- Email: [email protected]
Field-based evaluation of semipermeable membrane devices (SPMDs) as passive air samplers of polyaromatic hydrocarbons (PAHs)
ARTICLE IN PRESS
Atmospheric Environment 38 (2004) 5983–5990 www.elsevier.com/locate/atmosenv
Field-based evaluation of semipermeable membrane devices (SPMDs) as passive air samplers of polyaromatic hydrocarbons (PAHs) Michael E. Bartkowa,, James N. Huckinsb, Jochen F. Mu¨llera a
National Research Centre for Environmental Toxicology (ENTOX), University of Queensland, 39 Kessels Rd, Coopers Plains, Queensland 4108, Australia b Columbia Environmental Research Centre (CERC), US Geological Survey, 4200 New Haven Road, MO 65201, USA Received 4 May 2004; accepted 29 June 2004
Abstract Semipermeable membrane devices (SPMDs) have been used as passive air samplers of semivolatile organic compounds in a range of studies. However, due to a lack of calibration data for polyaromatic hydrocarbons (PAHs), SPMD data have not been used to estimate air concentrations of target PAHs. In this study, SPMDs were deployed for 32 days at two sites in a major metropolitan area in Australia. High-volume active sampling systems (HiVol) were codeployed at both sites. Using the HiVol air concentration data from one site, SPMD sampling rates were measured for 12 US EPA Priority Pollutant PAHs and then these values were used to determine air concentrations at the second site from SPMD concentrations. Air concentrations were also measured at the second site with co-deployed HiVols to validate the SPMD results. PAHs mostly associated with the vapour phase (Fluorene to Pyrene) dominated both the HiVol and passive air samples. Reproducibility between replicate passive samplers was satisfactory (CVo20%) for the majority of compounds. Sampling rates ranged between 0.6 and 6.1 m3 d1. SPMD-based air concentrations were calculated at the second site for each compound using these sampling rates and the differences between SPMD-derived air concentrations and those measured using a HiVol were, on average, within a factor of 1.5. The dominant processes for the uptake of PAHs by SPMDs were also assessed. Using the SPMD method described herein, estimates of particulate sorbed airborne PAHs with five rings or greater were within 1.8-fold of HiVol measured values. r 2004 Elsevier Ltd. All rights reserved. Keywords: Semivolatile organic compounds; Atmospheric sampling; Sampling rates
1. Introduction Polyaromatic hydrocarbons (PAHs) are potentially hazardous, semivolatile organic pollutants that enter the atmosphere from anthropogenic sources such as motor Corresponding author. Tel.: +61-7-3274-9147; fax: +61-73274-9003. E-mail address: [email protected] (M.E. Bartkow).
vehicle emissions and from natural sources such as bush fires (Smith and Harrison, 1996). Atmospheric transport of PAHs can result in the bioaccumulation of these compounds from the air into plants such as crops (Smith et al., 2001). Intake of PAHs from the diet is considered an important route of exposure in humans (Bostrom et al., 2002). Despite the importance of the atmospheric pathway for human exposure, atmospheric monitoring of PAHs
1352-2310/$ - see front matter r 2004 Elsevier Ltd. All rights reserved. doi:10.1016/j.atmosenv.2004.06.036
ARTICLE IN PRESS M.E. Bartkow et al. / Atmospheric Environment 38 (2004) 5983–5990
5984
is not routinely performed. This is mainly due to the requirements of expensive HiVol sampling equipment, which is reliant on continuous field maintenance and electricity to operate. To overcome some of the limitations of HiVol sampling, passive samplers such as semipermeable membrane devices (SPMDs) can be used to monitor semivolatile organic pollutants in the atmosphere. SPMDs do not require electricity or expensive equipment to operate in the field but must be deployed inside a protective chamber to minimize the effects of sunlight, rain and wind on sampler performance. SPMDs were introduced by Huckins et al. (1990) as a passive water sampler, but have since been used to sample both indoor and outdoor atmospheres (Petty et al., 1993; Prest et al., 1995; Strandberg et al., 1997; Ockenden et al., 1998a, b; Rantalainen et al., 1999; Caslavsky et al., 2000; Booij and van Drooge, 2001; Lohmann et al., 2001; Meijer et al., 2003; Soderstrom and Bergqvist, 2003). However, only limited data on sampling rates are available. SPMDs can be used to identify sources of pollutants and measure time-integrated air concentrations of pollutants. One way to measure air concentrations using SPMDs is to determine the rate at which pollutants accumulate in the sampler. Where uptake is linear with time and air concentrations and temperature is assumed to remain constant, sampling rates can be calculated for the target compounds according to the following: RS-cal ¼
N SPMD ; CVt
N SPMD : RS-cal t
C SPMD ¼ C O eðke tÞ ;
ð3Þ
where CO is the concentration of the PRC in the SPMD prior to deployment and ke is the elimination rate constant (d1). To our knowledge, only sampling rates for the uptake of airborne PCBs into SPMDs have been reported in the literature (Ockenden et al., 1998a; Shoeib and Harner, 2002). We used a field-based study to determine SPMD sampling rates for PAHs and to evaluate the use of PRCs. An outdoor calibration site was used to determine sampling rates. These data were then used to estimate air concentrations of target compounds for a second site where SPMDs were also deployed. Air concentrations estimated from the SPMD data were then compared with HiVol results obtained over the same sampling period.
ð1Þ
where RS-cal is the sampling rate (m3 d1) measured at the calibration site, NSPMD is the amount of compound accumulated in the sampler (ng), CV is the concentration of that compound in the vapour phase (ng m3) at the calibration site and t is time of exposure (days). A sampling rate can then be used to derive vapour phase air concentrations using identical SPMDs deployed at other sites, according to CV ¼
atmosphere (Ockenden et al., 2001). A known amount of PRCs are loaded into the SPMDs prior to deployment and the elimination rate constant of these compounds for the duration of the exposure is determined for each site. This in situ calibration approach is based on the theory that the elimination rate constants of PRCs and uptake rate constants of target compounds at a given site and are related. If the PRCs used in the SPMD can be assumed to be not present in the atmosphere then loss from the sampler can be described using a first-order decay kinetic equation (Huckins et al., 2002)
ð2Þ
Passive sampling theory and experimental evidence shows that the sampling rate can be influenced by environmental factors such as wind speed if uptake is air-side limited (Harner et al., 2003). Therefore, it is important that site conditions are similar or that sampling rates can be corrected for different environmental conditions between sites. Huckins et al. (2002) have shown the use of performance reference compounds (PRCs) can be useful in correcting for differences in exposure conditions where SPMDs are deployed in the water. PRCs are analytically non-interfering compounds which are similar to the target compounds. Theoretically, the same approach can be taken using PRCs in samplers deployed in the
2. Materials and methods 2.1. Air sampling Each SPMD was made of 82 cm 2.8 cm of lay-flat low-density polyethylene (LDPE) tubing (Brentwood Plastics, MO), with a membrane thickness of about 50 mm and an overall surface area of 460 cm2. The LDPE tubing of each SPMD had been pre-extracted in 180 mL of redistilled hexane for 24 hours followed by another 24-hour extraction period in fresh redistilled hexane. The tube was then filled with 1 mL of triolein, which had been previously spiked with a known amount of two PRCs (2D10-anthracene and 2D10-pyrene). SPMDs (including field blanks) were wrapped in aluminium foil immediately after fabrication and stored at 17 1C until deployment. The preparation and operation of the HiVol system is detailed in previous work (Bartkow et al., 2004). SPMDs and HiVols were deployed at two Queensland Environmental Protection Agency (QEPA) Ambient Air Monitoring Stations in metropolitan Brisbane, Queensland, Australia. One site was located at South Brisbane, adjacent to the M1 motorway, approximately 2 km south of the CBD. The other study site was situated at
ARTICLE IN PRESS M.E. Bartkow et al. / Atmospheric Environment 38 (2004) 5983–5990
