Toxicity of Oil from BP Deepwater Horizon Blowout on the Early Life Stage of Red Drum

Toxicity of Oil from BP Deepwater Horizon Blowout on the Early Life Stage of Red Drum,

Sciaenops ocellatus

A Thesis Presented to the Faculty of the College of Arts and Sciences,

Florida Gulf Coast University

In Partial Fulfillment Of the Requirement for the

Degree of Master of Science

Kelsey L. McEachern

July 2014

APPROVAL SHEET

This thesis is submitted in partial fulfillment of

the requirements for the degree of

Master of Science

______

Kelsey L. McEachern

Approved:

______

Darren Rumbold, Ph.D.

Committee Chair / Advisor

______

Aswani Volety, Ph.D.

______

Greg Tolley, Ph.D.

Florida Gulf Coast University, 10501 FGCU Boulevard South, Fort Myers, FL 33965

The final copy of this thesis has been examined by the signatories, and we find that both the content and the form meet acceptable presentation standards of scholarly work in the above mentioned discipline.

ACKNOWLEDGEMENTS

I would like to convey my deepest gratitude and thanks to my major advisor Dr. Darren

Rumbold for his unwavering support during the course of my research and composition of this thesis. His academic advice and guidance throughout this process was invaluable. I consider myself lucky to have had him as a friend and mentor over the past three years. I would also like to acknowledge and thank my committee members Dr. Aswani Volety and Dr. Greg Tolley for their insight, engagement and roles as mentors during my education at FGCU. Over the past three years the combination of these three committee members and the faculty and staff of the

College of Arts and Sciences all have shown me the importance of what research science should be, while maintaining the highest standards possible.

Additionally, I would like to recognize the funding grant from BP/ The Gulf of Mexico

Research Initiative through the Florida Institute of Oceanography for making this research possible. Another instrumental part of this process was led by Dr. Ai Ning Loh and Ian

Campbell and the Centre of Documentation, Research and Experimentation on Accidental Water

Pollution who assisted in the analysis of water chemistry. Finally, I would like to thank my peers, friends, and family for the support and memories shared throughout the course of this process.

TABLE OF CONTENTS

ACKNOWLEDGEMENTS ...... ii TABLE OF CONTENTS ...... iii LIST OF TABLES ...... v LIST OF FIGURES ...... vi ABSTRACT ...... 1 I. INTRODUCTION ...... 3 a. Background on Deepwater Horizon Blowout...... 3 b. Tourism & Economy ...... 3 c. Conceptual Diagram of Ecological Risk Resulting from the Macondo Well Blowout ...... 5 i. Resources at Risk ...... 6 ii. Exposure Characterization ...... 7 iii. Effects Characterization ...... 13 iv. Assessment Endpoint ...... 15 d. Red drum, Sciaenops ocellatus ...... 17 II. ANALYSISPHASE OF RISK ASSESMENT ...... 20 a. CEWAF Preparation and Test Conditions ...... 21 b. Response Endpoints and Statistical Analyses...... 23 III. RESULTS ...... 28 a. CEWAF Characterization ...... 28 b. Developmental Abnormalities Observed ...... 30 c. Graduated Severity Index (GSI) ...... 32 d. Length Metrics ...... 35 e. EC50 ...... 37 f. LC50 ...... 38 IV. DISCUSSION...... 39 a. Normal Finfish Embryo-Larval Development...... 40 b. Modes of Toxicity and PAH Specific Toxicity ...... 41 i. Cardiac Development ...... 45 ii. Skeletal Development and Length Metrics ...... 47 iii. Yolk Sac Development ...... 49

iii

iv. Finfold Development ...... 50 c. Need for Standardization in Toxicity Testing ...... 51 d. Toxic Units ...... 52 e. EC50 / LC50 Comparison ...... 54 f. Variability ...... 56 g. Competitive Disadvantage and Ecological Significance ...... 57 REFERENCES ...... 63

LIST OF TABLES

Table 1. Comparison of Finfish Toxicity Tests

LIST OF FIGURES

Figure 1. Conceptual diagram of ecological risk…………………………………………….pg 7

Figure 2. Satellite image of surface oil………………………………………………………pg 9

Figure 3. Image of red drum larvae showing various measurements taken………………..pg 29

Figure 4. Average relative concentration and composition of major PAH constituents …. pg 30

Figure 5. Estimated average toxic units (TU) from 4 CEWAF stock solutions…………... pg 31

Figure 6. Developmental abnormalities observed as a result of 24 hr acute exposure to MC252

CEWAF……………………………………………………………………….....pg 33

Figure 7. The relation between exposure concentration and mean rank score of GSI …….pg 35

Figure 8. Total length of post hatch yolk sac larvae as a function of exposure

concentration……………………………………………………………….……pg 36

Figure 9. Ratio of the length of snout to vent to the length from vent to tip of the

tail/notochord……………………………………………………………….……pg 37

Figure 10. Ratio of yolk length to diameter versus exposure concentration………………...pg 38

Figure 11. Viability of red drum embryos exposed to increasing concentrations of MC252

CEWAF……………………………………………………………………….…pg 39

Figure 12. Mortality of red drum embryo-larval stages exposed to increasing concentrations of

MC252 CEWAF…………………………………………………………….…...pg 40

ABSTRACT

On April 20, 2010, one of the largest offshore oil spills in history occurred with the blowout of the British Petroleum Deepwater Horizon (DWH) Macondo Prospect well. The well blowout and subsequent release of oil had the potential to impact the ecosystem of the Gulf of

Mexico that harbors many significant resources. Accordingly, this study attempts to systematically evaluate and organize data, information, assumptions, and uncertainties regarding the DWH blowout using an ecological risk assessment approach. The analysis phase included toxicity tests of chemically enhanced water accommodated fraction (CEWAF) of MC252 crude oil and Corexit on the early life stages (ELS) of the red drum, Sciaenops ocellatus. The objectives of this study were threefold: (1) to characterize the toxicity of CEWAF of MC252 crude oil to red drum ELS, (2) determine how the sensitivity of this species compares to other finfish ELS; and (3) to determine if these laboratory results might be used to predict effects from in situ exposure to other commercially, recreationally and ecologically important finfish in the wake of the DWH blowout. Red drum ELS exposed for 24 hours to CEWAF with total polycyclic aromatic hydrocarbons (PAHs) ranging from 0.25 mg/L to 5.5 mg/L presented with one or more gross abnormalities including: cardiac edema, skeletal abnormalities, yolk sac edema, finfold abnormalities and decreased growth. The median concentration at which 50% of the red drum larvae experienced abnormalities (EC50) and were considered non-viable after 24 hour exposure to CEWAF ranged from 0.38 mg/L to 1.63 mg/L (n=2). The median lethal concentration (LC50) ranged from 0.48 mg/L to 2.43 mg/L (n = 2). The results of this study are in agreement with other toxicity tests using various finfish ELS native to the GOM, and show that the effects of MC252 CEWAF toxicity can be debilitating to individual finfish ELS.

Population level effects to red drum and other sensitive finfish as a result of the DWH blowout

would be dependent on the spatiotemporal severity of exposure particularly in relation to the timing of natural density-dependent population regulation.

I. INTRODUCTION

a. Background on Deepwater Horizon Blowout

On April 20, 2010, one of the largest offshore oil spills in history occurred with the

blowout of the British Petroleum (BP) Deepwater Horizon (DWH) Macondo Prospect well at a

depth of 1,500 m. Over the following months, until July 15th when the well was temporarily

capped, over 4.9 million barrels of oil were released over 180,000 km2 (Camilli et al., 2010;

Crone and Tolstoy 2010; McNutt et al. 2011). Approximately 140 miles of Louisiana, Alabama

and Mississippi coastline and 80 miles of Florida coastline and estuarine habitat were exposed to

oil. The oil traveled from about 40 miles offshore across surface waters vital to the planktonic

communities in the Gulf of Mexico (GOM). The well blowout and subsequent release of oil had

the potential to severely impact the U.S. economy by jeopardizing the utilization of living and

non-living natural resources that are continually harvested or extracted from the GOM.

b. Tourism & Economy

Extraction of oil, harvesting of seafood and non-extractive uses, such as tourism and

shipping, are major economic drivers for the region and beyond. The GOM basin supplies a

significant portion of the country’s fossil fuel needs: one quarter of the U.S. domestic natural gas

and one-eighth of its oil (GOMPO 2011). Currently there are 3,500 oil and gas exploration

platforms in the western GOM, equating to about 50,000 wells drilled (Tunnell 2011). In

addition, the GOM contains approximately 1,000 natural, slow chronic natural hydrocarbon

seeps, where the surrounding biota such as natural petroleum-eating bacteria have to an extent

adapted to the presence of oil (Tunnell 2011). The BP Deepwater Horizon (DWH) platform was

located in the northern Gulf of Mexico in block 252 in the Mississippi Canyon of the Macondo

Prospect. Coastlines in jeopardy of being oiled in the wake of the blowout included Texas,

Louisiana, Mississippi, Alabama and the eastern panhandle of Florida.

In addition to the extraction of the abiotic natural resources, the GOM prospers from the utilization and harvest of other biotic resources. The GOM as a whole supports a $20 billion tourism industry (GOMPO 2011). The fishing of the marsh flats along the coasts of northern

Florida, Louisiana and Mississippi, and diving the Flower Garden Banks or Keys are some of the highly visited tourism attractions in the GOM that were put at risk by the blowout of the DWH well head. While a significant amount of recreational fishing occurs on the east coast, away from the threat of oil exposure, Florida is the fishing capital of the world and much of that fame is gained from the GOM commercial and recreational fisheries. At one point 84,101 square miles surrounding DWH was closed to all commercial and recreational fishing: this equated to 34.8% of the GOM exclusive economic zone (EEZ) (Tunnell 2011). A threat to the health and sustainability of abiotic and biotic natural resources of the GOM is a threat to the economy that is reliant on those resources. In 2010, GOM fisheries made a substantial contribution to the U.S. economy: 1.3 billion pounds of commercially caught fish and shellfish valued at $639 million,

177.2 million pounds of shrimp valued at $340 million, and 15.7 million pounds of oysters valued at $54.5 million (GOMPO 2011). Considering that one barrel of crude oil, once refined, makes about 19 gallons of gas, it can be calculated that around 93,100,000 gallons of potential gasoline was lost. At $4.00 per gallon, this equates to $3.72 million dollars lost at the pump. It is estimated that in the next seven years that the total economic impact will be $8.7 billion dollars as a result of the DWH disaster. Of the $3.7 billion loss in revenues, $1.6 billion will be in the commercial sector and $1.9 in the recreational sector. Of the $1.9 billion loss in profits, $0.8 billion is from the commercial sector and $1.1 billion is from the recreational sector. Finally,

$4.9 billion of the overall economic impact is from the commercial sector with the remaining

$3.5 billion from the recreational sector (Sumaila et al. 2012).

c. Conceptual Diagram of Ecological Risk Resulting from the Macondo Well Blowout

It would be imprudent to believe that we are capable of accounting for every possible environmental impact from an event like the DWH blowout. Accordingly, this study will follow a formal ecological risk assessment process in an attempt to “systematically evaluate and organize data, information, assumptions, and uncertainties” (USEPA 1998) regarding the DWH blowout. One of the first tasks as part of problem formulation is the development of the conceptual model that identifies and organizes various measures to evaluate the risk hypotheses.

Just because oil is detected in the environment does not necessarily mean it is bioavailable or will cause adverse effects to biota exposed (Boehm and Page 2007). For an organism or population of organisms to be affected even by a spill event, there needs to be an unequivocal link between exposure to the oil and the observed adverse effect. As shown in Figure 1, multiple exposure pathways are possible. The organization of relationships between anthropogenic and naturally occurring stressors within a population or community can be facilitated through the use of a conceptual diagram. The conceptual diagram is meant to: define spatio-temporal scales, inventory resource use activities, describe sources of natural and anthropogenic stressors, describe the mode of action of toxicity, identify ecological values and endpoints that need to be protected and finally determine ecologically significant measures of effect (Gentile et al. 2001).

Figure 1. Conceptual diagram of ecological risk resulting from the Deepwater Horizon well blowout.

i. Resources at Risk

Many biological resources of the GOM were also put at risk in the wake of the DWH blowout. The GOM is a unique ecosystem with increased biodiversity including 15,419 recorded species, of which 10% are endemic to the area (Felder and Earle 2009, Tunnell 2011). The GOM is home to 14 species listed under the Endangered Species Act, Marine Mammal Protection Act or Migratory Bird Treaty Act, with an additional 39 species under the International Union for

Conservation of Nature (IUCN) Red List of Threatened Species. Other large pelagic species of concern in the GOM include the whale shark, bluefin tuna, swordfish, greater amberjack, cobia, king mackerel and blue marlin (Grimes et al. 1990, Gentile et al. 2001, Franks and Brown-

Peterson 2002, Block et al. 2005, Murie and Parkyn 2008). Two separate spawning populations of bluefin tuna migrate to the waters of GOM annually in the summer (Muhling et al. 2012).

ii. Exposure Characterization

“Exposure characterization describes sources of stressors, their distribution in the environment, and their contact or co-occurrence with ecological receptors” (USEPA 1998).

In general three main factors contribute to the exposure profile of an oil spill: the amount and type of oil spilled, environmental conditions at the time of the spill and the kind of environments impacted by the spill (Tunnell 2011). As news spread of the DWH disaster, the media reported the quantity of oil being released from the well head and in to the ocean. Enough oil was released from the wellhead to fill about 12,000 average backyard pools. While this sheer volume of oil released is impressive, it is even more impressive because it was released over a short time period (i.e., as compared to the large amount of oil that is released slowly from the natural seeps) (Tunnell 2011). As oil boiled to the surface 1,500 m above the wellhead, a surface slick 1.6 km in diameter developed (Ryerson et al. 2012) and was then subject to the winds and surface currents for transport toward the estuaries along the coasts of the northern GOM. The range of surface oil was estimated using satellite images. It should be noted that the areal extent of the surface films may be underestimated, because the thinner more dispersed slicks were more difficult to detect (Frias-Torres and Bostater 2011). Using satellite imaging to track the surface oil released, it was estimated that oil at some point covered 100% of the northern GOM whale shark migratory area, 32.8% of bluefin tuna spawning area, and 38% of blue marlin spawning and larval area (Frias-Torres and Bostater 2011). In another study looking at the co- occurrence of surface oil and native range of GOM fauna, Chakrabarty et al. (2012) studied the ranges of 124 fish species including 77 that are endemic to the GOM. It was estimated that 64% of all species examined including 52% of the endemic species had population records in the reported area of the surface oil. One quarter of all the endemic fish species in the GOM were

placed in the highest potential impact category. Species populations that did not overlap with the presence of surface oil were not considered risk free, because there was still the possibility of exposure via subsurface plumes of oil and Corexit (Cooney and Council 1999, Chakrabarty et al.

