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Determining Air-Water Exchange, Spatial and Temporal Trends of Freely Dissolved PAHs in an Urban Estuary Using Passive Polyethylene Samplers Rainer Lohmann,*,† Meredith Dapsis,† Eric J. Morgan,†,‡ Victoria Dekany,† and Pamela J. Luey† †
Graduate School of Oceanography, University of Rhode Island, Narragansett, Rhode Island 02882, United States
bS Supporting Information ABSTRACT: Passive polyethylene (PE) samplers were deployed at six locations within Narragansett Bay (RI, USA) to determine sources and trends of freely dissolved and gas-phase polycyclic aromatic hydrocarbons (PAHs) from May to November 2006. Freely dissolved aqueous concentrations of PAHs were dominated by fluoranthene, pyrene, and phenanthrene, at concentrations ranging from tens to thousands of pg/L. These were also the dominant PAHs in the gas phase, at hundreds to thousands of pg/m3. All stations mostly followed the same temporal trends, with highest concentrations (up to 7300 pg/L for sum PAHs) during the second of 11 deployments, coinciding with a major rainstorm. Strong correlations of sum PAHs with river flows and wastewater treatment plant discharges highlighted the importance of rainfall in mobilizing PAHs from a combination of runoff and atmospheric washout. PAH concentrations declined through consecutive deployments III to V, which could be explained by an exponential decay due to flushing with cleaner ocean water during tides. The estimated residence time (tres) of the PAH pulse was 24 days, close to an earlier estimate of tres of 26 days for freshwater in the Bay. Air-water exchange gradients indicated net volatilization of most PAHs closest to Providence. Further south in the Bay, gradients had changed to mostly net uptake of the more volatile PAHs, but net volatilization for the less volatile PAHs. Based on characteristic PAH ratios, freely dissolved PAHs at most sites originated from the combustion of fossil fuels; only two sites were at times affected by fuel spill-derived PAHs.
’ INTRODUCTION Passive sampling has significant advantages for the detection of trace compounds, as it yields inexpensive, technically simple measurements and reduces detection limits. Passive sampling relies on diffusion to accumulate analytes of interests instead of pumping air or water through a filter.1 As a result, passive samplers only accumulate molecules which are freely dissolved in the water or the atmosphere. This avoids the need to correct active (total) sampling results for the analytical interference of, e.g., dissolved organic carbon (DOC). In general, passive samplers integrate freely dissolved concentrations over time. For compounds that reach equilibrium during the deployment period, passive samplers reflect the most recent exposure before retrieval, not a weighted average of total exposure.2 Recently, the use of low-density polyethylene (PE) as passive samplers of freely dissolved polycyclic aromatic hydrocarbons (PAHs) and polychlorinated biphenyls (PCBs) was reported as a function of temperature and salinity.3 PE-water equilibrium partitioning coefficients (KPE-W, in L/kg) were determined as: CPE KPE-W ¼ ð1Þ CW r 2011 American Chemical Society
where CPE and CW are the compound’s concentration in PE (ng/ kg PE) and its freely dissolved concentration (ng/L).3 PE has also been used as a passive samplers of atmospheric PAHs.4 KPE-air can similarly be defined as CPE divided by the concentration in air (Cair). In cases in which equilibrium between the analyte in the PE sampler and the surrounding medium (air or water) is not reached, the addition of performance reference compounds (PRCs) to the PE prior to deployment can be used to infer the exchange rate kinetics, ke. Technically, ke reflects the clearance rate of the PRC (t-1), which can be calculated as:5 CPE 0 ke ¼ ln CPE t
! 3t
-1
ð2Þ
Received: July 30, 2010 Accepted: January 19, 2011 Revised: January 2, 2011 Published: February 25, 2011 2655
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Figure 1. Map of sampling stations in Narragansett Bay in 2006 (A, Conimicut Point; B, Mount Hope Bay; C, North Prudence; D, Poppasquash Point; E, Mount View Bay; F, Quonset Point).
