Study of residual oil in Bay Jimmy sediment 5 years after the Deepwater Horizon oil spill: Persistence of sediment retained oil hydrocarbons and effect of dispersants on desorption

STOTEN-24122; No of Pages 10 Science of the Total Environment xxx (2017) xxx–xxx

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Study of residual oil in Bay Jimmy sediment 5 years after the Deepwater Horizon oil spill: Persistence of sediment retained oil hydrocarbons and effect of dispersants on desorption Jun Duan a, Wen Liu a, Xiao Zhao a, Yuling Han a, S.E. O’Reilly b, Dongye Zhao a,c,⁎ a b c

Environmental Engineering Program, Department of Civil Engineering, Auburn University, Auburn, AL 36849, USA Bureau of Ocean Energy Management, GOM Region, Office of Environment, New Orleans, LA 70123, USA Beijing University of Civil Engineering and Architecture (BUCEA), Beijing 100044, PR China

H I G H L I G H T S

G R A P H I C A L

A B S T R A C T

• Biomarker analysis showed the DwH oil persisted in Bay Jimmy sediment after 5 years. • The TPHs level remained orders of magnitude higher than in the pre-spill level. • Nearly all C9–C20 n-alkanes and 2-ring PAHs were degraded upon 5-year weathering. • Dispersants promote desorption of residual oil from sediment, especially for PAHs. • Over 57% of sediment adsorbed TPHs is resistant to desorption even with dispersants.

a r t i c l e

i n f o

Article history: Received 18 July 2017 Received in revised form 21 September 2017 Accepted 21 September 2017 Available online xxxx Editor: D. Barcelo Keywords: Oil residual Sediment Oil weathering Desorption Dispersant Oil spill

a b s t r a c t The 2010 Deepwater Horizon (DwH) oil spill contaminated ~1,773 km of the Gulf of Mexico shorelines. Yet, few field data are available on the long-term fate and persistency of sediment-retained oil. While an unprecedented amount of oil dispersants was applied, the effects of oil dispersants on desorption of field aged oil remain unknown. This study aimed to investigate the abundance, distributions and physico-chemical availability of the oil retained in Bay Jimmy sediment, Louisiana, five years after the DwH oil spill, and to determine the effects of two model oil dispersants on the desorption potential of the residual oil. Total petroleum hydrocarbons (TPHs), n-alkanes and polycyclic aromatic hydrocarbons (PAHs) in the sediment were analyzed and compared with those in the crude oil and the pre-DwH levels, and batch desorption kinetic tests were carried out to quantify the dispersant effects on the desorption rate and extent. The biomarker hopanes profile and diagnostic ratio were analyzed, which confirmed the origin and persistence of the sediment-retained oil. After five-year natural weathering, the oil level in the sediment remained orders of magnitude higher than the pre-spill level. Nearly all low-molecular-weight n-alkanes and 2-ring PAHs had been degraded. Oil dispersants, SPC 1000 and Corexit EC9500A, were able to enhance solubilization of the sediment-retained oil upon resuspension of the sediment. Successive desorption experiments indicated that 71.6% of TPHs, 74.8% of n-alkanes, and 91.9% of PAHs in the sediment remained highly stable and hardly desorbable by seawater; yet, addition of 18 mg/L of SPC 1000 enhanced the desorption and lowered these fractions to 57.3%, 68.1%, and 81.4%, respectively. The findings are important for understanding the natural weathering rate and persistence of oil residual and the effects of dispersants on the physical and biological availabilities of aged oil in coastal sediments. © 2017 Elsevier B.V. All rights reserved.

⁎ Corresponding author at: BUCEA, China and Auburn University, USA. E-mail address: [email protected] (D. Zhao).

http://dx.doi.org/10.1016/j.scitotenv.2017.09.234 0048-9697/© 2017 Elsevier B.V. All rights reserved.

Please cite this article as: Duan, J., et al., Study of residual oil in Bay Jimmy sediment 5 years after the Deepwater Horizon oil spill: Persistence of sediment retained oil hy..., Sci Total Environ (2017), http://dx.doi.org/10.1016/j.scitotenv.2017.09.234

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1. Introduction Oil spill has been a worldwide challenge in the modern society, which not only causes substantial economic loss, but also poses serious threats to the environmental and human health. For example, the 1978 Amoco Cadiz wreck contaminated some 360 km of shores, salt marshes and estuaries of northern Brittany, France, and caused severe mortality and growth impedance to various coastal species (Conan et al., 1982); the 1983 Castillo de Bellver incident in South Africa spilled 2.93 × 105 m3 of crude oil (Moldan et al., 1985); and the 1991 Gulf War discharged an estimated 160 to 340 million gal (0.6 × 106–1.3 × 106 m3) of Kuwait crude oil into the Gulf, which caused lingering environmental disasters long after the war (Sauer et al., 1993). From 1970 to 1972, the spills in the Nipisi area near Edmonton, Canada, spilled ~ 60,000 barrels over 25 acres of land (Wang et al., 1998). The 2010 Deepwater Horizon (DwH) oil spill in the Gulf of Mexico (GOM) is considered one of the largest accidental marine oil spill in the history of the petroleum industry (Redmond and Valentine, 2012). The spill released around 7.94 × 10 5 –1.11 × 10 6 m 3 crude oil into the waters of GOM (Reddy et al., 2012). To mitigate the impacts of the oil spill, a total of 1.84 million US gal (7,000 m3) of dispersants (Corexit EC9500A and Corexit EC9527A) were applied at the wellhead and on the ocean surface (Kujawinski et al., 2011). Around 2.9 × 10 6 gal (769 m 3 ) of Corexit EC9500A was used during the DwH oil spill. The dispersant contains one anionic and three nonionic surfactants, and three solvents (Gong et al., 2014a). Corexit EC9500A was formulated based on Corexit EC9527A, but with enhanced penetration and emulsion fighting properties and lower toxicity (Mitchell and Holdway, 2000). Although the oil well is about 41 miles (66 km) off the southeast coast of Louisiana, and despite various response measures implemented, for example, the use of oil dispersants, surface skimming and controlled burns, large amounts of oil still reached beaches, wetlands, marshes and estuaries, resulting in extensive damage to marine and wildlife habitats and fishing and tourism industries (Tangley, 2010; Juhasz, 2012). For instance, in Louisiana, 2,090 tons of oily material was removed from the beaches in 2013 (NPR, 2013). Over 1,773 km of the shoreline was contaminated, with most of the heavily and moderately oiled shoreline located in the state of Louisiana (Michel et al., 2013). Bay Jimmy salt marshes in the northern Barataria Bay experienced some of the largest amounts of oil accumulation (Silliman et al., 2012; Zengel and Michel, 2013). More than 79% of the marshes in the northern Barataria Bay including Bay Jimmy and nearby coastal area were found oiled in July–October 2010 (Kokaly et al., 2013). Before the DwH oil spill, the concentrations of the petroleum hydrocarbons in the sediment were quite low. Kırman et al. (2016) analyzed the pre-DwH spill sediments of Bay Jimmy area and reported the total petroleum hydrocarbon (TPHs) concentration was in the range from 4 to 90 mg/kg. Iqbal et al. (2007) evaluated the pre-DwH petroleum pollution in southern Louisiana, and they found the total (polycyclic aromatic hydrocarbons) PAHs concentration in 90% of the sampled sediments was b2.0 mg/kg, and the concentration of total saturated alkanes (C9–C35) was b17 mg/kg. However, seven months after the DwH spill, the TPHs concentration in the Bay Jimmy marsh sediment surged to as high as 510 mg/g (Lin and Mendelssohn, 2012), and the total PAHs concentration reached up to 18,279 mg/kg one year after the DwH spill, which is around 9000 times higher than the levels in the pre-contaminated sediments (Hatch et al., 2013). While oil in the water column appears to have disappeared several years after the DwH spill, residual oil continued to be found in beaches, wetlands and estuaries, and lingering adverse effects continue to be reported. Turner et al. (2014) reported that the concentrations of alkanes and PAHs in the GOM wetlands in June 2013 were about 100 and 20 times lower, respectively, than the levels measured in February 2011, but remained at 3.7 and 33 times higher than those in May 2010. The authors also estimated that the concentration of

