Field Validation of Polyethylene Passive Air Samplers for Parent and

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Field Validation of Polyethylene Passive Air Samplers for Parent and Alkylated PAHs in Alexandria, Egypt Mohammed A. Khairy*,†,‡ and Rainer Lohmann† †

Graduate School of Oceanography, University of Rhode Island, 215 South Ferry Road, Narragansett, Rhode Island 02882, United States ‡ Department of Environmental Sciences, Faculty of Science, Alexandria University, 21511 Moharam Bek, Alexandria, Egypt S Supporting Information *

ABSTRACT: Polyethylene samplers (PEs) were deployed at 11 locations in Alexandria, Egypt during summer and winter to test and characterize them as passive samplers for concentrations, sources, and seasonal variations of atmospheric concentrations of polycyclic aromatic hydrocarbons (PAHs). PE−air equilibrium was attained faster for a wider range of PAHs during the winter season possibly due to increased wind speeds. Calculated PE−air partitioning constants, KPE‑A, in our study [Log KPE‑A = 0.9426 × Log KOA − 0.022 (n = 12, R2 = 0.99, Std error = 0.053)] agreed with literature values within 8.6 in the summer season and KPE‑A > 10.7 in the winter season (Figure 2A3 and 2B3). Overall, there was good agreement between PRC-based equilibration predictions and those observed from the 1-, 2-, and 3-week uptake experiments. Deriving KPE‑A and Their Temperature Adjustment. After characterizing the stage of uptake for each of the target compounds, atmospheric concentrations measured with the codeployed high-volume air sampler were used to estimate KPE‑A values for compounds that reached equilibrium. We used concentrations of PAHs that attained equilibrium after 1 week deployment period (Figure 2A1 and 2B1). Calculated Log KPE‑A values (Table SI 5) were correlated against Log KOA and the following equations were obtained (Figure SI 4A, B):

the summer season, only PAHs up to fluorene reached equilibrium (Log KPE‑A = 6.4). The lower dissipation rates of d12-benz(a)anthracene and d12-benzo(a)pyrene in PEs deployed during July 2010 (summer) are probably related to the lower wind speed during the summer season. Chemical exchange between the surrounding air and the passive sampler is limited by the mass transfer through the airside boundary layer and/or the sampler membrane. If the chemical exchange is limited by the sampler membrane, then chemicals with widely different KPE‑A values will show similar clearance rate constants, whereas if the chemical exchange is limited by the air-side resistance, then the clearance rate constants of chemicals should decrease with increasing sampler−air partitioning coefficients (ke α 1/KPE‑A based on the following relation: ke = (ka·As)/(Vs·KPE‑A), where ka is the mass transfer coefficient through the air-side boundary layer and As and Vs are the sampler surface area and volume, respectively).33 Overall, PAH loss rate constants measured in PEs decreased with increasing log KPE‑A and increased with increasing wind speeds (loss rate constants during winter > summer). There was a statistically significant difference (loss rate constants significantly decrease with increasing KPE‑A) between kes of the three PRCs in each season (1-way ANOVA, Fsummer = 135.5, Fwinter = 127.5 at α = 0.05, p < 0.001). Moreover, the loss rate constant of BaP-d12 was significantly higher in the winter season (tcalc = −5.485, df = 20, α = 0.05, P < 0.001). These relationships confirm air-side resistance dominating loss from the PE samplers. Comparison of Predicted and Observed % Equilibration during Field Deployments. In the present study, three PEs were deployed simultaneously for 1, 2, and 3 weeks each at site 11 during each season. This was done to better understand the equilibration of PAHs in the PEs over time and validate PRC-based equilibration times. Results indicated that % equilibrium between samplers and the surrounding air generally decreased with increasing log KPE‑A (Figure 2A and B). Lower molecular weight PAHs attained equilibrium faster than higher molecular weight PAHs. Three different regimes could be observed for the accumulated PAHs in the codeployed PEs. In the first regime, PAHs in PEs were close to or reached

Log KPE−A(summer) = 0.9142 × Log KOA + 0.1097 (3)