2012) and other indirect exposure pathways. Also at risk were the many fish populations, including overfished grouper and red snapper (NMFS 2011), that have positively buoyant early life stage (ELS) that were at increased risk of exposure and negative toxic effects at the sea surface.

Figure 2. Satellite image of surface oil in northern Gulf of Mexico after Deepwater Horizon well blowout.

The PAH constituents of oil only have the potential to cause toxic effects if they are bioavailable to the receptor (Singer et al. 2000). When an organism encounters oil it can be through direct physical contact with large oil molecules or through the chemical PAHs that can be in the surrounding aqueous solution. Both physical coating of oil on the outside of the organism leading to mechanical toxicity and chemical toxicity from uptake are possible. Carls et al. (2008) demonstrated that chemical versus mechanical mode of action through PAH exposure was responsible for at least some of the embryotoxicity in fish. Tests in that study compared development of zebrafish embryos exposed to mechanical and chemical toxicity via exposure to

variable concentrations of mechanically dispersed oil and chemical diffusion of PAHs in the agarose matrix. Dose dependent biological effects in both treatments were identical, therefore confirming that toxicity is chemically driven and not mechanically or physically. Organisms within the spatial scope of the spill will most likely be directly exposed via ambient water.

Smaller and younger organisms that spend part or all of their lives as plankton do not have the capability to swim and avoid the oil and therefore are more likely to suffer the most toxicity via increased exposure to oil from the DWH blowout (Saco-Alvarez et al. 2008). In addition, pollutants like PAHs can be concentrated in the upper most millimeters of the water in the sea surface microlayer. Specialized plankton found in this area, termed neuston, therefore will likely suffer the most. The planktonic community is comprised of marine holoplankton like copepods, and the meroplankton is represented by the early life stage (ELS) of finfish and other invertebrates that will mature and most likely settle out of the upper water column in a matter of weeks to months after fertilization. In the wake of DWH blowout the surface waters in the northern GOM (with the exception of the immediate vicinity) could likely be characterized as an acute exposure, while the exposure at depth could be characterized as chronic due to a western moving subsurface plume (Diercks et al. 2010).

Indirect exposure to embryos and larvae of residues offloaded from directly exposed adults is also a potential risk. The accumulation of lipophilic contaminants such as PAHs in an adult female can be alleviated by elimination of those contaminants in the fat soluble, lipophilic yolk of its eggs (Sinderman 2006). This could eventually interfere with ELS development and result in decreased survival of offspring, if mortality does not initially occur. Indirect toxicity can also occur via ingestion of contaminated prey. One example of this possibility could be the feeding of whale sharks on contaminated plankton in the northern GOM (Frias-Torres and

Bostater 2011) at the sea surface where PAHs can be concentrated up to 500 times compared to the underlying water column (Wurl and Obbard 2004). Primary and secondary producers in the plankton community that survived exposure to these elevated PAH concentrations could then serve as a conduit of bioaccumulation in upper trophic levels. If zooplankton were susceptible to oil toxicity, Corexit toxicity or both and only presented sublethal effects, there could be the potential for trophic transfer of PAHs to the GOM whale shark populations. The amount of indirect exposure to PAHs and Corexit would be dependent on the diet composition and abundance of contaminated prey ingested. The whale shark, has a feeding rate of up to 1467 -

2763 g/hour of plankton (assuming a planktonic biomass of 4.5 g/m3 in the water column) (Motta et al. 2010). Regardless of whether organisms were directly or indirectly exposed, environmental conditions (e.g., weather conditions, temperature, pH) and organismal characteristics (e.g, prior exposure, age, gender) can influence the severity of oil and dispersant toxicity and its effects.

The type of oil released greatly influence the solubility and therefore behavior of the oil once it is released. Each oil product has a unique chemical composition of volatile organic compounds like benzene, toluene, ethylbenzene and xylene (BTEX) and heavier molecular weight 2 - 6 polycyclic aromatic hydrocarbons (PAHs) (Faksness et al. 2008, Reddy et al. 2011).

Boehm and Page (2007) categorize petroleum products under five different groupings. Group I products are typically very light gasoline that evaporates quickly. Group V products like asphalt are very heavy and are not water soluble. Most crude oils typically fall within Group III and will partially evaporate and can form emulsions in water. Once the oil is released the weathering process begins and is dependent on temperature, wind and wave action, and salinity. Solubility of BTEX and PAHs is directly correlated to the temperature and salinity. PAH solubility increases with increasing temperature whereas BTEX (as gasses) solubility decreases with

increasing temperature (Faksness et al. 2008, Whitehouse 1984). In general as oil is increasingly weathered the percentage of volatile 2-ring naphthalenes is reduced, while heavier 3-ring constituents are more resistant to degradation by physical processes (Incardona et al. 2004).

Furthermore, with increased wind and wave action more mixing will occur and result in a higher concentration of the water accommodated fraction (WAF) of oil in the water column. Another very important characteristic of exposure is the release location. Oil that is released subsurface at a wellhead for example will result in a greater exposure of oil in the water column via a greater partitioning of PAHs into the water column versus incorporation of PAHs into the underlying water column through wave action and sinking if the release were at the surface in the form of a spill from a ship’s hull (Reddy et al. 2011).

A number of different mitigation strategies were employed following the blowout. One quarter of the oil released was physically collected using booms and or burned (Schrope 2011).

In an attempt to reduce the quantities of oil reaching the shores a chemical dispersant, Corexit, was used to disperse large oil globules and slicks into smaller oil particles that could then be subject to increased abiotic and biotic degradation. Dispersants are used as a surfactant that orients itself at the oil water interface and lowers the interface tension to facilitate the formation of smaller oil micelles (Canevari 1978). Abiotic degradation includes a number of process including photooxidation, physical wind and wave action, dissolution into the water column, and evaporation into the atmosphere among others. (Liu et al. 2012), most of which occurs at the surface. Biodegradation can include microbial degradation or to a lesser extent ingestion and subsequent digestion within organisms. The use of dispersant enhances the amount of oil that physically mixes into the underlying water column, thereby reducing the amount in the surface slick. Dispersal into the water column alleviates some of the risk of a thick slick of oil on the

shoreline and in the estuary; however, increasing solubility increases the exposure and risk to the biota and communities beneath the surface in the water column. The use of dispersants creates a chemically enhanced water accommodated fraction (CEWAF) of oil in the water column.

Dispersant use depends on the sea state, weather, water depth, degree of turbulence, relative abundance and life stage of resident organisms, and other factors that influence an oil spill

(Tunnell 2011). Over one million gallons of Corexit 9527A and 9500A were released at the sea surface, and an additional 770 thousand gallons were released at the well head to control the crude oil (USCG 2011). COREXIT 9500A was later replaced with a less toxic COREXIT 9527, both produced by Nalco (Nalco Company, Sugar Land, TX). The comparison of toxicities for dispersants, in conjunction with CEWAF toxicities, will be discussed further later in the paper.

The MC252 oil contained polycyclic aromatic hydrocarbons (PAHs) and their alkylated homologues along with saturated n-alkanes, with 50% of that being low molecular weight

(LMW) hydrocarbons (methane, C2 - C11) (Ryerson et al. 2012). About 3.9% of the MC252 oil was comprised of PAHs, which equates to 2.1 x 1010 gallons by weight of PAHs released into the northern GOM (Reddy et al. 2011). Petrogenic PAHs, like those released in the DWH disaster, typically consist of a mixture of naphthalenes, fluorenes, phenanthrenes and dibenzothiophenes

(Anderson and Lee 2006). The chemical toxicity of crude oil (i.e., as opposed to mechanical) is thought to be a result of the unique mixtures of PAHs (Whitehead et al. 2012).

The DWH blowout was unique due to the presence of a continuous subsurface plume of oil more than 35 km in length at a depth of approximately 1,100 m that existed for months without evidence of significant degradation (Camilli et al. 2010). Evaporation could play a role in eliminating some of the aromatic hydrocarbons only within the upper 30 m of surface water

(Reddy et al. 2011). While this subsurface plume did not present a direct risk to the species of

concern in this study, there was a risk of mixing within the water column and introduction to the sea surface and/or exposure of adult teleost populations at depth. iii. Effects Characterization

“Ecological effects characterization evaluates stressor- response relationships or evidence that exposure to stressors causes an observed response” (USEPA 1998).

Although crude oil can be comprised of hundreds if not thousands of different organic compounds and some metals, polycyclic aromatic hydrocarbons (PAHs) have long been known to represent some of the most toxic constituents (Westernhagen 1988, Hose and Brown 1998).

The toxic effects of a pollutant are dependent on the exposure pathway, dose, chemical characteristics, environmental conditions and bioavailability, or a combination of these characteristics (Schofield et al. 2007). Exposure of these toxicants to developing ELS can possibly lead to a cascade of consequences (Dethlefsen and Tiews 1985, Barron et al. 2004,

Incardona et al. 2005, Carls et al. 2008). Most often measures of effect on organisms are dose dependent and increase in number and severity as toxicity increases (Carls et al. 1999). Oil toxicity can have adverse effects ranging from slight stress to rapid mortality. As discussed below, these organism-level effects can, if sufficient numbers of individuals are impacted

(particularly individuals with high reproductive value), lead to population level impacts.

Adverse effects of a stressor on an individual organism are largely dependent on the exposure profile (as was characterized previously), its life stage, and previous exposure to the same or similar stressor. While exposure to PAHs or Corexit alone may manifest toxic effects, the combination of crude oil and Corexit exposure may have additive effects with higher toxicity than exposure to either independently (Falk-Petersen et al. 1983, Holdway 2002). Lönning

(1976) reported the toxicity of Corexit and crude oil resulted in severe defects in fertilization and development and that the combined effects were much more detrimental than either alone.

Typically exposure to CEWAF is more toxic than either a dispersant or the oil alone, because of the increased exposure to concentrations of PAHs in the water column due to increasing PAH solubility and the suspension of smaller micelles (Singer et al. 2000, Barron and

Ka'aihue 2003, Ramachandran et al. 2004). Often PAH toxicity, Corexit toxicity or both will manifest in the form of gross abnormalities. Examples of gross physical abnormalities as a result of WAF exposure include skeletal curvature, reduction or malformation of the jaw, microcephaly, multiple types of organ edemas and lesions (Falk-Petersen et al. 1983, Holdway

2002). These gross abnormalities hinder the performance of an already difficult life stage for larval fish. Jaw reductions could lead to an alteration in feeding, starvation, and eventual death.

Effects other than gross abnormalities include decreased sperm viability, reduced ability to fertilize, changes in larval behavior, reduced competencies in food capture and predator avoidance, and increased susceptibility to disease in the early life stage (Sinderman 1994, 2006).

Even subtle, sub-lethal effects like the induction of stress proteins, reduced growth, reduced reproduction, reduced immunity and less severe variations of the effects listed above may eventually impair an organism’s ability to obtain sufficient food or avoid predation. The full impact of sublethal effects may not be evident until later in the organism’s life history; such as the inability of the individual to contribute to the breeding population.

The GOM ecosystem was subject to a variety of natural and anthropogenic stressors prior to the DWH blowout including nutrient loading, expansion of seasonal hypoxic zones, wetland loss, land subsidence, invasive species, climate change, fishing pressures, and effects of hurricane damage (Machlis and McNutt 2010). In the presence of multiple stressors, the natural

variability and fluctuations in abundance, age structure, etc. may increase in amplitude, and the population may suffer from reduced compensatory reserve. What needs to be determined in the wake of an event like the DWH blowout is how severe the negative impacts could be to the equilibrium of the system and its ability to compensate for detriment before the system reaches a tipping point and irreversible damage is done to the ecosystem structure. iv. Assessment Endpoint

“Assessment endpoints are ecological values defined by specific entities and their measurable attributes, providing a framework for measuring stress-response relationships”

(USEPA 1998).

In the wake of an oil well blowout like that of DWH, zooplankton and ELS of many species in the surface and upper layers of the water column were likely the most vulnerable because of the lack of avoidance techniques and sensitivity of early developmental stages

(Sinderman 2006, Tunnell 2011). Early onset toxicity and possible genetic or cell mutations are more detrimental in ELS compared to more developed juveniles or adults because embryos and larvae are relatively smaller, and it is this smaller size that facilitates the distribution of the mutagen rapidly throughout the body (Hose 1994). This characteristic, along with underdeveloped defense mechanisms such as immune responses and avoidance behaviors, is why finfish ELS are considered to be the one of the most sensitive groups when considering ecological effects in the wake of the DWH disaster. If the presence of PAHs, Corexit or both interfere with prey capture or predator avoidance, recruitment success could be affected. The ecological significance of the marine plankton group does not rely solely on individual species; it is the aggregate group that represents a food source and sanctuary for mesoplankton such as finfish ELS that are only temporary tenants in the zooplankton community. While a sharp

decrease in abundance to a population is significant on its own, one also needs to bear in mind that there is some minimum population level where stochastic events could easily lead to dramatic population swings and possible extirpation. Impacts to the plankton as a whole could also lead to cascading effects through altered community structure and food webs (Sinderman

2006). Finfish ELS act as primary consumers and are an essential link in the marine food web.