alkanes may return to near baseline values by the end of 2015. Viegas (2013) reported that in 2013 dolphins and other marine life in the affected area continued to die with infant dolphins dying at six times the normal rate. Once reached shorelines, oil hydrocarbons will undergo strong adsorption onto the sediments. It has been well known that sediment adsorbed petroleum hydrocarbons are more resistant to various weathering processes, such as biodegradation and photodegradation (Tansel et al., 2011; Yin et al., 2015). The slow dissipation and weathering rate of sediment-adsorbed oil may pose an extended period of risks to the marine ecosystem and human health since the sediment retained oil may be slowly released upon sediment resuspension or changes in the surrounding environmental conditions such as concentration gradient and the presence of surfactants or dispersants (Liu et al., 2012; Yin et al., 2015). Field data on the weathering of sediment-retained oil hydrocarbons have been very limited. In general, oil weathering, including physical processes, bio-degradation, and photo-oxidation, begins within the first few days of a spill and can continue for months to years (Sauer et al., 1993). As a rule, biodegradation typically occurs more slowly than physical-chemical weathering, and the low molecular weight alkanes (C10–C22) are metabolized most rapidly followed by low molecular weight aromatic hydrocarbons, higher molecular weight alkanes, the isoprenoids (branched alkanes, pristane and phytane), higher molecular weight PAHs and lastly the asphaltenes and resins (Sauer et al., 1993). A field study following the 1991 Gulf war oil spill showed that sediments from exposed habitats generally exhibited more oil weathering than the moderately exposed and sheltered habitat (Sauer et al., 1993, 1998). Two years after the spill, some sediments still contained relatively fresh oil, while other sediments showed varying stages of weathering (fresh-advanced). Seventy percent of the subsurface samples contained less weathered residual oil, compared to 20% for the surface samples (Sauer et al., 1998). A long-term study of the Nipisi spill site indicated that 15%–43% of residual oil in the surface samples (0–4 cm) was clearly degraded, while subsurface samples (10–40 cm) exhibited great quantities of oil even 25 years after the spills, indicating that the extent of contamination and degree of degradation correlated strongly with sample depth (Wang et al., 1998). While these limited field data indicate some general weathering trend for persistent oil components such as NC18 alkanes and PAHs in sediments (Turner et al., 2014), the long-term fate and persistence of the sediment-retained residual oil remain unknown. Consequently, there is a need to understand the status of the residual oil in the GOM sediments and to develop a systematic approach toward predicting the persistence of the residual oil. Whereas oil dispersants are used to disperse oil slicks in the water column and promote biodegradation, recent works showed that oil dispersants also facilitate transport of oil into sediments (Gong et al., 2014a; Zhao et al., 2015). For instance, Gong et al. (2014a) observed that not only can Corexit EC9500A enhance sediment uptake of dissolved oil PAHs and alkanes, the dispersant can also induce significant desorption hysteresis, i.e., the presence of the dispersant caused the desorption isotherm to be higher than the adsorption isotherm due to the added adsorption capacity from the adsorbed dispersant. Therefore, dispersants may greatly affect the physical, chemical and biological availabilities of oil. However, the effects of oil dispersants on desorption of oil, especially field aged oil, remain unknown. The overall goal of this study was to study the status and desorption potential with or without oil dispersants of 5-year aged oil from the DwH oil spill in Bay Jimmy sediment. The specific objectives were to: 1) identify the origin of the residual oil, 2) understand the abundance and distribution of the residual oil components in the sediment; 3) study the effects of model oil dispersants on the oil desorption process; and 4) determine the maximum desorption potential of the residual oil washout via successive desorption tests. The information would aid in our understanding of the natural weathering rate and persistence of oil residual in coastal sediments.

Please cite this article as: Duan, J., et al., Study of residual oil in Bay Jimmy sediment 5 years after the Deepwater Horizon oil spill: Persistence of sediment retained oil hy..., Sci Total Environ (2017), http://dx.doi.org/10.1016/j.scitotenv.2017.09.234

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2. Materials and methods 2.1. Materials The sediment was collected from a salt marsh area at Bay Jimmy, Louisiana, in March 2015, about five years after the 2010 DwH oil spill. Fig. 1 shows the map and approximate locations of the sampling site and spots. Five surface sediment samples (0–5 cm) were taken at an intertidal marsh site near N29°26′27.78″/W89°54′6.95″ (Fig. 1), where oil residual was still visible and some historical data are available. Based on preliminary analysis, the samples contained comparable levels of oil TPHs (b 30% differences), and thus were combined and used in this study. Earlier, Kırman et al. (2016) reported compositions and depth distribution of oil hydrocarbons before and after the DwH spill based on two sediment cores at this site. The site was selected because of the following considerations: 1) the site was known to be impacted by the DwH spill as stated earlier; 2) the site has been studied by several researchers and both pre- and post-DwH data have been reported (Kırman et al., 2016; Kokaly et al., 2013; Silliman et al., 2012; Zengel and Michel, 2013), which are essential for assessing long-term oil weathering; and 3) there has been no known major man-made or natural interference at the site. The sediment samples were stored in sealed bottles placed in a refrigerator (4 °C) according to the standard methods on sediment sample storage and handling (Zirbser et al., 2001), and were tested within one year. Our analysis confirmed that there was no significant change in the concentrations of oil hydrocarbons during the storage. Before use, the raw sediment samples (water content = 46 wt.%) were combined and well mixed. Sediments analysis was conducted by Soil Testing Laboratory at Auburn University, and the specific analytical methods have been described elsewhere (Gong et al., 2014a). Table S1 in the Supplementary Materials (SM) gives the salient properties of the sediments. In general, the sediment texture was classified to be silty loam, with a pH of 7.4 and a sediment organic matter (SOM) content of 8.8 wt.%. Seawater was collected from the top 30 cm of the water column from Grand Bay, Alabama (N30°22′43.4″/W88°18′24.4″), and used as a model ‘clean’ GOM seawater in the desorption experiments. The key seawater quality parameters included: pH = 7.9, dissolved organic matter (DOM) = 0.77 mg/L as total organic carbon (TOC), and salinity = 3.15 wt.%. The concentration of n-alkanes in the seawater was below 0.5 μg/L, and the total concentration of PAHs was below 0.013 μg/L (Gong et al., 2014a). The seawater sample was filtered through a membrane (0.45 μm, polyamide) and then sterilized by autoclaving at 121 °C