Springwood, a predominantly residential suburb, located 20 km south east of the Brisbane CBD. The station itself was located on the grounds of a school, approximately 20 m from a 4-lane roadway that services residential areas. Samplers were deployed at South Brisbane (calibration site) and Springwood in April 2002 for 32 days. Mean ambient air temperature for the duration of the deployment was 22 1C at both sites (QEPA, 2003). Wind speed was also measured at both sites by QEPA at 10 m height (see Results and discussion). A HiVol and four replicate SPMDs were deployed at each site. The SPMDs were deployed 1.5 m above the ground in galvanized iron chambers with an open bottom and louvers on all sides. The chambers were designed to ensure adequate air flow while minimizing the effect of direct sunlight, rain and wind on sampler performance and were similar to those used by Ockenden et al. (1998b). SPMD field blanks were taken into the field and then frozen until extraction. Duplicate HiVols were also deployed at a separate test site to quantify reproducibility for these systems. SPMDs and XAD cartridges were collected after 32 days of exposure. Glass fibre filters were regularly inspected and collected every 3 or 4 days, depending on the amount of accumulated particulate material. Passive samplers and filters were wrapped in foil on-site and the XAD cartridges were sealed with ground glass stoppers. Samples were transported to the laboratory and stored at 17 1C. 2.2. Analysis To investigate the potential for SPMD sampling of particle sorbed PAHs as well as vapour phase PAHs, the external surface of the SPMD was not cleaned prior to analyte extraction. For quantification, all samples were spiked with an internal standard mix containing known amounts of 10 deuterated PAHs (2D10-acenaphthene, 2D10-fluorene, 2D10-phenanthrene, 2 D10-fluoranthene, 2D12-benz[a]anthracene, 2D12-chrysene, 2D12-benzo[b]fluoranthene, 2D12-benzo[a]pyrene, 2 D12-indeno[1,2,3-c,d]pyrene and 2D12-benzo[g,h,i]perylene). The internal standard was spiked directly into the triolein of the SPMDs. The SPMDs were resealed and then extracted in the dark using 180 mL of redistilled hexane at room temperature (22 1C). After 24 hours the solvent was refreshed and the samplers extracted for a further 24 hours. The combined extracts from each sampler were reduced to about 1 mL using rotary evaporators. SPMD extracts were transferred into DCM and subjected to clean-up using gel-permeation chromatography (19 mm 150 mm guard column, followed by a 19 mm 300 mm main column, packed with Envirogel (100 A˚ pore size, 15 mm particle size, Waters) as the stationary phase and with DCM as the
5985
mobile phase). The flow rate was 4.5 mL min1 and the sample collected between 14 and 24 min. XAD cartridges and filter papers were soxhlet extracted separately for 10 h in toluene. All samples were subsequently cleaned using adsorption chromatography (refer Bartkow et al., 2004). The following PAHs were routinely quantified: Naphthalene (Nap), acenaphthylene (Acy), acenaphthene (Ace), fluorene (Flo), phenanthrene (Phe), anthracene (Ant), fluoranthene (Flu), pyrene (Pyr), benz[a]anthracene (B[a]A), chrysene (Chr), benzo[b+k]fluoranthene (B[b,k]F), benzo[e]pyrene (B[e]P), benzo[a]pyrene (B[a]P), perylene (Per), indeno[1,2,3-c,d]pyrene (I[c,d]P), dibenz[a,h]anthracene (D[a,h]A) and benzo[g,h,i]perylene (B[g,h,i]P). For details on the separation and quantification of the PAHs refer to Bartkow et al. (2004). Passive sampler results were reported as the total amount of each PAH in an SPMD (ng). HiVol results were reported as vapour phase air concentrations (CV) or particle phase air concentrations (CP) (ng m3). In some cases, the vapour phase concentration of the high molecular weight PAHs was very close to the detection limits, which increased the uncertainty associated with the values. 2.3. QA/QC The quantification criteria included confirmation of the retention times of the labeled standard and respective analyte. Routinely, the mass fragment with the highest intensity (base peak) was used for quantification. QC samples (field blanks, solvent blanks and spikes) constituted 410% of the total number of samples. SPMD field blanks were taken into the field when samplers were deployed and collected and were extracted and analysed in the same manner as deployed samples. The sample detection limits for individual compounds in a HiVol sample were defined as three times the analyte concentration measured in the field blank. Detection limits for the SPMD samples were calculated as the mean amount in the field blanks plus three times the standard deviation (Refer Table A in Supplementary Data). Where a compound could not be identified in the blank, the detection limit was set as three times the average noise range. Recoveries of the internal standards ranged between 93–122% for the HiVol samples and 45–105% for the SPMDs. Typically, the surfaces of the exposed SPMDs are cleaned prior to extraction, to ensure that any particulates in the lipid-associated surficial film present on SPMDs at the end of exposures do not inhibit dialytic recoveries of the target compounds (Huckins et al., 1999). The surfaces of the SPMDs were not cleaned and this could explain why the recoveries reported from the SPMDs were lower than for the HiVol samples.
ARTICLE IN PRESS M.E. Bartkow et al. / Atmospheric Environment 38 (2004) 5983–5990
5986
Table 2 Atmospheric concentrations of PAHs as determined by HiVol air sampling
3. Results and discussion 3.1. Sampler reproducibility Reproducibility was determined using coefficients of variation (expressed as %CV) for SPMDs (replicates=4) and using normalized differences for HiVol samples (replicates=2). The %-normalized differences were calculated for replicates A and B according to value A value B 100: ðvalue A þ value BÞ=2 Due to their relatively high volatility, reproducibility for Nap, Acy and Ace was consistently poor and results for these compounds were not included in any further analysis. Reproducibility between SPMDs for the remaining compounds was good, with most CVs o20%, except for Flo (34%) and B[a]P (41%) (Table 1). For the HiVol sampler data, the mean normalized difference only exceeded 25% for Flu (27%).
PAH
South Brisbane
Springwood
CV (ng m3) CP (ng m3) CV (ng m3) CP (ng m3) Flo Phe Ant Flu Pyr B[a]A Chr B[b,k]F B[e]P B[a]P Per I[c,d]P D[a,h]A B[g,h,i]P
1.96 4.80 0.82 2.86 3.05 0.19 0.24 0.03 0.01 N.D. N.D. 0.004 N.D. 0.009
0.20 0.09 0.03 0.11 0.18 0.32 0.22 0.57 0.25 0.12 0.02 0.27 0.05 1.05
1.51 2.53 0.31 0.86 0.76 0.04 0.06 0.02 0.01 N.D. N.D. N.D. N.D. 0.002
N.D. 0.09 0.04 0.02 0.03 0.03 0.03 0.10 0.04 0.02 0.00 0.05 0.01 0.16
N.D: not detected.
3.2. HiVol samples Based on total residues per unit volume (i.e., CV+CP), the atmospheric concentrations of PAHs were generally higher at South Brisbane than at Springwood (Table 2). This was not surprising considering the proximity of the South Brisbane site to the city centre and major roadways. The compounds Flo through Pyr were predominantly associated with the vapour phase (490% measured on XAD only), while
Table 1 Mean amounts of PAHs sampled by SPMDs and coefficients of variation (CV) (n=4) PAH
Flo Phe Ant Flu Pyr B[a]A Chr B[b,k]F B[e]P B[a]P Per I[c,d]P D[a,h]A B[g,h,i]P
3.3. SPMDs
Springwooda
South Brisbane
the compounds B[b,k]F through B[g,h,i]P were predominantly associated with particles. PAH profiles from each site were dominated by the compounds most associated with the vapour phase (Flo to Pyr). Phe was present at the highest concentrations with Per and D[a,h]A at the lowest concentrations. The concentration of PAHs in the atmosphere was generally much lower during this sampling campaign than measured by Mu¨ller et al. (1998) in the mid-1990s. In particular, the concentration of B[a]P at Springwood and South Brisbane was generally 10 times lower than B[a]P concentrations at similar sites, 8 years ago (Mu¨ller et al., 1998).
NSPMD (ng)
CV (%)
NSPMD (ng)
CV (%)
110 660 100 410 190 25 45 40 10 2 N.D. 15 N.D. 25
34 13 18 6 15 10 3 5 9 41 — 10 — 13
50 175 35 60 25 3 7 6 2 N.D. N.D. 2 N.D. 2
9 4 19 4 20 17 9 16 — — — — — —
N.D: not detected. a B[e]P, I[c,d]P and B[g,h,i]P only detected in a single SPMD.
Mean amounts of PAHs sampled by the SPMDs are presented in Table 1. PAHs associated with the vapour phase dominated the samples, with Phe contributing to approximately 40% of the total amount of the PAHs quantified. After 32 days of exposure, those PAHs that were present in the atmosphere at low total concentrations (CV+CPp50 pg m3) were not always detectable in a single SPMD. Per and D[a,h]A were not detected in SPMDs from either site, while at Springwood, B[a]P was not detected in any samplers and B[e]P, I[c,d]P and B[g,h,i]P were only detected in one of the four SPMDs. The detection limits could be improved by compositing SPMDs or by deploying the SPMDs for a longer time. Interestingly, compounds such as B[b,k]F, B[e]P, B[a]P, I[c,d]P and B[g,h,i]P that were more associated with particles were routinely detected in SPMDs from South Brisbane.