Oil derived PAHs and persistent organic pollutants (POPs) are characteristically hydrophobic, and as such the constituents of the zooplankton community are often used as a sentinel organisms to track and monitor anthropogenic marine pollution (Carls et al. 2006, Hallanger et al. 2011). It is hypothesized that the strength of a recruitment class is dependent on the magnitude of mortality during the ELS time period (Beck and Turingan 2007). Successful development of finfish ELS is crucial to survival to adulthood. If PAH toxicity interferes with the successful development of sensory organs, behavior, swimming and feeding mechanisms, finfish ELS are not likely to be able to feed and avoid predators past the yolk-stage ELS (Beck and Turingan 2007). Furthermore, the vulnerability of finfish recruitment classes is a function of each individual’s success at avoiding predators and prey capture (Bailey and Houde 1989). In a study by Benfield and Shaw in 2005 looking at the potential vulnerability of pelagic fish assemblages in the GOM to surface oil spills and slicks associated with deep water petroleum development, the fragile ELS of these pelagic finfish was deemed of serious importance. Some reasons they cited included:

“(1) most produce large numbers of small eggs with limited yolk reserves that

hatch into larvae dependent on plankton in the near-surface waters for nutrition;

(2) most fisheries target pelagic fish taxa; (3) oil is buoyant and will accumulate

in the neustonic zone; and (4) based on slicks formed by natural petroleum seeps,

even oil released from near the bottom will likely rise to the surface.

The Gulf of Mexico harbors many valuable species within its basin as previously detailed; however, the scope of this study will focus on the teleost constituents of the zooplankton community because finfish ELS represent an essential link in the Gulf of Mexico ecosystem and economy (Holdway 2002, Jiang et al. 2010). It is because of the importance of finfish ELS in the planktonic community and their sensitivity compared to other organisms in the

GOM, that they were chosen as an assessment endpoint to characterize the effects of DWH oil toxicity. More specifically the ELS of the red drum, Sciaenops ocellatus, were used to measure the effect of oil toxicity on the development of a finfish embryo to yolk sac larvae.

d. Red drum, Sciaenops ocellatus

Normal development and survival of red drum ELS were chosen as an assessment endpoint within this ecological risk assessment of the DWH blowout because of their ecological importance. Various aspects of its life history and the fishery may increase the vulnerability of red drum to oil from the blowout.

Red drum are highly sought after as a recreational species throughout the Gulf states. In the 1980s they supported a highly prized commercial fishery in the northern Gulf, reaching nearly 17.6 million pounds in 1986 (GMFMC 1987). After intense commercial and recreational pressure during that time the impact on the population was finally realized, and the commercial fishing of red drum was significantly reduced to what is now very limited industry (Davis 1990).

While recreational sport fishing for the species still goes on, it is heavily regulated with set bag and size limits. Recreational red drum harvests between 2008 and 2009 ranged from 11.7 to 15.3

thousand pounds per year in the GOM (NMFS). Currently anglers can keep 1 - 2 fish per day depending on region and the fish has to be between 18 and 27 inches total length.

Red drum can live on average to 40 years, but one at 60 years old has been recorded

(Davis 1990). Males typically reach maturity at 1 - 3 years of age; while females reach maturity at 3 - 6 years of age. Once maturity is reached they typically move to the near shore shelf waters and feed primarily on menhaden, anchovies and benthic crustaceans. Spawning typically occurs in inlets and passes in late summer typically from September to October. During this time males have been shown to court females with a drumming sound; hence the name, red drum. Females on average produce 500,000 eggs annually but can produce up to 3.5 million (Davis 1990). Like most other finfish, red drum are broadcast spawners and exhibit a type III survival curve where millions of offspring are released into the water column but only a relatively small percentage reach maturity (Davis 1990). This type of reproduction strategy includes three major periods

(Sinderman 2006). Period 1 represents the stage of eggs and larval fish, when the majority of density dependent mortality occurs. Period 2 represents the pre-recruit fish when mortality is reduced and the number of surviving fish is beginning to stabilize. The remainder of a fish’s life once it has reached maturity and is part of the fish stock is represented by Period 3 (Sinderman

2006). Density dependent resources, such as food, can only support a certain amount of individuals; therefore as a result of competition for resources, only a small percentage of a given year-class will be able to utilize those resources to grow and survive. As a result there is often a sharp decline in abundance of that year class when resources become limiting.

Red drum larvae have not been documented more than 12 miles from shore. Larval dispersal along the coasts is dependent on currents, larval behavior, vertical distribution, growth and mortality (Fogarty and Botsford 2007). Red drum occurrence ranges from the central GOM

up to Massachusetts in the Atlantic Ocean. At the larval stage they lack scales, pectoral or anal fins, mouth parts or full development of eyes, and rely on the yolk sac for nutrition (Davis 1990).

They also remain in the surface waters while the yolk is being absorbed and as they begin to feed. It is at this point in their life history that they are least likely to tolerate poor water conditions (Davis 1990). The characteristic of positively buoyant embryos and further larval development at the water’s surface presents the possibility for oil exposure, as it too predominately rises to the surface based on its physiochemical properties. Compared to more temperate species, red drum typically develop at temperatures above 20˚C, grow rapidly and have a higher energy demand for metabolic processes (Brightman et al. 1997). For example, red drum eggs reared at 25˚C hatch in 24 hours versus Atlantic cod Gadus morhua and winter flounder Pleuronectes americanus that develop at 4 - 8˚C and spend approximately 30 days as eggs (Hempel 1979). The yolk will provide nourishment for another 2 - 5 days while the mouth is developing; after which the fins and scales will develop and they will be considered a mesocarnivore (Hempel 1979, Davis 1990). This is considered the critical period when the transition from endogenous to exogenous feeding during early ontogeny occurs. Larvae have stage specific preferences for certain types of prey (Beck and Turingan 2007). Growth rates of

ELS could potentially have a 100 fold or greater impact on the variability in survival of a recruitment class (Bailey and Houde 1989). The availability and quality of these food items is essential to their continued development. Finfish ELS could potentially lose the battle of survival if their prey items are compromised by oil toxicity. Red drum ELS will then spend approximately another 20 days in the water column feeding on zooplankton such as copepods and amphipods. Soon thereafter they will settle out of the water column and become demersal.

Those finfish ELS that develop more quickly will have first access to newly available resources

like food.

During the ELS of red drum they still need to be wary of predators. Predation on fish

eggs and larvae as an ecological process is important to the health and dynamics of the

subsequent population. Predators of red drum ELS include gelatinous zooplankton, cyclopoid

copepods, chaetognaths, euphausids and parasitic amphipods (Bailey and Houde 1989). Unlike

the earlier stages of development for red drum these predators already have well developed

senses to detect prey, such as visual, mechano- or chemoreception and physical contact

responses (Bailey and Houde 1989). Although the odds seem to be against red drum ELS, they

do have some advantages in the fight to stay alive. Fish eggs are large relative to the prey utilized

by most planktonic invertebrates and have a comparably resilient chorion that makes the grasp

and capture by small invertebrate predators difficult. Also, once motile, the larva may evade

attack if it detects the predator first. Avoiding predation would be dependent, however, first on

successful development of sensory organs and, second on normal morphological development

and swimming speed.

II. ANALYSISPHASE OF RISK ASSESMENT

As part of a larger investigation exploring possible lingering effects of the oil on plankton

and neuston of the Gulf, this study was undertaken to begin to assess the risk that DWH oil

posed to ecologically important teleost species of the GOM. Specifically, embryo-larval toxicity

tests were conducted using red drum ELS in the presence of MC252 CEWAF. Although

development and survival of red drum ELS was chosen as an assessment endpoint in this study,

other pelagic GOM species have similar reproductive strategies, so the red drum can be

considered a model organism for these other species. When studying planktonic assemblages

Richards et al. (1993) suggests that, “… they can be studied as a unit whose individual taxa respond similarly to the environment without necessarily invoking emergent community properties”. The effects observed from the toxicity of CEWAF should draw parallels to what would be observed in other teleost ELS of the GOM that may have also been affected by DWH.

a. CEWAF Preparation and Test Conditions

CEWAF was prepared using protocols established by the Chemical Response to Oil Spills:

Ecological Effects Research Forum (CROSERF) (Aurand and Coelho 2005). The crude oil,

Source Oil B (A0031B) obtained from BP’s consultant, AECOM (Fort Collins, CO), was first artificially weathered, by heating to 90°C with stirring to speed volatilization, until the weight of oil was reduced by 33% +/-1.1% (cf. Incardona et al. 2014). CEWAF was created in a 2 L glass aspirator bottle using Corexit and the artificially weathered oil that was added at approximately

1:10 ratio by weight to 30 ppt artificial seawater that was spinning with a 25% vortex. Actual loading rates were 1849.07 ± 52 mg/L oil and 238 ± 29 mg/L Corexit (n=3). This mixture continued to spin for 18 hours and then was allowed to settle for 6 hours. At that time the

CEWAF was pulled from the bottom of the bottle so as not to disturb or collect the layer of oil at the surface. Oil is a complex mixture of many compounds with different solubilities. Variation in source material, dispersant used and even minor differences in the protocols used in the preparation of the CEWAF can result in highly variable PAH mixtures, and therefore variable toxicity since different PAHs have differing toxicity. Accordingly, nominal concentrations or oil loadings do not accurately represent the aqueous exposure media and the measurement of resulting concentrations of individual PAHs in the CEWAF is essential. Concentrations were calculated based on the average solubility of 0.29% (± 0.09%) that was measured in 4 identically produced CEWAFs. Individual PAHs were analytically determined in the CEWAF through a

collaborated effort with researchers at France’s Centre of Documentation, Research and

Experimentation on Accidental Water Pollution (CEDRE) using the Stir Bar Sorptive Extraction

(SBSE or Twister™) technology and thermal desorption GC-MS (Roy et al. 2005). Stock

CEWAF was then serially diluted to create 5 different treatment concentrations. Although

CROSERF strongly recommends using variable loading rather than serial dilution to prepare

WAF exposure concentrations, for CEWAF, “the Committee concluded that they could not recommend one method over the other” (Aurand and Coelho 2005, page 97). Therefore the use of serial dilutions for this study was considered acceptable.

For comparative purposes, the total amount of toxic units in the resulting CEWAF were also calculated as the sum of the fractional contribution of each individual PAH analyte in the mixture multiplied by its specific acute potency divisor (Langheinrich et al. 2003). The acute potency divisor is a number unique to each PAH analyte that is based on the amount of each analyte that can cause an adverse effect (Stene and Lønning 1984).

Red drum embryos used in repeated exposure tests were obtained from the Florida Fish and Wildlife Conservation Commission (FFWCC) Aquaculture Lab (Port Manatee, FL) from four different spawning events. These repeated tests were done to assess possible differences in

ELS sensitivity due to genetic makeup of spawning brood stock and to assess repeatability of

CEWAF preparation and artificial aging of oil. During one exposure a problem occurred while the CEWAF was being prepared, and the results were deemed invalid. Thus, the present study reports the results from three toxicity tests with overlapping ∑PAH concentrations.

Red drum embryos were exposed, within six hours of fertilization, as follows: 15 embryos were placed in each beaker containing 50 mL of either control water or one of five different CEWAF concentrations. Control and dilution water for CEWAFs was aged using

artificial seawater made from Instant Ocean Artificial Salt Mix and de-ionized water. Each treatment and control was carried out in triplicate. In exposure 1, embryos were exposed to

CEWAF concentrations of 0.25, 0.49, 0.99, 1.97 and 3.95 µg/L. Based on the results of exposure

1, CEWAF concentrations were increased slightly in exposures 2 and 3 and were 0.34, 0.69,

1.38, 2.75 and 5.5 µg/L. Beakers were placed in a temperature controlled water bath at 25˚C with a 12 hr photoperiod with fluorescent and UV overhead lighting (315-800 µWcm-2 UVA, 12-28

µWcm-2 UVB) for 24 hours.

b. Response Endpoints and Statistical Analyses

As Au (2004) outlines, the following criteria should be considered when choosing a biomarker of effect: ecological relevance, sensitivity, specificity and dose response relationship.

When designing these toxicity tests confounding factors, technical difficulties and cost effectiveness also need to be taken in to account (Au 2004). Characterization of survival and abnormalities in development were considered to be a suitable and important measure of effect for acute toxicity tests (Carls and Meador 2009, de Soysa et al. 2012). These tests have the potential to distinguish dose response relationships, are carried out with relative ease and are cost effective (Au 2004). The use of static toxicity tests and quantification of gross abnormalities was chosen as a sufficient measure of effect to characterize the toxicity of the DWH oil and the dispersant Corexit.

Effects can be assessed based on three dimensions: 1) the proportion of the exposed population affected (i.e., a quantal variable), 2) the severity of an effect (i.e., often a measurement of a continuous variable such as growth) and, 3) the type of effect, which might begin with behavioral changes and progress to reduced growth and finally death as toxicant concentration or exposure duration increases (Suter 2007). Several assessment endpoints were

utilized in the present study. After 24 hours of exposure, each larva was visually examined and categorized as live or dead (when they did not respond to repeated physical stimulation; i.e., prodding with transfer pipet) and as viable hatch or non-viable hatch. Viable hatched red drum possessed normal pigmentation, exhibited no spinal curvature and showed apparently even bilateral symmetry. Non-viable individuals were seen to have obvious severe abnormalities, had not hatched or were dead. Individuals from each treatment level were preserved in 5% buffered formalin and stored in glass vials for further morphological analyses (i.e., continuous variable to assess severity of sublethal effect).

Quantal data from the determination of live/dead and viable/non-viable hatch was then used to determine median lethal concentration (LC50) and median effective concentration

(EC50) using the Trimmed Spearman-Karber (TSK) method using computer software obtained from the U.S. Environmental Protection Agency. The TSK method was used as an alternative to the Probit or Logit analysis, because TSK is better suited for use in extended series of bioassay or toxicity experiments and it can better handle anomalous data (Hamilton et al. 1977). Because oil is a complex mixture that varies in toxicity depending on the relative amounts of individual

PAHs, a toxic unit approach was also undertaken to assess the expected toxicity from this unique mixture of PAHs (Barron et al. 2004). A toxic unit value was assigned to each analyte of the composition that made up the CEWAF to determine which if any particular analyte played a more significant role in the toxicity to the red drum ELS development.

Preserved larvae were later microscopically examined using a dissecting stereoscope to assess various morphometric characteristics. The developmental stage of the larvae from this study limited the amount of physical landmarks that could be used as metrics when determining the stress response of red drum embryos to CEWAF within 24 hr of exposure. To avoid

uncertainties inherent in feeding studies (e.g., quantity and quality), the present study was an acute exposure while larvae were lecithotrophic. The acute exposure duration limited the degree of development of more complex structures for analysis. A graduated severity index (GSI) supplemented by limited length metrics was considered to be the best approach to characterize the morphology of exposed larvae under this set of conditions and available possible metrics for analysis.

