Contamination and Human Health Risk Assessment

Arch Environ Contam Toxicol DOI 10.1007/s00244-017-0394-9

SPECIAL ISSUE: OCEAN SPILLS AND ACCIDENTS

Contamination and Human Health Risk Assessment of Polycyclic Aromatic Hydrocarbons (PAHs) in Oysters After the Wu Yi San Oil Spill in Korea Andrew Loh1,2 • Un Hyuk Yim1,2 Moonkoo Kim1,2

•

Sung Yong Ha1 • Joon Geon An1

•

Received: 30 November 2016 / Accepted: 14 March 2017 Ó Springer Science+Business Media New York 2017

Abstract After the collision of the Singapore-registered oil tanker M/V Wu Yi San into the oil terminal of Yeosu, Korea on January 31, 2014, approximately 900 m3 of oil and oil mixture were released from the ruptured pipelines. The oil affected more than 10 km of coastline along Gwangyang Bay. Emergency oil spill responses recovered bulk oil at sea and cleaned up the stranded oil on shore. As part of an emergency environmental impact assessment, region-wide monitoring of oil contamination in oyster had been conducted for 2 months. Highly elevated concentrations of polycyclic aromatic hydrocarbons (PAHs) were detected at most of the spill affected sites. Four days after the spill, the levels of PAHs in oysters increased dramatically to 627–81,000 ng/g, the average of which was 20 times higher than those found before the spill (321–4040 ng/g). The level of PAHs in these oysters increased until 10 days after the spill and then decreased. Due to the strong tidal current and easterly winter winds, the eastern part of the Bay—the Namhae region—was heavily contaminated compared with other regions. The accumulation and depuration of spilled oil in oyster corresponded with the duration and intensity of the cleanup activities, which is the first field observation in oil spill cases. Human health risk assessments showed that benzo[a]pyrene equivalent concentrations exceeded levels of

& Un Hyuk Yim [email protected] 1

Oil and POPs Research Group, South Sea Research Institute, KIOST, Geoje 53201, Republic of Korea

2

Marine Environmental Science Major, Korea University of Science and Technology, Daejeon 34113, Republic of Korea

concern in the highly contaminated sites, even 60 days after the spill.

On January 31, 2014, the Singapore-registered oil tanker Wu Yi San collided into the oil terminal of a petrochemical complex in Yeosu, Korea. The collision with the vessel impacted a bridge connecting a pier to the petrochemical facility, rupturing three floating oil pipelines. From the ruptured pipelines, approximately 900 m3 of Basrah Light crude oil, naphtha, and an oil–water mixture were spilled into the sea. The spilled oil mixture spread rapidly over 10 km of coastline along Gwangyang Bay. Directly after the spill, emergency oil spill responses removed the bulk of the oil at sea over 4 days, and stranded oil on the shoreline was cleaned extensively for a month after the spill. Cleanup using high-pressure sprays, flushing, absorption pads, snares, skimmers, and booms were applied to the affected shoreline areas. Stranded oil can release toxic oil components in the intertidal zone and can affect the surrounding environment, including oyster and mussel beds (Carls et al. 2001). Toxicity of petroleum related products often is caused by the presence of organic compounds, such as saturate, aromatics, and carboxylic acids (Neff 2004). Among them, polycyclic aromatic hydrocarbons (PAHs) have highly varying toxic potencies, and groups of parent and alkylated PAHs in multi-media samples are routinely monitored after oil spills. Main sources of PAHs can be divided as petrogenic or pyrogenic. Petrogenic PAHs originate from petroleum and their refined products, and pyrogenic PAHs are produced from oxygen-depleted, high-temperature combustion of fossil fuels and biomass. Petrogenic or pyrogenic sources of PAHs can be distinguished through interpretation of PAHs profiles. Generally, petrogenic PAH

