Use of Depuration Compounds in Passive Air Samplers: Results from

Mar 20, 2009 - A field study was performed on a vertical tower, where wind speed data were available at different heights with the same air concentrat...
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Environ. Sci. Technol. 2009, 43, 3227–3232

Use of Depuration Compounds in Passive Air Samplers: Results from Active Sampling-Supported Field Deployment, Potential Uses, and Recommendations CLAUDIA MOECKEL,† TOM HARNER,‡ LUCA NIZZETTO,† BO STRANDBERG,§ ANDERS LINDROTH,| AND K E V I N C . J O N E S †,* Centre for Chemicals Management, Lancaster Environment Centre, Lancaster University, Lancaster LA1 4YQ, U.K., Atmospheric Science and Technology Directorate, Environment Canada, Toronto M3H 5T4, Canada, Sahlgrenska Academy, Department of Public Health and Community Medicine, University of Gothenburg SE-40530 Gothenburg, Sweden, and Department of Physical Geography and Ecosystems Analysis, Lund University, So¨lvegatan 12, 22362 Lund, Sweden

Received October 13, 2008. Revised manuscript received February 24, 2009. Accepted February 26, 2009.

Depuration compounds (DCs) are added to passive air samplers (PAS) prior to deployment to account for the winddependency of the sampling rate for gas-phase compounds. This correction is particularly useful for providing comparable data for samplers that are deployed in different environments and subject to different meteorological conditions such as wind speeds. Two types of PASsthe polyurethane foam (PUF) disk sampler and semipermeable membrane devices (SPMDs)swere deployed at eight heights on a 100 m tower to test whether the DC approach could yield air concentrations profiles for PCBs and organochlorine pesticides and account for the wind speed gradient with height. Average wind speeds ranged from 0.3 to 4.5 m s-1 over the 40 day deployment, increasing with height. Two low volume active air samples (AAS), one collected at 25 m and one at 73 m over the 40 day deployment showed no significant concentration differences for target compounds. As expected, the target compounds taken up by PAS reflected the wind profile with height. This winddependency of the PAS was also reflected in the results of the DCs. A correction based on the DC approach successfully accounted for the effect of wind on PAS sampling rates, yielding a profile consistent with the AAS. Interestingly, in terms of absolute air concentrations, there were differences between the AAS and PAS-derived values for some target compounds. These were attributed to different sampling characteristics of the two approaches that may have resulted in slightly different air masses being sampled. Based on the results of this study, guidelines are presented for the use of DCs and for the calibration of PAS using AAS. * Corresponding author e-mail: [email protected]. † Lancaster University. ‡ Environment Canada. § University of Gothenburg. | Lund University. 10.1021/es802897x CCC: $40.75