The GSI used for this experiment was adapted from Middaugh et al. 1988, Hose et al. 1996 and Carls et al. 2000. Developmental stage was scored as follows 0 = yolk-sac larvae, 10 = late embryo, early embryo, morula/blastula. The developmental stage was scored as such to account for environmental death considering that larvae that had not reached the appropriate developmental stage at the 24 hr mark would not survive to adulthood and be an active participant in the recruitment class. Any ELS that had a developmental stage score of 10 were not able to be scored any further. Skeletal, finfold, cardiovascular and yolk development were scored as follows: 0 = normal development, 1 = slight defect in size or structure, 2 = moderate or multiple slight defects, 3 = severe defect or multiple moderate defects. Skeletal abnormalities were characterized by the presence and severity of contortion of the notochord from its primary axis, resulting in an L – or S – shaped curvature and even a moderate corkscrew effect of the trunk segment at times. The GSI scores of each individual within each treatment were then used to create a mean rank score (MRS) for each morphological characteristic for each treatment level. The MRS for each treatment of each characteristic was calculated as shown below. After these scores were plotted for each characteristic according to each treatment level the resulting curves were then compared.

Where:

= Score (0, 1, 2 or 3)

= the number of times the score, , was recorded N = the number of larvae within that treatment

This suite of characteristics was scored to reflect the probability that a larvae with improper development of the notochord axis, finfold, cardiovascular system and/or yolk sac will result in greater potential for mortality. A score closer to 0 can be interpreted as being closer to normal development in 24 hr, and will have a greater chance of surviving to adulthood and making a successful contribution that year’s recruitment. This analysis does not aim to suggest that proper morphological development results in successful reproduction, only that such an individual should have the potential to do so. An individual with a developmental stage score of 10 or an increased MRS would indicate that it is unlikely that the larva would survive to adulthood and would experience environmental death by predation or starvation.

Length metrics including total length, ratio of the length from the tip of the snout to the vent and the length from the vent to the tip of the tail (SV: VT) and the ratio of the yolk sac length to diameter provided the most ease to obtain during analysis and the most valuable numeric measurements. As stated previously, the development of more complex characteristics for analysis were not yet reached by 24 hr. For example, Beck and Turingam 2007 used red drum larvae that were between 3 and 30 days old when looking at the development of the mandible.

When examining more developed 8 day old herring larvae using the methods of a GSI in addition to length metrics, Carls et al. (2000) even noted, “…curved larvae could not be

accurately measured using the micrometer…”. It should be noted that they assumed skeletal curvature to be an artifact of preservation, and they had a large enough sample size of larvae that were not deformed to analyze. Skeletal curvature cannot be attributed to preservation artifact in this study considering the severity and high occurrence of gross skeletal abnormalities in multiple exposures across all treatment levels. Measurement therefore proved more difficult than anticipated due to the bending and angle of the larvae being examined. For example, the goal of a 1-D measurement of standard length was complicated by the 3-D nature of the larva. As a result the estimated perceived length was recorded for comparison. Lengths were recorded in a fashion similar to what is shown in Figure 3. The total length was calculated from the sum of the length of the fore section from snout to vent and the length from vent to tip of the tail (Figure 3

A B). Some larvae from the higher treatments only yielded a total length (measured from the tip of the snout to the tip of the tail) because all other landmarks used for the measurement of the yolk and SV:VT were not present or were poorly defined. The yolks of hatched larvae were measured for length and diameter (Figure 3 C D). A ratio of yolk length to diameter was then used to compare the relative shape of the yolk with respect to concentration. Only the data from exposures 2 and 3 were analyzed with respect to length metrics. Length metrics were compared among treatment levels using either the ANOVA or Kruskal Wallis analysis.

C B

Figure 3. Image of red drum larvae showing various measurements taken to assess normal growth: (A) The length from the tip of the snout to the vent. (B) The length from the vent to the tip of the tail/notochord. (C) The length of the yolk. (D) The diameter of the yolk.

III. RESULTS

a. CEWAF Characterization

The ∑PAH (sum of 42 PAH analytes) in 4 identically produced CEWAF stock solutions,

made from Source Oil B and Corexit 9500, was 5.88 ± 1.02 mg/L (mean ± 1 SD). Of the 42

analytes, naphthalenes (bar with dark blue fill) comprised 91.2%, of which C3-napthalene made

up 55% of the total naphthalenes, fluorenes (light blue fill) comprised 4.2%, and the phen/antra

(green fill) group made up 3.1% (Figure 4). The remaining 31 common PAH constituents of

crude oil analyzed contributed less than 1% to the total PAH concentration.

Napthalene C1-Napthalene C2-Napthalene C3-Napthalene

0 1000 2000 3000 4000 5000

Biphenyl Fluorene C1-Fluorene C2-Fluorene C3-Fluorene Anthracene Phenanthrene C1-Phenan/anthra C2-Phenan/anthra C3-Phenan/anthra Dibenzothiophene C1-Dibenzothiophene C2-Dibenzothiophene C3-Dibenzothiophene Pyrene Fluoranthene C1-Fluoranthenes / Pyrenes C2-Fluoranthenes / Pyrenes Benzo[a]anthracene Chrysene C1-Chrysenes C2-Chrysenes Benzo[e]pyrene

0 50 100 150 200 250 PAH Concentration (ppb, g/L)

Figure 4. Average relative concentration and composition of major PAH constituents in 4 CEWAF stock preparations. Note two different scales.

Because each of these different PAHs is known to have a different toxicity (Barron et al.

2013, Incardona et al. 2014), the toxic unit approach similar to Barron et al. (2004) was used to estimate the toxicity of the stock CEWAF solution made from MC252 Source B oil and Corexit

9500. Using this approach, the summed toxicity of the unique combination of these PAHs would equate to 98.7 ± 28.9 TU/L (n=4) in stock CEWAF solution made from MC252 Source Oil B and Corexit 9500. Naphthalenes (dark blue fill) comprised 78.4% of the toxic units of all the

PAH analytes present in the CEWAF stock solution, of which C3-napthalene made up 81.4% of the total naphthalene TUs, fluorenes (light blue fill) comprised 8.9%, and the phen/antra group

(green fill) made up 8.2%, chrysene (purple) comprised 2.5%, dibenzothiophene (yellow) 1.3%

(Figure 5). The remaining 23 common constituents of PAHs analyzed contributed less than 1% to the total TU.

Napthalene C1-Napthalene C2-Napthalene C3-Napthalene

0 20 40 60 80 100 Biphenyl Fluorene C1-Fluorene C2-Fluorene C3-Fluorene Anthracene Phenanthrene C1-Phenan/anthra C2-Phenan/anthra C3-Phenan/anthra Dibenzothiophene C1-Dibenzothiophene C2-Dibenzothiophene C3-Dibenzothiophene Pyrene Fluoranthene C1-Fluoranthenes / Pyrenes C2-Fluoranthenes / Pyrenes Benzo[a]anthracene Chrysene C1-Chrysenes C2-Chrysenes Benzo[e]pyrene

0 2 4 6 8 10 Toxic Units (TU) Figure 5. Estimated average toxic units (TU) from 4 CEWAF stock solutions based on unique acute potency divisor (APD) and concentration of individual PAHs. Note two different scales.

b. Developmental Abnormalities Observed

Figure 6 E shows a red drum larvae from the control after 24 hr. The notochord extends from the base of the head outward in a relatively straight line. While this was not a metric that was analyzed, it was qualitatively noticed that the yolk seemed to set further forward of the vent position than in any other treatments. In most of the other treatments of CEWAF the vent was almost immediately adjacent to the yolk. A difference between yolk positions relative to the vent can be observed in the comparison between D and E in Figure 6. Finfold abnormalities were

most noticeable on the dorsal finfold behind the head and above the anterior portion of the notochord, and at the developing caudal fin (see gray arrow heads in Figure 6 C D). Finfold abnormalities were characterized by what seemed to be the deterioration or stunted development of a smooth finfold edge, or there was an aggregated mass of cells typically at the end of the developing caudal fin. In addition, it was observed that many of the hatched larvae from the 1.97 mg/L and 2.75 mg/L treatments exhibited severe necrosis, truncated caudal sections and undeveloped vents (Figure 6-B). It was common to observe severe bends or deviations from a straight line in most larvae from any of the CEWAF treatments (see black arrowhead Figure 6

C). Edema was often observed associated with the yolk sac or the cardiac sinuses (Figure 6 C E; black arrows). Figure 6 A shows a very late stage embryo from the 5.5 mg/L exposure that did not hatch from the chorion within 24 hr. Many of the other larvae at this treatment level were less developed embryos with the yolk and notochord not discernible within the chorion.

Figure 6. Developmental abnormalities observed as a result of 24 hr acute exposure to MC252 CEWAF.

c. Graduated Severity Index (GSI)

Overall, the mean rank score (MRS) of the graduated severity index (GSI) for the four major characteristics examined increased as the exposure concentration increased (Figure 7). In general a higher MRS represents the increased occurrence of larvae categorized with a higher

GSI for that respective characteristic. Cardiac and yolk sac development had higher MRSs at lower exposure concentrations and increased more rapidly with increasing concentration than for other characteristics, suggesting cardiac and yolk development was most sensitive to CEWAF toxicity. Regression analysis on each exposure with respect to the MRS of cardiac development

concluded that the slope of the representative line for each exposure was significantly different than zero (Table 2). Regression analysis on each exposure with respect to the MRS of yolk sac development concluded that only the slope of the representative line for Exposure 1 was statistically significantly different than zero (Table 2). It is also noteworthy that MRS for cardiac development and yolk sac exhibited good repeatability among the three exposures. In contrast, embryos in Exposure 1 exhibited lower MRS values than the other two exposures for finfold and skeletal development for comparable concentrations (Figure 7). Regression analysis on each exposure with respect to the MRS of skeletal development concluded that only the slope of the representative line for Exposure 3 was significantly different than zero (Table 2). Regression analysis on each exposure with respect to the MRS of finfold development concluded that each slope of the representative line was significantly different than zero (Table 2). This variability may imply that red drum ELS from Exposure 1 could better tolerate the CEWAF toxicity with respect to skeletal and finfold abnormalities, and indicate that skeletal and finfold development were less sensitive than cardiac and yolk development to CEWAF toxicity.

3.5 3.5 Cardiac Skeletal

3.0 3.0

2.5 2.5

2.0 2.0

1.5 1.5

Mean Rank Score Rank Mean 1.0 Score Rank Mean 1.0

Exposure 1 Exposure 1 Exposure 2 0.5 0.5 Exposure 2 Exposure 3 Exposure 3

0.0 0.0 0.0 0.2 0.4 0.6 0.8 1.0 1.2 1.4 1.6 1.8 0.0 0.2 0.4 0.6 0.8 1.0 1.2 1.4 1.6 1.8

Log + 1 Concentration Log + 1 Concentration Regression E1; R2 = 0.77 ; p = 0.0211 Regression E3; R2 = 0.96 ; p = 0.0036 Regression E2; R2 = 0.89 ; p = 0.0156 Regression E3; R2 = 0.90 ; p = 0.0145

3.5 3.5 Yolk Finfold 3.0 3.0

2.5 2.5

2.0 2.0

1.5 1.5

Mean Rank Score Rank Mean Mean Rank Score Rank Mean 1.0 1.0 Exposure 1 Exposure 1 0.5 Exposure 2 0.5 Exposure 2 Exposure 3 Exposure 3

0.0 0.0 0.0 0.2 0.4 0.6 0.8 1.0 1.2 1.4 1.6 1.8 0.0 0.2 0.4 0.6 0.8 1.0 1.2 1.4 1.6 1.8

Log + 1 Concentration Log + 1 Concentration Regression E1; R2 = 0.80 ; p = 0.0166 Regression E1; R2 = 0.72 ; p = 0.0331 Regression E2; R2 = 0.77 ; p = 0.0489 Regression E3; R2 = 0.81 ; p = 0.0390

Figure 7. The relation between exposure concentration and mean rank score of GSI for cardiac, skeletal, yolk sac and finfold characterization. Regressions found to be statistically significant for each exposure respective of characteristic are shown beneath each graph. Data was transformed using Log (x + 1) for normalization purposes and to facilitate graphic presentation.

d. Length Metrics

Total length of hatched larvae generally decreased as exposure concentration increased

(Figure 8); however this relationship was not monotonic and there was no statistical significance with respect to a linear regression of the data. Total length of larvae differed among treatments from Exposure 2 and 3 (ANOVA, df = 4, F = 60.13, p < 0.0001), with all treatments being statistically different from the control (Dunnett’s, p ≤ 0.0013).

2.6 Total Length 2.4

2.2

2.0

1.8

1.6

Length (mm) Length 1.4

1.2 Exposure 1 Exposure 2 1.0 Exposure 3

0.8 0 1 2 3 4 5

PAH (mg/L) Figure 8. Total length of post hatch yolk sac larvae as a function of exposure concentration.

Visual inspection suggests a weak positive relationship between the ratio of the SV length to VT length in the red drum larvae and exposure concentration (Figure 9). An increased SV: VT ratio indicates the shortening of the trunk or tail section of the larvae relative to the length of the fore section surrounding the vital organs and yolk. The SV: VT ratio differed significantly among treatment levels (Kruskal-Wallis test, df =4, p <0.001). Nonparametric post hoc comparisons to the control revealed that only larvae from the 0.69 mg/L treatment level did not

statistically differ from controls (Steel Method, p<0.0001). This exception and lack of statistical difference was likely a result of low sample size of larvae within that treatment (n = 6).

Regression analysis on each exposure with respect to the total length concluded that only the slope of the representative line for Exposure 2 was statistically significantly different than zero

(Table 2).

1.6 SV:VT 1.4

1.2

1.0

0.8

0.6

SV:VT Length Ratio Length SV:VT 0.4 Exposure 1 Exposure 2 0.2 Exposure 3

0.0 0 1 2 3 4 5

PAH (mg/L)

Figure 9. Ratio of the length of snout to vent to the length from vent to tip of the tail/notochord (SV:VT) versus exposure concentration. Regression E2; R2 = 0.82; p = 0.0335

In general, the length to diameter ratio of the larval yolk sac decreased with increasing exposure concentration (Figure 10). The mean yolk length to diameter ratio of larvae from exposures 2 and 3 differed among treatments (ANOVA, df =4, F =6.86, p < 0.0001). Post hoc comparisons of all treatments revealed that the mean yolk length to diameter ratio for all treatments was statistically different from the control (Tukey-Kramer HSD, p ≤ 0.0275).

