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.
■
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
INTRODUCTION
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.
© XXXX American Chemical Society
Received: May 22, 2013
Accepted: November 12, 2013
A
dx.doi.org/10.1021/es4048945 | Environ. Sci. Technol. XXXX, XXX, XXX−XXX
Environmental Science & Technology
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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.
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)CHCH 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
■
B
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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 ×
⎟
⎢⎣⎝ VPSM
KPSM−A ⎠
⎝
⎤⎞
× 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
C
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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)
62
21
287
41
86
47
134
208
cPUF (pg m−3 disk)
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
6b
>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
log KPUF‑A and log QPUF‑Ab
b
>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.
Time required for equilibrium to be established varied
depending on the chain length for the PFCAs. For example, the
longer chained C12−C16 PFCAs had a longer linear or
curvilinear phase compared to the C4−C11 PFCAs (Figure 1).
A similar trend was previously observed for the uptake of
polychlorinated biphenyls (PCBs) in PUF-PAS showing a
longer linear uptake phase for the higher molecular weight
PCBs (i.e., penta- to heptachlorinated biphenyls).19
As mentioned previously, the uptake profile is chemical
dependent and also influenced by variability of the chemical
concentration during the deployment period and by meteorological conditions like the wind speed and temperature that also
vary with time.17−19,26 The average wind speed was relatively
constant throughout the deployment period with a mean of
about 16 km h−1 and did not show a temporal bias, whereas the
temperature was lower at the beginning and in the end of the
study (on average ∼14 °C) and higher in summer during the
middle of the uptake study (on average ∼22 °C) (Tables S5
and S6 in the SI). However, no significant correlation was
observed between the sampling rate and the range of wind
speeds, temperature, and other meteorological parameters (i.e.,
relative humidity, wind direction, and amount of rain) for the
study (p > 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
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 (<28 days) and equilibrated after
a few weeks. Although Me- and EtFOSA equilibrated faster
(i.e., equilibration after ∼56 days) in comparison to the Meand EtFOSE (i.e., equilibration after ∼120 days), the equivalent
air volume at equilibrium was similar for both PFAS classes in a
range of 60−70 m3, reflecting similar KPUF‑A values for these
two compound classes. In contrast, the PFSAs constantly
increased and the uptake profile was still in the curvilinear
phase until the end of the deployment period of 168 days (the
PUF disks deployed over 197 days were not analyzed for
PFSAs and PFCAs). It is important to note that 8:2 and 10:2
FTOH, 6:2 FTMAC, PFCAs, FTACs, and FOSA were not
detected in the PUF-PAS which shows that PUF-PAS have a
very low sorptive capacity for these compounds. A low uptake
capacity for the FTOHs in PUF disks was also observed in a
previous indoor uptake study with equilibration of FTOHs after
a few days resulting in an effective air volume of only a few
cubic meters.16
For the SIP-PAS, almost all analyzed PFAS classes (i.e.,
PFSAs, PFCAs, FTOHs, FOSAs, and FOSEs) were detected,
except the 6:2 FTMAC and FTACs (see Table S12 in the SI).
The uptake for the PFCAs and FTOHs was greatly improved
using SIP-PAS compared to the PUF-PAS for which no PFCAs
and FTOHs were detected. Furthermore, greater sorption
capacities were also observed for the FOSAs and FOSEs
showing longer linear uptake curves for these compounds in
SIP-PAS (>56 days) in comparison to the PUF-PAS (<28
days). In contrast, the uptake profile for the PFSAs was similar
for the PUF-PAS and SIP-PAS with equivalent air volume of
∼300 m3 after 160 days of deployment (Figure S6 in the SI).
■
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
D
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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)
76
35
511
20
3268
448
446
291
572
345
215
244
77
102
45
37
29
20
27024
7838
825
642
1493
840
cSIP (pg m−3 disk)
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
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
kA (m d−1)
R (m−3 d−1)
6b
>1.50 × 10
>1.26 × 106 b
>2.55 × 106 b
>1.07 × 106 b
3.25 × 106
1.67 × 106
4.68 × 106
b
>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
E
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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).
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).
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
⎛
⎡⎛ A
kA ⎞
VAIR = KPSM−A × VPSM × ⎜⎜1 − exp −⎢⎜ PSM ×
⎟
KPSM−A ⎠
⎣⎢⎝ VPSM
⎝
⎤⎞
× 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).
F
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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.
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
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
(6)
Overall, VAIR can be calculated by applying a R-value of 4 m3
d 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
−1
G
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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
H
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E. J. Comparison of annular diffusion denuder and high volume air
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Harner, T. Global pilot study of legacy and emerging persistent
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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.
■
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.
■
AUTHOR INFORMATION
Corresponding Authors
*E-mail: lutz.ahrens@slu.se; phone: +46 70-2972245; fax: +46
70-2972245.
*E-mail: tom.harner@ec.gc.ca; phone: +1 416-739-4837; fax:
+1 416-739-4281.
Notes
The authors declare no competing financial interest.
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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