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for 1 h. To inhibit any microbial activities, 200 mg/L anhydrous sodium sulfate and sodium azide (NaN3) was added into the seawater before the subsequent desorption experiments. All chemicals and organic solvents used in this study were of analytical or higher grades. Hexane, anhydrous sodium sulfate and NaN3 were purchased from Fisher Scientific (Fair lawn, New Jersey), and dichloromethane (DCM) from Alfa Aesar (Ward Hill, MA, USA). Silica gel (60– 200 μm) was acquired from Sigma-Aldrich (St. Louis, Missouri) and was activated prior to each use per a protocol by Wang et al. (1994). Briefly, the silica gel was rinsed consecutively with acetone, hexane and DCM. After air-dried, the silica gel was heated in the oven at 50 °C for 8 h and then at 180 °C for 20 h. A standard reagent consisting of the 16 EPA listed PAHs specified in EPA Method 610, a standard mixture of n-alkanes (C9–C40), Pristane (Pr), Phytane (Ph), and two internal standards (5α-androstane for n-alkanes and fluorene-d10 for PAHs) were purchased from Supelco (Bellefonte, PA, USA). Dispersants Corexit EC9500A was acquired per courtesy of Nalco Company (Naperville, IL, USA) and SPC 1000 was purchased from Polychemical Corporation (Chestnut Ridge, NY, USA). Both dispersants were used as received upon proper dilution. 2.2. Extraction of sediments Extraction of various oil components from the contaminated sediment was conducted per a modified protocol recommended by Wang and Fingas (1997). Based on a thorough review of several methods used for the characterization and quantification of petroleum hydrocarbons, such as EPA Methods 418.1, 602, 610 and 624, Wang and Fingas (1997) concluded that the protocol is generally more selective, sensitive, and more reliable. This method has been widely used (with over 190 citations) and remains being used by recent researchers (e.g., Mikkonen et al., 2012; Zhao et al., 2016). Our own results also showed that this method offered better mass recoveries and detection limits than the EPA methods for various oil components including TPHs, n-alkanes, PAHs and biomarkers. Briefly, each 2 g of a sediment sample was sequentially extracted by 30 mL of DCM and hexane mixture at a 1:1 volume ratio in a batch setting for 24 h at room temperature. Upon separation of the solid via centrifugation (5000 rpm, 10 min), the sediment was further extracted using 30 mL of DCM twice consecutively. Afterwards, the extracts were combined and passed through a column filled with 10 g of anhydrous Na2SO4 to remove moisture. The solvent was then concentrated under nitrogen purging to 5 mL for further processing or analysis.

Fig. 1. Approximate sediment sampling locations at Bay Jimmy, LA, USA. The coordinates in the inset indicate the sampling spot included in a previous study by Kırman et al. (2016), and the site was also covered by another study by Turner et al. (2014).

Please cite this article as: Duan, J., et al., Study of residual oil in Bay Jimmy sediment 5 years after the Deepwater Horizon oil spill: Persistence of sediment retained oil hy..., Sci Total Environ (2017), http://dx.doi.org/10.1016/j.scitotenv.2017.09.234

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The concentrated extract was then passed through a column packed with 3 g of the activated silica gel that was placed on top of 3 g of anhydrous sodium sulfate, which was supported by a layer of glass wool at the bottom. The silica gel was able to adsorb all the oil hydrocarbons. The column retained oil hydrocarbons were then fractionated through the following solvent elution procedures (Wang and Fingas, 1997; Yim et al., 2011; Zhao et al., 2016). First, the oil-laden silica gel column was eluted with 12 mL of hexane, and the eluted fraction was labeled as F1. One half of F1 was used to analyze the saturated hydrocarbons (n-alkanes) and the biomarker hopanes. Then, the column was further eluted using 15 mL of a mixture of DCM and hexane (volume ratio = 1:1) to get fraction F2, one-half of which was used to analyze parent and alkylated PAHs. Lastly, the other halves of F1 and F2 were combined (labeled as F3), and F3 was analyzed for TPHs. Throughout this paper, the oil component concentrations are given based on the air-dried sediment weight. 2.3. Desorption tests The desorption rate and extent of various oil components from the sediment were tested using plain seawater and in the presence of dispersants Corexit EC9500A or SPC 1000, respectively. SPC 1000 is a water-based dispersant that has been reviewed by EPA and is included in the National Contingency Plan (NCP) Product Schedule for use as a dispersant for oil spills (Jacob and Bergman, 2001). Dispersant SPC 1000 is a unique aqueous composition with highly effective emulsifiers, surfactants and a water-soluble coupling solvent. SPC 1000 contains no petroleum solvents. Dispersant SPC 1000 has an overall effectiveness of 73%, which is almost 50% better than the EPA requirement and is considered less toxic than the petroleum based dispersants (Jacob and Bergman, 2001). For each desorption batch test, 2 g of the wet sediment was mixed with 40 mL sterilized seawater in the presence of 0 or 18 mg/L of the dispersants in a 43 mL amber glass vials sealed with Teflon-lined caps, where almost no headspace was left to minimize volatilization loss of the oil compounds. The dispersant dosage (18 mg/L) was used based on previous studies (e.g., Gong et al., 2014a; Zhao et al., 2016). It corresponds to a dispersant-to-oil ratio (DOR) of 1:24, which is in the middle of the manufacturer recommended DORs (1:50 to 1:10) (Hemmer et al., 2011). Practically, this concertation may simulate an engineered remediation effort (e.g., sediment washing) or a scenario where a dispersant is freshly applied following an oil spill. The vials were tightly sealed and rotated on an end-to-end rotator (60 rpm, 25 °C) in the dark. At predetermined times (0.3, 0.5, 1, 2, and 4 days), the vials were sacrificially sampled and extracted with DCM three times consecutively (40 mL solution with 10 mL DCM in each extraction). The extracts were combined and cleaned up by passing through an anhydrous sodium sulfate column to remove moisture and fine suspended particles (Wang and Fingas, 1997; Zhao et al., 2016), and then concentrated under a gentle nitrogen blow to around 5 mL. Successive desorption tests were also conducted to determine the maximum desorbable oil components in the sediment. Following each apparent desorption equilibrium (which was reached in ~4 days), the vials were centrifuged and supernatants pipetted out, and replaced with either seawater only or seawater containing 18 mg/L of SPC1000. At predetermined times (0.3, 0.5, 1, 2, and 4 days), the vials were sacrificially sampled, and the supernatants were extracted and analyzed for the remaining oil components in the aqueous phase following the same procedures as described above. The successive desorption tests were carried out in triplicate to assure data quality. 2.4. Chemical analysis TPHs in the extracts were analyzed using GC-FID (Agilent 6890 Series GC System, USA) equipped with a DB5 column (30 m × 0.25 mm, 0.25 μm film thickness). The injection volume was 1 μL with a split ratio of 20. The column temperature was programed to ramp from 40