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exhibits a bell-shaped pattern of alkylated PAHs and depletion of HMW PAHs, and pyrogenic PAH exhibits a skewed pattern with predominance of unsubstituted compounds over their alkylated homologues (Douglas et al. 2007; Lake et al. 1979; Youngblood and Blumer 1975). Bivalves are known for their ability to accumulate significant amounts of contaminants through direct or indirect exposure from the surrounding environment (Farrington 1991). Their low metabolic capacity and high body burden allow them to accumulate wide ranges and high levels of contaminants (Cossa 1989). Since the first proposal of Goldberg’s ‘‘Mussel Watch’’ program in 1975, regional or international monitoring programs using mussel or oysters have been continuously conducted (Farrington et al. 2016). In the same context, bivalves have been extensively used as biomonitors for petroleum hydrocarbons and to measure the rate of recovery after oil spill accidents (Elmgren et al. 1983; MLTM 2013; Short and Heintz 1997; Tunnell et al. 1981). Right after the Wu Yi San oil spill, multi-media environmental monitoring, including oil, seawater, sediment, pore water, and bivalves, were conducted to evaluate acute environmental damages by the spilled oil. This study mainly focused on the oyster monitoring to identify (1) the spatial distribution and temporal changes of petroleum derived PAHs, and (2) the potential human health risks from consuming the contaminated oysters.

Experimental Methods Study Area and Sample Collection Gwangyang Bay is a semi-enclosed embayment on the southern coast of the Korean Peninsula (Fig. 1). The western part of the bay is relatively hydro-dynamically static with poor water exchange and sluggish circulation; the water depth is less than 5 m and the tidal exchange volume averages 4%. The Eastern part of the bay is the main channel for tidal exchange; water depth is more than 20 m and tidal exchange volume is on the average of 14%. Tidal exchange volume ranges from 20% at spring tide to 8% at neap tide, with an average of 14% (POSCO 1983). Gwangyang Bay is a highly industrialized area with steel and petrochemical complexes as the dominant industry; the area is thus prone to intentional or unintentional contamination of petroleum hydrocarbons into the surrounding environment. This area was designated as one of the Special Management Area (SMA), and as part of the SMA program, some of the priority pollutants have been routinely monitored in bivalves and sediments (Hong et al. 2011; KORDI 2003).

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One of the main types of spilled oil (Basrah Light crude oil) was collected from the tanks leading to the ruptured pipelines. Stranded oil was collected from the shorelines a day after the spill. Oyster samples (Crassostrea gigas) were collected from a total of 21 stations along the spill regions in Gwangyang Bay, which includes 8 stations in Yeosu, 12 stations in Namhae, and 1 station in Hadong (Fig. 1). Sampling was performed at intervals of 4, 10, 30, and 60 days after the spill for environmental impact assessments. Pre-spill monitoring had been conducted at seven stations (St.1, St.2, St.4, St.7, St.8, St.18, and St.21) in April 2012. More than 20 oysters were randomly collected from each station and were shucked from their shells then stored at -20 °C before analysis. Analytical Methods Our analysis of the petroleum hydrocarbons in the oil was based on the method described by Wang et al. (1994), with slight modifications (Yim et al. 2011). For the oil samples, approximately 20 mg of oil was dissolved in dichloromethane (GC2 grade, Burdick and Jackson) and fractionated using activated silica gel columns (100–200 mesh, Sigma-Aldrich). Saturate fractions (F1) were collected using hexane (GC2 grade, Burdick and Jackson) and aromatic fractions (F2) were eluted with a mixture of dichloromethane and hexane (1:1 volume). Equal volumes of fractions F1 and F2 were mixed to form a fraction denoted as F3, which were then used for the analysis of the total amount of petroleum hydrocarbons. Surrogate and GC-internal standards were added into each fraction for quantitation and quality controls. For the oyster samples, extraction and clean up procedures were based on the method described by Jin et al. (2008). Approximately 20 g of homogenized oyster tissue sample material was mixed with anhydrous sodium sulfate and then extracted for 16 h with a Soxhlet apparatus through 200-mL aliquots of GC-grade dichloromethane. Surrogate standards were added before the extraction. The extracts were then sequentially cleaned with silica gelalumina columns followed by high-pressure liquid chromatography equipped with a gel permeation column to remove interferences such as lipids. Cleanup columns were prepared by slurry packing of 10 g of alumina (deactivated with 1% water), followed by 20 g silica gel (deactivated with 5% water), and topped with 1 cm of anhydrous sodium sulfate. The eluates were concentrated to 0.5 mL and then spiked with internal standards and finally adjusted to an accurate preinjection volume of 1 mL for gas chromatography–mass spectrometry analysis. The saturated compounds and the total petroleum hydrocarbon (TPH) within the oil samples were analyzed using an Agilent 7890 GC equipped with a flame-