ARTICLE IN PRESS M.E. Bartkow et al. / Atmospheric Environment 38 (2004) 5983–5990
3.4. SPMD-derived air concentrations Sampling rates were calculated according to modeling theory, using the vapour phase air concentrations for most PAHs (Eq. (1)). However, the sampling rates for the five largest compounds (B[b,k]F to B[g,h,i]P) which were mostly bound to particles, were calculated using the total air concentration (CV+CP). This approach was taken because the vapour phase air concentrations were close to quantification limits and particulate sorbed compounds on the exterior surface of the SPMD were quantified as part of the amount accumulated by SPMDs. Uptake was assumed to be linear for all compounds during the exposure; however, a relatively low sampling rate for Flo (1.8 m3 d1) indicated that uptake for that compound may have reached the curvilinear phase. Sampling rates ranged between 0.6 and 6.1 m3 d1 (Table 3). Based on the two-film resistance model, if airside resistance dominates uptake then sampling rates are not directly related to KOA, whereas for sampler-side resistance, sampling rates would increase, proportionally, with KOA (Harner et al., 2003). Our data show similar sampling rates for all PAHs whereas KOA values vary by as much as six orders of magnitude for the compounds of interest. Hence, we can assume air-side resistance dominates uptake for most PAHs under the present conditions, although sampler-side resistance is likely to have contributed to uptake of the smaller PAHs. Sampling rates tended to be highest for the 3- and 4ring PAHs (1.8–6.1 m3 d1), whereas rates for the 5- and 6-ring PAHs were p2.1 m3 d1. Although there are no other directly measured sampling rates for PAHs reported in the literature, data for PCBs are available for comparison. Shoeib and Harner (2002) deployed SPMDs (surface area of 495 cm2) indoors and reported
Table 3 Mean SPMD sampling rates for PAHs at South Brisbane PAH
RS-cal (m3 d1)a
Flo Phe Ant Flu Pyr B[a]A Chr B[b,k]F B[e]P B[a]P I[c,d]P B[g,h,i]P
1.8 4.4 4.1 4.5 2.0 4.1 6.1 2.1 1.3 0.6 1.7 0.7
a
Sampling rates B[b,k]F to B[g,h,i]P calculated using CV+CP.
5987
sampling rates between 3.3 and 9.9 m3 d1. Although comparability between the two studies was limited due to differences in exposure conditions and the surface area of the SPMDs, the sampling rates measured in this study for the 3- and 4-ring PAHs compare closely with the PCB sampling rates. However, the sampling rates for the higher molecular weight PAHs appeared to be significantly lower compared to rates for the higher molecular weight PCBs. Mean elimination rate constants for each PRC were calculated from the SPMDs deployed at South Brisbane and Springwood using Eq. (3). Time zero concentrations for the first-order loss of PRCs were derived from unopened field blanks. Mean elimination rate constants were very similar (South Brisbane: 2D10-Ant: 0.090 d1 and 2D10-Pyr: 0.094 d1 and Springwood: 2D10-Ant: 0.091 d1 and 2D10-Pyr: 0.085 d1). These results were surprising for several reasons. Firstly, elimination rate constants for 2D10-Ant were similar to the elimination rate constants for 2D10-Pyr. However, even when airside resistance dominates exchange rates, elimination rate constants should be negatively related to KOA because ke ¼
kV A ; K SPMD-V V SPMD
ð4Þ
where kV is the mass transfer coefficient in the air boundary layer, A is the area of the exchanging surface, KSPMD-V is the SPMD-air partition coefficient, and VSPMD is the volume of the whole SPMD. Evidence suggests that KOA is EKSPMD-V, for compounds which are predominately in the vapour phase (unpublished data, USGS, Columbia, MO, USA). Secondly, according to t90%eq: ¼ 2:3=ke ; the magnitude of the elimination rate constants indicated that 90% of equilibrium concentrations were reached by the parallel native compounds. However, in other studies using SPMDs, workers have shown that accumulated PCBs with similar and higher KOA values were not within 90% of equilibrium (Shoeib and Harner, 2002). Furthermore, the influence of wind should not have contributed to the high loss rates because external (outside the exposure chambers) mean wind speeds were not very high (1.7 and 0.8 m s1 at South Brisbane and Springwood, respectively (QEPA, 2003)) and the deployment chamber should have minimized the effect of wind speed on sampler performance. A probable explanation for the lack of significant difference between the ke values for Ant and Pyr and also the overall high loss rates is that the sampling chambers did not exclude enough indirect sunlight, which resulted in photodegradation of the PRCs. Recently, trials undertaken at USGS have shown that significant photolysis of PAHs and polybrominated diphenyl ethers can occur in minutes to hours if SPMDs are not adequately shielded (C. Orazio, USGS, personal communication). If
ARTICLE IN PRESS M.E. Bartkow et al. / Atmospheric Environment 38 (2004) 5983–5990
5988
To estimate air concentrations at Springwood we used the unadjusted sampling rates measured at South Brisbane. The differences between SPMD-derived air concentrations and those measured using the HiVol system were within a factor of 2 for all compounds (Table 4). One interesting aspect of these data is the good agreement between predicted and measured air concentrations of PAHs mostly associated with the particulate phase. This finding elicits several questions. These include: (1) how much does PAH sorption to airborne particulates attenuate photolysis; (2) to what extent does particle deposition on the exterior surface of SPMDs contribute to the accumulation of particulate sorbed PAHs using our analytical methods; (3) is the mass of particles deposited on the SPMD membrane surface during a specified time interval approximately equivalent to the mass of particles present in the volume of air sampled by an SPMD during the same time interval; and (4) if the surficial lipid film is removed before dialysis, do particulate-associated PAHs desorb during an exposure and accumulate in the SPMD membrane and internal lipid phase? With the exception of question 2 the answers to these questions await further research. McLachlan (1999) developed a framework to identify the dominant uptake processes for SOCs into vegetation and it has been successfully applied to a range of SOCs (Bo¨hme et al., 1999; Mu¨ller et al., 2001). McLachlan makes use of two plots where he differentiates between three segments showing which process dominates uptake; (1) equilibrium partitioning of the vapour phase compounds; (2) kinetically limited vapour phase deposition; and (3) particle-bound deposition.
a significant portion of the PRCs were photodegraded, then the native PAHs may also have been affected. In future work, a photolysis sensitive marker compound with known degradation rates and very low SPMD fugacity could be added to SPMDs in order to gauge any effects of photodegradation on the target analytes. However, in this study our sampling rates do not account for any effects of photodegradation.
Table 4 Mean SPMD-derived PAH air concentrations and co-deployed HiVol-derived air concentrations at Springwood and the factor difference between the two measurements PAH
SPMD-based High volume- Factor difference CV or CV+CP based CV or CV+CP (ng m3)a (ng m3)a
Flo Phe Ant Flu Pyr B[a]A Chr B[b,k]F B[e]P I[c,d]P B[g,h,i]P
0.98 1.36 0.31 0.46 0.48 0.03 0.04 0.11 0.05 0.04 0.09
1.51 2.53 0.31 0.86 0.76 0.04 0.06 0.13 0.05 0.05 0.16
1.5 1.9 1.0 1.9 1.6 1.5 1.6 1.2 1.0 1.4 1.8
a Air concentrations are presumed to be representative of the vapour phase (CV) only for PAHs, Flo to Chr and total air concentration (CV+CP) for B[b,k]F to B[g,h,i]P.
10.0
I[c,d ]P
9.5
B[b,k ]F B[e ]P
9.0
log (CSPMD/CV)
8.5
B[a ]P
B[g,h,i ]P
Chr Flu
Phe
8.0
Ant
Flo
B[a ]A
Pyr
7.5 7.0
Kinetically limited gaseous deposition
6.5
Particle-bound deposition
6.0 5.5 5.0 6
7
8
9
10
11
12
13
log K OA
Fig. 1. Plot of log (CSPMD/CV) vs. log KOA (taken or predicted from Beyer et al. (2002)) for each of the PAHs quantified in the SPMDs at South Brisbane. For the vapour phase air concentration of B[a]P and I[c,d]P we used the detection limit of 0.003 and 0.001 ng m3, respectively.