Regression analysis on each exposure with respect to the total length concluded that only the slope of the representative line for Exposure 1 was statistically significantly different than zero.

2.0 Yolk Length:Diameter

1.5

1.0

0.5 Exposure 1

Yolk Length: Diameter Ratio Diameter Length: Yolk Exposure 2 Exposure 3

0.0 0 1 2 3 4 5

PAH (mg/L)

Figure 10. Ratio of yolk length to diameter versus exposure concentration. Regression E2; R2 = 0.66; p = 0.0481

e. EC50

None of the embryos incubated at the two highest concentrations in the repeated exposures, 3.95 mg/L and 5.5 mg/L, were considered viable (they had either not hatched, manifested severe spinal curvature, exhibited abnormal jaw formation, or were presumed dead when they did not respond to repeated physical stimulation). More specifically, all of the embryos in the highest concentration 5.5 mg/L were considered dead. Consequently, sublethal metrics (e.g., length measurements, GSI characterization) were not assessed for the 5.5 mg/L treatment. The median concentration at which 50% of the red drum larvae experienced abnormalities (EC50) and were considered non-viable after 24 hour exposure to CEWAF ranged from 0.38 mg/L to 1.63 mg/L (n=2). The EC50 could not be calculated for exposure 3, because

all treatment levels resulted in percentages of non-viable larvae that were greater than 60%.

Therefore, without having representative data in the lower range of toxicity, the EC50 could not be calculated. Regression analysis on each exposure with respect to the EC50 concluded that the slopes of the representative lines for Exposures 1, 2 and 3 were significantly different than zero.

100

Percent Non-Viable (%) Non-Viable Percent

20 Exposure 1 Exposure 2 Exposure 3 0 0.0 0.2 0.4 0.6 0.8 1.0 1.2 1.4 1.6 1.8 2.0

Log + 1 Concentration

Figure 11. Viability of red drum embryos exposed to increasing concentrations of MC252 CEWAF. Data was transformed by Log (x + 1) for normalization purposes and to facilitate graphic presentation. Regression E1: R2 = 0.76, p = 0.0238; Regression E2: R2 = 0.79, p = 0.0178 Regression E3: R2 = 0.83, p = 0.0110.

f. LC50

Considerable variation was observed in mortality among the three repeated exposures with very little mortality occurring at any of the CEWAF concentrations during Exposure 1 (Figure.

12). Thus, this endpoint exhibited poor repeatability. While these embryos had not died in 24 hr, as presented earlier, those in higher concentrations failed to hatch or developed severe abnormalities and were scored non-viable (Figure 11). Increasing the concentrations of the

CEWAF slightly in the higher treatment levels of Exposures 2 and 3 resulted in mortality (Figure

12). Because a review of the laboratory records found no clear reason to invalidate Exposure 1,

it was retained for EC50 and sublethal metrics. The 24-hr median lethal concentration (LC50)

calculated from Exposures 2 and 3 ranged from 0.48 mg/L to 2.43 mg/L; however, given

survivability in Exposure 1, the LC50 remains uncertain. Regression analysis on each exposure

with respect to the LC50 concluded that the slope of the representative line for Exposures 2 and

3 were significantly different than zero (Table 2).

100

Exposure 1 Exposure 3 80 Exposure 3

Percent Mortality (%) Mortality Percent

0 0.0 0.2 0.4 0.6 0.8 1.0 1.2 1.4 1.6 1.8 2.0

Log + 1 Concentration Figure 12. Mortality of red drum embryo-larval stages exposed to increasing concentrations of

MC252 CEWAF. Data was transformed by Log (x + 1) for normalization purposes and to facilitate 2 2 graphic presentation. Regression E2: R = 0.89, p = 0.0039; Regression E3: R = 0.72, p = 0.0323

IV. DISCUSSION

The objectives of this study were threefold: (1) to characterize the toxicity of CEWAF of

MC252 crude oil to red drum ELS, (2) determine how the sensitivity of this species compares to

other finfish ELS; and (3) to determine if these laboratory results might be used to predict effects from in situ exposure to other commercially, recreationally and ecologically important finfish in the wake of the DWH blowout in the GOM ecosystem. Each objective will be considered in turn in the following sections.

a. Normal Finfish Embryo-Larval Development

Embryonic development is a complex series of events that must occur in the proper sequence for it to proceed normally. If something disrupts that sequence, even slightly, it can result in gross deformations, decreased sensory function or suboptimal metabolism often leading to death before reaching the end of larval development (Hempel 1979, Falk-Petersen 2005). In normal finfish development the gastrulation process begins within a few hours of fertilization, the blastoderm begins to extend around the yolk, and the thickened ridge of cells along the yolk becomes the embryonic axis. Once the three germ layers, ectoderm, mesoderm and endoderm are established and the gastrulation process is complete, organogenesis may begin. At this point the notochord, neural tube, gut and primordial cone shaped heart are present. Organogenesis includes the further development of the vascular system, sensory – motor reflexive circuits and primordial otoliths, liver, swim bladder, pancreas and gallbladder.

By the time the larvae is ready to hatch from the chorion the segmental, v-shaped myotomes of the muscle and the kidney, excretory organ and primary lymphoid organs have also developed (Padrós and Crespo 1996). When the fish hatches the gills are not yet well developed, therefore all gas exchange is carried out cutaneously and the squamous cells within the epidermis are responsible for the osmotic balance with the surrounding environment. In addition the formation of the eyes and jaw, including mouth parts, has not yet been completed by the time the

larvae hatch. The larvae thereby rely on nourishment from the yolk until it can successfully begin to feed on smaller zooplankton.

b. Modes of Toxicity and PAH Specific Toxicity

Kuhnhold (1970) reported that embryos of cod, Gadus morhua, were more sensitive to the water soluble fractions of crude oils during the gastrula stage than at later stages of embryonic development. In the same study, early larvae were more sensitive to oil exposure than embryos. While the debate continues as to whether the embryos or larvae of finfish are more sensitive to oil exposure, there is a consensus that finfish ELS (stages during embryogenesis and feeding post hatch larvae) are the most sensitive compared to adult life stages (Falk-Petersen and

Kjorsvik 1987, Incardona et al. 2004). Thus, in addition to inter-species differences, sensitivity differs among life stages. This is due, in part, to the fact that the complex multiphase states of seawater, oil and dispersant can have different modes of action (MOA) for toxicity on the developing organs and tissues of both embryos and more developed yolk sac larvae (Rico-

Martínez et al. 2013). Different organs and tissues exhibit substantial variation in their sensitivities to PAH exposure, which is largely dependent on the developmental stage of the target organs (Goodbody-Gringley et al. 2013). There are at least three recognized embryo-larval toxicity models for PAHs (1) CYP1A induction, (2) the narcosis theory, and, (3) PAH toxicity mediated via cardiac dysfunction (Incardona et al. 2014, Barron et al. 2004, Barron and Ka'aihue

2003, van Wezel and Opperhuizen 1995). Furthermore these MOAs of PAH toxicity may not be mutually exclusive (i.e., the fourth model is the mixture model).

Initially the majority of PAH toxicity was believed to be a function of nonpolar narcosis, where the level of toxicity was largely based on the log octanol-water partition coefficient (log

Kow). More lipophilic compounds were presumed to result in increased toxicity (Holm et al.

2003). For example, according to the narcosis theory it would be expected that chrysene (Kow

5.3) would be more toxic than naphthalene (Kow 3.3) and phenanthrene (Kow 4.5). Contrary to the narcosis theory, Incardona et al. (2004) reported that phenanthrene caused the most toxicity to zebrafish larvae, while those exposed to naphthalene appeared to develop normally. Therefore,

3- and 4-ring PAH analytes are thought to function within a specific and unique mode of action different than that of the narcosis theory (Incardona et al. 2004).

The aryl hydrocarbon receptor (AhR) pathway is a ligand activated basic transcription factor that controls the expression of a group of genes encoding enzymes that have the capability to convert PAHs to water soluble derivatives that can then be excreted. Secondary metabolites of the PAH introduced into the system and processed by the enzymes of the cytochrome P450

(CYP450) superfamily are thought to cause the majority of the negative effects. More toxic responses are more common and severe when the pollutant introduced is more resistant to metabolism by CYP enzymes and the AhR pathway is continually activated: for example pyrene, dioxins, poly chlorinated biphenyls (PCBs) and other high molecular weight PAHs. This continued activation and CYP catalytic activity then leads to oxidative stress and cellular death that presents as acute toxicity in response to the presence of PAHs. Activation of AhR induces the production of CYP450 enzymes. The resulting inappropriate overabundance of CYP450 can lead to the conversion of naturally occurring levels of testosterone to estradiol at inappropriate times in an organism’s life cycle, thereby resulting in feminization of males. Overabundance of

CYP450 can also induce a carcinogenic intermediate in the conversion of benzo[a]pyrene

(B[a]P) to benzo[a]pyrene diol epoxide (BDPE) that can then bind to sulfur rich DNA and create accumulated DNA adducts in the form of cancer evident in the tissues (Incardona et al. 2005).

The PAH analysis in the present study suggested naphthalenes were a major component of the MC252 source oil and was in good agreement with others analyses of MC252 source oil

(Incardona et al. 2014, de Soysa et al. 2012). Naphthalenes were once thought to be a predominant source of PAH toxicity as part of the narcosis theory (Sharp et al. 1979, Coelho et al. 2013); however, more recent findings indicate they are not a major contributing factor to finfish ELS toxicity as compared to the three ring PAHs (Carls et al. 1999, Heintz et al. 1999,

Incardona et al. 2004, Incardona et al. 2013). Hatching rates and growth (measured by length and weight) in fathead minnows, Pimephales promelas, were not affected until concentrations of naphthalenes reached 0.85 mg/L (DeGraeve et al. 1982). Significant mortality of the minnow did not occur until concentrations of naphthalenes reached 4.38 mg/L (DeGraeve et al. 1982).

Similarly data from Anderson et al. (1977) on the embryonic exposure of WAF to the Gulf killifish, Fundulus similis, suggested that a 5 mg/L concentration of naphthalene is the upper limit of survival. The question of naphthalene toxicity is also crucial when interpreting the effect that weathering, more precisely loss of volatile components such as the 2-ring PAHs, has on oil toxicity. This is an issue central to a debate that has arisen in the literature recently regarding a report of mortality of pink salmon (Oncorhynchus gorbuscha) embryos in very weathered oil at the extremely low total PAH concentration of 18 µg/L (Heintz et al. 1999, Page et al. 2012,

Heintz et al. 2012). In explaining the differences observed in toxicity thresholds, Heintz et al.

(2012) argue that “including lower molecular weight PAHs (such as naphthalene) in their dose measures causes Page et al. [1] to inflate their doses.” Clearly, this has implications for the interpretation of toxicity in the present study of artificially weathered oil that continued to have a large fraction consisting of naphthalenes. During toxicity testing on tuna embryo larvae concentrations of naphthalene, fluorene and dibenzothiophene were significantly depleted by the

end of 24 hr exposure with MC252 HEWAF, while about 25% of phenanthrene still existed

(Incardona et al. 2014).

While the mode of toxicity is not yet fully understood for each component of oil, it is widely accepted that the specific mode is dependent on PAH weight and physical structure

(Incardona et al. 2005, Carls et al. 2008). PAH components of the DWH crude oil, as demonstrated by similar toxicity resulting from either CEWAF or HEWAF exposure, may interact with important molecular mechanisms to influence embryogenesis (de Soysa et al. 2012,

Brette et al. 2014, Incardona et al. 2014). Three and a half hours post fertilization (hpf) zebrafish,

Danio rerio, embryos exposed to WAF until 5 hpf resulted in some of the same gross morphological abnormalities (de Soysa et al. 2012) as seen in the present study: dorsal tail curvature and cardiac edema, with the addition of cyst formation, reduced head structures and brain hemorrhages. Studies examining the effects of each individual analyte lead researchers to believe that the majority of PAH toxicity is due to the unique composition of fluorene, dibenzothiophene and phenanthrene, more importantly the percentage of phenanthrene

(Incardona et al. 2004). Pyrene is documented to induce peripheral vascular effects within the circulatory system and neural cell death (Incardona et al. 2004). The aliphatic hydrocarbons are not thought to attribute as much to PAH toxicity. Although larvae did present with delayed or failed inflation of the swim bladder in a study by Incardona et al. (2004), zebrafish larvae exposed throughout embryogenesis to 9.99 mg/L naphthalene, 9.98 mg/L anthracene or 2.0 mg/L chrysene exhibited normal development of physical features. In another study, the effects of naphthalene exposure on zebrafish embryos were considerably less severe than exposure to phenanthrene (Kennedy et al. 2000). Zebrafish exposed to 9.97 mg/L fluorene, 9.95 mg/L dibenzothiophene or 9.98 mg/L phenanthrene all resulted in larvae that exhibited decreased

growth, skeletal abnormalities and the presence of cardiac edema (Incardona et al. 2004).

Dibenzothiophene and phenanthrene induced more severe cardiac edema than fluorene exposure and overall can result in a suite of abnormalities that closely resembles exposure to a mixture of

PAHs like crude oil WAFs (Incardona et al. 2004). Zebrafish larvae exposed to 1.01 mg/L pyrene presented with anemia, reduced cardiac circulatory function and cell death in the brain and trunk sections of the neural tube, but pyrene exposed larvae did not present with grossly noticeable cardiac edema like the 3-ring PAHs did (Incardona et al. 2004). Furthermore, the cardiotoxic potency of PAH exposure to yellowfin and bluefin tuna embryos correlated closely with the concentrations of the 3-ring PAHs more so than with total ∑PAH (Brette et al. 2014).

The alkylated homologs of the parent PAH analytes probably have cardiac specific effects similar to the nonalkylated homologs (Incardona et al. 2005).

i.Cardiac Development

Incardona et al. (2004, 2005) demonstrated that individual ELS exposure to three ring

PAHs such as fluorine, dibenzothiophene and phenathrene, which are typically found in weathered oil, resulted in many malformations in a dose dependent manner. Malformations include abnormal cardiac looping and formation, which can have significant impacts on optimal larval development and function. Cardiac arrhythmias were the earliest observed effect in response to North Alaskan Slope (NAS) crude oil WAF (0.028, 0.28, 0.56 and 1.4 mg/L), which, though slightly lower concentrations, are similar to the treatments used in the present study. Mild pericardial edema and reduced blood flow associated with poor cardiac contractility and bradycardia were evident in embryo-larvae from 33 hr of WAF exposure (Incardona et al. 2005).