Arch Environ Contam Toxicol Fig. 1 Location of sampling site in Gwangyang Bay. The oil spill site is represented by red star. Post-spill monitoring sites are represented by black circle. Pre-spill ? post-spill monitoring sites are represented by blue square

Hadong

Gwangyang

St.21 St.20

St.1

St.19 St.18 St.17 St.16

Spill Site St.3

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St.13

Yeosu

St.12 St.11

St.5 St.6

St.10 St.9 St.7

South Korea

St.8

ionization detector (FID) and a DB-5 capillary column (30 m 9 0.32 mm 9 0.25 lm; Agilent J&W, Santa Clara, CA). The oven temperature was programmed to start at an initial temperature of 50 °C (2 min) and then ramped to 300 °C at 10 °C/min and held for 16 min. The injector and detector temperatures were set at 300 °C. GC-detectable TPH, resolved peaks, and the unresolved complex mixture (UCM) were analyzed. The GC-resolved peaks ranged from n-C8 to n-C40 including the selected isoprenoids, pristine, and phytane. The PAHs were analyzed using an HP 5890 GC, equipped with an HP 5972 mass selective detector (MS); a DB-5MS (30 m 9 0.25 mm 9 0.25 lm; Agilent J&W) was used for analysis of the PAHs in both the oil and bivalve samples. The mass spectrometer was operated in a selected ion-monitoring (SIM) mode based on the molecular ions of the target PAHs. The oven temperature was

programmed to start at an initial temperature of 60 °C (1.5 min), then ramped up to 300 °C at 4 °C/min, and held for 10 min. The injector temperature was set at 300 °C and the transfer line was set at 280 °C. We analyzed 16 United States Environmental Protection Agency (US EPA) priority PAHs (16 PAHs) and alkylated PAHs. The alkylated PAHs included C1- to C4- naphthalene, C1- to C3-fluorene, C1to C4-phenanthrene, C1- to C3-dibenzothiophene, and C1to C3-chrysene. Total PAHs refer to the sum of the 16 PAHs, the alkylated PAHs, and an additional four other PAHs (dibenzothiophene, retene, benzo[e]pyrene, and perylene). Abbreviations of the target PAHs are as follows: naphthalene (Nap), acenaphthylene (Acl), acenaphthene (Act), fluorene (Flu), phenanthrene (Phe), anthracene (Ant), dibenzothiophene (DBT), fluoranthene (Flr), pyrene (Pyr), benzo[a]anthracene (BaA), chrysene (Chr), benzo[b]fluoranthene (BbF), benzo[k]fluoranthene (BkF),

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1.84 ± 0.04 1.46 ± 0.03 2.42 ± 0.03 1.88 ± 0.045 7.0 ± 2.0 17.0 ± 4.7 Values were based on the average of six analyzed stranded oil samples

107.5 ± 22.1 140.1 ± 31.5 19.4 ± 4.2 33.5 ± 3.3 22.2 ± 1.4 Strandeda

24.9 ± 5.8

21.8 Basrah Light

a

1.46 2.51 1.90 13.3 43 141 356 10.2

Asphatene (%) Resin (%) Aromatic (%) Saturate (%) Oil type

Table 1 Chemical properties of source and stranded oil

Of the three types of oil spilled, the highest amount released was Basrah Light crude oil (BLC), followed by naphtha and then the oil mixture, with a ratio of 4:2:1, respectively. BLC is categorized as a Group III oil according to the persistence classification guidelines, which have specific gravities between 0.85 and 0.95 and have tendency to lose up to 40% of volume through evaporation and, thus, have high potential for long-term environmental impact (ITOPF 2014). The percentage of aromatics (58.5%) in BLC was much higher than that of the saturate (21.8%), while percentages of resins and asphaltenes were relatively low (9.5 and 10.2%, respectively; Table 1). Chromatograms showed that n-alkanes ranged from C8 to C40 with C8 being most abundant. The PAHs and their alkyl homologues ranged from naphthalene to benzo(ghi)perylene, where alkyl homologues of naphthalene (C1- to C4-) and dibenzothiophene (C1- to C3-) were the most dominant (Fig. 2a). Six stranded oil samples collected one day after the spill showed compositions indicative of a slightly weathered state. Compared with spill source oil, the percentage of aromatics (24.9%) decreased, and those of resins (33.5%) and asphaltenes (19.4%) increased correspondingly, reflecting the initial weathering effects (Table 1). GC-FID chromatograms showed n-alkanes ranging from C11 to C40