(R2 = 0.993, n = 7, SE = 0.043, SE of intercept = 0.163, SE of slope = 0.024) Log KPE−A(winter) = 0.9711 × Log K OA − 0.1536

(4)

(R2 = 0.975, n = 12, SE = 0.088, SE of intercept = 0.334, SE of slope = 0.049) 3994

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Results from both deployments were averaged to calculate KPE‑As for all the other investigated PAHs according to the following equation: Log KPE−A = 0.9426 × Log KOA − 0.022

(5)

(R2 = 0.991, n = 12, SE = 0.052, SE of intercept = 0.197, SE of slope = 0.029) Calculated KPE‑A values based on eq 5 were compared with those calculated according to Bartkow et al.,23 Kennedy et al.,33 and Lohmann36 (Figure SI 6). Our KPE‑A values were in very good agreement with those obtained according to the equation proposed by Bartkow et al.23 with factor different less than 46% in all the investigated individual PAHs (Table SI 5), and similar to the prediction by Lohmann36 based on the partitioning of PAHs between PE and water. Calculated KPE‑A values in our study were temperature corrected using ΔHvap. The excellent agreement between eqs 3 and 4 (no statistical significant difference between the regression coefficients at p = 0.001) indicated that the use of ΔHvap accurately corrected KPE‑A for PAHs. In other words the enthalpy of PE−air partitioning (ΔHPE‑A) was dominated by changes in the PAH’s vapor pressure, not by changes in the PE uptake. As an alternative, we also temperature adjusted KPE‑A with the energy of octanol−air exchange (ΔUOA, kJ/mol) from Beyer et al.37 and compared it to the correction using ΔHvap. Good match (Figure SI 7) between calculated KPE‑As corrected using both enthalpies was observed with a factor difference 90% of the total detected concentrations. The most abundant PAHs were naphthalene (22−51%), phenanthrene (2.0−19%), and C1-phenanthrene/anthracene (4.0− 18%). Samples were more enriched with naphthalene during the winter season (39%) when compared with the summer season (29%). In the summer season, there was a marked increase in the enrichment of the samples with phenanthrene (14% compared to 9.0% in the winter samples) and C1phenanthrene (10%) which could be related to the increases in the evaporative emissions from petroleum products such as asphalt and coal tar sealant.35 The spatial distribution of PAHs in Alexandria City (Figure SI 10) indicated that higher PAH concentrations were observed at sites characterized by high traffic (St 11, 2, and 3) and