ARTICLE IN PRESS M.E. Bartkow et al. / Atmospheric Environment 38 (2004) 5983–5990
We have used one of these plots to elucidate which chemicals are predominantly sorbed on particulates, which appear to be deposited onto the surface of SPMDs. Fig. 1 shows the log ratio of the concentration of each PAH in the SPMD sample and vapour phase versus the respective log KOA for each compound. The dominant form of uptake for the compounds Flo through Chr is shown to be via kinetically limited deposition from the vapour phase, whereas uptake of the compounds B[b,k]F to B[g,h,i]P are shown in the segment where particle deposition dominates. This result is not surprising for two reasons, firstly, the larger PAHs were mostly particle-bound upon contact with the surface of the sampler, and, secondly, the SPMDs had developed an oily surficial film which accumulated particles, and was not removed prior to extraction. Separate quantification of this ‘surface fraction’ could further assist in understanding the contribution of particles to the total amount of PAHs accumulated by SPMDs. In conclusion, this field-based study shows that SPMDs can be used to estimate air concentrations of PAHs with reasonable accuracy. Deployment times can be lengthened to ensure that compounds present at relatively low concentrations can also be detected. Alternatively, individual SPMD samples from the same site can be combined to enhance detection limits. In order to account for any effects of photolysis, a compound with a high KOA (i.e., low SPMD fugacity), which is known to be very sensitive to photolysis, could be used as a marker for potential photodegradation of analytes accumulated in SPMDs. Experiments aimed at understanding the rate at which particles are deposited onto SPMDs would also be useful.
Acknowledgements The authors thank Darryl Hawker for comments on the manuscript, Ralph Riese, Mike King and Queensland EPA for access to monitoring stations and environmental data and Karen Kennedy and Neil Holling for assistance with lab work and analysis. This work and Ph.D. program was funded by an ARC SPIRT Linkage Grant, with industry support from Queensland EPA, Queensland Health Scientific Services and ERGO. MB receives an APAI scholarship. Queensland Health provides funding for The National Research Centre for Environmental Toxicology.
Appendix A. Supplementary data The online version of this article contains additional supplementary data. Please visit doi:10.1016/ j.atmosenv.2004.06.036.
5989
References Bartkow, M.E., Hawker, D.W., Kennedy, K.E., Mu¨ller, J.F., 2004. Characterizing uptake kinetics of PAHs from the air using polyethylene-based passive air samplers of multiple surface area-to-volume ratios. Environmental Science and Technology 38, 2701–2706. Beyer, A., Wania, F., Gouin, T., Mackay, D., Matthies, M., 2002. Selecting internally consistent physicochemical properties of organic compounds. Environmental Toxicology and Chemistry 21, 941–953. Bo¨hme, F., Welsch, P.K., McLachlan, M.S., 1999. Uptake of airborne semivolatile organic compounds in agricultural plants: field measurements of interspecies variability. Environmental Science and Technology 33, 1805–1813. Booij, K., van Drooge, B.L., 2001. Polychlorinated biphenyls and hexachlorobenzene in atmosphere, sea-surface microlayer, and water measured with semi-permeable membrane devices (SPMDs). Chemosphere 44, 91–98. Bostrom, C.E., Gerde, P., Hanberg, A., Jernstrom, B., Johansson, C., Kyrklund, T., Rannug, A., Tornqvist, M., Victorin, K., Westerholm, R., 2002. Cancer risk assessment, indicators, and guidelines for polycyclic aromatic hydrocarbons in the ambient air. Environmental Health Perspectives 110, 451–489. Caslavsky, J., Zdrahal, Z., Vytopilova, M., 2000. Application of SPMDs for PAH sampling in the Deza chemical factory. Polycyclic Aromatic Compounds 20, 123–141. Harner, T., Farrar, N.J., Shoeib, M., Jones, K.C., Gobas, F.A.P.C., 2003. Characterization of polymer-coated glass as a passive air sampler for persistent organic pollutants. Environmental Science and Technology 37, 2486–2493. Huckins, J.N., Tubergen, M.W., Manuweera, G.K., 1990. Semipermeable membrane devices containing model lipid: a new approach to monitoring the bioavailability of lipophilic contaminants and estimating their bioconcentration potential. Chemosphere 20, 533–552. Huckins, J.N., Petty, J.D., Orazio, C.E., Lebo, J.A., Clark, R.C., Gibson, V.L., Gala, W.R., Echols, K.R., 1999. Determination of uptake kinetics (sampling rates) by lipidcontaining semipermeable membrane devices (SPMDs) for polycyclic aromatic hydrocarbons (PAHs) in water. Environmental Science and Technology 33, 3918–3923. Huckins, J.N., Petty, J.D., Lebo, J.A., Almeida, F.V., Booij, K., Alvarez, D.A., Cranor, W.L., Clark, R.C., Mogensen, B.B., 2002. Development of the permeability/performance reference compound approach for in situ calibration of semipermeable membrane devices. Environmental Science and Technology 36, 85–91. Lohmann, R., Corrigan, B.P., Howsam, M., Jones, K.C., Ockenden, W.A., 2001. Further developments in the use of semipermeable membrane devices (SPMDs) as passive air samplers for persistent organic pollutants: field application in a spatial survey of PCDD/Fs and PAHs. Environmental Science and Technology 35, 2576–2582. McLachlan, M.S., 1999. Framework for the interpretation of measurements of SOCs in plants. Environmental Science and Technology 33, 1799–1804. Meijer, S.N., Ockenden, W.A., Steinnes, E., Corrigan, B.P., Jones, K.C., 2003. Spatial and temporal trends of POPs in Norwegian and UK background air: implications for
ARTICLE IN PRESS 5990
M.E. Bartkow et al. / Atmospheric Environment 38 (2004) 5983–5990
global cycling. Environmental Science and Technology 37, 454–461. Mu¨ller, J.F., Hawker, D.W., Connell, D.W., 1998. Polycyclic aromatic hydrocarbons in the atmospheric environment of Brisbane, Australia. Chemosphere 37, 1369–1383. Mu¨ller, J.F., Hawker, D.W., McLachlan, M.S., Connell, D.W., 2001. PAHs, PCDD/Fs, PCBs and HCB in leaves from Brisbane, Australia. Chemosphere 43, 507–515. Ockenden, W.A., Prest, H.F., Thomas, G.O., Sweetman, A., Jones, K.C., 1998a. Passive air sampling of PCBs: field calculation of atmospheric sampling rates by trioleincontaining semipermeable membrane devices. Environmental Science and Technology 32, 1538–1543. Ockenden, W.A., Sweetman, A.J., Prest, H.F., Steinnes, E., Jones, K.C., 1998b. Toward an understanding of the global atmospheric distribution of persistent organic pollutants: the use of semipermeable membrane devices as timeintegrated passive samplers. Environmental Science and Technology 32, 2795–2803. Ockenden, W.A., Corrigan, B.P., Howsam, M., Jones, K.C., 2001. Further developments in the use of semipermeable membrane devices as passive air samplers: application to PCBs. Environmental Science and Technology 35, 4536–4543. Petty, J.D., Huckins, J.N., Zajicek, J.L., 1993. Application of semipermeable membrane devices (SPMDs) as passive air samplers. Chemosphere 27, 1609–1624. Prest, H.F., Jacobson, L.A., Huckins, J.N., 1995. Passive sampling of water and coastal air via semipermeable membrane devices. Chemosphere 30, 1351–1361.
QEPA, 2003. South East Queensland Ambient Air Quality Monitoring Program. Air Services, Environmental Sciences Division, Queensland Environmental Protection Agency 2003, electronic database. Rantalainen, A., Hyotylainen, T., Saramo, M., Niskanen, I., 1999. Passive sampling of PAHs in indoor air in Nepal. Toxicological and Environmental Chemistry 68, 335–348. Shoeib, M., Harner, T., 2002. Characterization and comparison of three passive air samplers for persistent organic pollutants. Environmental Science and Technology 36, 4142–4151. Smith, D.J.T., Harrison, R.M., 1996. Concentrations, trends and vehicle source profile of polynuclear aromatic hydrocarbons in the UK atmosphere. Atmospheric Environment 30, 2513–2525. Smith, K.E.C., Thomas, G.O., Jones, K.C., 2001. Seasonal and species differences in the air-pasture transfer of PAHs. Environmental Science and Technology 35, 2156–2165. Soderstrom, H.S., Bergqvist, P.A., 2003. Polycyclic aromatic hydrocarbons in a semiaquatic plant and semipermeable membrane devices exposed to air in Thailand. Environmental Science and Technology 37, 47–52. Strandberg, B., Wagman, N., Bergqvist, P.A., Haglund, P., Rappe, C., 1997. Semipermeable membrane devices as passive samplers to determine organochlorine pollutants in compost. Environmental Science and Technology 31, 2960–2965.