TPH lg/mg

Chemical Properties of Source and Stranded Oil

9.5

UCM lg/mg

Results and Discussion

58.5

R Alkane lg/ mg

R PAHs lg/ mg

2-/3-mD/1mD

4-mD/1mD

C31S/R hopane

C29/C30 hopane

benzo[e]pyrene (BeP), benzo[a]pyrene (BaP), perylene (Per), indeno(1,2,3-cd)pyrene (IP), dibenz(a,h)anthracene (DA), benzo(ghi)perylene (BP), and retene (Ret). Quality control samples were processed in a manner identical to the actual samples. A minimum of one procedural blank was used for each batch of 14 samples. Blank levels were no more than three times higher than the detection limit. The method detection limit for target PAHs ranged from 0.46 to 0.92 ng/g, except for naphthalene, which had a detection limit of 3.36 ng/g. Surrogate standards were used to determine the recovery of each sample and quantify the analytes. The surrogate standard recoveries satisfied acceptable ranges (40–120%). The certified reference materials (National Institute of Standards and Technology 1974b) were analyzed to monitor the performance of the analytical methods. The recoveries of the PAHs in the certified materials were within an acceptable range (certified values ±20%). The results reported in this study were based on dry weight. The wet:dry ratio was determined by calculating the weight loss after drying of approximately 2 g of homogenized sample at 105 ± 5 °C for more than 30 min.

1.84



(a) 35

Composition (%)

30 25 20 15 10 5

C0Nap C1Nap C2Nap C3Nap C4Nap Acl Act C0Flu C1Flu C2Flu C3Flu C0DBT C1DBT C2DBT C3DBT C0Phe Ant C1Phe C2Phe C3Phe C4Phe Ret Flr Pyr BaA C0Chr C1Chr C2Chr C3Chr BbF BkF BeP BaP Per IP DA BP

0

Compounds

(b) 35

Composition (%)

30 25 20 15 10 5


0

Compounds Fig. 2 PAHs composition profiles of spilled oil (a) and stranded oil (b). Refer to ‘‘Analytical Methods’’ section for abbreviation of PAHs

with C15 being most abundant. In contrast to the loss of low molecular weight alkyl homologues (C0- and C1-) of naphthalene, an increase of alkyl homologues (C2- and C3-) of dibenzothiophene was observed in the stranded oil (Fig. 2b). Representative source fingerprinting indices of the PAHs and biomarkers were well matched between the source oil and the stranded oil (Table 1). Temporal Changes and Spatial Distribution of the PAHs Concentration of the total PAHs in the oysters collected before the spill ranged from 210 to 4040 ng/g with an average of 1200 ng/g (Fig. 3a). Stations in close proximity to the petrochemical and steel complexes showed relatively higher concentrations compared with other stations. PAH levels drastically increased (by nearly 20 times at highest contaminated site) 4 days post-spill (627–81,000 ng/g with an average of 33,800 ng/g), and increased even higher over

the first 10 days post-spill (2000–110,000 ng/g with an average of 35,000 ng/g; Fig. 3b, c). PAHs levels then decreased over the following 30 days (923–63,000 ng/g with an average of 15,200 ng/g) and then to even lower levels over the following 60 days after the spill (515–41,000 ng/g with an average of 5900 ng/g; Fig. 3d, e). These temporal trends were similar with those previously found after the Hebei Spirit oil spill, where total PAHs in oyster right after the spill were approximately 40–500 times higher than pre-spill and decreased exponentially over time (Yim et al. 2012). The movement and stranding of spilled oil contaminated the entire Gwangyang Bay coastline to varying degrees. Notably, massive amounts of oil moved into the Sindeok beach (St. 4) near the spill site, and caused severe contamination in intertidal areas. Station 4 showed the highest PAH concentration in oysters (110,000 ng/g), and there were prolonged effects from the residual oil for 2 months after the spill. Due to strong tidal movements and easterly

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Total PAHs (ng/g dw.)