Based on the results of their ELS exposures, Incardona et al. (2005) concluded that the mode of toxicity for cardiac dysfunction differed from syndromes that arise later in the larval period, such

as CYP1A induction or narcosis. More recently, de Soysa et al. (2012) speculated that cardiac edema, heart morphogenesis defects and reduced circulatory function could be a result of a disruption in the proper development of cranial neural crest cells. Cranial neural crest cells are stem cells that differentiate into cell types that contribute to the development of pigment cells, peripheral nervous system, head cartilage, endothelial and smooth muscle vasculature and portions of the heart (reviewed in de Soysa et al. 2012). Endothelial vasculature and proper heart development and function have all been reported to be compromised in WAF treated embryos

(de Soysa et al. 2012). Cardiac dysfunction resulted in blood taking nearly 3 times longer to travel the distance of seven somites (which give rise to skeletal muscle, cartilage, tendons, endothelial cells and dermis) in WAF exposed larvae as compared to larvae from the control group (de Soysa et al. 2012). This slowed circulation of blood likely stemmed from decreased contraction ability of the heart. The authors further concluded that the decreased blood flow to the developing somites could impair the integrity of the entire peripheral nervous system and continued growth and development to adulthood (de Soysa et al. 2012). After observing severe cardiac toxicity as a result of exposure to MC252 oil, Brette et al. (2014) hypothesized that the

MOA of PAH composition of the MC252 oil must affect the ion channels involved in the EC coupling which links electrical excitation to the physical contraction of cardiomyocytes. They went on to explain further that the basic function of contraction in cardiomyocytes is highly conserved across all vertebrates, and therefore the MOA of PAH toxicity should be reflected in a similar fashion in other finfish as a result of similar oil exposure.

Subtle sublethal effects on embryonic heartbeat of finfish ELS can cause permanent secondary changes in heart shape and cardiac output (Hicken et al. 2011). The indicator of this cardiac induced syndrome is most often the presence of a pericardial edema or accumulation of

fluid in the cardiac sinus and can result from ∑PAH concentrations from DWH crude oil as low as 1-15 µg/L (Incardona et al. 2014). Strong evidence supports that the presence of severe cardiac edema in the yolk-sac larvae will ultimately result in death before the ELS can reach the feeding stage (Hicken et al. 2011, Incardona et al. 2013, Jung et al. 2013). In general, while the severity of crude oil cardio toxicity may vary, it is still observed across all PAH exposures regardless of oil origin or WAF preparation method (Incardona et al. 2014). While the present study did not attempt to identify mode of toxicity at the tissue level, the observed cardiac and other morphological abnormalities were in good agreement with de Soysa et al. (2012) and

Incardona et al. (2014). It should be noted that while Incardona et al. (2014) followed protocols similar to the present study to artificially weather oil, they created a mechanically dispersed high energy WAF (HEWAF) without the use of chemical dispersants and observed an increase in C2-

C4 naphthalenes and C1-C4 phenanthrenes relative to unweathered fresh source oil that was similar to the present study.

ii. Skeletal Development and Length Metrics

Skeletal abnormalities observed in the present study, which included dorsal or upward curvature of the body axis, were also reported in greater amberjack, bluefin tuna and yellowfin tuna larvae exposed to MC252 HEWAF (Incardona et al. 2014). Skeletal abnormalities are thought to be a toxic effect secondary to cardiac dysfunction likely resulting from a functional defect rather than a deficit in the structural integrity of the skeleton (Incardona et al. 2004).

Incardona et al. (2004) even reported that dorsal curvature could be reversed with PAH depuration as long as the defect was not too severe. Embryos exposed for 48 hrs to a concentration of 0.68 mg/L WSF of Prudhoe Bay crude oil resulted in gross abnormalities of the notochord that reduced, and to some degree prevented, locomotion once the larvae had hatched

(Smith and Cameron 1979). In addition, similar to the chemical composition of the MC252

Source B oil used in this study, the composition of PAHs in Prudhoe Bay WSF contained high amounts of naphthalenes and comparable compositions of phenanthrene (Smith and Cameron

1979, Middaugh et al. 1988).

Previous embryo larval toxicity studies have reported decreased larval length as a result of oil exposure (Smith and Cameron 1979, Carls and Rice 1990, Hatlen et al. 2010, and de Soysa et al. 2012). Similarly, the total length of red drum larvae in the present study appeared to be inversely related to exposure concentration; however, this relationship was non-monotonic and not statistically significant. Red drum larvae from Exposure 1 deviated from the trends in

Exposures 2 and 3 with respect to total length. While the trend for decreasing length is still present, the lengths of larvae from Exposure 1 appeared to be consistently longer than those from similar treatments in Exposures 2 and 3. With respect to the results of the SV: VT for all concentrations less than 1.38 mg/L, the ratio of 1 or less indicates that the position of the vent was no more than half the total length from the snout: meaning the position of the vent was closer to the yolk. This could also indicate that the decrease in total length is occurring as a result of a deficit in the length from the vent to the tip of the tail.

De Soysa et al. (2012) hypothesized that the reduced growth of the tail could indicate reductions in cell abundance from reduced cell proliferation or increased programmed cell death.

Furthermore they demonstrated that exposure of embryos to WAF caused a statistically significant increase in the number of cells undergoing apoptosis along the trunk of larvae compared to that of the controls (de Soysa et al. 2012). PAHs have been documented to up- regulate proteins in juvenile cod that characteristically induce apoptosis (Bohne-Kjersem et al.

2009). The necrosis observed in the present study that may have attributed to decreased growth,

was similar to what was observed in zebrafish embryos after exposure to intermediate fuel oil

(IFO) followed by exposure to sunlight (Hatlen et al. 2010). Carls and Rice (1990) reported that walleye Pollock, Theragra chalcogramma, embryos exposed to water soluble fractions of oil

(WSF) had reduced larval lengths. Smith and Cameron (1979) also observed deficiencies in growth (as indicated by shorter length) in Pacific herring, Clupea harengus pallasi, embryos exposed to water soluble fraction (WSF) of Prudhoe Bay crude oil.

Smaller larvae, even if categorized here as viable hatch, would likely have a significant competitive disadvantage later in life compared to normal-sized conspecifics (Bailey and Houde

1989). Detriment to swimming ability can manifest as a result of decreased length or skeletal defect. Bailey and Houde (1989)state that the burst swimming speed of finfish larvae is a function of length, developmental stage and feeding condition. In addition, the percentage of larvae that successfully escape attack is positively related to length. Finfish larvae with skeletal deformities would most likely be unable to evade predators, because the fast start response that is dependent on the C-shaped contortion of the body axis and rapid acceleration would be compromised (Bailey and Houde 1989). iii. Yolk Sac Development

A decrease in the ratio of yolk length: diameter as CEWAF exposure concentration increased could indicate that the yolk shape was becoming less elliptical and more round with increasing concentration. This characteristic could be a result of significant edema constricting the yolk within the yolk sac or increased absorption of nutrients to meet increased energy demand resulting from metabolic stress in response to CEWAF exposure (Bailey and Houde

1989). Due to the lipophilic nature of most PAHs, the yolk sac could be the initial site of toxicant

uptake and storage. As the nutritional reserves of the yolk sac are utilized and metabolized as the larvae develops the larvae may be continually exposed to PAHs even though, in the wild, the ambient environment may be free of PAHs (Smith and Cameron 1979). Furthermore, depending on the distribution of the metabolites of the PAHs from the yolk sac, enhanced phototoxicity

(e.g., UV radiation of the epithelial cells containing PAH metabolites) may result (Hatlen et al.

2010). iv. Finfold Development

As in other studies of CEWAF toxicity on the development of finish (Hatlen et al. 2010,

Incardona et al. 2005, Incardona et al. 2014), significant finfold abnormalities were observed across all concentrations in the present study. Greater amberjack, bluefin tuna and yellowfin tuna larvae exposed to MC252 HEWAF manifested finfold abnormalities (Incardona et al. 2014) identical to those observed in the present study. Abnormalities where characterized by the lack of actinotrichia or fin ray precursors, reduced growth of the finfold and blisters on the leading edge of the finfolds, particularly noticeable in the front dorsal area and caudal region (Incardona et al.

2014). Likewise, zebrafish embryos exposed to North Alaskan Slope oil, and in another case to

IFO, also presented with finfold defects consisting of irregular edges or blisters involving all fins

(Incardona et al. 2005, Hatlen et al. 2010). Incardona et al. (2014) speculated that the finfold abnormalities may be unique direct effect of the PAHs as opposed to a secondary effect such as development delay because of previously compromised cardiac function. In addition the apparent deterioration of the developing finfold may be photo enhanced via sunlight exposure. Hatlen et al. (2010) accounted a rapid and severe lytic deterioration of the finfold from IFO exposure in combination with subsequent UV exposure. The present study was also carried out under artificial UV light; however, further tests would need to be conducted to determine if the finfold

developed normally and then began to deteriorate or if abnormalities (e.g., blisters) occurred at the leading edge of the developing finfold. The potential effect of a contaminant, such as PAHs, on the epithelium and mucus membranes of finfish can make these tissues more susceptible to microbial infections and continued deterioration of motor skills (Au 2004).

Based on the observations and data from this study, it is believed that the mode of toxicity resulting from MC252 exposure is most likely a mixed model where specific PAHs target specific organs.

c. Need for Standardization in Toxicity Testing

Clearly the response of the red drum ELS to oil exposure was highly variable in the present study both within a given test run or exposure and among the three repeated exposures using eggs from different spawning events. Variability in observed responses, and especially poor inter-laboratory reproducibility and frequent absence of monotypic dose response, has been previously reported for oil toxicity studies (Aurand and Coelho 2005). In part, this variability was thought to arise as a result of variations in protocols and is the basis for standardization for oil toxicity testing (Singer et al. 2000, Aurand and Coelho 2005). Aurand and Coelho (2005) stated that, “…it is not surprising that toxicity test results can be affected by a host of factors having to do with test conditions. Therefore, the standardization of as many test parameters as possible is of paramount importance.”

Aurand and Coelho (2005) further state “The lack of standardization and incomplete documentation on methods has been a serious problem with much of the early research on dispersants and dispersed oil.” When conducting laboratory toxicity tests emphasis should be placed on adherence to the standards suggested by CROSERF to improve analytical chemistry

protocols, media preparation standards, exposure regimes and integrated data sets. This comprehensive document encourages study designs where data can be collected under realistic exposure scenarios to better facilitate an oil spill response decision process when needed and where any proposed deviations from the standard must be considered against the loss of comparability (Aurand and Coelho 2005, Coelho et al. 2013). First, while used in early toxicity testing, nominal concentrations based on loading rates of oil should no longer be tolerated as an accurate description of the concentrations of PAHs in exposure effluent (Hicken et al. 2011).

Further, because of the variability in toxicity of the different PAHs, it is recommended that PAH concentrations in exposure solutions be characterized, at least for 32 of the major analytes (using

GC/MS), which can be summarized as the sum of PAHs (∑PAH) , rather than simply determining the total petroleum hydrocarbons (TPH, using Gas Chromatograph/Flame Ionization

Detection) (Barron and Ka'aihue 2003, Aurand and Coelho 2005, Barron et al. 2013, Coelho et al. 2013).

d. Toxic Units

Again recognizing PAH-specific toxicity and that the resulting composition will always vary due to differences in source oil composition, dissimilarities in solubility in different surfactants and variations in mixing protocols, a number of authors recommend using a toxic unit approach (Di Toro et al. 2007, Hansen et al. 2003, Lee et al. 2001, Redman et al. 2012). This approach is based on the presumption that toxicity of a PAH mixture like crude oil as a whole can best be considered as the sum of the individual analyte toxicities and relies heavily on an acute potency divisor (APD; OSAT 2 2011). One concern about the derivation of these APDs is that they are currently based on the narcosis model of toxicity (OSAT 2 2011) and, thus, may not be representative of the PAH specific toxicity that is actually carried out on finfish ELS. The

derivation of APDs also hinge on a model that relies heavily on the partitioning coefficients of each PAH analyte. The USEPA (Hansen et al. 2003) derived their suggested APD (a.k.a. final acute value, FAV) based on the LC50 concentrations of 77 acute toxicity tests in seawater primarily using individual exposures to acenaphthrene, fluoranthene, naphthalene, phenanthrene and pyrene. Of the 30 saltwater species tested (most commonly annelid worms, mysids, grass shrimp, pink salmon and sheepshead minnows) only one test used the early life stage of a finfish.

In addition, to determine the APDs for individual PAHs that did not have the confirmation of results from a toxicity test, Hansen et al. (2003) relied on the partitioning coefficients of each analyte based on the Log Kow values and aqueous solubilities of each chemical which utilizes the chemical’s structure to estimate these various properties (Bohne-Kjersem et al. 2009). Petersen and Kristensen (1998) reported that Kow values are not suitable to characterize the affinity of

PAH analytes for bioaccumulation in fish tissues based on a difference in the lipid-normalized bioconcentration factor (BCFL) and the octanol water coefficient (Kow) for naphthalene, phenanthrene, pyrene, benzo(a)pyrene and polychlorinated biphenyl (PCB) in the ELS of finfish species. It is for the reasons previously detailed, that it should be emphasized that the use of partitioning coefficients should not be given as much weight when determining the APD for the use of toxic units when studying individual PAH toxicity to finfish ELS. A more reliable way of quantifying the APD would be to establish the baseline toxicities of each of the more commonly analyzed 42 PAH analytes with respect to the most sensitive organisms that may be at risk of exposure (e.g., finfish ELS).The solubility, and therefore degree of bioavailability, of PAH analytes is directly related to temperature and salinity (Faksness et al. 2008, Petersen and

Kristensen 1998). Baseline studies to determine specific APDs could then be refined even further to accommodate for changes in habitat: for example, determining the APD for phenanthrene

exposure to tropical finfish ELS versus exposure to temperate finfish ELS. If APDs were reported in this fashion, in the wake of a disaster such as DWH the field reported concentrations of PAHs could supply a wealth of information about the toxicity of the unique PAH profile to the organisms that are at risk of exposure when a complete suite of toxicity tests is not feasible

(Barron and Ka'aihue 2003). Currently there is not a valid APD for unilateral use in the estimation of PAH toxicity to specific marine biota.

e. EC50 / LC50 Comparison

Clearly, given the natural variability in MOA, PAH specificity in toxicity and solubility, the natural variability in PAH composition of source oil, differences in protocols in artificially weathering oil, preparing WAF or CEWAF and measuring various response endpoints, caution is warranted when making comparisons between studies, especially when results have been simplified to a single number such as EC50 or LC50. Beyond that, natural variability in sensitivities among species must also be taken into account. With these caveats, Table 1 summarizes results from previously reported oil toxicity tests to provide context in assessing

MC252 oil toxicity.