where C0PE is the concentration of the PRC in the PE before deployment, and CtPE is the concentration of the PRC in the PE after deployment time t (days). Assuming that uptake and loss rates are equivalent, dissolved (or gas-phase) concentrations of compounds that are not equilibrated can be deduced as6 CWðairÞ ¼
CPE 1 - e-ke t KPE-WðairÞ
ð3Þ
Different passive samplers have been used in harbors, rivers, and urban waterways to determine freely dissolved concentrations of various hydrophobic organic contaminants (HOCs).7-10 Often, several passive samplers are deployed such that the gradient of HOCs can be determined across interfaces, such as from sedimentto-water, water-to-air, or within water.11-13 Narragansett Bay (NB) is a temperate estuary on the Atlantic coast of Rhode Island, USA (Figure 1). It is a relatively small, heavily urbanized estuary characterized by a long history of pollution, including historical deposition of PAHs and other HOCs.14,15 In previous work, we reported that PCBs were volatilizing from the Bay for most of the summer of 2006,12 possibly linked to their release from historically contaminated sediments.16 This prompted us to investigate whether ongoing primary sources (i.e., urban runoff, atmospheric deposition, fuel spills) were the dominant sources of PAHs into NB, or their release from historically contaminated sediments. We, therefore, undertook a major sampling campaign of 11 deployments from May to October 2006 across six sites to (i) establish baseline concentrations of truly dissolved PAHs in NB; (ii) determine the main sources of PAHs to the estuary; (iii) establish air-water exchange gradients to deduce whether the Bay was volatilizing PAHs or taking them up; and (iv) elucidate the main factors controlling the fate of PAHs in the estuary.
’ MATERIALS AND METHODS Site Description. Narragansett Bay is fairly shallow, with a mean depth of 8.3 m, and has an area of ∼330 km2. Its long-term average hydraulic residence time is generally reported as 26 days.17 The bay consists of two major passages (Figure 1), both of which receive freshwater input from urban rivers. Nevertheless, NB is fairly saline (∼25-31 psu), as it is strongly
influenced by advection of oceanic waters from Rhode Island Sound. Stratification can occur in the summer months, during a neap tide coupled with high freshwater input or strong solar warming, but the estuary is generally considered to be partially mixed.17 NB has a mixed semidiurnal tide and circulation is influenced by wind.18 The general circulation is counterclockwise in the main part of the Bay, i.e., Atlantic water enters in the East passage, moves past Poppasquash (site D) toward Conimicut Point (A), and exits again through the West passage, via North Prudence (C), Mount View (E), and Quonset Point (F; Figure 1). Freshwater comes mainly from rivers in the North: the Blackstone and Pawtuxet Rivers toward sites A and C, and the Taunton River toward site B in Mount Hope Bay. Deployment of PEs. PEs were deployed in the surface water off six monitoring buoys that are part of the NB fixed-site water quality monitoring network operated by the Office of Water Resources (RI Department of Environmental Management), and the Marine Ecosystems Research Laboratory at the University of RI, Graduate School of Oceanography (URI-GSO) (Figure 1). At sites A and F, additional PEs were placed in inverted stainless steel bowls ca. 1 m above the water, providing protection from direct radiation and rainfall. PEs were deployed in the surface water and air for an average period of 15 days, although deployment times of individual PEs varied from 11 to 22 days (for more details, see Table SI 2). A surface sonde (Yellow Springs Incorporated, YSI, 600XL) was attached to each buoy, providing measurements of, among others, salinity, temperature, and chlorophyll a concentrations. Wastewater treatment plant (WWTP) discharge data was obtained from the Narragansett Bay Commission. Preparation of Polyethylene Passive Samplers. PEs were prepared as detailed elsewhere.12 PEs were cut from commercial sheeting (Carlisle Plastics, Inc., Minneapolis, MN) with a thickness of 51 μm, yielding a 10 30 cm strip of ∼1-2 g each. PEs were precleaned twice overnight in dichloromethane (DCM). After precleaning, 12-14 PEs were immersed in an 80:20 (v/v) methanol/water solution spiked with PRCs at a nominal concentration of 5 μg per sampler in methanol for four days, according to the method detailed in ref 5. Two deuterated PAHs were selected to serve as PRCs: d10-anthracene and d12-benz[a]anthracene. Once spiked, PEs were strung on stainless steel wires, placed in precleaned aluminum foil packets, numbered, and frozen until the time of deployment in either water or air. A small snippet (ca. 1 10 cm, ∼0.15 g) of each PE was taken to determine initial PRC concentrations. Extraction of PEs. After addition of 50 μL of a surrogate standard containing deuterated PAHs (d10-acenaphthene, d10phenanthrene, d12-chrysene, and d12-perylene; 5 ng/μL in nonane), PEs were extracted twice in DCM overnight. The resulting extracts were combined and concentrated to ∼1 mL on a rotary evaporator, solvent exchanged to hexane, and concentrated to ∼50 μL. Twenty μL of d14-terphenyl (5 ng/μL), was added as an injection standard before analysis. Water Filtrations. In a follow-up study, active and passive sampling were coperformed to verify that PE sampler-derived dissolved concentrations of PAHs in the water were broadly comparable to more established water sampling methods. Three 25-μm thick PEs were deployed for two, four, and eight days each at the GSO dock, and the results were compared to water filtered through a combination of glass fiber filter (GFF, Whatman, Piscataway, NJ, USA) and polyurethane foam (PUF, Tisch Environmental, Cleves, OH, USA) plugs. Deuterated d12-benzo(a)pyrene was added as a third PRC to account for nonequilibrium 2656
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Figure 2. Average freely dissolved PAH profile at six sites in NB in 2006.
