Effects of Macondo Canyon 252 oil (naturally and chemically dispersed) on larval Crassostrea virginica (Gmelin, 1791).
KEY WORDS: Crassostrea virginica, oil spills, Deepwater Horizon, MC252, water-accommodated fraction, chemically enhanced water-accommodated fraction, Corexit EC9500A
On April 20, 2010, an explosion on the Deepwater Horizon (DWH) rig caused a release of 4.9 million barrels (200 million gallons) of oil into the Gulf of Mexico before the well was capped on July 15 (Kelly 2010, National Commission on the BP Deepwater Horizon Oil Spill and Offshore Drilling 201 lb). Up to 100,000 [km.sup.2] of the Gulf of Mexico was affected by the spill for at least 1 exposure (Kelly 2010), including more than 1,000 total linear mi. of coastlines in Louisiana, Alabama, Mississippi, and Florida, and the key harvesting areas of the eastern oyster, Crassostrea virginica (Gmelin 1791) (Upton 2011).
In an effort to contain the oil and prevent it from reaching the shoreline, a variety of methods such as use of booms, skimmers, burning, direct recovery, and dispersants were used (National Commission on the BP Deepwater Horizon Oil Spill and Offshore Drilling 2011a). The use of dispersants, used within 48 h of the spill, was the most controversial of these methods and it has been calculated that 43,884 barrels (1.9 million gallons) of dispersant were used (Federal Interagency Solutions Group 2010). Dispersants are typically used when other methods are not found to be adequate. Chemical dispersants do not remove oil from water but merely accelerate its natural dispersion. Although the potential benefits of chemical dispersants include decreasing the impact of oil to shorelines and surface-dwelling organisms, dispersed oil remains in the water column where it threatens pelagic and benthic organisms (National Research Council 2005). The potential for decreased survival and delayed or abnormal development of exposed embryonic and larval oysters may have resulted in decreased population numbers in affected areas, with resulting ecological and economic consequences.
A review of oiling maps (National Oceanic and Atmospheric Administration GeoPlatform) shows that various amounts of dispersed and undispersed oil reached oyster fishing grounds in Louisiana and other Gulf states, although the amount of oil affecting specific oyster beds is hard to discern. Undispersed oil particles tend to remain near the surface of the water, and the most apparent effect on oyster reefs is the smothering of juvenile and adult shells (Tunnell 2011). Dispersed oil generates a large number of small particles that increase bioavailability through enhanced diffusion of hydrocarbons across cellular membranes (National Research Council 2005). Oil releases may affect marine organisms in a variety of ways, including physically (e.g., asphyxia), through habitat modification (e.g., decreased food, decreased oxygen), or through toxic effects caused by exposure to polycyclic aromatic hydrocarbons (PAHs) that are known to produce carcinogenic and mutagenic effects. The DWH event coincided with the spring spawning season for a number of aquatic organisms, including the eastern oyster. Exposure to oil (dispersed or undispersed) in the water column not only affects survival, but also may affect oyster health by reduced feeding, respiration, gametogenesis, and spawning (Suchanek 1993). Several studies have shown negative impacts of oil exposure on various bivalves, gastropods, copepods, nematodes, and sea urchins (Temara et al. 1999, Lee et al. 2002, Le Hir & Hily 2002, Martinez-Jeronimo et al. 2005, Fernandez et al. 2006, Lewis et al. 2008). In addition to acute effects, predator avoidance and food intake may be impacted by a reduction in swimming (Wu et al. 1997, Barron et al. 2003, Georgiades et al. 2003, Greco et al. 2006).
The aim of this study was to establish what level of exposure (time, concentration) to Macondo Canyon (MC) 252 oil (chemically and naturally dispersed) from the DWH well is necessary to affect larval eastern oysters. Acute exposure laboratory studies were conducted to assess fertilization success, development, survival, and swimming behavior, and sublethal exposure laboratory experiments were conducted to assess the impact of brief exposures on growth and survival.
