Lipid-free tubing PSDs were constructed from low-density polyethylene tubing and fortified with perdeuterated performance/permeability reference compounds (PRCs) using methods described elsewhere [29
]. Briefly, additive-free tubing was cleaned with hexanes to remove potential analytic interferences, the tubing was heat sealed at one end, then the PRCs were injected into the interior of tubing and it was sealed at the other end. The final dimensions of each sampler are 2.6 cm wide by 1 m long. PSDs were stored in sealed Teflon bags until use and storage stability quality control samples were maintained throughout the study.
Stainless steel cages containing five PSDs were deployed from piers in coastal marine waters. One set of samplers, infused with PRCs, was deployed at each site during each sampling event. Sampling cages were suspended in the water column, at least 1 m above the bottom. Water depths varied by site and tide cycle between 2–8 m. PSDs can sequester contaminants that are not freely dissolved in the water column if the sampler membrane comes in direct contact with non-aqueous phase media, such as oil sheens or droplets. To avoid this, precautionary measures were taken to prevent contact of the sampling material with crude oil floating on the surface of the water during deployments. When surface oil or sheen was visible, samplers were lowered into the water sealed in plastic bags that were removed after the samplers were secured below the surface in the water column. Based on visual inspection of the samplers upon retrieval from the water and results that show dissolved concentrations below solubility limits for all analytes in all samples, there is no evidence that the samplers in this study were ever superficially contaminated by oil.
Samplers were deployed at four sites: Grand Isle, Louisiana, Gulfport, Mississippi, Gulf Shores, Alabama and Gulf Breeze, Florida (). The site at Grand Isle, LA was located the closest to the source of the spill and had little natural or human-devised physical protection from the influence of Gulf of Mexico waters during the oil spill. The site in Gulfport, MS was afforded limited protection from oiling by offshore barrier islands, which were heavily oiled during the spill. The sampling sites in Bon Secour, AL and Gulf Breeze, FL were at the mouths of Mobile Bay and Pensacola Bay, respectively. They were more protected from direct oiling than the other sites because of the natural peninsulas that delimit the bays as well as the booms that were put in place to protect those areas. All of the sites are impacted, to varying degrees, by local urban and industrial activities, which highlights the importance of having pre-oiling baseline data for comparison when determining the impacts of the spill on these coastal waters.
Sampling locations at four states in the Gulf of Mexico
These sites were chosen based on the Mississippi Canyon Trajectory Forecasts that the National Oceanic and Atmospheric Administration (NOAA) began producing after oil was observed on the surface of the ocean. Basic geography and information about dominant ocean currents in the Gulf of Mexico were also taken into consideration. Additional criteria for site selection included their accessibility to researchers and protection for sampling equipment from theft and vandalism. All four research sites are located in shallow, coastal waters and are accessed by piers or docks.
The first sampling event began on May 10th, 2010 and was completed prior to shoreline oiling at any of the study sites. A total of nine sampling events were conducted over the course of more than a year (names in quotes refer to the shorthand used in text and figures): ‘May 2010’ (May 10–13, 2010), ‘June (1)’ (June 8–11, 2010), ‘June(2)’ (June 11–July 7, 2010), ‘July’ (July 7–August 5, 2010), ‘August’ (August 5–September 8, 2010), ‘September’ (September 8–October 13, 2010), ‘March’ (February 9–March 15, 2011), ‘April’ (March 15–April 29, 2011), ‘May 2011’ (April 29–June 8, 2011). Samplers were not recovered from Grand Isle, LA in July, 2010. Sampling deployment times varied throughout the study due to practical considerations involving travel, weather and site accessibility, as well as other factors. The first deployment, in May, 2010, was limited to four days because of projected impending shoreline oiling at the site in Grand Isle, LA and an interest in obtaining pre-oiling data at all sites. The sampling period in early June was limited to four days because of concern about the samplers becoming saturated at sites that were being heavily oiled. Analysis of those samples indicated that longer deployment periods could be used for the remainder of the study. Because PRCs were used to calculate uptake rates it was not necessary for the deployment times to be the same throughout the study or for the samplers to reach equilibrium.