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(d) 120000 100000 80000 60000 40000 20000 0


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Fig. 3 Spatiotemporal changes of PAHs in bivalves after the Wu Yi San oil spill. Stations marked with ‘‘NS’’ represents stations that were not sampled. Each station number corresponds to the station shown in Fig. 1. a Pre-spill, b 4 days, c 10 days, d 30 days, e 60 days

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Temporal Changes in PAH Composition The characteristics of a bell-shaped pattern within the alkyl homologue series of PAHs (3- to 4-ring PAHs) and some pyrogenic 4- to 6-ring PAHs suggests a mixed source of pyrogenic and petrogenic in the pre-spill oysters (Fig. 5a;

(a)

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Target PAHs (ng/g dw.)

winter winds, the spilled oil spread rapidly, mainly into the eastern part of the bay. Compared with the western part of the bay, the eastern coastline was more highly contaminated. Twelve sites in Namhae showed highly elevated PAH concentrations; seven of these sites exceeded 40,000 ng/g. Other regions in Gwangyang Bay also exhibited significant oiling. Seven sites in Yeosu and one site in Hadong showed elevated PAHs in bivalves and similar temporal changes at the heavily contaminated sites. Station 8 which is located at the southern most of Gwangyang Bay showed no significant changes in PAH concentrations and is regarded as a reference site. The temporal trends of PAHs in bivalves after the spill corresponded with the timeline of the emergency oil spill responses. Initial cleanup activities to recover bulk oil at sea ended 4 days after the spill. Because most of the oil slick was stranded at shorelines, extensive shoreline cleanup had been performed for 10 days after the spill. The oiling status of the spilled oil on the affected areas had been monitored daily by a shoreline cleanup assessment team (SCAT), and the amount of visible oil on the shoreline was reported to be significantly decreased 10 days post spill. Secondary cleanup to remove residual oil continued for 30 days after the spill (KCG 2014). During extensive oil spill cleanups such as high- or lowpressure water flushing, surf washing, and tilling, stranded oil is resuspended, thus facilitating dispersion and dissolution (Lee et al. 1997; Lunel et al. 1996). Filter-feeding bivalves such as oysters and mussels are susceptible to water-column disturbances caused by cleanup events (Baumard et al. 1999). In this case, it could be inferred that the accumulation and depuration of petrogenic PAHs in oysters were largely governed by the duration and extent of the shoreline cleanup activities. Thus, the initial increasing trend of PAHs in oysters for 10 days post spill could be interpreted to be the accumulation phase, and decreasing trends for 2 months after the spill might represent the depuration phase. This was well demonstrated in temporal changes of PAHs at St. 4 and St. 18 (Fig. 4). In the accumulation phase, PAH concentrations drastically increased to as high as 45,000 and 80,000 ng/g, respectively, followed by a further increase to 65,000 and 110,000 ng/g, respectively in the 10 days post-spill. After that, PAHs were depurated exponentially to 7000 and 40,000 ng/g, respectively over 2 months, and they are expected to further decrease to pre-spill levels over time.

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Fig. 4 Temporal changes of total PAHs accumulated in oysters compared with pre-spill data. a St. 4, b St. 18

Douglas et al. 2007; Stout 2007). Compared with pre-spill data, elevated proportions of naphthalene and dibenzothiophene were found in oysters collected four days after the spill (Fig. 5b). The bell-shaped dibenzothiophene and phenanthrene alkyl homologue series and the elevated proportions of naphthalene suggest a petrogenic source of contamination (Stout et al. 2002; Wang and Fingas 2003). The proportion of naphthalene subsequently decreased by an order of magnitude and that of dibenzothiophene increased significantly 10 days after the spill (Fig. 5c). Concentrations of the less persistent C1- and C2- alkyl homologues of dibenzothiophene and phenanthrene were substantially lower than the C3- and C4- homologues after 30 days (Fig. 5d). These temporal changes were more apparent in 60 days after the spill (Fig. 5e). Diagnostic ratios often are used as indicators for source identifications after oil spills. For example, alkylated dibenzothiophene over alkylated phenanthrene are degradation-resistant, which make them useful source-specific

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Composition (%)


Composition (%)