Atmospheric Environment 38 (2004) 5983–5990 www.elsevier.com/locate/atmosenv
Field-based evaluation of semipermeable membrane devices (SPMDs) as passive air samplers of polyaromatic hydrocarbons (PAHs) Michael E. Bartkowa,, James N. Huckinsb, Jochen F. Mu¨llera a
National Research Centre for Environmental Toxicology (ENTOX), University of Queensland, 39 Kessels Rd, Coopers Plains, Queensland 4108, Australia b Columbia Environmental Research Centre (CERC), US Geological Survey, 4200 New Haven Road, MO 65201, USA Received 4 May 2004; accepted 29 June 2004
Abstract Semipermeable membrane devices (SPMDs) have been used as passive air samplers of semivolatile organic compounds in a range of studies. However, due to a lack of calibration data for polyaromatic hydrocarbons (PAHs), SPMD data have not been used to estimate air concentrations of target PAHs. In this study, SPMDs were deployed for 32 days at two sites in a major metropolitan area in Australia. High-volume active sampling systems (HiVol) were codeployed at both sites. Using the HiVol air concentration data from one site, SPMD sampling rates were measured for 12 US EPA Priority Pollutant PAHs and then these values were used to determine air concentrations at the second site from SPMD concentrations. Air concentrations were also measured at the second site with co-deployed HiVols to validate the SPMD results. PAHs mostly associated with the vapour phase (Fluorene to Pyrene) dominated both the HiVol and passive air samples. Reproducibility between replicate passive samplers was satisfactory (CVo20%) for the majority of compounds. Sampling rates ranged between 0.6 and 6.1 m3 d1. SPMD-based air concentrations were calculated at the second site for each compound using these sampling rates and the differences between SPMD-derived air concentrations and those measured using a HiVol were, on average, within a factor of 1.5. The dominant processes for the uptake of PAHs by SPMDs were also assessed. Using the SPMD method described herein, estimates of particulate sorbed airborne PAHs with five rings or greater were within 1.8-fold of HiVol measured values. r 2004 Elsevier Ltd. All rights reserved. Keywords: Semivolatile organic compounds; Atmospheric sampling; Sampling rates
1. Introduction Polyaromatic hydrocarbons (PAHs) are potentially hazardous, semivolatile organic pollutants that enter the atmosphere from anthropogenic sources such as motor Corresponding author. Tel.: +61-7-3274-9147; fax: +61-73274-9003. E-mail address: [email protected] (M.E. Bartkow).
vehicle emissions and from natural sources such as bush fires (Smith and Harrison, 1996). Atmospheric transport of PAHs can result in the bioaccumulation of these compounds from the air into plants such as crops (Smith et al., 2001). Intake of PAHs from the diet is considered an important route of exposure in humans (Bostrom et al., 2002). Despite the importance of the atmospheric pathway for human exposure, atmospheric monitoring of PAHs
1352-2310/$ - see front matter r 2004 Elsevier Ltd. All rights reserved. doi:10.1016/j.atmosenv.2004.06.036
ARTICLE IN PRESS M.E. Bartkow et al. / Atmospheric Environment 38 (2004) 5983–5990
5984
is not routinely performed. This is mainly due to the requirements of expensive HiVol sampling equipment, which is reliant on continuous field maintenance and electricity to operate. To overcome some of the limitations of HiVol sampling, passive samplers such as semipermeable membrane devices (SPMDs) can be used to monitor semivolatile organic pollutants in the atmosphere. SPMDs do not require electricity or expensive equipment to operate in the field but must be deployed inside a protective chamber to minimize the effects of sunlight, rain and wind on sampler performance. SPMDs were introduced by Huckins et al. (1990) as a passive water sampler, but have since been used to sample both indoor and outdoor atmospheres (Petty et al., 1993; Prest et al., 1995; Strandberg et al., 1997; Ockenden et al., 1998a, b; Rantalainen et al., 1999; Caslavsky et al., 2000; Booij and van Drooge, 2001; Lohmann et al., 2001; Meijer et al., 2003; Soderstrom and Bergqvist, 2003). However, only limited data on sampling rates are available. SPMDs can be used to identify sources of pollutants and measure time-integrated air concentrations of pollutants. One way to measure air concentrations using SPMDs is to determine the rate at which pollutants accumulate in the sampler. Where uptake is linear with time and air concentrations and temperature is assumed to remain constant, sampling rates can be calculated for the target compounds according to the following: RS-cal ¼
N SPMD ; CVt
N SPMD : RS-cal t
C SPMD ¼ C O eðke tÞ ;
ð3Þ
where CO is the concentration of the PRC in the SPMD prior to deployment and ke is the elimination rate constant (d1). To our knowledge, only sampling rates for the uptake of airborne PCBs into SPMDs have been reported in the literature (Ockenden et al., 1998a; Shoeib and Harner, 2002). We used a field-based study to determine SPMD sampling rates for PAHs and to evaluate the use of PRCs. An outdoor calibration site was used to determine sampling rates. These data were then used to estimate air concentrations of target compounds for a second site where SPMDs were also deployed. Air concentrations estimated from the SPMD data were then compared with HiVol results obtained over the same sampling period.
ð1Þ
where RS-cal is the sampling rate (m3 d1) measured at the calibration site, NSPMD is the amount of compound accumulated in the sampler (ng), CV is the concentration of that compound in the vapour phase (ng m3) at the calibration site and t is time of exposure (days). A sampling rate can then be used to derive vapour phase air concentrations using identical SPMDs deployed at other sites, according to CV ¼
atmosphere (Ockenden et al., 2001). A known amount of PRCs are loaded into the SPMDs prior to deployment and the elimination rate constant of these compounds for the duration of the exposure is determined for each site. This in situ calibration approach is based on the theory that the elimination rate constants of PRCs and uptake rate constants of target compounds at a given site and are related. If the PRCs used in the SPMD can be assumed to be not present in the atmosphere then loss from the sampler can be described using a first-order decay kinetic equation (Huckins et al., 2002)
ð2Þ
Passive sampling theory and experimental evidence shows that the sampling rate can be influenced by environmental factors such as wind speed if uptake is air-side limited (Harner et al., 2003). Therefore, it is important that site conditions are similar or that sampling rates can be corrected for different environmental conditions between sites. Huckins et al. (2002) have shown the use of performance reference compounds (PRCs) can be useful in correcting for differences in exposure conditions where SPMDs are deployed in the water. PRCs are analytically non-interfering compounds which are similar to the target compounds. Theoretically, the same approach can be taken using PRCs in samplers deployed in the
2. Materials and methods 2.1. Air sampling Each SPMD was made of 82 cm 2.8 cm of lay-flat low-density polyethylene (LDPE) tubing (Brentwood Plastics, MO), with a membrane thickness of about 50 mm and an overall surface area of 460 cm2. The LDPE tubing of each SPMD had been pre-extracted in 180 mL of redistilled hexane for 24 hours followed by another 24-hour extraction period in fresh redistilled hexane. The tube was then filled with 1 mL of triolein, which had been previously spiked with a known amount of two PRCs (2D10-anthracene and 2D10-pyrene). SPMDs (including field blanks) were wrapped in aluminium foil immediately after fabrication and stored at 17 1C until deployment. The preparation and operation of the HiVol system is detailed in previous work (Bartkow et al., 2004). SPMDs and HiVols were deployed at two Queensland Environmental Protection Agency (QEPA) Ambient Air Monitoring Stations in metropolitan Brisbane, Queensland, Australia. One site was located at South Brisbane, adjacent to the M1 motorway, approximately 2 km south of the CBD. The other study site was situated at
ARTICLE IN PRESS M.E. Bartkow et al. / Atmospheric Environment 38 (2004) 5983–5990