to 280 °C at a rate of 8 °C min−1 and held at 280 °C for 60 min (Liu et al., 2012). Oil n-alkanes (C9–C40), PAHs (the 16 parent PAHs specified in EPA Method 610 and the corresponding alkylated PAHs), pristane, phytane and hopanes were analyzed using GC–MS (Agilent 7890A GC coupled with the 5975C Series MS, Agilent Technologies Inc., Santa Clara, CA, USA). The analytical method was optimized and the selected ion monitoring (SIM) mode was set up based on the previous reports and the NIST library in GC–MS (Wang et al., 2007; Liu et al., 2012). A DB-EUPAH column (length 20 m; inner diameter 0.18 mm; film thickness 0.14 μm) was used to separate the analytes. The front inlet temperature was set at 250 °C. The GC oven temperature was programmed as follows: 50 °C (hold for 0.8 min), ramp to 180 °C at 40 °C min− 1, ramp to 230 °C at 7 °C min−1 (hold for 0.5 min), and ramp to 335 °C at 15 °C min−1 (hold for 5 min). The sample injection volume was 2 μL. Standard curves were developed for each PAH following the EPA Method 610 and the alkylated homologs were quantified according to a standard method using straight baseline integration of various levels alkylated PAHs (Wang et al., 2007). The quantification of the biomarker hopanes was achieved by using a characteristic ion m/z of 191 in the SIM mode, and the specific hopanes were determined based on the standard retention times (Mulabagal et al., 2013; Peters et al., 2005). 3. Results and discussion 3.1. Source identification In addition to the aforementioned evidence documenting the impacts of the DwH oil spill on the TPHs levels in the Bay Jimmy sediment, direct fingerprinting evidence was sought to confirm the residual oil was indeed from the oil spill. In general, crude oil formed under different geological conditions possesses distinctive biomarker fingerprints and these biomarkers are relatively more resistant to weathering compared to other oil components (such as n-alkanes and PAHs) (Mulabagal et al., 2013; Wang et al., 2006). Therefore, the chemical fingerprints of biomarkers in oil samples can help identify the source of oil hydrocarbons. Hopanes and steranes are the most cited biomarkers due to their high abundance and resistance to weathering. Earlier, researchers developed chromatographic fingerprints of hopanes and steranes of DwH oil using GC–MS in the SIM mode and using the characteristic ions at m/z 191 and 217, respectively (John et al., 2016; Mulabagal et al., 2013; Yin et al., 2015). In this study, the fingerprints of hopanes were compared between the MC252 reference crude oil (the DwH oil) and the oil from the Bay Jimmy sediment. Fig. 2 compares the characteristic chromatograms of hopanes from the MC252 reference oil and those from the Bay Jimmy sediment oil. It is evident that the chromatographic profiles nearly resemble each other, indicating a very high probability that the Bay Jimmy oil originated from the DwH oil spill. This is also backed by the following evidences: 1) the historical data before and after the DwH spill, e.g., those reported by Kırman et al. (2016), provided compelling evidence that the oil hydrocarbons at the site are associated with the DwH spill; 2) studies following the DwH event showed that the Macondo Well was not leaking after the shut in on July 15, 2010 (Hickman et al., 2012); and 3) the site was quite isolated from human activities, and thus, no opportunistic or man-made spills are evident. The relative ratios of source-specific hopanes have been used to identify oil origins (Mulabagal et al., 2013; Wang et al., 2006). Therefore, seven hopane diagnostic ratios were further analyzed to reveal the sample origin, including the ratios of T s (18α(H)22,29,30-Trisnorneohopane)/Tm (17α(H)-22,29,30-trisnorhopane), C29 (17α(H),21β(H)-30-norhopane)/C30 (17α(H),21β(H)-hopane), C31(22S)/C31(22S + 22R), C32(22S)/C32(22S + 22R), C33(22S)/C33(22S + 22R), C34(22S)/C34(22S + 22R) and C35(22S)/C35(22S + 22R) (John et al., 2016; Mulabagal et al., 2013). Table 1 gives these diagnostic ratios, and Fig. S1 in SM shows the radar plots of the ratios. The

Please cite this article as: Duan, J., et al., Study of residual oil in Bay Jimmy sediment 5 years after the Deepwater Horizon oil spill: Persistence of sediment retained oil hy..., Sci Total Environ (2017), http://dx.doi.org/10.1016/j.scitotenv.2017.09.234

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Table 2 Concentrations of TPHs extracted from the Bay Jimmy sediment via sequential extraction. Extraction time

1st 2nd 3rd Sum

Fig. 2. Chromatographic fingerprints of hopanes (m/z of 191) in MC252 crude oil and in 5-year aged oil in Bay Jimmy sediment. Ts:Ts (18α(H)-22,29,30-Trisnorneohopane), Tm:Tm (17α(H)-22,29,30-trisnorhopane), C29:C29 (17α(H),21β(H)-30-norhopane), C30:C30 (17α(H),21β(H)-hopane), C31S:C31 (17α,21β-Homohopane (22S)), C31R:C31 (17α,21β-Homohopane (22R)), C32S:C32 (17α,21β-Bishomohopane (22S)), C32R:C32 (17α,21β-Bishomohopane (22R)), C33S:C 33 (17α,21β-Trisnornohomohopane (22S)), C33R:C 33 (17α,21β- Trisnornohomohopane (22S)), C34S:C 34 (17α,21βTetrahomohopane (22S)), C34R:C34 (17α,21β-Tetrahomohopane (22R)), C35S:C35 (17α,21β-Pentahomohopane (22S)), C35R:C 35 (17α,21β-Pentahomohopane (22R)). All peaks are normalized to the respective C30 value.

diagnostic ratios for the two oil samples well matched each other, which confirmed that the Bay Jimmy oil was from the 2010 DwH oil spill. The carbon preference index (CPI) is defined as the relative abundance of odd-numbered n-alkanes to that of even-numbered ones (Ehrhardt and Petrick, 1993). The CPI of the Bay Jimmy oil was measured to be 0.97, which is close to the reported CPI value of 1 for MC252 oil (Liu et al., 2012). The Pr/Ph ratio was calculated to be 0.91, which is also consistent with that for MC252 oil (0.90) (Liu et al., 2012). 3.2. Residual oil components in Bay Jimmy sediment after 5 years of weathering 3.2.1. Residual TPHs in sediment The type and concentrations of various residual oil components in the Bay Jimmy sediment were determined. The concentration of TPHs in the sediment was determined to be 2.31 mg/g-sediment (Table 2). It was shown that 2.03 mg/g-sediment or 87.7 wt.% of the TPHs was recovered from the first extraction, 10.2 wt.% more from the second

Table 1 Diagnostic ratios of hopanes in MC252 oil and 5-year aged oil from Bay Jimmy sediment. Diagnostic ratio

Ts/Tm C29/C30 C31S/C31(S + R) C32S/C32(S + R) C33S/C33(S + R) C34S/C34(S + R) C35S/C35(S + R)

Oil source MC252 oil

5-year aged oil from Bay Jimmy

0.99 0.33 0.63 0.62 0.61 0.61 0.57

0.94 0.36 0.65 0.62 0.61 0.61 0.57

Note: Ts: Ts (18α(H)-22,29,30-Trisnorneohopane), Tm: Tm (17α(H)-22,29,30trisnorhopane), C29: C29 (17α(H),21β(H)-30-norhopane), C30: C30 (17α(H),21β(H)hopane), C31S: C31 (17α,21β-Homohopane (22S)), C31R: C31 (17α,21β-Homohopane (22R)), C32S: C32 (17α,21β-Bishomohopane (22S)), C32R: C32 (17α,21β-Bishomohopane (22R)), C33S: C33 (17α,21β-Trisnornohomohopane (22S)), C33R: C33 (17α,21βTrisnornohomohopane (22S)), C34S: C34 (17α,21β-Tetrahomohopane (22S)), C34R: C34 (17α,21β-Tetrahomohopane (22R)), C35S: C35 (17α,21β-Pentahomohopane (22S)), C35R: C35 (17α,21β-Pentahomohopane (22R)).