industrial (St 8 and 4) activities during the summer season (see Text SI 9 for more details). During the winter season, higher PAH concentrations were observed at sites characterized by industrial activities (St 8 and St 4) and heavy traffic composed mainly of heavy trucks. In all the other samples, detected PAH concentrations were relatively stable with minor variations (410−630 ng/m3). This finding supports the suggestion that vehicle emissions are the major source of PAHs in the atmospheric environment of Alexandria. No statistically significant difference was observed between detected concentrations of PAHs in the investigated samples during both seasons (t = 0.325; p = 0.748; α = 0.05). Although combustion-derived PAH emissions may be elevated during the colder months, PAH gas-phase concentrations will be reduced by partitioning to particles which is enhanced at cold temperature. This effect will be greatest in urban areas where particle concentrations are highest. Detected concentrations of PAHs during both seasons suggest that vehicle emissions are the major sources of PAHs in Alexandria. Although unleaded petrol has recently been used, leaded gasoline is still in use, especially in old vehicles. At the same time, catalytic converters have not yet been extensively introduced to Egypt. Major contributors of vehicular pollution are four-cylinder gasoline vehicles (e.g., cars and minibuses), diesel-powered heavy vehicles (e.g., transportation buses and mini-trucks), and twostroke-engine-powered motor cycles. The latter are powered with mixed gasoline-oil fuel to compensate for the lack of wet sump. Traffic intensities are too high for Alexandria’s inadequate road network causing congestion on busy narrow streets. Other sources of PAHs may include diesel trains, natural gas, open burning of municipal wastes, and vehicle tires. PAH Source Characteristics. Concentrations and patterns of PAHs quantified in a given environment have been used to reflect their possible sources. Various molecular diagnostic ratios of PAH concentrations have been used for qualitative and quantitative characterization of sources in environmental samples.42,49,50 However, PAH isomer ratios show substantial intrasource variability and intersource similarity51 and should be interpreted with caution based on previous knowledge about possible sources of PAHs that exist in a given area. Molecular ratios used in this study included Ant/Ant+Phn, Phn/Ant, alkylated Phn (sum of all methylated Phns) to Phn (MPhn/ Phn), MDBT/MPhn, Phn+Ant/Phn+Ant+C1-Phn/Ant, Flra/ Flra+Pyr, BaA/BaA+Chry, InP/InP+BghiP, and BaP/BghiP. Calculated diagnostic ratios are presented in Figure SI 11. All the calculated ratios indicated that vehicular emissions are the major source of PAHs in the atmospheric environment of Alexandria, Egypt. This is also in-line with the rather uniform concentrations across wide swaths of this major city (see Text SI 10 for more details). Implications. The results of this study indicate that lowdensity polyethylene sheets can be successfully used to assess spatial and temporal trends of PAHs in an arid urban city. This is of great significance as it indicated how a simple, accessible, and cost-effective sampling medium can yield valuable data in developing countries because of the problems associated with conventional active air sampling techniques. Accordingly, databases regarding PAHs can be acquired to develop pollution control and management plans on the national scale. The results obtained from this study also highlighted the role of urban cities as sources of PAHs to the regional environment. This is of great importance for cities like Alexandria located on the Mediterranean coast as substantial loading to the coastal 3996