Published on Web 03/20/2009

 2009 American Chemical Society

Introduction Sampling and analysis of persistent semivolatile organic compounds (SVOCs) (e.g., polychlorinated biphenyls (PCBs), polybrominated diphenyl ethers (PBDEs), and organochlorine pesticides (OCPs)) is required for national and international monitoring programs, and in compliance of international protocols such as the Stockholm Convention (1). In addition, because global distribution occurs via atmospheric transport, numerous research studies have focused on elucidating the atmospheric transport processes affecting these chemicals (e.g., 2-4). Active sampling of SVOCs (i.e., using a pump to draw known volumes of air through filters and sorbents) is the most accurate method to monitor airborne concentrations but cannot be conducted at a large number of sites simultaneously because of cost and logistical limitations. To overcome this, a number of different passive air samplers (PAS) have been developed, tested, and used in many studies (e.g., 4-8). Their basic principle is to trap SVOCs that have reached the sampling medium passively, i.e., by advection and diffusion. Diffusion through the laminar air layer adjacent to the sampling medium has been identified as the rate limiting step for uptake of many chemicals (5). Since the thickness of this boundary layer decreases with increasing wind speed, the uptake rate may depend on the wind speed at the sampling site, unless the passive sampler operates purely under fixed diffusion path lengths. To dampen this effect, passive sampling media are usually deployed in protective chambers (6, 9-13) that also shield them from direct sunlight and deposition of coarse particulate matter. Tuduri et al. (14) have tested the influence of wind speed on a widely used PAS design, based on a polyurethane foam (PUF) disk protected by a chamber made from two stainless steel bowls. Their study suggested a gradual increase in uptake rates with a relatively small slope over an air velocity range of 0-0.9 m s-1 inside the chamber, corresponding to an ambient wind speed of about 0-3.5 m s-1. Uptake rates increased sharply above this range. It was concluded that this will not cause problems in many studies as wind speeds are typically within the “safe” range. Klanova et al. (10) compared PUF-disk and active air samples (AAS) that were codeployed over 3 years over varying meteorological conditions at a meteorological station and determined that sampling rates of predominantly gas-phase chemicals increased by about a factor of 2 with wind speed, over the range 1.5-5.5 m s-1. The study also showed that predominantly particle-associated chemicals were taken up at a rate approximately 10% the gas-phase sampling rate. In many cases, however, samplers are deployed at locations where high winds are common yet wind speed is not monitoredsfor instance, samplers deployed in mountainous or remote coastal sites. In these instances a direct approach is required to account for the variable sampling rates and to provide better quantitative estimates of the true air concentration. One approach that is now commonly used is to spike PAS with depuration compounds (DCs) (sometimes also called performance reference compounds or PRCs) prior to deployment (e.g., 4, 6, 11, 12, 15). DCs are semivolatile chemicals that cannot be found in the environment and do not interfere with the analysis of target compounds. They will volatilize into the atmosphere if exposed to air or dissolve into the aqueous phase if deployed in water during the deployment period. The amount lost will depend on their physicochemical properties, exposure time, and wind speed or water flow rates (11, 16). However, although DCs were used in a number of air studies to either demonstrate that VOL. 43, NO. 9, 2009 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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wind speed effects were negligible, or to correct for them, their performance in PAS has not been validated in laboratory or in field experiments. The aim of the present study was therefore to test how useful DCs are for normalizing results derived from two widely used PAS designs (PUF disks and SPMDs). A field study was performed on a vertical tower, where wind speed data were available at different heights with the same air concentration of target compounds, as monitored using active air samplers.

Materials and Methods Sampling was conducted from May 30 to July 9, 2007 at Norunda Common, a well studied forest site 30 km north of Uppsala, Sweden (60°05′ N, 17°29′ E), equipped with a 100 m high tower mainly used for measuring fluxes of trace gases. Table SI-1 (Supporting Information) summarizes temperature and wind speed, which are assumed to be the most influential meteorological parameters during the sampling period. The SPMDs used in this study were of standard specifications (90 cm long and 2.75 cm wide tubes of low-density polyethylene filled with 1 mL of 95% pure triolein (lipid); surface area: 0.0495 m2; mass (triolein + membrane): 4.6 g; volume: 5.08 × 10-6 m3; density: 900 000 g m-3) purchased from Environmental Sampling Technologies (EST Laboratories, St. Joseph, MO). Properties of the PUF disks (PacWill Environmental Stoney Creek, ON) were as follows: diameter: 0.14 m; thickness: 0.0135 m; surface area: 0.0365 m2; mass: 4.40 g; volume: 2.07 × 10-4 m3; density: 21 300 g m-3. SPMDs were deployed on steel holders (“spider” carriers, EST Laboratories, St. Joseph, MO). To protect the passive sampling media from rain, wind, sunlight, and mechanical damage, both the spiders with SPMDs and PUF disks were mounted inside protective chambers made from stainless steel as described by Chaemfa et al. (9) (Lancaster design), but based on the original configuration of Shoeib and Harner (5). Metal tubes were used as spacers to define the position of the “spider” or PUF disk inside the chamber, to minimize variability between samplers. Passive samplers were installed at eight different elevations along the 100 m high tower. To monitor variations between samples from the same height, SPMDs were deployed in triplicate at two heights on the tower and duplicate PUF disks were deployed at all heights (each in separate chambers). Table SI-2 gives the number of SPMDs and PUF disks at each height. For active air sampling, low volume pumps connected to a filter holder were used (Figure SI-1). The air was drawn at a flow rate of ca. 11 L min-1 (resulting in a face velocity of about 14.5 cm s-1) from the bottom to the top of the holder so the 9 cm long metal tube sheltered the filter from wind and rain (i.e., the sampling head was pointing downward). Particles and SVOCs associated with particulate matter were trapped on a glass fiber filter (GFF); for gas-phase compounds PUF plugs were used. To detect breakthrough of vapor-phase compounds, two PUF plugs were installed in a sequence and analyzed separately. The AAS (one at 25 m, just at the top of the canopy, and one at 75 m, well above the canopy) were taken continuously over the whole passive air sampling period. Both PUF disks and plugs were precleaned by Soxhlet extraction using DCM for 16 h, GFFs were baked at 450 °C overnight, SPMDs were used as purchased from the supplier. Prior to deployment of the passive samplers 5 µL of a DCs standard (each DC ca. 300 pg µL-1) was added to the SPMD lipid with a GC-syringe and evenly distributed inside it by moving the lipid from one end to the other several times after heat resealing the SPMD plastic tube. PUF disks were also spiked with DCs by adding 25 µL of a DC standard (each DC ca. 400 pg µL-1) to approximately 15 mL of hexane, which was then evenly applied to both sides of the PUF disk using a Pasteur pipet. After the solvent was evaporated, spiked disks were stored 3228