MATERIALS AND METHODS
Oil and dispersed oil solutions for all experiments were prepared with MC252 oil (British Petroleum Company. BP PLC, London, UK) and/or Corexit 9500A dispersant (Nalco/Exxon Energy Chemicals, Sugarland, TX). Solutions were prepared according to Chemical Response to Oil Spills: Ecological Research Forum procedures (Clark et al. 2001, Singer et al. 2001, Barron & Ka'aihue 2003). Crude oil was weathered physically in the laboratory for 24 h prior to solution preparation by placement on a stir plate and mixing with a magnetic stir bar in the dark in a fume hood. Water accommodated fractions (WAFs) of physically weathered oil and chemically enhanced water-accommodated fractions (CEWAFs; 1:10 ratio) were prepared in 2-L flasks. All solutions used filtered (0.2 [micro]m), UV-treated saltwater (salinity, 28). Flasks were covered then mixed at a moderate mixing energy (25% vortex) for 24 h. Solutions were allowed to settle for 3 h prior to use. Static toxicity tests were conducted for all experiments using nominal WAF concentrations of 100-1,200 mg/L and CEWAF concentrations of 6.25-200 mg/L, with 5 replicates per treatment. Oil and dispersed oil (2 g/L) preserved with dichloromethane (1:10 v/v) were extracted using modified EPA method 3510C and were analyzed (Mote Marine Laboratory, Sarasota, FL) (1) for polycyclic aromatic hydrocarbons (PAHs; parent compounds and their homologs) using GC/MS (Agilent 7890A/5975C) and modified EPA method 8260 and (2) for total petroleum hydrocarbons (TPLls; n-C9-n-C42) using a GC with a flame ionization detector (Agilent 7890A) as described in Goodbody-Gringley et al. (2013).
Ripe oysters (Crassostrea virginica) were collected from non-DWH-affected East Coast Florida waters and were strip spawned. Gametes were examined microscopically, and egg quality, sperm motility, and gamete quantity were assessed. Eggs were sieved on a 200-[micro]m screen to remove debris. Pooled eggs from 2-3 females were collected on a 20-[micro]m screen, rinsed with UV-treated saltwater (salinity, 28), filtered to 1 [micro]m, and fertilized with sperm from 1-3 males for each assay (50 sperm/egg).
Fertilization success experiments were conducted in 15-mL test tubes, to which 4.5 mL saltwater, oil, or dispersed oil had been added. Gametes were added to each tube in the following manner: 0.490 mL filtered saltwater containing unfertilized eggs (800/mL) followed immediately by the addition of 0.01 mL sperm (50/egg). Four hours postfertilization, solutions were fixed by the addition of 10% neutral buffered formalin (100 [micro]L). One-milliliter aliquots were examined microscopically and the embryonic stage was recorded. Percent fertilization was calculated for all treatment groups.
For trochophore development experiments, percent fertilization was recorded 4 h postfertilization, and 0.5 mL fertilized eggs were placed into 15-mL tubes containing 4.5 mL saltwater, oil, or dispersed oil. On reaching trochophore stage (12 h postfertilization), solutions were formalin fixed as outlined previously. One-milliliter aliquots were examined microscopically. The number of oysters at each developmental stage was recorded and the percent that attained trochophore stage was calculated. For D-stage development experiments, fertilized gametes were allowed to reach trochophore stage before placement into 1 of the treatment groups described earlier. After 24 h, solutions were formalin fixed. One-milliliter aliquots were examined microscopically. The number of trochophores and D-stage larvae were recorded. D-stage abnormalities (e.g., convex hinge, indented shell margin and protruded mantles) as described by His et al. (1997) were recorded, and percent of normal and abnormal larvae were compared. Significant differences were determined by repeated-measures analysis of variance (data log + 1 transformed), followed by a Tukey's "Studentized" range test for comparative differences between treatments using SAS v. 9.2 (SAS Institute, Cary, NC). Normality and equal variance were tested to ensure the validity of using a parametric test versus a nonparametric test.
Short-term survival experiments were conducted with D-stage and eyed larvae. Fifteen D-stage larvae were placed in finger bowls containing 50 mL WAF or CEWAF solution as outlined earlier. Fifteen eyed larvae were placed in 1-L beakers containing 600 mL of the appropriate solution. All containers were placed in incubators (27[degrees]C, 12 h/12 h light/dark cycle). Oysters were fed Isochrysis galbana (50,000 cells/mL) once per day. Survival was assessed at 24, 48, 72, and 96 h. Lethal concentrations (L[C.sub.50]) were determined using the trimmed Spearman-Karber method (ToxCalcv5.0). D-stage swimming behavior was assessed concurrent with survival. Swimming ability was scored on a scale of 1-4 points: 1 point, highly active; 2 points, moderately active; 3 points, alive but not swimming; or 4 points, nonresponsive.