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(8) Lohmann, R.; Corrigan, B. P.; Howsam, M.; Jones, K. C.; Ockenden, W. A. 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. Environ. Sci. Technol. 2001, 35 (12), 2576−2582. (9) Harner, T.; Farrar, N. J.; Shoeib, M.; Jones, K. C.; Gobas, F. A. P. C Characterization of polymer-coated glass as a passive air sampler for persistent organic pollutants. Environ. Sci. Technol. 2003, 37 (11), 2486−2493. (10) Meijer, S. N.; Ockenden, W. A.; Steinnes, E.; Corrigan, B. P.; Jones, K. C. Spatial and temporal trends of POPs in Norwegian and UK background air: Implications for global cycling. Environ. Sci. Technol. 2003, 37 (3), 454−461. (11) Jaward, F. M.; Farrar, N. J.; Harner, T.; Sweetman, A. J.; Jones, K. C. Passive air sampling of PCBs, PBDEs, and organochlorine pesticides across Europe. Environ. Sci. Technol. 2004, 38 (1), 34−41. (12) Jaward, F. M.; Meijer, S. N.; Steinnes, E.; Thomas, G. O.; Jones, K. C. Further studies of the latitudinal and temporal trends of persistent organic pollutants in Norwegian and U.K. background air. Environ. Sci. Technol. 2004, 38 (9), 2523−2530. (13) Shen, L.; Wania, F.; Lei, Y. D.; Teixeira, C.; Muir, D. C. G.; Bidleman, T. F. Hexachlorocyclohexanes in the North American atmosphere. Environ. Sci. Technol. 2004, 38 (4), 965−975. (14) Bohlin, P.; Jones, K. C.; Tovalin, H.; Strandberg, B. Observations on persistent organic pollutants in indoor and outdoor air using passive polyurethane foam samplers. Atmos. Environ. 2008, 42 (31), 7234−7241. (15) Determination of polycyclic aromatic hydrocarbons (PAHs) in ambient air using gas chromatography/mass spectrometry (GC/MS); Compendium Method TO-13A, 2nd ed.; U.S. Environmental Protection Agency: Cincinnati, OH, 1999; http://www.epa.gov/ ttnamti1/files/ambient/airtox/to-13arr.pdf. (16) Bartkow, M. E.; Booij, K.; Kennedy, K. E.; Müller, J.; Hawker, D. Passive sampling theory for atmospheric semivolatile organic compounds. Chemosphere 2005, 60 (2), 170−176. (17) Fernandez, L. A.; MacFarlane, J. K.; Tcaciuc, A. P.; Gschwend, P. M. Measurement of Freely Dissolved PAH Concentrations in Sediment Beds Using Passive Sampling with Low-Density Polyethylene Strips. Environ. Sci. Technol. 2009, 43 (5), 1430−1436. (18) Friedman, C.; Burgess, R. M.; Perron, M. M.; Cantwell, M. G.; Ho, K. T.; Lohmann, R. Comparing polychaete bioaccumulation and passive sampler uptake to assess the effects of sediment resuspension on PCB bioavailability. Environ. Sci. Technol. 2009, 43 (8), 2865−2870. (19) Tomaszewski, J. E.; Luthy, R. G. Field deployment of polyethylene devices to measure PCB concentrations in pore water of contaminated sediment. Environ. Sci. Technol. 2008, 42 (16), 6086− 6091. (20) Adams, R. G.; Lohmann, R.; Fernandez, L. A.; Macfarlane, J. K.; Gschwend, P. M. Polyethylene devices: Passive samplers for measuring dissolved hydrophobic organic compounds in aquatic environments. Environ. Sci. Technol. 2007, 41 (4), 1317−1323. (21) Morgan, E.; Lohmann, R. Detecting Air-Water and SurfaceDeep Water Gradients of PCBs Using Polyethylene Passive Samplers. Environ. Sci. Technol. 2008, 42 (19), 7248−7253. (22) Sacks, V. P.; Lohmann, R. Development and use of polyethylene passive samplers to detect triclosans and alkylphenols in an urban estuary. Environ. Sci. Technol. 2011, 45 (6), 2270−2277. (23) Bartkow, M. E.; Hawker, D. W.; Kennedy, K. E.; Muller, J. F. Characterizing uptake kinetics of PAHs from the air using polyethylene-based passive air samplers of multiple surface area-to-volume ratios. Environ. Sci. Technol. 2004, 38 (9), 2701−2706. (24) Bartkow, M. E.; Jones, K. C.; Kennedy, K. E.; Holling, N.; Hawker, D. W.; Muller, J. F. Evaluation of performance reference compounds in polyethylene-based passive air samplers. Environ. Pollut. 2006, 144 (2), 365−370. (25) Lohmann, R.; Burgess, R. M.; Cantwell, M. G.; Ryba, S. A.; MacFarlane, J. K.; Gschwend, P. M. Dependency of polychlorinated biphenyl and polycyclic aromatic hydrocarbon bioaccumulation in Mya

waters could occur both through atmospheric deposition and surface runoff. Results obtained from our investigation highlighted the need for performing more studies to better understand the partitioning and equilibriation mechanisms of alkylated PAHs between PEs and the surrounding atmosphere as most studies concentrated on the parent PAHs despite the higher detectable concentrations of alkylated PAHs in the atmospheric environment of urban cities worldwide.

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ASSOCIATED CONTENT

S Supporting Information *

Details on deployments, methods, PAH concentrations, and their selected physicochemical properties, plus additional correlations. This information is available free of charge via the Internet at http://pubs.acs.org.

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AUTHOR INFORMATION

Corresponding Author

*E-mail: [email protected]; phone: 401-8746765; fax 401-874-6811. Notes

The authors declare no competing financial interest.

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ACKNOWLEDGMENTS We thank Mr. Dave Adelman (GSO, URI) for his effort in the preparation of the PEs. Special thanks are also due to Prof. Dr. Alaa Mostafa for his help in the deployment of the PEs. Dr. Mohammed Khairy acknowledges the Fulbright Foundation for offering him the opportunity to carry out this research in the United States. Dr. Rainer Lohmann acknowledges funding from EPA’s Great Lakes Restoration Initiative Award GLAS # 00E00597-0 and Great Lakes Air Deposition Program Award # GLAD 2010-5 supporting passive sampler research at URI.

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