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in amber glass bottles until deployment. Spiking was conducted 10 days before deployment, to encourage a homogeneous distribution of the DCs within the disk. During the deployment period DCs volatilize from the passive sampler according to their partitioning properties and, assuming an air-side controlled loss (as well as air-side controlled uptake), their loss rate is inversely proportional to the thickness of the laminar air layer surrounding the sampler. The thickness of this layer is reduced under higher wind speeds resulting in faster mass transfer (10, 12). Usually, several compounds (e.g., 13C labeled or deuterated analogues of target compounds) that cover a range of volatility are used to ensure that at least one of them experiences losses between 20 and 80%. This is required to distinguish differences in the loss from the analytical variability (16). All blanks need to be spiked in the same way as the samples to determine the concentration of DCs at the start of the experiment. To allow for comparison with losses of DCs in earlier studies PUF disks were spiked with PCB congeners 54, 104, and 188 as DCs and SPMDs were spiked with 13C12 labeled PCB congeners 28, 52, 101, 138, 153, and 180 (4). According to preliminary estimates, assuming a typical uptake rate of 4 m3 day-1 (5), only the most volatile of these DCs were supposed to experience suitable losses within a deployment period of 40 days. However, since high wind speeds were likely to result in elevated uptake rates (10), less volatile DCs were used as well. To minimize the risk of contamination, all sampling media were handled with gloves (changed between each sample when collecting exposed samplers) or solvent-rinsed tweezers and stored in solvent-rinsed amber glass jars (for PUF disks and plugs) or lever lid tins (for the SPMDs). After collection, samples were stored sealed in a cool place and quickly shipped to Lancaster where they were kept at -20 °C until being processed. Details of the cleanup and analytical methods used in the present study including a list of all target compounds are given in the Supporting Information (Text SI-1).

Results and Discussion Quality Assurance/Quality Control. Field blanks were included at a rate of 1 per 4 passive air samples (i.e., 5 PUF disk and 4 SPMD field blanks) and additionally 1 GFF and 2 PUF plug blanks were analyzed. Blanks consisted of preextracted PUF disks and plugs and SPMDs as purchased from the supplier, respectively, extracted and analyzed in the same way as samples. Depending on the compound, blank levels ranged from “not detected” (n.d.) to 65 pg sample-1 in PUF disks, n.d. to 30 pg sample-1 for PUF plugs, and n.d. to 200 pg sample-1 in SPMDs. No analytes could be found in the GFF blank. The limit of detection (LOD) was calculated as three times the standard deviation of the mean blank of individual target compounds or the amount per sampler corresponding to the lowest calibration standard, whichever was the greater. Detection limits ranged between 25 and 130 pg sample-1 for PUF disks, 25 and 50 pg sample-1 for PUF plugs (corresponding to approximately 0.04-0.08 pg m-3), 25 and 50 pg sample-1 for GFFs, and 25 and 260 pg sample-1 for SPMDs. If the concentration of a compound after blank correction was below the LOD, the sample was ignored regarding this compound in later calculations rather than assuming the concentration being zero or half of the LOD. No breakthrough was observed in the AAS for the majority of target compounds except HCB where complete breakthrough occurred and PCBs 18 and 22, which experienced breakthrough to the back-up plug of 39% and 11% (PCB 18) and 5.1% and 2.4% (PCB 22) in 25 and 73 m height, respectively. Recovery of surrogate compounds was 60-140% for PUF disks and plugs and GFFs and 52-150% for SPMDs but the range was much narrower in most instances for single com-