D-stage oyster larvae were exposed to dispersed oil concentrations (CEWAF) of 16 mg/L for 24 h and monitored until settlement to determine long-term effects of short-term exposure. The concentration used represents the L[C.sub.10] value derived from the ToxCalc analysis. D-stage larvae were obtained from strip spawns as described previously. A total of 5,530 eggs/mL were obtained from 5 females and were fertilized at a 1:100 ratio with sperm from 5 males. After 2 h, fertilized eggs were placed in a 400-L larval rearing tank containing UV-filtered saltwater (salinity, 28; temperature, 26[degrees]C; photoperiod, 12 h light/12 h dark). After 24 h, D-stage oysters were sieved and counted. Approximately 200.000 D-stage oysters were placed into 1 of 6 13-L buckets, provided with minimal aeration, containing saltwater or 16 ppm CEWAF (3 replicates per treatment) for 24 h (2610/6/2014C). Oysters were fed Isochrysis galbana (50,000 cells/mL) during exposure. After exposure, surviving oysters from each treatment group were combined and transferred into 1 of4 400-L larval rearing tanks. Tanks were drained down and sieved each day, and larvae were transferred to a newly filled larval rearing tank and fed I. galbana as outlined earlier.
Every other day, sieved oysters from each tank were suspended in 1-L beaker, and a 1-mL sample was collected and fixed in 10% buffered formalin. The number of larvae present in the 1-mL sample was counted and survival was assessed. Larvae were examined microscopically, photographed (Infinity 2 digital camera; Lumenera Co.), and shell length was measured using Infinity Analyze (Lumenera Co.).
Fertilization was reduced significantly ([F.sub.15,64] = 7.86, P < 0.0001) by exposure to CEWAFs (100 mg/L), whereas exposure to WAFs had no impact on fertilization success (Fig. 1).
Fewer trochophores developed after exposure of fertilized eggs to dispersed oil (25 mg/L) or oil (100 mg/L; [F.sub.15,64] = 50.16, P < 0.0001) compared with controls (Fig. 1). Chemically enhanced WAF concentrations less than 25 mg/L (6.25 mg/L and 12.5 mg/L) did not affect trochophore development significantly. Water-accommodated fraction concentrations less than 100 mg/L were not tested.
Fewer D-stage larvae developed after exposure of trochophores to CEWAFs (12.5 mg/L) or WAFs (200 mg/L; [F.sub.15,64] = 24.91, P < 0.0001) than controls (Fig. 1). Increasing concentrations resulted in further reduction of D-stage development in both solutions. No D-stage development was seen in trochophores exposed to 200 mg/L CEWAFs.
D-stage abnormalities (e.g., convex hinge, indented shell margin and protruded mantles) were seen after exposure to dispersed oil and oil (Fig. 2). Exposure to CEWAFS [greater than or equal to] 100 mg/L resulted in a significantly higher percent of abnormal D-stage larvae ([F.sub.5,24] = 4.41, P < 0.0072). A significant difference in D-stage abnormalities occurred at all WAF exposures (100-1.200 mg/L) except 400 mg/L ([F.sub.5,23] = 5.04, P < 0.0029). The anomaly seen at 400 mg/L may be explained by noting that the number of trochophores that developed into D stage varied between treatment groups, and most abnormalities seen at lower concentrations had concave or convex hinges, in contrast to the more apparent mantle protrusion seen at higher (800 mg/L and 1.200 mg/L) exposures.
Exposure to CEWAFs had a significant effect on survival of D-stage and eyed larvae (178 mg/L and 82 mg/[L.sup.-1], L[C.sub.50], 24 h; 44 mg/L L[C.sub.50], 48 h), whereas exposure to WAFs affected D-stage larvae only (Table 1). The acute toxicity for D-stage larvae exposed to oil decreased over time from 1,093 mg/L (L[C.sub.50], 24 h) to 262 mg/L (L[C.sub.50], 96 h); however, no observed toxicity effects were seen with eyed larvae, and no L[C.sub.50] value (>1,200 mg/L,96 h) was obtained (Table 1).
Significant differences (P [less than or equal to] 0.01) were seen in swimming behavior after exposure to CEWAFs at 24 h ([F.sub.5,19] = 21.09, P < 0.0001), 48 h ([F.sub.5,19] = 37.30. P < 0.0001), 72 h([F.sub.4,18] = 20.18, P < 0.0001), and 96 h (F519 = 5.50, P = 0.0027) of exposure as indicated by superscript letters (Table 2). Exposure to 100 mg/L CEWAFs affected swimming behavior from 48-96 h of exposure. (No activity was recorded for 100-200 mg/L at 24 h.) Larvae exposed to 12.5 mg/L CEWAF showed decreased activity; however, this was not seen in the 25-mg/L treatment group. This treatment difference was seen across all replicates and was consistent with higher mortalities seen in that treatment group compared with 25 mg/L (data not shown). It is possible that the 2 treatment groups may have been mislabeled inadvertently or that some other issue (e.g., water quality) arose in the 12.5-mg/L treatment group that was responsible for affecting activity as well as survival. Significant differences (P [less than or equal to] 0.01) were likewise seen in swimming behavior after exposure to WAFs at 24 h ([F.sub.5,21] = 5.50, P < 0.0021), 48 h ([F.sub.5,21] = 13.44, P < 0.0001), 72 h ([F.sub.5,21] = 12.46, P< 0.0001), and 96 h ([F.sub.5,21] = 6.20, P = 0.0011) of exposure (Table 2). Exposure to WAFs of 400 mg/ L resulted in decreased swimming activity at 24 h and 48 h, and exposure to 200 mg/L resulted in decreased swimming activity at 72 h and 96 h.