of PAHs in the PE samplers. Three active samples of ca. 10 L each were collected on day eight. GFFs were used to remove particles, while PUF plugs collected dissolved PAHs. Only the PUFs were fortified with internal standards (see above) and extracted in DCM in a Soxhlet overnight. The resulting extract was dried by filtering it through a funnel filled with sodium sulfate, concentrated, and prepared for analysis, as above. Instrument Analyses and Quality Control. Analysis for PAHs was conducted on an Agilent 6800 gas chromatograph (GC) coupled to an Agilent 5973N mass spectrometer (MS) operated in the negative electron ionization mode. PAHs were quantified by isotope dilution relative to the nearest surrogate standard. The most abundant molecular ions were analyzed in selected ion monitoring (SIM). PEs were analyzed for 26 PAHs (SI Table 1). For more details, see the SI. Physico-Chemical Properties. To obtain a full set of KPE-Ws, those measured by Adams et al.3 were correlated against octanol-water partitioning coefficients, Kows, to obtain values for all PAHs (Table SI 1). Internally consistent Kow and octanol-air partitioning constants, Koa, were chosen from Ma et al.19 KPE-air were derived based on a correlation with Koa as proposed by Bartkow et al.20 These KPE-air were also consistent with those derived via KPE-air = KPE-w/Ka-w, with the air-water partitioning constants, Ka-w, taken from ref 19. Temperature and Salinity-Dependency of PEs. Freely dissolved PAH concentrations were calculated by adjusting the mass of PAHs in the PE sampler for nonequilibrium (according to the PRC data) and by using temperature-adjusted KPE-Ws in eq 3. KPE-Ws were T-corrected using the internal energy 4UW of aqueous dissolution for subcooled liquid PAHs (see SI). Where available, 4UW were taken from ref 21; otherwise, an average value of 25 kJ/mol was used. KPE-Ws were not salt corrected, as increased salinity in the field would enhance water-to-air exchange but not lead to lower dissolved concentrations.
’ RESULTS AND DISCUSSION PAH Equilibrations in Field-Deployed PE Samplers. Both PRCs were mostly dissipated across all sites and deployments
(d10-anthracene at 100%, d12-benz(a)anthracene >90%), implying that PAHs up to benz(a)anthracene were at or close to equilibrium in the PEs. At site A, d12-benz(a)anthracene was only dissipated to 0.70 for most mid MW PAHs (fluorenebenzofluoranthenes), with the exception of phenanthrene and retene. During this deployment, freely dissolved PAHs were dominated by one source (rainstorm event) that overwhelmed other, minor sources and nonconservative behavior of the individual PAHs. Two of the samplers were lost during the following deployment (III), preventing us from verifying whether the correlation between salinity and Cw continued to be significant. Influence of Rain/Runoff/Wastewater Treatment. We investigated the correlation of Σ17 PAHs with flows from the major rivers supplying freshwater into NB (Blackstone, Pawtuxet, and Taunton Rivers), and inflow via NB’s two biggest WWTPs, Fields Point and Bucklin Point, as potential point sources of PAHs. Σ17 PAHs at site A were significantly correlated to Blackstone and Pawtuxet River flows, and to in- and outflows from the Bucklin Point WWTP, and to a lesser degree to flow from the Field’s Point WWTP and the Taunton River (see Figure 4, SI Table 16). Water flows from the Blackstone and Pawtuxet Rivers and Bucklin Point WWTP inflow were all highly correlated with Σ17 PAHs at site A (r2 > 0.90%, p < 0.05). The better correlations with Bucklin Point WWTP flows were surprising, as Fields Point is the bigger WWTP, and is closer to 2658
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Figure 3. Changes in freely dissolved Σ17 PAHs over time at six sites in NB in 2006.