Concentration (mg/g-sediment) Sample 1

Sample 2

Mean

2.005 0.251 0.056 2.312

2.053 0.223 0.041 2.317

2.029 0.237 0.048 2.315

Extraction efficiency (%)

87.7 10.2 2.1 100

extraction, and 2.1 wt.% more from the third extraction (Table 2). It should be noted that although the standard extraction method is valid for extracting less aged oil hydrocarbons from sediment samples (Wang and Fingas, 1997; Yim et al., 2011), it may be less effective for more aged TPHs. Therefore, the total recovery should be envisioned as somewhat operationally defined. Earlier, Kırman et al. (2016) reported that the concentration of TPHs in the sediment from the same area (N29°26′27.3″/W89°54′07.0″) but aged for only 1.5 years reached up to 77 mg/g. This level is N 33 times higher than that in the 5 years weathered oil, indicating the sediment sorbed oil underwent significant weathering over the past five years. However, the TPHs concentration remained orders of magnitude higher than that in the pre-DwH oil spill sediment samples (0.004 to 0.090 mg/g), indicating the strong resistance of residual oil components to natural weathering. Zhao et al. (2016) observed that both n-alkanes and PAHs in GOM sediments can be photodegraded under solar light. Yin et al. (2015) studied the longterm PAHs weathering rate and found that the rate decreased significantly when the oil residual was buried. The surface (0–5 cm) sediment in this work was taken from the intertidal marsh that is subject to strong solar radiation. Therefore, the observed oil weathering is, at least in part, attributed to photochemical degradation. 3.2.2. Residual n-alkanes in sediment The concentration of total extractable n-alkanes in the sediment was determined to be 1.14 mg/g-sediment (Fig. 3). Nearly all the n-alkanes (99.2%) were extracted in the first two extractions. After 5 years of weathering, the concentration of residual n-alkanes of low molecular weight (LMW) (C9–C20) in the sediment accounted for only 2.2% of the total n-alkanes; the medium-molecular-weight (MMW) n-alkanes (C21–C30) made up 76.6%, and the high-molecular-weight (HMW) nalkanes (C31–C40) amounted to 21.2% of the total n-alkanes. It is noteworthy that C27 alone constituted 13.7% of the total n-alkanes. Zhao et al. (2016) quantified the various fractions of n-alkanes in the same Louisiana Sweet Crude oil, and determined the LMW, MMW and HMW fractions account for 67.0%, 30.9% and 2.1% of the total nalkanes, respectively, with C13 constituting 9.1% of the total n-alkanes. Comparing the changes of these fractions, it is evident that the LMW fraction was largely degraded through the 5-year weathering, the MMW fraction was about 16% higher and the HMW fraction was ~17% lower than their respective levels if there were no degradation for MMW and HMW, indicating that a large fraction of the HMW as converted to MMW n-alkanes due to natural weathering.

Fig. 3. Distributions of n-alkanes in the Bay Jimmy sediment five years after the DwH oil spill.

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The results agree with the findings reported by Vandermeulen and Singh (1994), who studied the persistence of 20-year weathered bunker C fuel oil following the Arrow oil spill, and observations by Wang et al. (1998), who studied the persistence of the 25-year-old Nipisi oil spill. The loss of the lighter fraction of n-alkanes is attributed to their relatively lower adsorption affinity to the SOM, and thus relatively higher availability to evaporation, dissolution, photochemical and biological degradation processes (Vandermeulen and Singh, 1994; Reddy et al., 2002). For instance, Zhao et al. (2016) observed that upon 14 days solar irradiation, the photodegradation transformed 66% of C11–C20, 34% of C21–C30, but only 10% of C31–C40 nalkanes in a loamy sand sediment from Grand Bay, AL. However, it should be noted that these data were obtained through bench scale experiments using sediment-seawater slurries under mixing. Consequently, the actual photodegradation rate can be much slower under the field conditions. The high content of C21–C30 can be attributed to several factors. First, the crude oil contained a relatively higher fraction (30.9%) of the MMW n-alkanes than the larger n-alkanes (2.1%). Second, the Corexit dispersants are much more preferential to disperse LMW and MMW than HMW n-alkanes (Zhao et al., 2016), and thus, more dispersed nalkanes of b C30 reached the wetland and retained by the sediment. The observed high percentile of the MMW fraction also suggested the important role of oil dispersants in transporting different fractions of oil components into coastal wetlands. Third, the MMW fraction is more adsorbable and more recalcitrant than the LMW fraction (Yin et al., 2015; Zhao et al., 2016), thus inhibiting evaporation, dissolution and natural weathering. Forth, compared to the HMW fraction nalkanes (NC30), the MMW fraction n-alkanes offer faster mass transfer rates for their smaller size, and thus can penetrate into deeper sediment and into deeper pores in the SOM, resulting in lower physical and biological availabilities. Lastly, partial weathering of HMW n-alkanes led to production and accumulation of more MMW n-alkanes during the weathering process. The percentage of HMW n-alkanes in the sediment (21.2%) was higher than that in the crude oil (2.1%). This is reasonable because the DwH crude oil consists mainly of light petroleum components (Coates et al., 1997; Rahman et al., 2003; Yin et al., 2015). In addition, HMW nalkanes are more strongly adsorbed by the sediment because of the higher hydrophobicity and lower solubility and volatility (Liu et al., 2012). 3.2.3. Residual PAHs in sediment The concentration of total parent PAHs in the sediment was measured to be 11.9 mg/g-sediment, with 92% recovered in the first extraction. Fig. 4 shows the distribution of various PAHs extracted from the sediment. After 5 years of weathering, pyrene was the most abundant

Fig. 4. Distributions of PAHs in the Bay Jimmy sediment upon sequential extractions using DCM-hexane. Nap: Naphthalene, C1-Nap: C1-Naphthalene, C2-Nap: C2Naphthalene, C3-Nap: C3-Naphthalene, Acl: Acenaphthylene, Ace: Acenaphthene, Flu: Fluorene, C1-Flu: C1-Fluorene, Phen: Phenanthrene, C1-Phen: C1Phenanthrene, C2-Phen: C2-Phenanthrene, An: Anthracene, Fl: Fluoranthene, Py: Pyrene, BaA: Benz(a)anthracene, Chry: Chrysene, BbF: Benzo(b)fluoranthene, BkF: Benzo(k)fluoranthene, BaP: Benzo(a)pyrene, BgP: Benzo(g,h,i)perylene, IP: Indeno(1,2,3-cd)pyrene, DA: Dibenzo(a,h)anthracene.