There was no significant difference (P > 0.05) in survival between control and exposed (16 mg/L, CEWAF, L[C.sub.10], 24 h) D-stage larvae (Fig. 3). Survival of D-stage larvae exposed to 16mg/L (L[C.sub.10], 24 h) CEWAFs 4 days postexposure was similar to the control (55 [+ or -] 32% and 57 [+ or -] 19%, respectively). At week 1, survival was 52 [+ or -] 19% for controls and 22 [+ or -] 12% for exposed groups; at week 3, survival was 19.3 [+ or -] 3% and 6.6 [+ or -] 5%, respectively.
There was no significant difference (P > 0.05) in growth between control and exposed larvae (Fig. 4). Control larvae increased from a length of 129 [+ or -] 9.1 (SD) [micro]m on day 1 to a length of 501 [+ or -] 75.8 [micro]m by day 19 of larval culture. Exposed larvae increased from a length of 129 [+ or -] 9.1 [micro]m on day 1 to a length of 509 [+ or -] 55.3 [micro]m by day 19.
Polycyclic Aromatic Hydrocarbon and TPH Analysis
The total PAH level in the CEWAF 2-g/L stock solution (1,429 [micro]g/L) was 3 times that of the 2-g/L WAF stock solution (452 [micro]g/L), whereas the TPH level (62 mg/L) in the CEWAF solution was 25 times greater than the WAF stock solution (2.5 mg/L; Table 3). The predominant compound in both solutions was naphthalene, which made up 65% of the compounds in the CEWAF and 83.5% of the compounds in the WAF stock solution. Levels of compounds containing 3 and 4 carbon rings (e.g., anthracene, fluorene, pyrene, chrysene, and phenanthrene) were approximately 2 times greater in CEWAFs than WAFs.
Chemically enhanced WAFs exhibited greater toxicity than WAFs if nominal (6-12-fold) or PAH (2-4-fold) values were compared, but were less toxic if TPH values were used (Table 3). Water-accommodated fractions were 2.5-4 times more toxic than CEWAFs if TPH values were considered (Table 3).
The water column represents the most likely route of uptake of oil contaminants by free-swimming marine larval organisms immediately after an oil release. In this study, exposure of Crassostrea virginica embryos and larvae to MC252 oil (WAFs) and dispersed oil (CEWAFs) affected fertilization, larval development, behavior, and survival adversely. Exposure to CEWAFs typically had a more pronounced effect than WAFs. This was especially evident with regard to fertilization success and eyed larval survival. Dispersants work by reducing the interfacial tension between oil and water, resulting in the creation of a large number of small oil droplets (National Research Council 2005). Although this action breaks up surface
oil slicks, the increase in the number of particulates enhances the bioavailability of hydrocarbons in the water column that would normally remain near the surface (Singer et al. 1998). Although dispersant use after the DWH event likely resulted in decreased amounts of oil reaching shorelines, its use likely increased the toxicity of that oil by enhancing its bioavailability to organisms living in the water column. Early life stages tend to be more susceptible to toxic compounds than adults, which may have negative consequences on populations inhabiting affected areas, especially in the short term. In addition, early larval life stages of many invertebrate species, including oysters, are part of the planktonic food web--an important food source for other organisms, such as fish.
Survival is often used as the end point to determine oil toxicity effects. In the current study, it was noted that oil exposure impacted fertilization success, normal development and behavior of free-swimming oyster larvae, and survival negatively. Effects were seen after as little as 4 h of exposure (fertilization success). Early larval stages (fertilized eggs, trochophores) showed enhanced sensitivity compared with more developed (eyed) larval stages.