Figure 4. Temporal trend of freely dissolved concentrations of Σ17PAH at site A (Conimicut Point), discharges from Bucklin Point wastewater treatment plant (million gallons per day), and Blackstone River flow (m3/s).
site A (Figure 1). The Bucklin Point WWTP might be a more important source of PAHs, due to higher PAH loading and/or less efficient removal of PAHs. Further south, Σ17 PAHs at sites D and E had weak correlations with flow data from the WWTPs but not the Blackstone River, whereas no significant correlation was left for any of the other sites (Table SI 16). The strong correlation of Σ17 PAH concentrations with river flows and WWTP discharges highlights the importance of rainfall in mobilizing PAHs from a combination of runoff and atmospheric washout. In other words, WWTPs need not be sources per se, but their flow rates are good proxies for general rainfall and stormwater impact in the watershed of Narragansett Bay. At site A, the mid MW PAHs (phenanthrene to benzo[b, k]fluoranthenes) displayed significant correlations with Blackstone River flow and wastewater inflow at Bucklin Point, but not the most volatile PAHs (naphthalenes to fluorene), nor the highest MW PAHs (benzo[a,e]pyrenes to perylene) (SI Table 17). Phenanthrene and anthracene displayed better correlations with the Buckling point WWTP inflow, which could imply it as a point source of these PAHs. These correlations imply that the most volatile PAHs probably have other pathways to enter Narragansett Bay, notably via air-to-water partitioning (see below). The heavier MW PAHs are preferentially sorbed to particles, and might enter NB in a sorbed state, which could explain the lack of correlation of their freely dissolved concentration with freshwater inflow. Residence Time of PAHs at Site A. After the pulse of PAHs linked to the rainstorm (period II), concentrations of PAHs decreased over the next four sampling periods III-VI (although the sampler was lost during period V). For Σ17 PAHs, the
decrease could be explained by an exponential decay due to flushing with cleaner ocean water during tides. Estimated tres for the PAH pulse was 24 days (time to decrease to 1/e), or a half-life (t1/2) of 17 days. In fact, a closer look revealed that the most volatile PAHs up to fluorene displayed no decreasing trend during this time period, nor did the highest MW PAHs, such as benzo[a]pyrene (SI Table 18). As discussed earlier, these PAHs were not significantly correlated to freshwater flow, indicating that they have other pathways to enter NB. Yet PAHs from phenanthrene to benz[b,k]fluoranthenes all displayed highest concentrations during period II, and lower ones thereafter. Fitting exponential curves to their decreasing concentrations in tres of these PAHs ranging from 19 to 43 days (t1/2 of 13-30 days). Residence times seemed to depend mostly on MW. Phenanthrene, anthracene, and methyl-phenanthrenes displayed somewhat slower tres of 29-43 days. This could have been due to net deposition of gas-phase phenanthrene, etc., counteracting the flushing of PAHs out of the Bay (see below). In contrast, pyrene, fluoranthene, chrysene, and benz[a]anthracene all displayed tres of 19-27 days. The closeness of the different tres of various PAHs suggests that they were all affected by the same process, in this case tidal flushing, which seemed more important than reactions, settling, or volatilization that would affect individual PAHs differently. These results also suggest the release of PAHs from the sediments was of minor importance for surface water freely dissolved concentrations. The estimation of tres based on PAH dispersion agrees nicely with a prior estimate of a hydraulic tres of 26 days (12-40 days, depending on freshwater flow) by Pilson17 for an average freshwater flow of ∼105 m3/s. 2659
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Figure 5. Air-water exchange gradients at site A (Conimicut Point, in blue) and site F (Quonset Point, in red) for (a) phenanthrene, (b) methylphenanthrenes, (c) pyrene, and (d) chrysene (a positive value indicates net volatilization of PAHs from the water; a negative value indicates net deposition).
