.92 (p < 0.05, Pearson Correlation), but the concentrations were slightly higher indicating a higher collection efficiency of PFCAs by SIP-PAS compared to HV-AAS. However, the PFOA concentrations derived from the SIP-PAS were generally lower compared to HV-AAS which might be due to concentrations close to the detection limit for SIP-PAS (see Figure 4). Overall, the difference for individual PFASs was within a factor of 2 using PUF-PAS and SIP-PAS which can be considered to be good agreement, especially considering that some variability is expected, due to the HV-AAS not operating 100% of the time. Atmospheric Composition and Seasonal Trends of PFASs. Overall, all of the 29 targeted PFASs were detected in air samples (Table S12 in the SI). The most abundant PFAS class for the total air concentration (sum of gas and particle phase measured by HV-AAS) was the FTOHs representing on average ∼80% of the ΣPFASs, followed by PFCAs (∼7%) and FTACs/fluorotelomer methacrylates (FTMACs) (∼7%) (Figure S12 in the SI). The other PFAS classes represented less than 3% of the ΣPFASs. Total air concentrations (HV-AAS) for ΣFTOHs ranged from 20−182 pg m−3 with 8:2 FTOH as the 14031



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Canada: Implications for human exposure. Environ. Sci. Technol. 2011, 45, 7999−8005. (7) Ahrens, L.; Shoeib, M.; Harner, T.; Lee, S. C.; Guo, R.; Reiner, E. J. Wastewater treatment plant and landfills as sources of polyfluoroalkyl compounds to the atmosphere. Environ. Sci. Technol. 2011, 45, 8098−8105. (8) Wang, Z.; Scheringer, M.; MacLeod, M.; Bogdal, C.; Müller, C. E.; Gerecke, A. C.; Hungerbühler, K. Atmospheric fate of poly- and perfluorinated alkyl substances (PFASs): II. Emission source strength in summer in Zurich, Switzerland. Environ. Pollut. 2012, 169, 204−209. (9) Müller, C. E.; Gerecke, A. C.; Bogdal, C.; Wang, Z.; Scheringer, M.; Hungerbühler, K. Atmospheric fate of poly- and perfluorinated alkyl substances (PFASs): I. Day-night patterns of air concentrations in summer in Zurich, Switzerland. Environ. Pollut. 2012, 169, 196−203. (10) Barber, J. L.; Berger, U.; Chaemfa, C.; Huber, S.; Jahnke, A.; Temme, C.; Jones, K. C. Analysis of per- and polyfluorinated alkyl substances in air samples from Northwest Europe. J. Environ. Monit. 2007, 9, 530−541. (11) Arp, H. P. H.; Goss, K. U. Irreversible sorption of trace concentrations of perfluorocarboxylic acids to fiber filters used for air sampling. Atmos. Environ. 2008, 42, 6869−6872. (12) Ahrens, L.; Shoeib, M.; Harner, T.; Lane, D. A.; Guo, R.; Reiner, E. J. Comparison of annular diffusion denuder and high volume air samplers for measuring per- and polyfluoroalkyl substances in the atmosphere. Anal. Chem. 2011, 83, 9622−9628. (13) Pozo, K.; Harner, T.; Wania, F.; Muir, D. C. G.; Jones, K. C.; Barrie, L. A. Toward a global network for persistent organic pollutants in air: Results from the GAPS Study. Environ. Sci. Technol. 2006, 40, 4867−4873. (14) Genualdi, S.; Lee, S. C.; Shoeib, M.; Gawor, A.; Ahrens, L.; Harner, T. Global pilot study of legacy and emerging persistent organic pollutants using sorbent-impregnated polyurethane foam disk passive air samplers. Environ. Sci. Technol. 2010, 44, 5534−5539. (15) Pozo, K.; Harner, T.; Lee, S. C.; Wania, F.; Muir, D. G.; Jones, K. C. Seasonally resolved concentrations of persistent organic pollutants in the global atmosphere from the first year of the GAPS study. Environ. Sci. Technol. 2009, 43, 796−803. (16) Shoeib, M.; Harner, T.; Lee, S. C.; Lane, D.; Zhu, J. Sorbentimpregnated polyurethane foam disk for passive air sampling of volatile fluorinated chemicals. Anal. Chem. 2008, 80, 675−682. (17) Shoeib, M.; Harner, T. Characterization and comparison of three passive air samplers for persistent organic pollutants. Environ. Sci. Technol. 2002, 36, 4142−4151. (18) Klánová, J.; Èupr, P.; Kohoutek, J.; Harner, T. Assessing the influence of meteorological parameters on the performance of polyurethane foam-based passive air samplers. Environ. Sci. Technol. 2008, 42, 550−555. (19) Chaemfa, C.; Barber, J. L.; Gocht, T.; Harner, T.; Holoubek, I.; Klanova, J.; Jones, K. C. Field calibration of polyurethane foam (PUF) disk passive air samplers for PCBs and OC pesticides. Environ. Pollut. 2008, 156, 1290−1297. (20) Moeckel, C.; Harner, T.; Nizzetto, L.; Strandberg, B.; Lindroth, A.; Jones, K. C. Use of depuration compounds in passive air samplers: Results from active sampling-supported field deployment, potential uses, and recommendations. Environ. Sci. Technol. 2009, 43, 3227− 3232. (21) Zhang, X.; Tsurukawa, M.; Nakano, T.; Lei, Y. D.; Wania, F. Sampling medium side resistance to uptake of semivolatile organic compounds in passive air samplers. Environ. Sci. Technol. 2011, 45, 10509−10515. (22) Ahrens, L.; Harner, T.; Shoeib, M.; Lane, D. A.; Murphy, J. G. Improved characterization of gas-particle partitioning for per- and polyfluoroalkyl substances in the atmosphere using annular diffusion denuder samplers. Environ. Sci. Technol. 2012, 46, 7199−7206. (23) United Nations Environment Programme (UNEP), Geneva hosts Stockholm Convention on Persistent Organic Pollutants from 4 to 8 May, 2009, from http://www.unep.org/newscentre/ (22/09/ 2010).