The CEWAF and WAF exposure levels used in these experiments represent moderate to high concentrations compared with those measured in the environment. Dispersed oil levels of 25-75 mg/L and oil levels of 20-600 mg/L have been reported in the water column 24 h after a spill and may be a magnitude greater immediately after a spill (Law et al. 1997, National Research Council 2005 Bellas et al. 2008). The exposure times used in the current experiments (4-96 h) were likewise representative of realistic scenarios. Oil contaminates persist for various lengths of time, depending on environmental conditions. More persistent levels occur in sheltered environments, such as estuaries, as a result of the lack of the strong wave action needed to remove them (Kingston 2002). After the Prestige spill, PAH levels along the Galician coast were 4-30 times higher than background levels and remained greater than background for approximately a year (Franco et al. 2006, Gonzalez et al. 2006, Neito et al. 2006). After the DWH event, coastal regions showed PAH levels higher than background for as little as 1 mo to as long as 1 y (Allan et al. 2012).
Fertilization success was affected negatively by exposure to CEWAFs but not WAFs. In contrast, both contaminants affected larval development, although the impact was seen at lower concentrations for CEWAFs. Oil has been shown previously to affect embryogenesis in various marine organisms. Decreased hatch rate and motility of sperm have been reported in fish and polychaetes (Wilson 1976, Lewis et al. 2008). Reduced fertilization success has likewise been reported in sea urchins and polychaetes (Beiras & Saco-Alvarez 2006, Fernandez et al. 2006, Lewis et al. 2008). In contrast, Baussant et al. (2011) saw no fertilization or development impacts in mussels. However, differences between these studies may be explained by the fact that mussel larvae were exposed to lower concentrations and that a flow-through, rather than a static exposure, system was used.
Abnormal D-stage development was seen after exposure to CEWAFs and WAFs. Similar results were reported for mussels (Mytilus galloprovincialis) and oysters (Crassostrea gigas) after exposure to oil from the Erika spill (Geffard et al. 2004), to mussels after exposure to Prestige and Marine WAFs (Saco-Alvarez et al. 2008), and to mussels after exposure to North Sea crude oil (Baussant et al. 2011). Not all species are affected similarly, however. Lewis et al. (2008) reported differences in developmental effects between polychaete worms exposed to the same concentrations of oil; a significant effect was seen with respect to Nereis virens development, but not for Arenicola marina. Most studies have focused on the effects of exposure to oil only; few have compared the effects of exposure to both oil and chemically dispersed oil on fertilization and development. One early study (Singer et al. 1998) compared the development of abalone (Haliotis rufescens) embryos exposed to CEWAFs and WAFs; the authors of that study reported shell deformation abnormalities in individuals after exposure to CEWAFs but not WAFs, with similar concentrations of TPH.
Increased mortality has also been reported for a variety of larval marine organisms after exposure to oil and dispersed oil (Long & Holdway 2002, Lewis et al. 2008, Saco-Alvarez et al. 2008). Most studies focus on only 1 life stage--with results that may or may not be applicable to all life stages. D-stage larvae were affected negatively by exposure to WAFs, whereas there was no impact on the survival of eyed larvae exposed to high WAF concentrations (1,200 mg/L) even after 96 h. In contrast, the survival of both larval stages was affected by exposure to CEWAFs at similar and lower concentrations (<200 mg/L). Although more mortalities after exposure to CEWAFS compared with WAFs have been reported previously in fish, snails, and coral (Gulec et al. 1997, Couillard et al. 2005, Shafir et al. 2007), other researchers have reported no increase in toxicity with dispersant use in a variety of species (Singer et al. 1998, Mitchell & Holdway 2000, Long & Holdway 2002, Fuller et al. 2004).
A through literature survey by Fincas (2008) reported that 75% of researchers found that chemically dispersed oil was more toxic than physically dispersed oil. More than 50% of those researchers concluded that this toxicity was the result of the increased availability of PAHs or oil in the water column itself (Fincas 2008).