Water-Air Exchange Gradients. Water-to-air exchange gra-
dients were established for two sites: A, nearest Providence (RI), and site F, closest to the Atlantic Ocean. Gradients were calculated as the ratio of the concentrations in PE, corrected for nonequilibrium, minus one Air - water gradient : CPE, waterðeqÞ =CPE, airðeqÞ - 1 ð4Þ such that a positive value indicates net volatilization of PAHs, and a negative value indicates net deposition. A temperature correction was not performed, so that the gradients reflected conditions in the field. In an estuary, in situ chemical activity gradients can also be caused by thermal disequilibrium or differing salinity gradients. At site A, fluorene and phenanthrene displayed air-water gradients generally 0, suggesting net volatilization (Figure 5, SI Table 19). Exceptions were deployment periods III, when fluorene and phenanthrene showed net volatilization, and period VI, when pyrene and benz[a]anthracene displayed net deposition. Further south at site F, gradients had changed to mostly net deposition of the more volatile PAHs, such as phenanthrene, methyl-phenanthrenes, fluoranthenes, and pyrene (Figure 5, SI Table 20). In contrast, air-water gradients indicated net volatilization for the less volatile PAHs, such as chrysene, benz[a]anthracene, benzo-fluoranthenes, and benzo-pyrenes. This indicates that site A received sufficient PAHs in the water, possibly from runoff and the rivers during deployment period II, to cause net volatilization of most PAHs. These results also suggest that the runoff or river input were more important for the northern part of the Bay than atmospheric deposition. As atmospheric transport is faster than dilution in the water column, the volatile PAHs are experiencing net deposition over the cleaner southern part of NB. In contrast, the less volatile PAHs are settling with particles out of the atmosphere, causing a net waterto-air gradient further away from urban centers.
Similar to our results for urban site A, Gigliotti et al.23 reported net volatilization for most PAHs from Raritan Bay and New York Harbor in 1998, with the exception of net deposition for phenanthrenes and methyl-phenanthrenes. Bamford et al.22 also reported net volatilization for fluorene, anthracene, pyrene, fluoranthene, and chrysene, but net absorption of phenanthrene in the Patapsco River in 1997/98. Source Indication by Molecular Ratios. We investigated the distribution of different PAHs (“molecular ratios”), which can give clues as to their sources.27 Two ratios were investigated, fluoranthene over fluoranthene and pyrene (fl/flþpyr) and phenanthrenes over the sum of methyl-phenanthrenes (phen/ Σme-phen). They provide a general assessment whether the PAHs stem from petroleum spills or from the combustion of fossil fuels. In general, fl/flþpyr ratios were >0.5 (SI Figure 1), indicating the PAHs mainly stem from the combustion of fossil fuels. Several samples from sites B and C were often below the threshold, possibly indicating a contribution from oil spills to the water. The phen/Σme-phen ratio of most samples was above 1.0, also indicating combustion-derived PAHs were dominating dissolved concentrations. However, according to the phen/Σme-phen ratio, most samples from sites C and D displayed a fuel signature. These results suggest that during several deployments at sites BD, fuel spill-derived PAHs contributed to total PAHs. Freely dissolved PAHs at the other sites seemed to originate mostly from the combustion of fossil fuels. Assessing Sources of PAHs to Narragansett Bay. Overall, wet weather events dominated the loading of PAHs to Narragansett Bay during the summer of 2006, with minor contributions from atmospheric deposition and local sources, such as PAHs from spilled oil. Rainstorms increased the flux of mid MW freely dissolved PAHs to the estuary, but not of volatile compounds (MW up to fluorene), nor of high MW PAHs (benzo[a]pyrene and higher), possibly due to their sorption to solids (which we did not assess in this study). Finally, contaminated sediments can also be a source of PAHs to the water column, especially during resuspension events.28 2660
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Environmental Science & Technology Our data do not really allow assessing this contribution. Other work suggests that deeper waters in NB are affected by diffusion of PAHs from sediments, but not surface waters (Brandeis and Lohmann, unpublished data). However, the correlation of high concentration events with river flow, WWTP discharges, and the subsequent decreases in concentrations suggests that ongoing primary discharges dominate PAH concentrations in Narragansett Bay.
’ ASSOCIATED CONTENT
bS
Supporting Information. Additional information on deployments, 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.
’ AUTHOR INFORMATION Corresponding Author
*Phone: 401-874-6612; fax: 401-874-6811; e-mail: lohmann@ gso.uri.edu. Present Addresses ‡
Boston University, 675 Commonwealth Ave., Boston, MA 02215.
’ ACKNOWLEDGMENT Kevyn Bollinger, Bridget Reaney, and Carey Friedman assisted with PE preparations, Deanna Silva and Julia Sullivan (all URI) assisted with PE analysis. Heather Stoffel, Ed Requintina (both URI), Patrick McGuire of RI-DEM OWR and MERL generously assisted us in PE deployment and retrieval. This study was supported by the Rhode Island Department of Environmental Management, Contract MPA R74A050017. Victoria Dekany acknowledges a NSF-sponsored Summer Undergraduate Research Fellowship in Oceanography (SURFO) at URIGSO. We thank Michael Pilson (URI) for helpful comments.
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