FIGURE 1. Height profiles of amounts of selected PCBs and OC pesticides taken up by PAS in comparison to the wind speed profile. (a) Concentrations in PUF disks and SPMDs normalized by Ci,act-average (filled circles represent individual samples; lines connect the averages of replicate samples at each height or individual samples if no replicates were taken). (b) Wind speed (full circles show average wind speed measurements during the PAS deployment period, with error bars representing the standard deviation, solid line shows estimated wind speed profile (19), used to derive wind speeds at the passive air sampling heights). pounds (refer to SI Text SI-1 for details on procedures and compounds used for recovery rate determination). Recovery statistics can be found in Table SI-3. All reported values are field blank corrected but not corrected for recovery rates.

Initial Comments on Results Air concentrations from the low volume AAS collected for the duration of the study at two heights (25 and 73 m) are presented in Table SI-4. Although values for the 73 m sample are higher for most compounds, these apparent differences are not significant (p ) 0.05). Almost identical recoveries of the 13C PCB surrogate standards in the two front PUF plugs increase confidence in this finding. This suggests that either (i) the expected gradient due to the forest filter effect (e.g., 18, 19) was too weak to be detected, or (ii) gradients, possibly in opposing directions may have existed but have been canceled out during the time-course of the sampling. In contrast to the active measurements, vertical gradients in accumulated target compounds were observed in both PAS types (Figure 1a) when expressed as ng sampler-1, normalized by the average concentration obtained from AAS

for individual compounds (Ci,act-average) and not yet corrected for approach to equilibrium (equivalent air volumes) or DC results. The observed gradient, particularly for the PUF disks, tracks the measured wind profile in Figure 1b that ranges from 0.3 to 4.5 m s-1 (see also Table SI-5) and is consistent with the wind-effect on PUF-disk sampling rate as reported by Klanova et al. (10). In contrast to the present study MoreauGuigon et al. (15) experienced rather uniform rates for the uptake of organochlorine pesticides by PUF disk based PAS at different elevations on the CN Tower in Toronto, Canada. However, this could be explained by the wind-shielding effect of the tower construction. Uptake Rate Estimation Using DCs. To investigate the potential of DCs to account for the effect of wind speed differences, sampler specific uptake rates R were calculated from the loss of these compounds during deployment utilizing eq 14 from ref 20 and assuming that (i) uptake is air-side controlled, (ii) the mass transfer coefficient of DCs through the air boundary layer is independent from their KPAS-A, (iii) CDC in the air is zero, (iv) uptake and loss mass transfer directions are opposite to each other, and (v) R equals VOL. 43, NO. 9, 2009 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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the mass transfer coefficient k multiplied by the passive sampler surface area: -ln R)

( )

corr CDC · KPAS-A · FPAS ·V CDC,0 t

(1)

with corr CDC )

CDC CDC-stable CDC-stable,0

(2)

CDC,0 and CDC are the concentrations of the DC [ng sample-1] at the beginning and the end of the deployment period, respectively. CDC values are corrected based on recoveries of an involatile (stable) DC that will not volatilize from the PAS. This correction accounts for miscellaneous sources of variability for all DCs (e.g., loss during spiking, loss of pieces of sampler material during collection, losses during the cleanup, matrix effects). KPAS-A is the chemical’s PAS-air partition coefficient with units of m3 g-1, FPAS is the PAS bulk density [g m-3], V is the volume of the PAS [m3], and t is the deployment period in days. According to the regression given by Shoeib and Harner (5) KPAS-A were calculated as: KPUF-A ) 100.6366log KOA-3.1774