PAH in the sediment, representing 23.7% of the total parent PAHs, followed by fluoranthene, phenanthrene and fluorene, which accounted for 14.2%, 14.1% and 12.5% of the total parent PAHs, respectively. In the crude oil, the most abundant PAHs were phenanthrene (22.9%), acenaphthene (18%), anthracene (12.8%) and fluorene (11.1%) (Zhao et al., 2016). The differences in the PAHs fractions between the crude and the weathered oil can be attributed to: 1) the different solubility and adsorbability, and thus, availabilities for weathering reactions, 2) different dispersant effects, i.e., the dispersants have varied effectiveness for dispersing different PAHs (Gong et al., 2014a; Zhao et al., 2016), resulting in different amounts of PAHs reaching the wetland, and 3) degradation of alkylated PAHs in the crude led to accumulation of the corresponding parent PAHs. Tansel et al. (2011) studied the persistence of anthracene, fluoranthene, pyrene, and chrysene in GOM water column and sediments shortly after the DwH spill. They reported that the half-lives of the PAHs in the sediments were many orders of magnitude longer than in the water column, and the half-lives increased with the water depth. For instance, the half-lives of pyrene in the shallow and deep sediments were about 9 and 16 years, respectively, compared to 70.83 days in water. Among the PAHs studied, pyrene was the most persistent in the sediments, whereas chrysene was the most persistent PAH in the water column. In general, SOM is considered the key sediment component in facilitating adsorption of hydrophobic organic compounds (Zhao et al., 2002), and SOM interacts more strongly with more hydrophobic PAHs, and the more strongly bonded PAHs are less available for weathering (Gong et al., 2014b; Yin et al., 2015). Moreover, Zhao et al. (2015) compared the effects of Corexit EC9500A on the uptake of the two-ring PAHs (naphthalene and 1-methylnaphthalene) and pyrene by sediments, and revealed that the dispersant more profoundly enhanced pyrene adsorption. Therefore, both the stronger affinity toward SOM and the more enhanced dispersant effect caused the high abundance of pyrene in the residual oil. The 3-ring and 4-ring PAHs accounted for 44.8% and 42.8% of the total parent PAHs in the weathered sediment, respectively, while 2ring, 5-ring and 6-ring PAHs contributed 1.4%, 10.5 and 0.5%, respectively (Fig. 4). Though 2-ring PAHs (e.g., naphthalene) dominated in the crude MC252 oil (~64%), the much higher solubility and volatility than the heavier PAHs caused the much more profound loss of these lighter PAHs (Liu et al., 2012; Yin et al., 2015). In the water column, the weathering rate toward PAHs decreases with the increase of benzene rings (Haritash and Kaushik, 2009; Sauer et al., 1998); in the sediment phase, however, adsorption and intraparticle mass transfer may control the weathering rate. The concentration of total alkylated PAHs was 8.1 mg/g-sediment. More than 99% of alkylated PHAs was recovered from the sediment in the first two extractions. The 3-ring alkylated PAHs (C1-fluorene, C1phenanthrene and C2-phenanthrene) were the most abundant alkylated PAHs, which accounted for 99.0% of the total alkylated PAHs. Compared to the parent PAHs, alkylated PAHs are more hydrophobic and thus are more strongly adsorbed in the sediment, resulting in the slower weathering rate (Liu et al., 2012; Yin et al., 2015). The ratio of alkylated PAHs to parent PAHs in the crude oil was 3.8 (Zhao et al., 2016), which was lowered to 2.4 in the weathered oil in the sediment. This observation indicates that the alkylated PAHs were degraded even faster than the parent PAHs in the sediment. Earlier, Zhao et al. (2016) observed that both parent and alkylated PAHs are more photochemically reactive than n-alkanes owing to their ability to absorb light in the UV region. In addition, the researchers observed that more alkylated PAHs (94%) were degraded than parent PAHs (74%) over 14 days of solar irradiation. Therefore, it can be inferred that photodegradation played an important role in the weathering of alkylated PAHs, which are considered more photoactive than parent PAHs. It is noted that while the changing concentrations of various oil components offered robust oil weathering patterns, a complementary approach may be needed to determine the type and concentration of

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degradation intermediates. The information will aid in further understanding of the weathering pathway and reaction mechanism. 3.3. Effects of oil dispersants on desorption of aged oil in sediment Fig. 5a shows the desorption kinetics of the 5-year aged TPHs, nalkanes and PAHs from the Bay Jimmy sediment with or without the two model dispersants, simulating re-suspended sediment-seawater

Fig. 5. Desorption kinetics of TPHs (a), n-alkanes (b) and parent PAHs (c) from Bay Jimmy sediment five years after the DwH oil spill with or without an oil dispersant (Corexit 9500A or SPC 1000). Experimental conditions: pH = 7.6–8.1, salinity = 3.15%, Corexit EC9500A = 0 or 18 mg/L, SPC 1000 = 0 or 18 mg/L, and temperature = 25 ± 0.2 °C. Mt: mass remaining in sediment at time t, Minitial: total initial mass of TPHs, n-alkanes or parent PAHs.