Springwood, a predominantly residential suburb, located 20 km south east of the Brisbane CBD. The station itself was located on the grounds of a school, approximately 20 m from a 4-lane roadway that services residential areas. Samplers were deployed at South Brisbane (calibration site) and Springwood in April 2002 for 32 days. Mean ambient air temperature for the duration of the deployment was 22 1C at both sites (QEPA, 2003). Wind speed was also measured at both sites by QEPA at 10 m height (see Results and discussion). A HiVol and four replicate SPMDs were deployed at each site. The SPMDs were deployed 1.5 m above the ground in galvanized iron chambers with an open bottom and louvers on all sides. The chambers were designed to ensure adequate air flow while minimizing the effect of direct sunlight, rain and wind on sampler performance and were similar to those used by Ockenden et al. (1998b). SPMD field blanks were taken into the field and then frozen until extraction. Duplicate HiVols were also deployed at a separate test site to quantify reproducibility for these systems. SPMDs and XAD cartridges were collected after 32 days of exposure. Glass fibre filters were regularly inspected and collected every 3 or 4 days, depending on the amount of accumulated particulate material. Passive samplers and filters were wrapped in foil on-site and the XAD cartridges were sealed with ground glass stoppers. Samples were transported to the laboratory and stored at 17 1C. 2.2. Analysis To investigate the potential for SPMD sampling of particle sorbed PAHs as well as vapour phase PAHs, the external surface of the SPMD was not cleaned prior to analyte extraction. For quantification, all samples were spiked with an internal standard mix containing known amounts of 10 deuterated PAHs (2D10-acenaphthene, 2D10-fluorene, 2D10-phenanthrene, 2 D10-fluoranthene, 2D12-benz[a]anthracene, 2D12-chrysene, 2D12-benzo[b]fluoranthene, 2D12-benzo[a]pyrene, 2 D12-indeno[1,2,3-c,d]pyrene and 2D12-benzo[g,h,i]perylene). The internal standard was spiked directly into the triolein of the SPMDs. The SPMDs were resealed and then extracted in the dark using 180 mL of redistilled hexane at room temperature (22 1C). After 24 hours the solvent was refreshed and the samplers extracted for a further 24 hours. The combined extracts from each sampler were reduced to about 1 mL using rotary evaporators. SPMD extracts were transferred into DCM and subjected to clean-up using gel-permeation chromatography (19 mm 150 mm guard column, followed by a 19 mm 300 mm main column, packed with Envirogel (100 A˚ pore size, 15 mm particle size, Waters) as the stationary phase and with DCM as the
5985
mobile phase). The flow rate was 4.5 mL min1 and the sample collected between 14 and 24 min. XAD cartridges and filter papers were soxhlet extracted separately for 10 h in toluene. All samples were subsequently cleaned using adsorption chromatography (refer Bartkow et al., 2004). The following PAHs were routinely quantified: Naphthalene (Nap), acenaphthylene (Acy), acenaphthene (Ace), fluorene (Flo), phenanthrene (Phe), anthracene (Ant), fluoranthene (Flu), pyrene (Pyr), benz[a]anthracene (B[a]A), chrysene (Chr), benzo[b+k]fluoranthene (B[b,k]F), benzo[e]pyrene (B[e]P), benzo[a]pyrene (B[a]P), perylene (Per), indeno[1,2,3-c,d]pyrene (I[c,d]P), dibenz[a,h]anthracene (D[a,h]A) and benzo[g,h,i]perylene (B[g,h,i]P). For details on the separation and quantification of the PAHs refer to Bartkow et al. (2004). Passive sampler results were reported as the total amount of each PAH in an SPMD (ng). HiVol results were reported as vapour phase air concentrations (CV) or particle phase air concentrations (CP) (ng m3). In some cases, the vapour phase concentration of the high molecular weight PAHs was very close to the detection limits, which increased the uncertainty associated with the values. 2.3. QA/QC The quantification criteria included confirmation of the retention times of the labeled standard and respective analyte. Routinely, the mass fragment with the highest intensity (base peak) was used for quantification. QC samples (field blanks, solvent blanks and spikes) constituted 410% of the total number of samples. SPMD field blanks were taken into the field when samplers were deployed and collected and were extracted and analysed in the same manner as deployed samples. The sample detection limits for individual compounds in a HiVol sample were defined as three times the analyte concentration measured in the field blank. Detection limits for the SPMD samples were calculated as the mean amount in the field blanks plus three times the standard deviation (Refer Table A in Supplementary Data). Where a compound could not be identified in the blank, the detection limit was set as three times the average noise range. Recoveries of the internal standards ranged between 93–122% for the HiVol samples and 45–105% for the SPMDs. Typically, the surfaces of the exposed SPMDs are cleaned prior to extraction, to ensure that any particulates in the lipid-associated surficial film present on SPMDs at the end of exposures do not inhibit dialytic recoveries of the target compounds (Huckins et al., 1999). The surfaces of the SPMDs were not cleaned and this could explain why the recoveries reported from the SPMDs were lower than for the HiVol samples.
ARTICLE IN PRESS M.E. Bartkow et al. / Atmospheric Environment 38 (2004) 5983–5990
5986
Table 2 Atmospheric concentrations of PAHs as determined by HiVol air sampling
3. Results and discussion 3.1. Sampler reproducibility Reproducibility was determined using coefficients of variation (expressed as %CV) for SPMDs (replicates=4) and using normalized differences for HiVol samples (replicates=2). The %-normalized differences were calculated for replicates A and B according to value A value B 100: ðvalue A þ value BÞ=2 Due to their relatively high volatility, reproducibility for Nap, Acy and Ace was consistently poor and results for these compounds were not included in any further analysis. Reproducibility between SPMDs for the remaining compounds was good, with most CVs o20%, except for Flo (34%) and B[a]P (41%) (Table 1). For the HiVol sampler data, the mean normalized difference only exceeded 25% for Flu (27%).
PAH
South Brisbane
Springwood
CV (ng m3) CP (ng m3) CV (ng m3) CP (ng m3) Flo Phe Ant Flu Pyr B[a]A Chr B[b,k]F B[e]P B[a]P Per I[c,d]P D[a,h]A B[g,h,i]P
1.96 4.80 0.82 2.86 3.05 0.19 0.24 0.03 0.01 N.D. N.D. 0.004 N.D. 0.009
0.20 0.09 0.03 0.11 0.18 0.32 0.22 0.57 0.25 0.12 0.02 0.27 0.05 1.05
1.51 2.53 0.31 0.86 0.76 0.04 0.06 0.02 0.01 N.D. N.D. N.D. N.D. 0.002
N.D. 0.09 0.04 0.02 0.03 0.03 0.03 0.10 0.04 0.02 0.00 0.05 0.01 0.16
N.D: not detected.
3.2. HiVol samples Based on total residues per unit volume (i.e., CV+CP), the atmospheric concentrations of PAHs were generally higher at South Brisbane than at Springwood (Table 2). This was not surprising considering the proximity of the South Brisbane site to the city centre and major roadways. The compounds Flo through Pyr were predominantly associated with the vapour phase (490% measured on XAD only), while
Table 1 Mean amounts of PAHs sampled by SPMDs and coefficients of variation (CV) (n=4) PAH
Flo Phe Ant Flu Pyr B[a]A Chr B[b,k]F B[e]P B[a]P Per I[c,d]P D[a,h]A B[g,h,i]P
3.3. SPMDs
Springwooda
South Brisbane
the compounds B[b,k]F through B[g,h,i]P were predominantly associated with particles. PAH profiles from each site were dominated by the compounds most associated with the vapour phase (Flo to Pyr). Phe was present at the highest concentrations with Per and D[a,h]A at the lowest concentrations. The concentration of PAHs in the atmosphere was generally much lower during this sampling campaign than measured by Mu¨ller et al. (1998) in the mid-1990s. In particular, the concentration of B[a]P at Springwood and South Brisbane was generally 10 times lower than B[a]P concentrations at similar sites, 8 years ago (Mu¨ller et al., 1998).
NSPMD (ng)
CV (%)
NSPMD (ng)
CV (%)
110 660 100 410 190 25 45 40 10 2 N.D. 15 N.D. 25
34 13 18 6 15 10 3 5 9 41 — 10 — 13
50 175 35 60 25 3 7 6 2 N.D. N.D. 2 N.D. 2
9 4 19 4 20 17 9 16 — — — — — —
N.D: not detected. a B[e]P, I[c,d]P and B[g,h,i]P only detected in a single SPMD.
Mean amounts of PAHs sampled by the SPMDs are presented in Table 1. PAHs associated with the vapour phase dominated the samples, with Phe contributing to approximately 40% of the total amount of the PAHs quantified. After 32 days of exposure, those PAHs that were present in the atmosphere at low total concentrations (CV+CPp50 pg m3) were not always detectable in a single SPMD. Per and D[a,h]A were not detected in SPMDs from either site, while at Springwood, B[a]P was not detected in any samplers and B[e]P, I[c,d]P and B[g,h,i]P were only detected in one of the four SPMDs. The detection limits could be improved by compositing SPMDs or by deploying the SPMDs for a longer time. Interestingly, compounds such as B[b,k]F, B[e]P, B[a]P, I[c,d]P and B[g,h,i]P that were more associated with particles were routinely detected in SPMDs from South Brisbane.