PSDs were retrieved from the field and transported to the laboratory at Oregon State University. Samplers were transported at ambient temperature in sealed Teflon bags. Storage stability studies, conducted prior to this research, verified that transport of PSDs in sealed Teflon bags at ambient temperatures (up to 50 °C) for up to two weeks does not result in a significant loss of PAH analytes. Transportation quality control samples, fortified with PRCs and PAH analytes, were used throughout the study. Recoveries of PAH analytes and PRCs from storage stability samples and fortified trip blanks did not exceed ±10% of the true value.
In the laboratory, the samplers were cleaned with hydrochloric acid and isopropanol to remove superficial fouling, mineral salts and water [13
]. Perdeuterated PAH surrogate recovery standards were spiked on the PSD samplers prior to extraction to allow for verification of extraction efficiency and recovery correction. The 5 PSDs from each cage were extracted together as one sample to increase detection capabilities. Samplers were extracted by dialysis in n
-hexane using methods detailed elsewhere [26
]. Briefly, samplers were immersed in 200 mL of n
-hexane for 4 hours, the dialysate was decanted, then dialysis was repeated for 2 hours and the dialysates were combined. Samples were quantitatively concentrated to a final volume of 1 mL.
Solvents used for pre-cleaning, clean-up and extraction were Optima® grade or better (Fisher Scientific, Pittsburgh, PA). The following 33 PAH analytes were included in analyses: naphthalene, 1-methylnaphthalene, 2-methylnaphthalene, 1,2-dimethylnaphthalene, 1,6-dimethylnaphthalene, acenaphthylene, acenaphthene, fluorene, dibenzothiophene, phenanthrene, 1-methylphenanthrene, 2-methylphenanthrene, 3,6-dimethylphenanthrene, anthracene, 2-methylanthracene, 9-methylanthracene, 2,3-dimethylanthracene, 9,10-dimethylanthracene, fluoranthene, pyrene, 1-methylpyrene, retene, benz(a)anthracene, chrysene, 6-methylchrysene, benzo(b)fluoranthene, benzo(k)fluoranthene, benzo(e)pyrene, benzo(a)pyrene, indeno(1,2,3-cd)pyrene, dibenz(ah)anthracene, benzo(ghi)perylene and dibenzo(al)pyrene. The perdeuterated PAH compounds used as PRCs were fluorene-D10, p, p′-DDE-D8and benzo(b)fluoranthene-D10. The following perdeuterated PAHs were used as surrogate recovery standards: naphthalene-D8, acenaphthylene-D8, phenanthrene-D10, fluoranthene-D10, pyrene-D10, benzo(a)pyrene-D12 and benzo(ghi)perylene-D12; and perylene-D12 was the internal standard.
PSD extracts were analyzed using Agilent 5975B Gas Chromatograph-Mass Spectrometer (GC-MS); with a DB-5MS column (30 m × 0.25 mm × 0.25 μm) in electron impact mode (70 eV) using selective ion monitoring (SIM). The GC parameters were as follows: injection port maintained at 300 °C, 1.0 mL min−1 helium flow, 70 °C initial temperature, 1 min hold, 10 °C min−1 ramp to 300 °C, 4 min hold, 10 °C min−1 ramp to 310 °C, 4 min hold. The MS temperatures were operated at 150, 230 and 280 °C for the quadrupole, source and transfer line respectively. Sample concentrations were determined by the relative response of the deuterated surrogate to the target analyte in a nine point calibration curve with a correlation coefficient greater than 0.98.
Over 30% of the total number of samples analyzed in this study corresponded to quality control samples, which included PSD fabrication blanks, field and trip blanks for each deployment/retrieval, laboratory clean-up blanks and reagent blanks. All target compounds were below the detection limit in all blank quality control samples.