Conflicting research results are often reported with the same species across studies. There are a variety of explanations for this, including types of oil used, methods of exposure (static, static renewal, flow through), and the tendency to report results based on nominal concentrations. It has been shown that methods of reporting (nominal, PAHs, TPHs) may affect conclusions (Perkins et al. 2003, Bejarano et al. 2014). In the current study the concentration PAHs and TPHs in the stock solutions was measured, and all 3 concentrations were reported when comparing L[C.sub.50] values of exposed D-stage and eyed larvae. Results showed that WAFs had less impact on survival than CEWAFs if either nominal or PAH values were compared. However, there were distinct differences with regard to magnitude when reporting nominal versus PAH concentrations. Results reported using nominal concentrations suggest that CEWAFs had a 10-fold greater toxicity than WAFs, whereas measured PAH levels showed that CEWAFs had only a 2-3-fold greater toxicity than WAFs. Chemically enhanced WAF stock solutions contained a 3-fold greater concentration of PAHs than WAFs. In contrast, WAFs had a greater toxicity than CEWAFs if TPH values were considered. This is of particular interest in that the CEWAF stock solutions had a 25-fold greater concentration of TPHs than WAFs. Total petroleum hydrocarbons are a mixture of many different compounds found in crude oil whereas PAHs are considered to contain the most toxic components of light crude oil. During and after the spill, the amount of TPHs, total PAHs, and various classes of PAHs were measured. Reported concentrations of total PAHs varied depending on location, with a high of 146 mg/L at the wellhead (Boehm et al. 2011), and reported highs of 0.895 mg/L in field-collected samples (Echols et al. 2015) and 0.00170 mg/L in coastal waters (Allan et al. 2012). As expected, as a result of settlement and bioaccumulation, sediment and flora and fauna contained higher levels of oil contaminants than seawater, with TPH concentrations as high as 200 mg/L and total PAH levels of less than 0.1 mg/L, with phenanthrene/anthracene having the highest (1.1 mg/L) concentrations (Sammacaro et al. 2013). Negative impacts were evident in the present study at TPH values of 1.3-5 mg/L and 0.02-0.2 mg/L, similar to or below reported TPH values and similar to or slightly above PAH values reported during the DWH event. In both the CEWAF and WAF solutions, naphthalene was the predominant (65%-84%, CEWAF, WAF) PAH compound, but anthracene also made a significant contribution to the total PAH (16.5%-6.5%, CEWAF, WAF), and phenanthrene made a smaller but still measurable component (2.4%-1.7%, CEWAF, WAF).
Biological end points other than mortality need to be considered when assessing contaminant effects, because the ability to move and thereby feed or escape predation likely affects survival. Some researchers have reported behavioral inhibitory (I[C.sub.50]) effects after exposure to lower levels of oil contaminants than that at which mortality (L[C.sub.50]) occurs. Effects on herring swimming behavior and asteroid prey localization behavior have been reported at concentrations half of the L[C.sub.50] values for both oil and dispersed oil (Barron et al. 2003, Georgiades et al. 2003). Other researchers have reported no differences in L[C.sub.50] and E[C.sub.50] values. Similar values were seen in Barnacle nauplii when comparing survival, swimming, and settlement behavior (Wu et al. 1997, Greco et al. 2006), as did amphipod survival and burying behavior (Gulec et al. 1997). In the current study, exposure of D-stage larvae to WAFs caused a decrease in swimming ability at one third of the 24-h L[C.sub.50] and one half of the 48-h LC[5.sub.0], after which time no differences were seen, and no differences were seen in concentrations causing decreased swimming ability or survival after exposure to CEWAFs. Rice et al. (1976) concluded that decreased swimming behavior is likely a result of narcosis after short-term high-level exposures to naphthalene. In the current study, although CEWAFs contained a 3-fold greater concentration of naphthalene, WAFs had a greater percentage of naphthalene (83%) compared with CEWAFs (65%). Regardless of the explanation, decreased swimming ability after exposures to lower levels of WAFs than induces mortality could still have an impact on survival by reducing the ability to escape predation or to capture food.
Although some oyster reefs in the Gulf of Mexico may have experienced prolonged exposure to high or moderate concentrations of oil or dispersed oil, most areas that were exposed would have experienced limited exposure to low levels of contaminants (Boehm et al. 2011, Allan et al. 2012). Studies on the effects of short-term sublethal exposures are lacking in the literature; therefore, short-term (24-h) sublethal exposures (L[C.sub.10]) of CEWAFs to D-stage larvae were conducted in the current study. A trend was noted toward decreased survival in the exposed group 3 wk postexposure, but the difference was not significant because of the low number of replicates and a high SD. There were no differences in growth or development, which suggests that short-term sublethal exposure has no effect on survival or growth of D-stage oyster larvae. Baussant et al. (2011) reported a decrease in mussel growth after exposure of fertilized eggs to lower concentrations than those used in this study at day 7, but not at day 21. On the other hand, Shafir et al. (2007) reported reduced survival in coral after short-term (24-h) exposure to dispersed oil for 50 days. Baca et al. (2005) followed the effects of exposure to mangroves, seagrass, and coral to dispersed and nondispersed Prudhoe Bay oil for 20 y. Although short-term mortality was seen at both sites, long-term effects were noted only at oiled sites.
Gulf of Mexico studies show that PAH levels in the water decreased dramatically after the well was capped (Allan et al. 2012, Echols et al. 2015), rendering long-term effects for most species unlikely. Carmichael et al. (2012) reported no evidence of oil consumption in subadult oysters exposed to possible coastal contaminated waters during and after the spill. Soniat et al. (2011) reported that sampled adult oysters from oil-exposed areas had tissue PAH levels indistinguishable from background levels 6 mo after the well was capped.