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Fig. 5 Temporal changes of PAHs compositions in oysters within the spill region. Refer to ‘‘Analytical Methods’’ section for abbreviation of PAHs. a Prespill, b 4 days, c 10 days, d 30 days, e 60 days

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1978). Just like other marine organisms, oysters undergo both phase I and II enzymatic biotransformation of xenobiotics (Livingstone 1985; Moore 1980; Buhler and Williams 1989; Foureman 1989). During phase I enzymatic biotransoformation, xenobiotics that are insufficiently polar cannot be excreted and therefore need to be conjugated before excretion. The presence of sulfur in dibenzothiophene allows it to be excreted more easily from the organism without the need for conjugation processes in phase II of enzymatic biotransformation (Campo et al. 2010).

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C3DBT/C3Phe ratio

7 6 5 4 3

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Fig. 6 Ratios showing source recognition indices

markers in abiotic samples (Douglas et al. 1996; Wang et al. 1999; Yim et al. 2011). These double ratios were plotted to see their applicability in oyster samples (Fig. 6). Double ratios (C2D/C2P vs C3D/C3P) gradually decrease with time. Interestingly, the temporal changes of these ratios lie on a line connecting two points, which start from stranded oil to pre-spill. This suggests the vulnerability of dibenzothiophene to depuration and/or metabolization compared with phenanthrenes in the oyster, which will be further discussed. Selective Depuration of Dibenzothiophene Although the fate of dibenzothiophene and phenanthrene is generally known to be similar in the abiotic environment, once they are accumulated in biota, their metabolic pathways are unclear. Accumulation rate of these two compounds were reported similar in previous studies (Eastmond et al. 1984; Herbes and Risi 1978), but the rates of depuration of these two compounds were different (Eastmond et al. 1984). Similar to reports by Eastmond et al. (1984), this study observed that dibenzothiophene was depurated more rapidly than phenanthrene from the organism. The more rapid depuration of dibenzothiophene than phenanthrene in this study might be due to (1) slight difference in polarity between dibenzothiophene and phenanthrene, and (2) difference in metabolic pathways in the organism. Parent and their respective alkyl homologues of dibenzothiophene and phenanthrene are known to have slight polarity differences (Meador et al. 1995), but this slight difference can significantly influence the enzymatic biotransformation processes in an organism (Moore 1980). Although oysters are known to have little metabolic capacities, these can significantly influence the rates of elimination of xenobiotics from the organism (Anderson

Depuration Rate and Half-Life of PAHs in Oysters The accumulation of organic contaminants is governed by type and duration of exposure, as well as by the metabolic activity of the organism (Djomo et al. 1996; McKim and Goeden 1982). The levels of contamination in an organism are driven by its ability to bioaccumulate contaminants through all routes of exposure from the ambient environment (Arnot and Gobas, 2006). When the organism is not consistently exposed to a source, the accumulated contaminants can be released, thus resulting in decreased levels of contamination. This process, also known as depuration, occurs when organisms resume normal activities, such as filtration, respiration, and reproduction (Cusson et al. 2005; Lee et al. 2008). The apparent first-order depuration of 16 PAHs, alkylated PAHs, and total PAHs were calculated by fitting the data into a linear least square analysis (Branson et al. 1975; Djomo et al. 1996; Neely 1979). In this study, depuration rate and half-life was determined using only substantially detected priority PAHs, such as naphthalene, fluorene, phenanthrene, fluoranthene, pyrene, benz[a]anthracene and chrysene for PAHs, and alkylated homologues of naphthalene, fluorene, phenanthrene, and chrysene for alkylated PAHs. ½Concentration at any given time ¼ ½Concentration at beginning of depuration Kd t ½lnCa ¼ ½lnCt Kd t Ca = concentration of contaminants at any given time, Ct = concentration of contaminants at the beginning of depuration, Kd = depuration rate constant, and t = time in terms of days. The depuration half-life (t1/2) also was computed. t1=2 ¼ ln 2=Kd Average depuration-rate constants for the target compound groups at all of the stations varied from -0.015 to -0.038; 16 PAHs showed the lowest rates, and the alkylated PAHs showed the highest rates (Table 2). As alkyl groups of

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Arch Environ Contam Toxicol Table 2 Depuration-rates and half-lives of PAHs in oysters