ARTICLE IN PRESS M.E. Bartkow et al. / Atmospheric Environment 38 (2004) 5983–5990
3.4. SPMD-derived air concentrations Sampling rates were calculated according to modeling theory, using the vapour phase air concentrations for most PAHs (Eq. (1)). However, the sampling rates for the five largest compounds (B[b,k]F to B[g,h,i]P) which were mostly bound to particles, were calculated using the total air concentration (CV+CP). This approach was taken because the vapour phase air concentrations were close to quantification limits and particulate sorbed compounds on the exterior surface of the SPMD were quantified as part of the amount accumulated by SPMDs. Uptake was assumed to be linear for all compounds during the exposure; however, a relatively low sampling rate for Flo (1.8 m3 d1) indicated that uptake for that compound may have reached the curvilinear phase. Sampling rates ranged between 0.6 and 6.1 m3 d1 (Table 3). Based on the two-film resistance model, if airside resistance dominates uptake then sampling rates are not directly related to KOA, whereas for sampler-side resistance, sampling rates would increase, proportionally, with KOA (Harner et al., 2003). Our data show similar sampling rates for all PAHs whereas KOA values vary by as much as six orders of magnitude for the compounds of interest. Hence, we can assume air-side resistance dominates uptake for most PAHs under the present conditions, although sampler-side resistance is likely to have contributed to uptake of the smaller PAHs. Sampling rates tended to be highest for the 3- and 4ring PAHs (1.8–6.1 m3 d1), whereas rates for the 5- and 6-ring PAHs were p2.1 m3 d1. Although there are no other directly measured sampling rates for PAHs reported in the literature, data for PCBs are available for comparison. Shoeib and Harner (2002) deployed SPMDs (surface area of 495 cm2) indoors and reported
Table 3 Mean SPMD sampling rates for PAHs at South Brisbane PAH
RS-cal (m3 d1)a
Flo Phe Ant Flu Pyr B[a]A Chr B[b,k]F B[e]P B[a]P I[c,d]P B[g,h,i]P
1.8 4.4 4.1 4.5 2.0 4.1 6.1 2.1 1.3 0.6 1.7 0.7
a
Sampling rates B[b,k]F to B[g,h,i]P calculated using CV+CP.
5987
sampling rates between 3.3 and 9.9 m3 d1. Although comparability between the two studies was limited due to differences in exposure conditions and the surface area of the SPMDs, the sampling rates measured in this study for the 3- and 4-ring PAHs compare closely with the PCB sampling rates. However, the sampling rates for the higher molecular weight PAHs appeared to be significantly lower compared to rates for the higher molecular weight PCBs. Mean elimination rate constants for each PRC were calculated from the SPMDs deployed at South Brisbane and Springwood using Eq. (3). Time zero concentrations for the first-order loss of PRCs were derived from unopened field blanks. Mean elimination rate constants were very similar (South Brisbane: 2D10-Ant: 0.090 d1 and 2D10-Pyr: 0.094 d1 and Springwood: 2D10-Ant: 0.091 d1 and 2D10-Pyr: 0.085 d1). These results were surprising for several reasons. Firstly, elimination rate constants for 2D10-Ant were similar to the elimination rate constants for 2D10-Pyr. However, even when airside resistance dominates exchange rates, elimination rate constants should be negatively related to KOA because ke ¼
kV A ; K SPMD-V V SPMD
ð4Þ
where kV is the mass transfer coefficient in the air boundary layer, A is the area of the exchanging surface, KSPMD-V is the SPMD-air partition coefficient, and VSPMD is the volume of the whole SPMD. Evidence suggests that KOA is EKSPMD-V, for compounds which are predominately in the vapour phase (unpublished data, USGS, Columbia, MO, USA). Secondly, according to t90%eq: ¼ 2:3=ke ; the magnitude of the elimination rate constants indicated that 90% of equilibrium concentrations were reached by the parallel native compounds. However, in other studies using SPMDs, workers have shown that accumulated PCBs with similar and higher KOA values were not within 90% of equilibrium (Shoeib and Harner, 2002). Furthermore, the influence of wind should not have contributed to the high loss rates because external (outside the exposure chambers) mean wind speeds were not very high (1.7 and 0.8 m s1 at South Brisbane and Springwood, respectively (QEPA, 2003)) and the deployment chamber should have minimized the effect of wind speed on sampler performance. A probable explanation for the lack of significant difference between the ke values for Ant and Pyr and also the overall high loss rates is that the sampling chambers did not exclude enough indirect sunlight, which resulted in photodegradation of the PRCs. Recently, trials undertaken at USGS have shown that significant photolysis of PAHs and polybrominated diphenyl ethers can occur in minutes to hours if SPMDs are not adequately shielded (C. Orazio, USGS, personal communication). If
ARTICLE IN PRESS M.E. Bartkow et al. / Atmospheric Environment 38 (2004) 5983–5990
5988
To estimate air concentrations at Springwood we used the unadjusted sampling rates measured at South Brisbane. The differences between SPMD-derived air concentrations and those measured using the HiVol system were within a factor of 2 for all compounds (Table 4). One interesting aspect of these data is the good agreement between predicted and measured air concentrations of PAHs mostly associated with the particulate phase. This finding elicits several questions. These include: (1) how much does PAH sorption to airborne particulates attenuate photolysis; (2) to what extent does particle deposition on the exterior surface of SPMDs contribute to the accumulation of particulate sorbed PAHs using our analytical methods; (3) is the mass of particles deposited on the SPMD membrane surface during a specified time interval approximately equivalent to the mass of particles present in the volume of air sampled by an SPMD during the same time interval; and (4) if the surficial lipid film is removed before dialysis, do particulate-associated PAHs desorb during an exposure and accumulate in the SPMD membrane and internal lipid phase? With the exception of question 2 the answers to these questions await further research. McLachlan (1999) developed a framework to identify the dominant uptake processes for SOCs into vegetation and it has been successfully applied to a range of SOCs (Bo¨hme et al., 1999; Mu¨ller et al., 2001). McLachlan makes use of two plots where he differentiates between three segments showing which process dominates uptake; (1) equilibrium partitioning of the vapour phase compounds; (2) kinetically limited vapour phase deposition; and (3) particle-bound deposition.
a significant portion of the PRCs were photodegraded, then the native PAHs may also have been affected. In future work, a photolysis sensitive marker compound with known degradation rates and very low SPMD fugacity could be added to SPMDs in order to gauge any effects of photodegradation on the target analytes. However, in this study our sampling rates do not account for any effects of photodegradation.
Table 4 Mean SPMD-derived PAH air concentrations and co-deployed HiVol-derived air concentrations at Springwood and the factor difference between the two measurements PAH
SPMD-based High volume- Factor difference CV or CV+CP based CV or CV+CP (ng m3)a (ng m3)a
Flo Phe Ant Flu Pyr B[a]A Chr B[b,k]F B[e]P I[c,d]P B[g,h,i]P
0.98 1.36 0.31 0.46 0.48 0.03 0.04 0.11 0.05 0.04 0.09
1.51 2.53 0.31 0.86 0.76 0.04 0.06 0.13 0.05 0.05 0.16
1.5 1.9 1.0 1.9 1.6 1.5 1.6 1.2 1.0 1.4 1.8
a Air concentrations are presumed to be representative of the vapour phase (CV) only for PAHs, Flo to Chr and total air concentration (CV+CP) for B[b,k]F to B[g,h,i]P.
10.0
I[c,d ]P
9.5
B[b,k ]F B[e ]P
9.0
log (CSPMD/CV)
8.5
B[a ]P
B[g,h,i ]P
Chr Flu
Phe
8.0
Ant
Flo
B[a ]A
Pyr
7.5 7.0
Kinetically limited gaseous deposition
6.5
Particle-bound deposition
6.0 5.5 5.0 6
7
8
9
10
11
12
13
log K OA
Fig. 1. Plot of log (CSPMD/CV) vs. log KOA (taken or predicted from Beyer et al. (2002)) for each of the PAHs quantified in the SPMDs at South Brisbane. For the vapour phase air concentration of B[a]P and I[c,d]P we used the detection limit of 0.003 and 0.001 ng m3, respectively.