In the current study, short-term (24-96-h) acute exposure of oyster larvae to WAFs or CEWAFs from the DWH oil affected fertilization, development, survival, and swimming ability negatively. Although oil concentrations and duration of exposure at many oyster reef locations in the Gulf of Mexico are unknown, concentrations that affected oyster larvae adversely in this study were at maximum or higher reported values for water samples collected in May 2010 and June 2010 (Allan et al. 2012, Echols et al. 2015), and were unlikely to have occurred after the capping of the well in July 2010, when reported total PAH values of water samples decreased dramatically (Allan et al. 2012, Echols et al. 2015). Still, the timing of the DWH event coincided with the spring spawning season, and therefore potential negative effects on oyster populations--particularly short-term effects on recruitment--may have occurred in some areas. However, oysters are a resilient species. If adult populations were not affected adversely, time to recovery for most affected areas has likely already occurred.
The authors thank Sloane Wendell. BP GCRO Reference Material Account Manager, for providing us with surrogate MC252 oil #SO-20110212-HMPA9-008; David Giddings, Strategic Technology Manager. Nalco Co. (Sugarland, TX), for providing us the Corexit 9500A dispersant; Fred Prahl, Biological
Scientist, HBOI-FAU, and Heather Kalisz, student at the FAU Honors College for providing assistance with experiments and data collection. This research was made possible by a grant from BP/The Gulf of Mexico Research Initiative through the Florida Institute of Oceanography #4710-1101-00B. This is HBOI-FAU publication no. 1939.
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SUSAN LARAMORE, (1) * WILLIAM KREBS (2) AND AMBER GARR (3)
(1) Florida Atlantic University, Harbor Branch Oceanographic Institute, 5600 US 1 North, Fort Pierce, FL 34946; (2) Colorado Catch, LLC, PO Box 210, Sanford, CO 81151; (3) EA Engineering, Science and Technology, 225 Schilling Circle, Suite 400, Hunt Valley, MD 21031
* Corresponding author. E-mail: email@example.com
TABLE 1. Median lethal concentration (L[C.sub.50]) values for oysters exposed to oil (WAF) and chemically dispersed oil (CEWAF) as determined by a trimmed Spearman-Karber method (ToxCalc). Stage Time WAF PAH (mg/L) (h) TPH (mg/L) D Stage 24 1,092.8 (962.4-1,240.8) 0.247 1.348 Eyed > 1200 -- -- D stage 48 554.2 (315.6-973.4) 0.125 0.684 Eyed > 1,200 -- -- D stage 72 289.2 (205.6-406.8) 0.065 0.384 Eyed > 1,200 -- -- D Stage 96 261.8 (126-544.1) 0.059 0.323 Eyed > 1200 -- -- Stage Time CEWAF PAH TPH (h) D Stage 24 177.6 (146.2-215.8) 0.127 5.560 Eyed 81.9 (72.4-92.7) 0.059 2.964 D stage 48 44.7 (36.9-54.2) 0.032 1.399 Eyed 44.2 (40.5-48.3) 0.032 1.384 D stage 72 33.8 (26.9-42.6) 0.024 1.058 Eyed 21.6 (18.7-25.2) 0.016 0.676 D Stage 96 24.8 (18.9-32.5) 0.018 0.776 Eyed 14.5 (12.7-16.6) 0.010 0.454 Reported values include nominal L[C.sub.50] (95% CL) and corresponding PAH and TPH levels (in milligrams per liter). Mortalities were not observed at a [greater than or equal to] 1,200 mg-L. CEWAF, chemically enhanced water-accom-modated fractions; PAH, polycyclic aromatic hydrocarbon; TPH, total petroleum hydrocarbon; WAF, water-accommodated fraction. TABLE 2. Behavioral response: swimming activity. Time WAF (h) Concentration % Active [+ or -] 95% CL 24 0 46.67 [+ or -] 13.33 (a,b) 100 48.00 [+ or -] 10.95 (a,b) 200 58.67 [+ or -] 18.50 (b) 400 30.67 [+ or -] 16.06 (a) 800 -- 1,200 29.33 [+ or -] 7.60 (a) 48 0 53.33 [+ or -] 14.14 (a) 100 54.67 [+ or -] 10.95 (a) 200 34.73 [+ or -] 14.41 (a,b) 400 10.67 [+ or -] 3.65 (a) 800 14.67 [+ or -] 5.58 (b,c) 1,200 29.33 [+ or -] 8.94 (b,c) 72 0 48.00 [+ or -] 14.45 (a) 100 37.33 [+ or -] 15.35 (a,b) 200 13.33 [+ or -] 4.71 (c) 400 12.00 [+ or -] 2.98 (c) 800 16.00 [+ or -] 5.96 (c) 1,200 21.33 [+ or -] 5.58 (b,c) 96 0 56.00 [+ or -] 18.01 (a) 100 44.00 [+ or -] 7.06 (a) 200 13.33 [+ or -] 11.55 (a,b) 400 17.33 [+ or -] 13.00 (a,b) 800 8.00 [+ or -] 5.58 (b) 1,200 33.33 [+ or -] 4.71 (a) Time CEWAF (h) Concentration % Active [+ or -] 95% CL 24 0 46.67 [+ or -] 13.33 (a,b) 6.25 65.33 [+ or -] 12.82 (b,c) 12.5 26.67 [+ or -] 10.54 (a) 25 72.00 [+ or -] 9.89 (c) 100 -- 200 -- 48 0 53.33 [+ or -] 14.14 (a) 6.25 58.67 [+ or -] 10.95 (a) 12.5 18.67 [+ or -] 17.26 (a,b) 25 54.67 [+ or -] 8.69 (a) 100 16.00 [+ or -] 10.11 (a,b) 200 0.00 [+ or -] 0.00 (c) 72 0 48.00 [+ or -] 14.45 (a,b) 6.25 54.67 [+ or -] 8.69 (a) 12.5 29.33 [+ or -] 7.60 (b) 25 53.33 [+ or -] 10.54 (a) 100 1.33 [+ or -] 2.98 (c) 200 96 0 56.00 [+ or -] 18.01 (a) 6.25 56.00 [+ or -] 5.96 (a) 12.5 24.00 [+ or -] 7.60 (a,b) 25 53.33 [+ or -] 10.54 (a) 100 0.00 [+ or -] 0.00 (b) 200 Percent of oyster larvae actively swimming exposed to short-term (24-96 h) acute concentrations of oil (WAF) and chemically dispersed oil (CEWAF). Concentrations represent nominal values (milligrams per liter). Activity was not recorded at 24 h for WAFs at 800 mg-L or CEWAFs at 100 mg-L and 200 mg-L. At 72 h and 96 h, all larvae exposed to 200 mg-L were dead. Different superscript letters within each time period in each column indicate significant differences between treatment groups (P < 0.005). CEWAF, chemically enhanced water-accommodated fraction; CL, confidence level; WAF, water-accommodated fraction. TABLE 3. Individual PAH, and total TPH and PAH concentrations of 2-g/L stock solutions of WAFs and CEWAFs used to prepare working solutions used in the acute toxicity experiments. Target compounds C rings CEWAF WAF ([micro]g/L) ([micro]g/L) Naphthalene (C0-C4) 2 925.96 377.66 Acenaphthylene 2 6.40 0.05 Acenaphthene 2 0.61 0.67 Fluorene (C0-C4) 3 102.92 14.65 Anthracene (C0-C4) 3 235.38 30.14 Phenanthrene 3 34.79 8.36 Fluoranthene 3 1.53 0.18 Chrysene (C0-C4) 4 32.6 4.02 Pyrene (C0-C4) 4 61.5 6.71 Benzo[a]anthracene 4 0.21 0.14 N aphthobenzo thiophene 4 1.32 0.16 (C0-C4) Dibenzothiophene 5 5.47 5.6 (C0-C4) Benzo[b]fluorene 5 5 0.72 0.09 Benzo[b]fluoranthene 5 0.00 0.07 Benzo[k]fluoranthene 5 0.43 0.00 Benzo[e]pyrene 5 0.61 0.11 Benzo[a]pyrene 5 0.00 0.02 Perylene 5 0.77 0.13 Dibenzo[a,h]anthracene 5 0.00 0.01 Indeno[ 1,2,3-cd]pyrene 6 0.01 0.00 Benzo[g,h,i]perylene 6 0.00 0.02 Total PAH ([micro]g/L) 1428.64 451.92 TPH C9-C42 ([micro]g/L) 62,613.50 2,466.57 CEWAFs, chemically enhanced water-accommodated fractions; PAH, polycyclic aromatic hydrocarbon; TPH, total petroleum hydrocarbons; WAFs, water-accommodated fractions.
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|Author:||Laramore, Susan; Krebs, William; Garr, Amber|
|Publication:||Journal of Shellfish Research|
|Date:||Dec 1, 2014|
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