(3)

KSPMD-A ) 100.8113log KOA-4.8367

(4)

where KOA is the average of temperature-adjusted KOA values calculated in 30 min steps (KOA,30min) over the whole deployment period. KOA,30min values for DCs and target compounds were calculated using temperature information at each sampling height from the relation given by Harner and Bidleman (21). As noted earlier, only DCs that experienced losses of 20-80% should be used to estimate uptake rates to minimize the influence of the analytical uncertainty (16). However, Pozo et al. (6) indicated that low DC recoveries are more useful than high ones because they allow for more accurate determinations of R as they are less subject to analytical uncertainty. The revised criteria for the use of DCs under the GAPS Network require now that R is based on results for several DCs that agree with each other and for which recoveries are less than 60%. Figure 2a shows plots of R calculated from the loss of DCs. The results meet the original DC requirement of 20-80% at most heights but fail, at all heights, the stricter requirement adopted under the GAPS Network for the PUF disk samplers. PCB-188 and 13C12 PCB180 were used as stable DCs for correcting losses of DCs in PUF disks and SPMDs, respectively. The plot of R vs height for DCs (Figure 2a), PCB-54 (PUF), and 13C12 PCB-28 (SPMDs), are of a shape similar to the uncorrected PAS profiles in Figure 1. R calculated from the loss of PCB-54 from PUF disks show a highly significant exponential correlation with wind speed (p < 0.001) (see Figure 2b and Table SI-6 for more details on KPAS-A · FPAS, the recovery of all DCs, and resulting uptake rates). R calculated from the loss of DCs from SPMDs shows no significant dependence on wind speed (see Figure 2b). The weaker wind speed dependence for SPMDs may be due to the way the SPMD is positioned/coiled in the sampling chamber, with much of the surface area further protected from direct air flow. However, the unexpectedly high losses 13 C12 PCB-101 experienced in SPMDs at some heights could not be explained. 3230

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FIGURE 2. Uptake rates R calculated from the loss of DCs: (a) against sampling height, (b) against wind speed. In general, uptake rates of PUF disks obtained from the loss of DCs up to the 25 m sampling point agree well with earlier findings (3.5-4 m3 d-1) (5, 9) but exceed them above this heightsreaching as high as 6-8 m3 s-1, consistent with wind-effect studies by Tuduri et al. (14) and results from Klanova et al. (10). Estimated uptake rates of SPMDs were on average about twice as high as those of PUF disks at all heights. Slightly higher values for SPMDs compared to PUF disks were also observed by Shoeib and Harner (5) and can be explained by their greater surface area (ASPMD/APUF ≈ 1.4). Estimation of Concentrations of POPs in Air. Uptake rates obtained from DCs can now be used to calculate concentrations of POPs in the air which can be compared to AAS results. Since the more volatile target compounds reach equilibrium (and therefore the curvilinear uptake phase) sooner than the less volatile ones, compound specific effective air volumes Veff [m3 sample-1], referring to the volume of air the PAS has effectively sampled, were estimated (13):

(

Veff ) KPAS-A · FPAS · V · 1 - exp

[

-t · R V · KPAS-A · FPAS

])

(5)

Concentrations in the air Ci,air could then be calculated as Ci,air )

Ci,PAS Veff

(6)

where Ci,PAS is the concentration of a target compound (i) in passive samples [ng sample-1]. Figure 3 shows Ci,act-average-normalized Cair of selected PCBs and OC pesticides derived from both PUF disks and SPMDs against the sampling height (full data set in Table SI-7). For the uptake rate correction of the data presented in Figure 3, the most volatile DCs were used (i.e., PCB-54 for PUF disks and 13C12 PCB-28 for SPMDs), as they experienced the highest losses during deployment and were therefore considered most accurate. Losses of other DCs were partly inconsistent or too low to be used in this calculation. However, using 13 C12 PCB-52 (SPMDs) would have resulted in similar but slightly higher uptake rates and thus lower concentration estimates. As discussed previously, it is preferable to use several DCs (in the correct volatilization range) to obtain the best estimate of R. The profile in Figure 3appears to remove the wind-effect on sampling rate and thus represent true air concentrations over the deployment period. The profile in Figure 3 lacks a strong gradient and is more consistent with results of the AAS.