ARTICLE IN PRESS M.E. Bartkow et al. / Atmospheric Environment 38 (2004) 5983–5990
We have used one of these plots to elucidate which chemicals are predominantly sorbed on particulates, which appear to be deposited onto the surface of SPMDs. Fig. 1 shows the log ratio of the concentration of each PAH in the SPMD sample and vapour phase versus the respective log KOA for each compound. The dominant form of uptake for the compounds Flo through Chr is shown to be via kinetically limited deposition from the vapour phase, whereas uptake of the compounds B[b,k]F to B[g,h,i]P are shown in the segment where particle deposition dominates. This result is not surprising for two reasons, firstly, the larger PAHs were mostly particle-bound upon contact with the surface of the sampler, and, secondly, the SPMDs had developed an oily surficial film which accumulated particles, and was not removed prior to extraction. Separate quantification of this ‘surface fraction’ could further assist in understanding the contribution of particles to the total amount of PAHs accumulated by SPMDs. In conclusion, this field-based study shows that SPMDs can be used to estimate air concentrations of PAHs with reasonable accuracy. Deployment times can be lengthened to ensure that compounds present at relatively low concentrations can also be detected. Alternatively, individual SPMD samples from the same site can be combined to enhance detection limits. In order to account for any effects of photolysis, a compound with a high KOA (i.e., low SPMD fugacity), which is known to be very sensitive to photolysis, could be used as a marker for potential photodegradation of analytes accumulated in SPMDs. Experiments aimed at understanding the rate at which particles are deposited onto SPMDs would also be useful.
Acknowledgements The authors thank Darryl Hawker for comments on the manuscript, Ralph Riese, Mike King and Queensland EPA for access to monitoring stations and environmental data and Karen Kennedy and Neil Holling for assistance with lab work and analysis. This work and Ph.D. program was funded by an ARC SPIRT Linkage Grant, with industry support from Queensland EPA, Queensland Health Scientific Services and ERGO. MB receives an APAI scholarship. Queensland Health provides funding for The National Research Centre for Environmental Toxicology.
Appendix A. Supplementary data The online version of this article contains additional supplementary data. Please visit doi:10.1016/ j.atmosenv.2004.06.036.
5989
References Bartkow, M.E., Hawker, D.W., Kennedy, K.E., Mu¨ller, J.F., 2004. Characterizing uptake kinetics of PAHs from the air using polyethylene-based passive air samplers of multiple surface area-to-volume ratios. Environmental Science and Technology 38, 2701–2706. Beyer, A., Wania, F., Gouin, T., Mackay, D., Matthies, M., 2002. Selecting internally consistent physicochemical properties of organic compounds. Environmental Toxicology and Chemistry 21, 941–953. Bo¨hme, F., Welsch, P.K., McLachlan, M.S., 1999. Uptake of airborne semivolatile organic compounds in agricultural plants: field measurements of interspecies variability. Environmental Science and Technology 33, 1805–1813. Booij, K., van Drooge, B.L., 2001. Polychlorinated biphenyls and hexachlorobenzene in atmosphere, sea-surface microlayer, and water measured with semi-permeable membrane devices (SPMDs). Chemosphere 44, 91–98. Bostrom, C.E., Gerde, P., Hanberg, A., Jernstrom, B., Johansson, C., Kyrklund, T., Rannug, A., Tornqvist, M., Victorin, K., Westerholm, R., 2002. Cancer risk assessment, indicators, and guidelines for polycyclic aromatic hydrocarbons in the ambient air. Environmental Health Perspectives 110, 451–489. Caslavsky, J., Zdrahal, Z., Vytopilova, M., 2000. Application of SPMDs for PAH sampling in the Deza chemical factory. Polycyclic Aromatic Compounds 20, 123–141. Harner, T., Farrar, N.J., Shoeib, M., Jones, K.C., Gobas, F.A.P.C., 2003. Characterization of polymer-coated glass as a passive air sampler for persistent organic pollutants. Environmental Science and Technology 37, 2486–2493. Huckins, J.N., Tubergen, M.W., Manuweera, G.K., 1990. Semipermeable membrane devices containing model lipid: a new approach to monitoring the bioavailability of lipophilic contaminants and estimating their bioconcentration potential. Chemosphere 20, 533–552. Huckins, J.N., Petty, J.D., Orazio, C.E., Lebo, J.A., Clark, R.C., Gibson, V.L., Gala, W.R., Echols, K.R., 1999. Determination of uptake kinetics (sampling rates) by lipidcontaining semipermeable membrane devices (SPMDs) for polycyclic aromatic hydrocarbons (PAHs) in water. Environmental Science and Technology 33, 3918–3923. Huckins, J.N., Petty, J.D., Lebo, J.A., Almeida, F.V., Booij, K., Alvarez, D.A., Cranor, W.L., Clark, R.C., Mogensen, B.B., 2002. Development of the permeability/performance reference compound approach for in situ calibration of semipermeable membrane devices. Environmental Science and Technology 36, 85–91. Lohmann, R., Corrigan, B.P., Howsam, M., Jones, K.C., Ockenden, W.A., 2001. Further developments in the use of semipermeable membrane devices (SPMDs) as passive air samplers for persistent organic pollutants: field application in a spatial survey of PCDD/Fs and PAHs. Environmental Science and Technology 35, 2576–2582. McLachlan, M.S., 1999. Framework for the interpretation of measurements of SOCs in plants. Environmental Science and Technology 33, 1799–1804. Meijer, S.N., Ockenden, W.A., Steinnes, E., Corrigan, B.P., Jones, K.C., 2003. Spatial and temporal trends of POPs in Norwegian and UK background air: implications for
ARTICLE IN PRESS 5990
M.E. Bartkow et al. / Atmospheric Environment 38 (2004) 5983–5990
global cycling. Environmental Science and Technology 37, 454–461. Mu¨ller, J.F., Hawker, D.W., Connell, D.W., 1998. Polycyclic aromatic hydrocarbons in the atmospheric environment of Brisbane, Australia. Chemosphere 37, 1369–1383. Mu¨ller, J.F., Hawker, D.W., McLachlan, M.S., Connell, D.W., 2001. PAHs, PCDD/Fs, PCBs and HCB in leaves from Brisbane, Australia. Chemosphere 43, 507–515. Ockenden, W.A., Prest, H.F., Thomas, G.O., Sweetman, A., Jones, K.C., 1998a. Passive air sampling of PCBs: field calculation of atmospheric sampling rates by trioleincontaining semipermeable membrane devices. Environmental Science and Technology 32, 1538–1543. Ockenden, W.A., Sweetman, A.J., Prest, H.F., Steinnes, E., Jones, K.C., 1998b. Toward an understanding of the global atmospheric distribution of persistent organic pollutants: the use of semipermeable membrane devices as timeintegrated passive samplers. Environmental Science and Technology 32, 2795–2803. Ockenden, W.A., Corrigan, B.P., Howsam, M., Jones, K.C., 2001. Further developments in the use of semipermeable membrane devices as passive air samplers: application to PCBs. Environmental Science and Technology 35, 4536–4543. Petty, J.D., Huckins, J.N., Zajicek, J.L., 1993. Application of semipermeable membrane devices (SPMDs) as passive air samplers. Chemosphere 27, 1609–1624. Prest, H.F., Jacobson, L.A., Huckins, J.N., 1995. Passive sampling of water and coastal air via semipermeable membrane devices. Chemosphere 30, 1351–1361.
QEPA, 2003. South East Queensland Ambient Air Quality Monitoring Program. Air Services, Environmental Sciences Division, Queensland Environmental Protection Agency 2003, electronic database. Rantalainen, A., Hyotylainen, T., Saramo, M., Niskanen, I., 1999. Passive sampling of PAHs in indoor air in Nepal. Toxicological and Environmental Chemistry 68, 335–348. Shoeib, M., Harner, T., 2002. Characterization and comparison of three passive air samplers for persistent organic pollutants. Environmental Science and Technology 36, 4142–4151. Smith, D.J.T., Harrison, R.M., 1996. Concentrations, trends and vehicle source profile of polynuclear aromatic hydrocarbons in the UK atmosphere. Atmospheric Environment 30, 2513–2525. Smith, K.E.C., Thomas, G.O., Jones, K.C., 2001. Seasonal and species differences in the air-pasture transfer of PAHs. Environmental Science and Technology 35, 2156–2165. Soderstrom, H.S., Bergqvist, P.A., 2003. Polycyclic aromatic hydrocarbons in a semiaquatic plant and semipermeable membrane devices exposed to air in Thailand. Environmental Science and Technology 37, 47–52. Strandberg, B., Wagman, N., Bergqvist, P.A., Haglund, P., Rappe, C., 1997. Semipermeable membrane devices as passive samplers to determine organochlorine pollutants in compost. Environmental Science and Technology 31, 2960–2965.