Table 1. Comparison of EC50s and LC50s from different types of oil and exposure methodologies.

EC50/LC50 Species Test Conditions Response Toxicant /Origin Reference (ppm) WAF Bluefin Tuna embryos 24 h static Cardiac Edema MC252 Weathered Source 0.0008 Incardona et al. 2014 (Thunnus thynnus) Yellowfin Tuna embryos 24 h static Cardiac Edema MC252 Weathered Source 0.0023 Incardona et al. 2014 (Thunnus albacares) Greater Amberjack embryos 24 h static Cardiac Edema MC252 Weathered Source 0.0124 Incardona et al. 2014 (Seriola dumerili) Pacific Herring embryos 16 d (checked Weathered Alaska North Slope Survival 53.3 Carls et al. 1999 (Clupea pallasi ) daily) crude oil on gravel Pacific Herring embryos 16 d (checked Spinal Weathered Alaska North Slope 33.5 (LWO) - Carls et al. 1999 (Clupea pallasi ) daily) Abnormality crude oil on gravel 3.6 (MWO) Killifish embryos Survival #2 Fuel Oil (API Reference Oil III) 1.5 Sharp et al. 1979 (Fundulus heteroclitus)

CEWAF Red Drum embryos MC252 Weathered Source & 24 h static Viability 0.38 - 1.63 Present Study (Sciaenops ocellatus) Corexit 9500 Red Drum embryos MC252 Weathered Source & 24 h static Mortality 0.48 - 2.43 Present Study (Sciaenops ocellatus) Corexit 9500 Red Drum embryos Western Gulf of Mexico Oil WAF & 48 h static Survival >100 Fucik et al. 1995 (Sciaenops ocellatus ) Corexit 9527 Red Drum embryos Central Gulf of Mexico Oil WAF & 48 h static Survival >100 Fucik et al. 1995 (Sciaenops ocellatus ) Corexit 9527 Atlantic Herring embryos Blue Sac disease Medium South American Crude 24 h static 8.5 McIntosh et al. 2010 (Clupea harengus ) (BSD) Oil & Corexit 9500 Turbot yolk-sac larvae Fresh Kuwait Crude Oil & Corexit 48 h static Survival 2 Clark et al. 2001 (Scophthalmus maximus) 9527 Turbot yolk-sac larvae Fresh Forties Crude Oil & Corexit 48 h static Survival 0.44 Clark et al. 2001 (Scophthalmus maximus) 9500 Inland Silverside juveniles Fresh Kuwait Crude Oil & Corexit 96 h static Survival 0.55 Clark et al. 2001 (Menidia beryllina ) 9527 Inland Silverside juveniles Weathered Kuwait Crude Oil & 96 h static Survival 1.09 Clark et al. 2001 (Menidia beryllina ) Corexit 9527 Inland Silverside juveniles Fresh Forties Crude Oil & Corexit 96 h static Survival 0.49 Clark et al. 2001 (Menidia beryllina ) 9500 Inland Silverside adults EPA National Contingency Plan 96 h static Survival No. 2 Fuel Oil & Corexit 9500 2.61 (Menidia beryllina ) Schedule Toxicity Summary

The exposure of killifish, Fundulus heteroclitus, embryos to a concentration of 2.1 mg/L

WSF of #2 fuel oil had a profound effect on time of hatching, hatching rate and hatching success

(Sharp et al. 1979). In concurrence with those results, red drum in present study not only delayed hatching but also completely arrested all further development, leading to death at elevated concentrations of 3.95 mg/L and 5.5 mg/L CEWAF. As already discussed, de Soysa et al.

(2012) reported that acute WAF exposure to MC252 crude oil did not arrest embryo development during early cleavage and gastrulation, although it did present multiple gross abnormalities later in embryogenesis. Although the actual exposure concentrations were not reported, based on the description of methods for WAF preparation these exposure concentrations were considered to be less than in the present study. As noted, results in Table 1 must be compared with caution, because only results from toxicity tests that created exposure

CEWAF using artificially weathered MC252 Source B oil and Corexit 9500 under the same conditions (e.g., mixing time, speed of mixing, settling time) can be compared directly without caveats. Despite the complication that arises when trying to compare the results from other toxicity tests, the consensus is that the unique PAH composition of the MC252 crude oil released during the DWH blowout directly negatively affects the development of multiple vital organs in finfish ELS and, depending on the exposure concentrations immediate death.

f. Variability

Despite our best efforts to standardize test conditions, oil toxicity test results can vary for all the reasons discussed above and, in addition, variable light and temperature controls, experience and skill of the laboratory analyst, test organism condition and sensitivity, dilution water quality and, if food is provided, quality of the food. With the exception of food quality (because larvae were not fed), none of these factors can be ruled out as the source of variability in the present

study that resulted in the poor repeatability in certain endpoints, particularly mortality, among the three different exposures. Assessing the influence of the these issues is a problem common to all toxicity testing and is the basis for many high throughput laboratories maintaining control charts using reference toxicants (Cowgill 1986, USEPA 1998). Small sample size within some of the treatments may have also added to the observed variability. Although the number of eggs placed into each exposure chamber was consistent, the number of larvae that could be scored at the end of the test for length metrics and GSI was different for each treatment. This was because only larvae that had hatched (whether classified as viable or nonviable) were able to be scored using these metrics for reasons previously discussed. Many larvae that were considered nonviable after 24 hr were scored relatively low for the GSI, meaning that they did not present as many or as severe morphological abnormalities, while something in their behavior caused them to be categorized as non-viable at the end of exposure (e.g., not responding to physical stimulation or decreased swimming ability). While the reason for this occurrence was not clear, there must have been underlying toxic effects that we were not able to assess with the suite of metrics currently employed to characterize the stress response to 24 hr CEWAF exposure. One possible explanation may have to do with the narcosis theory; which would support a mixed model for the mode of toxicity. On the other hand, the variability in response of the embryos may have also been due, in part, to differences in genetic diversity. The red drum eggs for the three repeated tests came from three different spawning events at the Port Manatee hatchery. The number of individuals (males or females) that participated in each of the spawns or their condition at that time of the spawn remains uncertain. Regardless of its source, the variability and resulting uncertainty cannot be ignored.

g. Competitive Disadvantage and Ecological Significance

The toxic response of finfish ELS to PAH exposure is highly conserved; where the same type of effects are observed across many different taxa (Carls and Meador 2009, Hatlen et al.

2010, Jung et al. 2013, Brette et al. 2014, Incardona et al. 2014). As such, the toxic effects of the exposure of red rum embryos to MC252 source oil can be considered comparable to other tropical finfish of the GOM in the wake of the DWH blowout. Subtle interspecific differences in the level of effect may occur between finfish of different temperate zones, due to the initial size of embryo and development rates at the time of exposure (Incardona et al. 2014). Most, if not all, of the sublethal effects discussed above would likely ultimately lead to decreased individual fitness by altering homeostasis and proper functioning of biological processes like respiration, detoxification, endocrine function, osmoregulation and nutrient absorption. Other sublethal effects of PAH exposure could be negative effects on DNA/RNA including heritable mutations that could increase or decrease genetic diversity in a breeding population. To measure the change in genetic diversity at the population level due to genotoxicity resulting from PAH exposure would be difficult to confidently correlate (White 2002, Au 2004, Coelho et al. 2013).

An individual with sublethal effects that has continued to grow to adulthood but cannot successfully reproduce still competes for limited resources that could be utilized by healthy individuals with the potential to propagate. The size of a finfish year class is not a necessarily a direct function of egg production (Cameron and Berg 1992). Decreases in abundance would be especially detrimental to the health of populations such as finfish ELS that experience periods of naturally high density-dependent mortality at some point in its life history. While initially appearing to have low significance, detriment to essential functions and capabilities such as growth and swimming could eventually be fatal via environmental death; e.g., decreased swimming performance could lead to the inability of the individual to evade a predator. The two

most probable causes of natural death in larval finfish are starvation and predation. These two may also be linked at times considering that a starving finfish larvae may be more susceptible to predation (Bailey and Houde 1989). If exposure to oil toxicity occurs and results in excess mortalities after this natural high density-dependent (regulation) mortality period it would likely have a larger effect on the recruitment than if the excess mortality had occurred prior to the density-dependent mortality (Goodyear 1985; Sinderman 1994, for review of density dependent processes see Rose et al. 2001). Finfish ELS that endure even acute exposure to concentrations above the documented threshold concentrations for developmental abnormalities in the wake of

DWH could experience increased mortality, even though it may be late onset. Pink salmon fry subjected to chronic exposure of oil up to 20 ppb total PAH in the wake of the Exxon Valdez oil spill survived at only half the rate of those not exposed over the next one and a half years

(Peterson et al. 2003). Mortality of incubating pink salmon eggs in oil exposed streams was still evident up to 4 years after the spill (Bue et al. 1998). Increased individual mortality, as a result of direct mortality from PAH toxicity or environmental death as a result of decreased competitive advantage because of sublethal effects, could result in a population level ecosystem response

(Hicken et al. 2011, Incardona et al. 2014).

Reported concentrations of TPH and sum of PAHs vary greatly in the literature due to spatial (both horizontal and vertical) and temporal differences in sample collection. On June 28th

2010 concentrations of naphthalene, phenanthrene, fluorene, fluoranthene and pyrene in the waters of Grande Terre in Barataria Bay, LA ranged from 0.5085 µg/L to 6.015 µg/L (Whitehead et al. 2012). By September 30th 2010 there was no evidence of PAHs in the water at Grande

Terre (Whitehead et al. 2012). Sammarco et al. (2013) reported the sum of PAHs concentration to range as high as 1.23 ug/L in offshore seawater samples collected from May through

November 2010. Rumbold et al. (2013) reported sum of PAHs concentrations as high as 2 ug/L in seawater and microlayer samples collected from in the northern GOM from 2011 to 2012.

Incardona et al. (2014) reviewed the reported field concentrations from early monitoring and found ΣPAH concentrations in the range of 3–14 μg/L. However, at least one review by Boehm et al. (2011) reported Total Polycyclic Aromatic Hydrocarbon (TPAH) in whole, unfractionated water samples at concentrations ranging up to 146,000 ug/L (parts per billion) in water samples taken between May and October 2010. They cautioned that concentrations of TPAH decreased with distance from the site of the blowout, down to <1.0 ppb within 15-20 miles in all directions except southwest, where a small number of samples exceeded 1 ppb out to a distance of 40 miles

(Boehm et al. 2011). However, these are highly weathered TPAHs, more weathered than the artificially weathered oil used in the present study as indicated by compositional differences. For examples, the C-3 naphthalenes that dominated the CEWAFs in the present study were mostly absent in the field collected samples (Sammarco et al. 2013, Incardona et al. 2014).

Negative effects on individual finfish ELS and potential linkage to cascading effects at the population or community level, as alluded to in the conceptual diagram (Figure 1), must be considered. Each substantial release of oil to an ecosystem is unique. The Braer spill in Scotland in 1993 presented more ecological risk to the biota of the water column surrounding the ship, whereas the Exxon Valdez oil spill in 1989 presented more of a risk to the shores of Prince

William Sound (Boehm and Page 2007). After the 1989 Exxon Valdez oil spill, many believed that Prince William Sound (PWS) could be characterized as having undergone a major ecosystem shift (Cooney and Council 1999). There was a dramatic loss of cover for the intertidal rockweed that triggered the establishment of opportunistic species like green algae and barnacles as well as the subsequent decline in important gastropods such as periwinkles, limpets and

predatory whelks. Additional evidence of a major shift was the 75% population crash of the

Pacific herring and poor recruitment of the Pink Salmon in 1993, both of which constitute a significant component of the PWS ecosystem and food web (Cooney and Council 1999, Carls et al. 2002). Although the exact trigger or combination of factors that caused this crash are unknown, Carls et al. (2002) attributed the crash of the adult herring population to increased population size, disease, suboptimal nutrition and the possibility of indirect links to the Exxon

Valdez oil spill. While there was a small increase in abundance due to fishery closures during the wake of the oil spill, the Pacific herring population was already at historically high levels from

1989 to 1992 and near its carrying capacity (Carls et al. 2002). In addition, decreased body weight of the harlequin duck, important intertidal foragers, was correlated to chronic exposure to

PAHs up to 9 years after the spill (Iverson and Esler 2010). Finally, by the year 2000 the sea otter population was still at half the estimated pre-spill numbers (Dean et al. 2000). All of the above species specific negative effects from the spill and exposure to Exxon Valdez oil eventually had a cascading effect on the ecosystem as a whole and resulted in a diversion from pre-spill conditions and altered trophic dynamics. A recent analysis of valued ecosystem components within PWS deemed the ecosystem recovered from the effects caused by the Exxon

Valdez oil spill, but not necessarily other stressors (Harwell and Gentile 2006). Although the

Exxon Valdez spill resulted in changes in ecosystem dynamics in the wake of spill, this may not be the case for DWH effects and the GOM.

Ecological significance of the toxic effects of CEWAF on developing red drum and ultimately all finfish ELS in the sea surface of the GOM remains unresolved. The data presented in this study provides evidence that there is obvious detriment to red drum embryos in the presence of MC252 CEWAF during these 24 hr of critical developmental. This interpretation

does not take in to account any other significant stochastic obstacles or stressors like hurricane events or other anthropogenic stressors a cohort may face. The heavily exploited finfish populations of the GOM should be monitored more closely considering that they are facing multiple stressors once exposed from the DWH blowout, both at an individual level and possibly in the form of cascading effects to the population level via increased individual mortality or recruitment, and at the community level through altered trophic interactions (Campagna et al.