FIGURE 3. Ci,act-average-normalized concentrations of selected PCBs and OC pesticides in air derived from amounts taken up by PAS, corrected for sample specific uptake rates obtained from the loss of DCs. Filled circles represent individual samples; lines connect the averages of replicate samples at each height or individual samples if no replicates were taken. †Corrected for uptake rates obtained from the loss of PCB-54. ‡Corrected for uptake rates obtained from the loss of 13C PCB-28. PAS results reflected the observation made with the AAS collected at elevations of 25 and 73 m, showing no significant concentration difference (p ) 0.05) with deployment height. However, for a few congeners (i.e., PCB-44, -52, -138, -149, and -153) concentrations obtained using PUF disk samplers were significantly higher (p ) 0.05) at sampling heights well above the canopy than close to/within the canopy. This may be explained by POPs partitioning into vegetation surfaces resulting in the so-called “forest filter effect” (17, 18, 22). To test for significance Ci,act at 25 and 73 m were compared, regarding both as average results. To account for this simplification, a coefficient of variance (COV) of 15% as reported by Gioia et al. (23) was applied. In the same way, the average of Ci,air, obtained at elevations of 2.5 to 25 m, was compared with the average of Ci,air observed at 44, 73, and 96 m, i.e., well above the canopy, but a COV of 35% has been applied to account for the variability associated with passive air sampling using DCs to determine the effective air volume (12). However, given that measurements were conducted in one sampling period only and no significant differences were found for the active sample and the SPMD based passive samples, the trend observed in PUF disk-based passive samples for some congeners should not be overinterpreted. Although DCs are able to correct for the wind effect on sampling rate, there seems to be a discrepancy/offset between the AAS and PAS-derived air concentration values for some target compounds. Air concentrations derived from AAS (Ci,act) and PAS (Ci,air) are compared in Figure 3 by plotting Ci,actnormalized Ci,air against the sampling height. For PUF disks, ratios between Ci,air and Ci,act-average were 1 for the penta- to hepta-CBs and OCPs except for HCB, with good agreement for the tetra-CBs and HCB. For SPMDs Ci,air/ Ci,act-average < 1 were observed for tri- and tetra-CBs while Ci,air agreed well with Ci,act-average for higher chlorinated congeners and OCPs. At first glance, the large offset in the AAS/PAS comparison seems to suggest that in the PAS, the low molecular weight PCBs are sampled at much lower rates compared to the high molecular weight PCBs. However, this is contrary to several field and laboratory investigations, which show no consistent and/or large bias between PUF-disk sampling rates for PCB