7

systems. In all cases, faster desorption was observed in the first two days, and became extremely slow thereafter; and the presence of the dispersants accelerated the desorption rate and extent. Generally, the amount of oil components desorbed followed the sequence of: SPC 1000 N Corexit EC9500A N no dispersant. In the absence of a dispersant, the final concentration of TPHs in seawater amounted to 34.7 mg/L, i.e., 22.4% of the aged TPHs was desorbed when the sediment was resuspended with seawater. This observation indicates that a portion of the aged TPHs in the unsaturated sediment, which was in equilibrium with the surrounding conditions, can be easily released upon resuspension by seawater. Practically, the seawater-leachable portion of TPHs represents primarily the lighter oil components, and TPHs that were absorbed in the pore water or trapped in the sediment pores, and those partitioned in the soft domain of the SOM according to the dualmode sorption theory (Zhao and Pignatello, 2004). The desorption can be attributed to: 1) the mechanical dispersion caused during the mixing (Delvigne and Sweeney, 1988; Schein et al., 2009), and 2) breakdown of the pre-existing equilibrium due to removal of TPHs from the pore water and the soft-SOM domain. In addition, TPHs that are adsolubilized due to the uptake of the surfactants in the oil dispersants can more desorbable when the external TPHs-laden seawater was replaced with oil-free seawater. The presence of 18 mg/L SPC 1000 or Corexit EC9500A increased both the desorption rate and extent of TPHs from the re-suspended sediment. The final amount of TPHs desorbed was increased by 38.2% by SPC 1000 and by 31.9% by Corexit EC9500A compared to the case with plain seawater. Similar dispersant enhanced desorption was also observed for n-alkanes and PAHs (Fig. 5b and 5c), where the final mass of n-alkanes desorbed was increased by 29.7% and 26.9% by SPC 1000 and Corexit EC9500A, respectively, and that of PAHs by 18.2% and 12.0%. The results reveal that both dispersants can promote the desorption of the aged oil components from the sediment, and SPC1000 appeared more effective than Corexit EC9500A. Moreover, the dispersants were more effective for desorption of PAHs than nalkanes. The application of dispersants may pose two contrasting effects on the oil-sediment interactions due to the amphiphilic properties of the surfactants (Gong et al., 2014a; Gong et al., 2014b). On the one hand, the dispersant components break down the large oil slicks into microscale droplets by lowering the interfacial tension, thereby decreasing the interfacial accumulation (adsorption) of TPHs and making TPHs more soluble with the water column (Lessard and DeMarco, 2000; Reed et al., 2004; Gong et al., 2014a; Gong et al., 2014b). On the other hand, oil dispersants are subject to sorption by the sediment, the sorbed dispersant components can facilitate adsolubilization of oil components onto the sediment (Gong et al., 2014a; Gong et al., 2014b; Zhao et al., 2015). For instance, Gong et al. (2014a) observed that the presence of 18 mg/L of Corexit EC9500A increased the equilibrium uptake of phenanthrene by 7% by a loamy sand sediment. Zhao et al. (2015) confirmed that the dispersant increased the equilibrium uptake of naphthalene, 1methylnaphthalene and pyrene by up to 85.3%. The overall effect of a dispersant depends on the competition of the solubilization and adsolubilization effects, the dispersant concentration, adsorbability of the dispersant and environmental conditions (Zhao et al., 2015). The observed dispersant-enhanced desorption/removal of oil TPHs from the sediment can be attributed to the following three key factors. First, the solubilization effect of the dispersants appeared to have overweighed the adsolubilization effect for the sediment-sorbed TPHs. Second, the dispersants facilitated dissolution of the oil droplets that were trapped in the sediment. And third, the dispersants caused partial dissolution of the SOM, the major sink for the oil TPHs, resulting in reduced sorption capacity for the TPHs. Zhang et al. (2011) reported that a surfactant solution leached out N 10 times more SOM from a potting soil than DI water. Gong et al. (2014a) and Zhao et al. (2015) also observed that the presence of Corexit 9500A leached out more SOM than plain seawater from a GOM sandy loam sediment. In fact, upon

Please cite this article as: Duan, J., et al., Study of residual oil in Bay Jimmy sediment 5 years after the Deepwater Horizon oil spill: Persistence of sediment retained oil hy..., Sci Total Environ (2017), http://dx.doi.org/10.1016/j.scitotenv.2017.09.234

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centrifugation of the seawater-sediment mixtures, the supernatant without dispersant appeared clear with some visible oil droplets on the water surface (Fig. S2 in SM) and with a TOC level of 24.8 mg/L in the water column; in contrast, the supernatants with SPC 1000 or Corexit EC9500A appeared light yellowish without visible oil droplets (Fig. S2 in SM) and with a TOC level of 45.1 mg/L (up to 6.6 mg/L from the dispersant) and 35.7 mg/L (up to 4.9 mg/L from the dispersant), respectively. As a result, more residual oil components were desorbed in the presence of the dispersants. The fact that SPC 1000 leached more SOM is also consistent with the observation that more TPHs were desorbed by SPC 1000. It should be noted though that when the SOM is not leachable by the dispersants or when leachable SOM is pre-removed from the sediment, the presence of the dispersants may inhibit the desorption of TPHs due to the surfactants-facilitated adsolubilization of oil components (Gong et al., 2014a; Zhao et al., 2015). Fig. 6 presents the relative concentrations of various fractions of nalkanes and PAHs desorbed into seawater from the sediment after 4 days of the desorption tests. It is evident that most desorbed fractions of the oil compounds were MMW and HMW n-alkanes and PAHs of 3–4 rings, as these were also the main components of the residual oil. Furthermore, the oil dispersants were more effective in enhancing desorption of the larger and more hydrophobic n-alkanes and PAHs, which, however, is also related to the dispersant facilitated dissolution of the SOM. SPC 1000 was more effective in desorbing HMW n-alkanes and 5–6 ring PAHs. The detailed mechanisms are yet to be investigated.

Fig. 6. Distributions of 5-year aged n-alkanes (a) and PAHs (b) that were desorbed from the Bay Jimmy sediment with or without dispersant. Experimental conditions: pH = 7.6–8.1, salinity = 3.15%, Corexit EC9500A = 0 or 18 mg/L, SPC 1000 = 0 or 18 mg/L, desorption time = 4 days and temperature = 25 ± 0.2 °C. Cfraction: Concentration of a given fraction of n-alkanes or PAHs desorbed into seawater, Ctotal: total n-alkanes or PAHs desorbed in seawater.

3.4. Successive desorption kinetics To probe the maximum possible desorption of the residual oil components, successive desorption experiments were carried out. Fig. 7 shows the successive desorption kinetics of the 5-year aged TPHs, nalkanes and PAHs from the sediment with or without 18 mg/L of SPC 1000 over 3 consecutive desorption runs. In all cases, the most desorption occurred in the first desorption run, and the desorption rate became extremely slow after the third run. In the absence of the dispersant, the amounts of TPHs, n-alkanes and parent PAHs desorbed from the sediment reached 28.4%, 25.2% and 8.1%, respectively, after the third run. In the presence of 18 mg/L of SPC 1000, the final

Fig. 7. Successive desorption kinetics of TPHs (a), n-alkanes (b) and parent PAHs (c) from the Bay Jimmy sediment five years after the DwH oil spill with or without dispersant. Experimental conditions: pH = 7.6–8.1, salinity = 3.15%, SPC 1000 = 0 or 18 mg/L, and temperature = 25 ± 0.2 °C. Mt: mass of TPHs, n-alkanes or parent PAHs remaining in sediment at time t, Minitial: total initial mass of TPHs, n-alkanes or parent PAHs.

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Table 3 Extraction of the sediment after the third run of successive desorption with or without 18 mg/L SPC1000. Dispersant

18 mg/L SPC1000 Seawater

TPHs (%)

n-alkanes (%)

PAHs (%)

Mt/M0

ME/M0

R

Mt/M0

ME/M0

R

Mt/M0

ME/M0

R

81.4 91.9

81.1 89.9

99.6 97.8

68.1 74.8

61.5 73.1

90.3 97.7

57.3 71.6

61.5 68.2

107.3 95.3

Note: Mt: mass adsorbed at time t, M0: the initial component mass in the sediment, ME: component mass extracted from the sediment after three successive desorption runs; Recovery E ðRÞ ¼ M Mt  100%.