Mean surrogate standard recoveries varied between 48–102% for naphthalene-D8 and benzo(g, h, i)perylene-D12 respectively. Lower recoveries were observed for 2–3 ring PAHs, which are relatively volatile, due to losses during sample preparation, especially sample concentration. Target analytes were recovery corrected, following analysis, based on the measured recovery of the surrogate with the most similar structure. The average relative standard deviation (RSD) for all analytes from replicate samples was 7.5%. Naphthalene had the highest RSD; averaging 21%. This variability is attributed to differences in losses during the sample concentration steps and is not significantly different from RSDs observed during the 500:1 concentration of surrogates overspiked in n-hexanes.
The method detection limits for PAH analytes in samples obtained from a composite of 5 PSDs was 10 pg/uL. This translates into detection limits ranging between 0.001–0.05 ng/L for individual PAH compounds in water. Calibration curves had a fit that was greater than 0.98 for all analytes in the method. Calibration verification standards for target analytes, surrogates and PRCs were analyzed at least every 10 samples and reported values within ±15% of the true value were considered acceptable. Only results from samples run between two calibration verifications that met the quality control criteria were accepted.
Water Concentration Calculation
Water concentrations were calculated using the empirical uptake model with PRC-derived sampling rates [13
]. The equations used to calculate the water concentrations presented in this study are detailed in the Supporting Information
. This model is based on uptake kinetics and does not require any assumptions about individual analytes being at equilibrium or in the linear uptake range. The use of PRCs allows for an accurate determination of in situ
, site-specific sampling rates under variable exposure conditions, including variable temperatures, flow rates and biofouling [31
]. Additionally, it is not necessary for the analytes of interest to reach equilibrium with the sampler in order to determine sampling rates [13
]; therefore, variable sampling deployment times are feasible. Fluorene-D10, benzo(b)fluoranthene-D10 and p, p′-DDE-D8 PRCs were used in the calculations. PAH compounds and p, p′-DDE have similar compound specific effects on sampling rates [13
]; therefore water partitioning coefficients for these compounds can be calculated with the same equation, based on log Kow
. These PRCs cover a range of log Kow
values that makes them adequate for deriving the uptake rates of the PAH analytes included in this study [31
]. When PRC recoveries were below 20% or above 80%, the sampling rates were determined using an improved model for calculating in situ
sampling rates when recoveries approach 0 or 100% [14
] (details provided in Supporting Information
For comparisons of total PAHs, all 33 PAH analytes were summed. Two-way comparisons between different sampling events at the same site were performed using the Wilcoxon signed-rank test. For sums and two-way comparisons, analyte concentration values below the detection limit were equal to zero. Probabilities less than p=0.05 were considered significant.
Confidence intervals were calculated for graphical representation of the data and are reported in the results. Numerous replication studies performed by this laboratory, in diverse environments, over the last ten years have demonstrated that PSDs give highly reproducible results; the variability in the data is consistent and predominantly a result of sample processing and analysis. Due to practical considerations it is often unfeasible to deploy replicate samplers. In these cases, the measured concentrations of analytes are representative only of the specific sampling location; the confidence interval is calculated based on pooled variance from replication pilot studies and represents a statistically defensible measure of variance around the reported value. In this study, the reported values were measured in the environment and the confidence intervals are calculated measures of methodological variance. The interpretation of the results should be limited to the areas where direct measurements were made.
For other analyses, data were standardized to avoid a magnitude bias when analyzing chemical profiles. Sample measurements were scaled to reflect relative abundances by representing individual analyte concentrations as percentages of the total PAHs measured in a given sample.
Principal component analysis (PCA) is a multivariate variable reduction technique in which principal components (PCs) are calculated as combinations of the original variables. The goal of PCA is to express as much of the total variation as possible with a few uncorrelated PCs. Use of PCA can reveal important features obscured within the original data and has been applied to PAH fingerprinting and allocation studies [23
]. In this case PCA was used to explore similarities, differences and changes in chemical profiles of samples obtained from the study sites over the course of more than a year. PCA was performed using all of the analytes from each sample. The resulting PCs were analyzed graphically for apparent similarities and differences between samples including spatial and temporal tendencies.
SAS 9.2 (SAS Institute Inc.) was used for statistical analyses and modeling. Graphics were created using Sigma Plot 11.0 (Systat Software Inc.).