homologues (9, 10, 12). A more likely explanation is that the PAS and AAS are representative of a different sample volume given the different sampling characteristics of the low volume AAS and the PAS. Other studies have shown that poor agreement may arise between active and passive air samplers when air concentrations for target compounds are highly variable during the deployment period (12). In these instances, good agreement between the two approaches would require that exactly the same air masses are sampled. However, this is not the case as PAS collect more air on windy days. This may introduce a bias if ambient air concentrations also vary with wind speed, i.e., being either much higher or much lower on windier days. However, if this was the cause for the offset between AAS/PAS, the deviations should show a wind speed- and thus height-dependence as wind-effects are minimal at low heights on the tower and greater at higher elevations; however, this was not the case. Another observation is that PCBs showed unusually high AAS-derived concentrations compared to yearly averages from other background sites (24) and were highly enriched in lower molecular weight compounds, specifically the trichlorinated congeners. PCB-28 contributed only about 25% to the sum of the PCB congeners 28, 52, 101, 118, 138, 153, and 180 for yearly averages at EMEP sites (24), compared to almost 60% in the present study. Possible reasons for these discrepancies are differences in the dominating origin of air masses. A PCB concentration peak in the air in spring as reported by Backe et al. (25) and Gouin et al. (26) may also explain the observation. This spring pulse was attributed to the rising temperature that causes volatilization of compounds that have deposited to surfaces during the colder months. It may favor lighter congeners as they are more likely to have approached equilibrium between the air and top litter layer material. If passive samplers approach equilibrium for lighter compounds at the end of the sampling period, compounds taken up earlier may revolatilize from the sampler when burdens in the air decrease after such a pulse. However, as Figure SI-2a illustrates, even the tri-CBs are still far from air-PAS equilibrium after 40 days and the amounts accumulated were therefore believed to have deviated only slightly from those resulting from linear uptake VOL. 43, NO. 9, 2009 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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(Figure SI-2b), i.e., no losses due to this phenomenon were expected. For detailed information on these figures refer to the Supporting Information (Text SI-2). This discrepancy cannot be resolved here, given that only one air sample was collected. Guidance on the Use of DCs. Based on the results of this investigation and a review of previous studies where DCs were used, the following guidelines or best practices for using DCs are proposed: (a) DCs should cover the compound class of the target analytesseither being labeled counterparts and or native species (e.g., for mixtures) that do not exist in the technical formulation; (b) several DCs should be used that cover a wide range of volatilities; (c) of these, some should be in the range that would result in losses exceeding 40%, so that accurate sampling rates can be calculated. Ideally, sampling rates will be confirmed based on results for 3 or more DCs that fall in this range. Large discrepancies between DC-based R-values may signal an analytical artifact. However, even if such strict conditions can not be met, DC losses may still give valuable information about relative differences between uptake rates at different sites that can be used for semiquantitative interpretation; (d) some DCs should be in the range of volatility where they will experience negligible loss and thus allow them to be used as an internal recovery check for each sample. (e) In calibrating/testing passive samplers over long deployment periods at sites where high winds are common, it is recognized PAS may show a sampling bias toward the windiest days. It is therefore useful to characterize this variability by collecting several AAS during the deployment period as this would allow for meteorological influences on PAS sampling rates to be investigated.

(7)

(8)

(9)

(10)

(11)

(12)

(13)

(14) (15)

(16)

Acknowledgments We are grateful to the European Union Framework-VI project AQUATERRA for funding this research, to the Marie Curie Research Training Networks Programme under the European Commission for financially supporting L.N. and to the technicians at the Norunda research site, particularly Anders Båth for technical support.

Supporting Information Available Details about experimental methods, meteorological conditions, the full concentration data set for PAS (pg sampler-1), DC recovery and derived uptake rates, concentrations in the air obtained from AAS and PAS and further details supporting the comparison between these results. This material is available free of charge via the Internet at http://pubs.acs.org.

Literature Cited (1) UNEP. Final Act of the Conference of Plenipotentiaries on The Stockholm Convention On Persistent Organic Pollutants; United Nations Environment Program: Geneva, Switzerland, 2001. (2) Wania, F. Assessing the potential of persistent chemicals for long-range transport and accumulation in polar regions. Environ. Sci. Technol. 2003, 37, 1344–1351. (3) Wania, F.; Mackay, D. Global fractionation and cold condensation of low volatility organochlorine compounds in polarregions. Ambio 1993, 22, 10–18. (4) Gioia, R.; Steinnes, E.; Thomas, G. O.; Meijer, S. N.; Jones, K. C. Persistent organic pollutants in European background air: derivation of temporal and latitudinal trends. J. Environ. Monit. 2006, 8, 700–710. (5) Shoeib, M.; Harner, T. Characterization and comparison of three passive air samplers for persistent organic pollutants. Environ. Sci. Technol. 2002, 36, 4142–4151. (6) Pozo, K.; Harner, T.; Wania, F.; Muir, D. C. G.; Jones, K. C.; Barrie, L. A. Toward a global network for persistent organic

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(17) (18) (19) (20) (21) (22)

(23)

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