Apparent depuration-rate constant (K) PAHs

-0.015

47.54

Alkylated PAHs

-0.038

18.23

Target ? alkylated PAHs

-0.037

18.72

naphthalenes include almost 50% of the initial volume of the alkylated PAHs, loss of naphthalene alkyl group series accounted for the high depuration rates of alkylated PAHs compared with the 16 PAHs. The half-life estimates for all of the stations within the spill region ranged from 47.5 days for the 16 PAHs, 18.7 days for the TPAHs, and 18.2 days for the alkylated PAHs. Because alkylated PAHs comprised more than 90% of the TPAHs, the TPAHs depuration rate and half-life was highly dependent on those of the alkylated PAHs. Human Health Risk Assessment As a basis for human health risk assessment, PAHs are separated into two subclasses based on their carcinogenicity: carcinogenic and noncarcinogenic (Reilly and York 2001). In application of the cancer slope factor, the toxic equivalency factor (TEF) approach often is used to estimate the carcinogenic activity of PAHs relative to BaP. Of the wide range of PAHs, only BaP has been well characterized in terms of toxicology and has been commonly used to determine ‘‘Levels of Concern’’ through the use of BaP equivalents (BaPE) (Yender et al. 2002). The following equation was used to determine levels of concern for public health (LoC: in ng/g lipid wt.) for carcinogenic PAH compounds (BaPE) potentially found in oysters (Reilly and York 2001; Yender et al. 2002). LoC ðBaPEÞ ¼ ðRL BW ATÞ=ðCSF CR EDÞ RL is the risk level. BW is the average consumer body weight. AT is the averaging time (i.e., life expectancy). CSF is the cancer slope factor of BaP. CR is the consumption rate (daily amount of seafood consumed). ED is the assumed exposure duration. Using regional statistics on the Korean lifestyle, LoC was calculated to be 3.35 ng/g lipid wt. Levels of TEF for the PAHs range from 0.001 to 5; BaP represents 1 as recommended by Nisbet and Lagoy (1992). Tissue concentrations of PAHs other than BaP were multiplied by their respective TEF and added to the tissue concentrations of BaP for the determination of BaPE. Cancer risk interpretations were performed by comparing LoC and BaPE values from each station (Fig. 7). BaPE concentrations exceeded the LoC for oysters in five stations (St. 4, 16, 17, 18, and 19) over the 10 days directly after the spill, which highlights the potential for human health risks. This exceedance was prolonged even after 60 days postspill. Korea has the most productive shellfish aquaculture

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Half-life (t1/2; days)

farming areas in the world and supplies large amounts of the shellfish products to local and nonlocal consumers (Kwon et al. 2004; Yoon et al. 2003), and hence persistence of contaminants in shellfish suggests a need for further long-term investigation and health risk assessments in this region. Comparison with Previous Oil Spills To briefly compare the levels of contamination in oysters, we selectively compared several components between the Wu Yi San oil spill and previous large spills (Table 3). It is common to find a rapid increase followed by a decrease in petroleum derived PAHs in bivalves after an oil spill. However, this is exceptional for the Deepwater Horizon oil spill in the sense that it was an offshore deep-sea blowout. Unlike other spills, the oil released from the Deepwater Horizon blowout allows it to undergo more weathering processes before being stranded on shorelines. The increase in petroleum derived PAHs in bivalves after an oil spill is mainly due to dissolved and dispersed oil to which they were exposed, but in the Deepwater Horizon oil spill, no significant differences of petroleum derived PAHs were observed in bivalves that were exposed to the highly weathered emulsions. After the largest oil spill in Korea—the Hebei Spirit oil spill in 2007—advances in the fields of environmental health, ecological risk assessment, and development of new technologies provided better preparedness for responses to similar accidents that occur. Compared with other large spills, emergency oil spill responses were deployed within 24 h after the Wu Yi San oil spill. Right after deployment, multimedia samples were collected and analyzed for short- and long-term effects. The rapid emergency environmental impact assessment and consecutive post-spill monitoring along with prespill data provided better information of the fate and effects of the spill. This also allows the identification of accumulation and depuration phases in oyster which is the first field observation in oil spill cases, and contributes to protect human health from consuming contaminated seafood.