percentages of TPHs, n-alkanes and parent PAHs desorbed were increased to 42.7%, 31.9% and 18.6%, respectively. The results indicate that 71.6% of TPHs, 74.8% of n-alkanes, and 91.9% of PAHs in the sediment would remain highly stable (or hardly desorbed by seawater) even when the sediment was re-suspended by seawater without dispersant. However, when the sediment was re-suspended and subjected to the dispersant, the residual TPHs, n-alkanes and parent PAHs were lowered to 57.3%, 68.1%, and 81.4%, respectively. The dispersant effect was most evident in the first run for all cases, which is attributed to the initial dispersant-facilitated leaching of SOM and dispersion of the residual oil droplets in addition to the dispersion effects of the dispersant as discussed above. In the second and third desorption runs, however, the presence of the dispersant actually showed an inhibitive effect on the desorption (Fig. 7). This is reasonable because, after the first run, nearly all the leachable SOM was removed from the sediment. In the subsequent runs, more dispersant would be adsorbed to the sediment when the fresh dispersant solution was added, leading to dispersant-facilitated adsolubilization of the oil components or inhibited desorption (Gong et al., 2014b; Zhao et al., 2015). In addition, the diffusion rate or the difference in the interfacial chemical potential can also contribute to the rapid initial desorption of TPHs, which is more clearly evidenced for the desorption runs with seawater only. Solvent extraction of the final sediment sample after the third run recovered N90% of the TPHs (Table 3), confirming the mass balance was maintained during the experiments. 4. Conclusions This study investigated the persistence of the DwH oil in a model GOM sediment that was subjected to five years of natural weathering, and studied the effects of oil dispersants on the desorption potential of the residual oil. The key findings are recapped as follows: 1) Based on analyses of the key biomarkers, CPI and Pr/Ph ratio, the residual oil extracted from the Bay Jimmy sediment was confirmed to be from the 2010 DwH oil spill. 2) After 5 years of natural weathering, the concentration of TPHs in the sediment was determined to be 2.310 mg/g-sediment, with the total n-alkanes accounting for ~ 50% (1.14 mg/g-sediment). MMW nalkanes (C21–C30) made up 76.6% of the total n-alkanes, with C27 alone accounting for 13.7%. The total parent and alkylated PAHs concentrations were determined to be 0.012 and 0.008 mg/g-sediment, respectively. The TPHs concentration in the sediment remained orders of magnitude higher than the pre-DwH oil spill level. 3) More than 97% of LMW n-alkanes (C9–C20) were degraded, leaving the MMW n-alkanes the most predominant n-alkanes (76.6%) in the weathered oil; and N98% of 2-ring PAHs were degraded, leaving 3and 4-ring PAHs the most abundant PAHs (44.8% and 42.8%) in the weathered sediment. Alkylated PAHs were degraded faster than the parent PAHs. 4) Oil dispersants can promote solubilization of sediment retained oil when the sediment is re-suspended. SPC 1000 was found more effective in facilitating desorption of the aged oil components than Corexit EC9500A, and the dispersants were more effective toward PAHs than n-alkanes. The enhanced desorption is attributed to dispersant-facilitated leaching of SOM from the sediment,

dissolution of sediment-sorbed TPHs, and dispersion of sedimentretained oil droplets. 5) Successive desorption experiments indicated that 71.6% of TPHs, 74.8% of n-alkanes, and 91.9% are physically unavailable or extremely slow to be desorbed without dispersant, and the presence of 18 mg/L of SPC 1000 lowered the residual fractions to 57.3%, 68.1%, and 81.4%, respectively. The results are useful for understanding the persistence of oil residuals in the GOM sediments and estimate the long-term weathering rate of sediment adsorbed/retained oil. The information may also help with assessing the long-term environmental impacts of the oil spill based on the desorption potential, and with understanding the effects of oil dispersants on the physical and biological availabilities of aged oil in GOM sediments. While each oil spill can be unique and as more field weathering data are yet to be collected, the information on the general weathering patterns and impacts of dispersants may serve as a useful guide for studying sediment-retained oil on the global scale. Acknowledgements The research was partially supported by the U.S. Department of the Interior Bureau of the Ocean Energy Management (M12AC00013). We are grateful to Zhengqing Cai and Haodong Ji for their assistance with some of the analytical work and valuable discussion. We also thank Dr. Ronald Delaune of Louisiana State University for discussing the sediment site and samples. Appendix A. Supplementary data Supplementary data to this article can be found online at https://doi. org/10.1016/j.scitotenv.2017.09.234. References Coates, J.D., Woodward, J., Allen, J., Philp, P., Lovley, D.R., 1997. Anaerobic degradation of polycyclic aromatic hydrocarbons and alkanes in petroleum-contaminated marine harbor sediments. Appl. Environ. Microbiol. 63, 3589–3593. Conan, G., Dunnet, G.M., Crisp, D.J., 1982. The long-term effects of the Amoco Cadiz oil spill. Philos. T. Roy. Soc. B. 323–333. Delvigne, G.A.L., Sweeney, C., 1988. Natural dispersion of oil. Oil Chem. Pollut. 4, 28–310. Ehrhardt, M., Petrick, G., 1993. On the composition of dissolved and particle-associated fossil fuel residues in Mediterranean surface water. Mar. Chem. 42, 57–70. Gong, Y., Zhao, X., Cai, Z., O'reilly, S.E., Hao, X., Zhao, D., 2014a. A review of oil, dispersed oil and sediment interactions in the aquatic environment: influence on the fate, transport and remediation of oil spills. Mar. Pollut. Bull. 79, 16–33. Gong, Y., Zhao, X., O'Reilly, S.E., Qian, T., Zhao, D., 2014b. Effects of oil dispersant and oil on sorption and desorption of phenanthrene with Gulf Coast marine sediments. Environ. Pollut. 185, 240–249. Haritash, A.K., Kaushik, C.P., 2009. Biodegradation aspects of polycyclic aromatic hydrocarbons (PAHs): a review. J. Hazard. Mater. 169, 1–15. Hatch, R.S., Yeager, K.M., Brunner, C.A., Wade, T.L., Briggs, K.B., Schindler, K.J., 2013. Salt Marsh Sediment Mixing Following Petroleum Hydrocarbon Exposure from the Deepwater Horizon Oil Spill, in AGU Fall Meeting Abstracts. Hemmer, M.J., Barron, M.G., Greene, R.M., 2011. Comparative toxicity of eight oil dispersants, Louisiana sweet crude oil (LSC), and chemically dispersed LSC to two aquatic test species. Environ. Toxicol. Chem. 30, 2244–2252. Hickman, S.H., Hsieh, P.A., Mooney, W.D., Enomoto, C.B., Nelson, P.H., Mayer, L.A., Weber, T.C., Moran, K., Flemings, P.B., McNutt, M.K., 2012. Scientific basis for safely shutting in the Macondo Well after the April 20, 2010 Deepwater Horizon blowout. P. Natl. Acad. Sci. USA. 109, 20268–20273. Iqbal, J., Gisclair, D., McMillin, D.J., Portier, R.J., 2007. Aspects of petrochemical pollution in southeastern Louisiana (USA): pre-Katrina background and source characterization. Environ. Toxicol. Chem. 26, 2001–2009.

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Please cite this article as: Duan, J., et al., Study of residual oil in Bay Jimmy sediment 5 years after the Deepwater Horizon oil spill: Persistence of sediment retained oil hy..., Sci Total Environ (2017), http://dx.doi.org/10.1016/j.scitotenv.2017.09.234