in PUF-PAS, results are presented for 29 consecutive weeks from the start of the study. The individual PFAS concentrations based on HV-AAS varied over time by up to a factor of 5. This variability is lower than previously reported at a site close to Hamburg, Germany.41 In particular, we did not see peak events with extremely high PFAS concentrations (i.e., 1 order of magnitude higher than baseline levels) as reported previously.41 The main factor governing the variability in air concentrations of PFASs over time proved to be temperature. The majority of PFASs classes measured by HV-AAS was significantly correlated with ambient temperature (p < 0.05, Pearson Correlation) (for details, see Table S15 in the SI). This suggests an important influence from local/regional sources that exhibit seasonality which may be partly attributed to temperature (i.e., enhanced volatilization). Generally, the PFAS concentrations decreased in the order of summer, spring, fall, and winter. Potential emission sources in Toronto for PFASs include inter alia WWTPs and landfills, which are considered point sources,42 and residential homes, which can be considered as diffuse sources.6 Ultimately, all four sampling approaches (i.e., HV-AAS, LV-AAS, SIP-PAS, and PUF-PAS) are deemed suitable for capturing temporal trends of PFASs in air.

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ASSOCIATED CONTENT

S Supporting Information *

Additional details on sampling sites, meteorological data, QA/ QC data, PFAS concentrations, and predicted log KOA for individual PFAS. This material is available free of charge via the Internet at http://pubs.acs.org.

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AUTHOR INFORMATION

Corresponding Authors

*E-mail: [email protected]; phone: +46 70-2972245; fax: +46 70-2972245. *E-mail: [email protected]; phone: +1 416-739-4837; fax: +1 416-739-4281. Notes

The authors declare no competing financial interest.

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ACKNOWLEDGMENTS Partial funds for this work were provided through the Chemicals Management Plan (Government of Canada), the Chemicals Management Division (Environment Canada), and the United Nations Environment Programme (UNEP).

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REFERENCES

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(42) Ahrens, L.; Shoeib, M.; Harner, T.; Lee, S. C.; Guo, R.; Reiner, E. J. Wastewater treatment plant and landfills as sources of polyfluoroalkyl compounds to the atmosphere. Environ. Sci. Technol. 2011, 45, 8098−8105.

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Characterization of Two Passive Air Samplers for Per-and

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