Conclusions After the Hebei Spirit oil spill on December 7, 2007, advances in national preparedness have been elevated, including prevention, response, and restoration of marine


(a) 5

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Arch Environ Contam Toxicol b Fig. 7 Temporal changes of benzo[a]pyrene equivalent concentra-

tions. Level of concern (LoC) was calculated based on statistics of Korean lifestyle. Each station number corresponds to the station shown in Fig. 1. a Pre-spill, b 4 days, c 10 days, d 30 days, e 60 days

Table 3 Comparison of the Wu Yi San oil spill with previous large spills Spill case

Wu Yi San

Date of spill

31st Jan 2014

24th March 1989

7th Dec 2007

20th April 2010

Spill type

Coastal pipeline leakage

Grounding of oil tankerm,r

Oil tanker collisionw,v

Deep-sea blowoutb,c,e

Type of oil spilled

Medium crude oil (Basrah Light) 900 m3

Medium crude oil (Alaskan North Slope)m,r 42,000 m3;m,r

Mixture of middle and heavy crude oil (Kuwait Export, Iranian Heavy, and Upper Zakum)v,w 12,547 m3;w

Light crude oil (Louisiana Sweet)b 500,000 m3;b

Application of cleanup

Mechanical only

Mechanical and chemicall,m

Mechanical and chemicalw,x

Mainly chemicalb,u

State of oil stranded on shorelines right after spill

Fresh to slightly weathered Right after spill

Slightly to highly weatheredo,p

Fresh to slightly weatheredv,w,

Highly weathereds,t

Approximately 1–3 dayso,r

Right after spillw,x

Approximately 25–40 daysj

Duration of cleanup for stranded oil

10–30 days

2 monthsh

1–2 yearsf,w

1–2 yearsi,k,n

Form of oil potentially exposed to bivalves (along/stranded on shorelines) Bivalve species

Dissolved and dispersed

Dissolved and dispersedo,p

Dissolved and dispersedf,x

Mainly emulsiona,d,g,t

Crassostrea gigas

Mytilus trossuluso,p

Crassostrea gigas

Crassostrea virginicab,d,q

2000–110,000

5000–234,000o

8000–100,000w

No significant changesd,q

Amount of oil spilled

Time for oil to reach shorelines (stranded)

Level of contamination in oysters post-spill (PAHs; ng/g dw)

Exxon Valdez

Hebei Spirit

Deepwater Horizon

a

Aeppli et al. (2012); bBeyer et al. (2016); cCamilli et al. (2010); dCarmichael et al. (2012); eDiercks et al. (2010); fHong et al. (2014); gLiu et al. (2012); hMearns (1996); iMichel et al. (2013); jOSAT (2011); kOwens et al. (2011); lPeterson (2001); mPetterson et al. (2003); nSantner et al. (2011); oShort and Babcock (1996); pShort and Harris (1996); qSoniat et al. (2011); rSpies et al. (1996); sSteffy et al. (2013); tStout et al. (2016); u Thibodeaux et al. (2011); vYim et al. (2011); wYim et al. (2012); xYim and Shim (2017)

oil spills. Despite these national efforts, negligence with regard to safety caused the Wu Yi San oil spill. This spill has reawakened awareness of the importance of emergency responses for marine disasters, and this study was conducted as a part of such emergency environmental responses. Multimedia, environmental monitoring has been conducted for 2 months; oysters clearly demonstrated that the temporal changes of petroleum hydrocarbons contamination in the intertidal environment. Levels of contamination in oysters corresponded with the time and duration of removal of residual oils during cleanup processes. These temporal changes of contaminants were identified as accumulation and depuration phases, which is the first field observation in oil spill cases.

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Human health risk assessment demonstrated actual risks of consuming contaminated oysters produced in these regions. Risk levels decreased over time, but one site exceeded the LoC 2 months after the spill, suggesting the necessity of further, long-term investigations of shellfish products in these regions. Temporal changes of bioaccumulated PAHs in oyster and their depuration rates all indicated that there remained persistent effects from the residual oil after the Wu Yi San oil spill, thus emphasizing the importance of a long-term monitoring study. Acknowledgements This work was supported by the Project entitled ‘‘Oil Spill Environmental Impact Assessment and Environmental Restoration (PM59291)’’ funded by the Ministry of Oceans and Fisheries of Korea.


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