2011). The true impact of the BP DWH oil spill on the GOM ecosystem may never be wholly understood. As discussed previously, natural adverse impacts and fishing mortality are nearly impossible to separate from the additional mortality or impact from the exposure to crude oil and

Corexit in the wake of the disaster in 2010. Extensive research, monitoring and continued diligent management will show in time the cumulative response of GOM fish populations as a function of abundance and growth rate to natural stressors, fishing pressure, and crude oil/dispersant exposure in the aftermath of the BP Deepwater Horizon oil spill.

REFERENCES

Anderson, J. W. and R. F. Lee. 2006. Use of biomarkers in oil spill risk assessment in the marine environment. Human and Ecological Risk Assessment: An International Journal 12:1192- 1222. Au, D. W. 2004. The application of histo-cytopathological biomarkers in marine pollution monitoring: a review. Marine Pollution Bulletin 48:817-834. Aurand, D., G. Coelho, and (Editors). 2005. Cooperative aquatic toxicity testing of dispersed oil and the “Chemical response to oil spills: Ecological effects research forum (CROSERF)”. Technical Report 07-03. Ecosystem Management & Associates, Inc., Lusby, MD. 105 pp.+ Appendices. Bailey, K. M. and E. D. Houde. 1989. Predation on eggs and larvae of marine fishes and the recruitment problem. Advances in Marine Biology 25:1. Barron, M. G., M. G. Carls, R. Heintz, and S. D. Rice. 2004. Evaluation of fish early life-stage toxicity models of chronic embryonic exposures to complex polycyclic aromatic hydrocarbon mixtures. Toxicological Sciences 78:60-67. Barron, M. G., M. J. Hemmer, and C. R. Jackson. 2013. Development of aquatic toxicity benchmarks for oil products using species sensitivity distributions. Integrated Environmental Assessment and Management. 9.4: 610-615. Barron, M. G. and L. Ka'aihue. 2003. Critical evaluation of CROSERF test methods for oil dispersant toxicity testing under subarctic conditions. Marine Pollution Bulletin 46:1191- 1199. Beck, J. and R. Turingan. 2007. The effects of zooplankton swimming behavior on prey-capture kinematics of red drum larvae, Sciaenops ocellatus. Marine Biology 151:1463-1470. Benfield, M. C. and R. F. Shaw. 2005. Potential spatial and temporal vulnerability of pelagic fish assemblages in the Gulf of Mexico to surface oil spills associated with Deepwater petroleum development. Page xv+158 pp in M. M. S. U.S. Dept. of Interior, Gulf of Mexico OCS Region, editor, New Orleans, LA. Block, B. A., S. L. Teo, A. Walli, A. Boustany, M. J. Stokesbury, C. J. Farwell, K. C. Weng, H. Dewar, and T. D. Williams. 2005. Electronic tagging and population structure of Atlantic bluefin tuna. Nature 434:1121-1127. Boehm, P. D. and D. S. Page. 2007. Exposure elements in oil spill risk and natural resource damage assessments: A review. Human and Ecological Risk Assessment 13:418-448. Boehm, P. D., D. Turton, A. Raval, D. Caudle, D. French, N. Rabalais, R. Spies, and J. Johnson. 2001. Deepwater Program: Literature review, environmental risks of chemical products used in Gulf of Mexico Deepwater oil and gas operations. Page 326 pp in U.S. Dept. of Interior, Minerals Management Service, Gulf of Mexico OCS Region. New Orleans, LA. Bohne-Kjersem, A., A. Skadsheim, A. Goksøyr, and B. E. Grøsvik. 2009. Candidate biomarker discovery in plasma of juvenile cod (Gadus morhua) exposed to crude North Sea oil, alkyl phenols and polycyclic aromatic hydrocarbons (PAHs). Marine Environmental Research 68:268-277. Brette, F., C. Cros, B. Machado, J. P. Incardona, N. L. Scholz, and B. A. Block. 2014. Crude oil impairs cardiac excitation-contraction coupling in fish. Biophysical Journal 106:732a. Brightman, R. I., J. J. Torres, J. Donnelly, and M. E. Clarke. 1997. Energetics of larval red drum, Sciaenops ocellatus. Part II: Growth and biochemical indicators. Fishery Bulletin 95:431- 444.

Bue, B. G., S. Sharr, and J. E. Seeb. 1998. Evidence of damage to pink salmon populations inhabiting Prince William Sound, Alaska, two generations after the Exxon Valdez oil spill. Transactions of the American Fisheries Society 127:35-43. Cameron, P. and J. Berg. 1992. Morphological and chromosomal aberrations during embryonic development in dab Limanda limanda. Marine Ecology Progress Series 91:163-169. Camilli, R., C. M. Reddy, D. R. Yoerger, B. A. S. Van Mooy, M. V. Jakuba, J. C. Kinsey, C. P. McIntyre, S. P. Sylva, and J. V. Maloney. 2010. Tracking hydrocarbon plume transport and biodegradation at Deepwater Horizon. Science 330:201-204. Campagna, C., F. T. Short, B. A. Polidoro, R. McManus, B. B. Collette, N. J. Pilcher, Y. S. de Mitcheson, S. N. Stuart, and K. E. Carpenter. 2011. Gulf of Mexico oil blowout increases risks to globally threatened species. BioScience 61:393-397. Canevari, G. 1978. Some observations on the mechanism and chemistry aspects of chemical dispersion. Chemical Dispersants for the Control of Oil Spills. . In Chemical Dispersants for the Control of Oil Spills (L. T. McCarthy, G. P. Lindblom, and H. F. Walter, Eds.), STP 659, p. 517. American Society for Testing and Materials, Philadelphia. Carls, M. G., L. Holland, M. Larsen, T. K. Collier, N. L. Scholz, and J. P. Incardona. 2008. Fish embryos are damaged by dissolved PAHs, not oil particles. Aquatic Toxicology 88:121- 127. Carls, M. G., J. E. Hose, R. E. Thomas, and S. D. Rice. 2000. Exposure of pacific herring to weathered crude oil: Assessing effects on ova. Environmental Toxicology and Chemistry 19:1649-1659. Carls, M. G., G. D. Marty, and J. E. Hose. 2002. Synthesis of the toxicological impacts of the Exxon Valdez oil spill on Pacific herring (Clupea pallasi) in Prince William Sound, Alaska, U.S.A. Canadian Journal of Fisheries and Aquatic Sciences 59:153-172. Carls, M. G. and J. P. Meador. 2009. A perspective on the toxicity of petrogenic PAHs to developing fish embryos related to environmental chemistry. Human and Ecological Risk Assessment: An International Journal 15:1084-1098. Carls, M. G., S. D. Rice, and J. E. Hose. 1999. Sensitivity of fish embryos to weathered crude oil: Part I. Low-level exposure during incubation causes malformations, genetic damage, and mortality in larval pacific herring (Clupea pallasi). Environmental Toxicology and Chemistry 18:481-493. Carls, M. G., J. W. Short, and J. Payne. 2006. Accumulation of polycyclic aromatic hydrocarbons by Neocalanus copepods in Port Valdez, Alaska. Marine Pollution Bulletin 52: 1480-1489. Chakrabarty, P., C. Lam, J. Hardman, J. Aaronson, P. House, and D. Janies. 2012. SpeciesMap: a web-based application for visualizing the overlap of distributions and pollution events, with a list of fishes put at risk by the 2010 Gulf of Mexico oil spill. Biodiversity and Conservation 21:1865-1876. Clark, J. R., G. E. Bragin, E. J. Febbo, and D. J. Letinski. 2001. Toxicity of physically and chemically dispersed oils under continuous and environmentally realistic exposure conditions: Applicability to dispersant use decisions in spill response planning. Pages 1249-1255 in International Oil Spill Conference. American Petroleum Institute. Coelho, G., J. Clark, and D. Aurand. 2013. Toxicity testing of dispersed oil requires adherence to standardized protocols to assess potential real world effects. Environmental Pollution. 177: 185-188.

Cooney, R. T. and Exxon Valdez Oil Spill Trustee Council (EVOSTC). 1999. Sound Ecosystem Assessment (SEA): An Integrated Science Plan for the Restoration of Injured Species in Prince William Sound, Alaska. Exxon Valdez Oil Spill Trustee Council. Crone, T. J. and M. Tolstoy. 2010. Magnitude of the 2010 Gulf of Mexico oil leak. Science 330:634. Davis, J. T. 1990. Red Drum: Biology and Life History.in S. R. A. Center, editor. The Texas A&M University System, College Station, Texas. de Soysa, T. Y., A. Ulrich, T. Friedrich, D. Pite, S. L. Compton, D. Ok, R. L. Bernardos, G. B. Downes, S. Hsieh, R. Stein, M. C. Lagdameo, K. Halvorsen, L. R. Kesich, and M. J. Barresi. 2012. Macondo crude oil from the Deepwater Horizon oil spill disrupts specific developmental processes during zebrafish embryogenesis. BMC Biology 10:40. Dean, T. A., J. L. Bodkin, S. C. Jewett, D. H. Monson, and D. Jung. 2000. Changes in sea urchins and kelp following a reduction in sea otter density as a result of the Exxon Valdez oil spill. Marine Ecology Progress Series 199:281-291. DeGraeve, G., R. Elder, D. Woods, and H. Bergman. 1982. Effects of naphthalene and benzene on fathead minnows and rainbow trout. Archives of Environmental Contamination and Toxicology 11:487-490. Dethlefsen, V. and K. Tiews. 1985. Review on the effects of pollution on marine fish life and fisheries in the North Sea. Journal of Applied Ichthyology 1:97-118. Di Toro, D. M., J. A. McGrath, and W. A. Stubblefield. 2007. Predicting the toxicity of neat and weathered crude oil: toxic potential and the toxicity of saturated mixtures. Environmental Toxicology and Chemistry 26:24-36. Diercks, A.-R., R. C. Highsmith, V. L. Asper, D. Joung, Z. Zhou, L. Guo, A. M. Shiller, S. B. Joye, A. P. Teske, N. Guinasso, T. L. Wade, and S. E. Lohrenz. 2010. Characterization of subsurface polycyclic aromatic hydrocarbons at the Deepwater Horizon site. Geophysical Research Letters 37. Faksness, L.-G., P. J. Brandvik, and L. K. Sydnes. 2008. Composition of the water accommodated fractions as a function of exposure times and temperatures. Marine Pollution Bulletin 56:1746-1754. Falk-Petersen, I. B. 2005. Comparative organ differentiation during early life stages of marine fish. Fish and Shellfish Immunology 19:397-412. Falk-Petersen, I. B. and E. Kjorsvik. 1987. Acute toxicity tests of the effects of oils and dispersants on marine fish embryos. Sarsia 72:411-413. Falk-Petersen, I. B., S. Lonning, and R. Jakobsen. 1983. Effects of oil and oil dispersants on plankton organisms. Astarte 12:45-47. Felder, D. L. (Ed.). (2009). Gulf of Mexico origin, waters, and biota: Biodiversity. Texas A&M University Press. Fogarty, M. J. and L. W. Botsford. 2007. Population connectivity and spatial management of marine fisheries. Oceanography 20:112-123. Franks, J. S. and N. J. Brown-Peterson. 2002. A review of age, growth, and reproduction of cobia, Rachycentron canadum, from US waters of the Gulf of Mexico and Atlantic Ocean. Pages 552-569 in Proceedings of the Gulf and Caribbean Fisheries Institute. Frias-Torres, S. and J. C. R. Bostater. 2011. Potential impacts of the Deepwater Horizon oil spill on large pelagic fishes. in Remote Sensing of the Ocean, Sea Ice, Coastal Waters, and Large Water Regions 2011, edited by Charles R. Bostater Jr., Stelios P. Mertikas, Xavier Neyt, Miguel Velez-Reyes, Proc. of SPIE Vol. 8175,81750F

Fucik, K. W., K. A. Carr, and B. J. Balcom. 1995. Toxicity of oil and dispersed oil to the eggs and larvae of seven marine fish and invertebrates from the Gulf of Mexico. The Use of Chemicals in Oil Spill Response, American Society for Testing and Materials Selected Technical Paper 1252. Gentile, J., M. Harwell, W. Cropper Jr, C. Harwell, D. DeAngelis, S. Davis, J. Ogden, and D. Lirman. 2001. Ecological conceptual models: a framework and case study on ecosystem management for South Florida sustainability. Science of the Total Environment 274:231- 253. GMFMC (Gulf of Mexico Fishery Management Council). 1987. Amendment number 1 and environmental assessment and supplemental regulatory impact review and initial regulatory flexibility analysis to the secretarial fishery management plan for the red drum fishery of the Gulf of Mexico.in G. o. M. F. M. Council, editor., Tampa, FL. GOMPO (Gulf of Mexico Program Office). 2011. General Facts about the Gulf of Mexico. J. Binninger and J. Allen, editors. http://www.epa.gov/gmpo/about/facts.html Goodbody-Gringley, G., D. L. Wetzel, D. Gillon, E. Pulster, A. Miller, and K. B. Ritchie. 2013. Toxicity of Deepwater Horizon Source Oil and the Chemical Dispersant, Corexit® 9500, to Coral Larvae. PloS one 8:e45574. Grimes, C. B., J. H. Finucane, A. L. Collins, and D. A. DeVries. 1990. Young king mackerel, Scomberomorus cavalla, in the Gulf of Mexico, a summary of the distribution and occurrence of larvae and juveniles, and spawning dates for Mexican juveniles. Bulletin of Marine Science 46:640-654. Hallanger, I. G., N. A. Warner, A. Ruus, A. Evenset, G. Christensen, D. Herzke, G. W. Gabrielsen, and K. Borgå. 2011. Seasonality in contaminant accumulation in Arctic marine pelagic food webs using trophic magnification factor as a measure of bioaccumulation. Environmental Toxicology and Chemistry 30:1026-1035. Hamilton, M. A., R. C. Russo, and R. V. Thurston. 1977. Trimmed Spearman-Karber method for estimating median lethal concentrations in toxicity bioassays. Environmental Science and Technology 11:714-719. Hansen, D. J. 2003. Procedures for the derivation of equilibrium partitioning sediment benchmarks (ESBs) for the protection of benthic organisms: PAH mixtures. US Environmental Protection Agency, Office of Research and Development, National Health and Environmental Effects Research Laboratory, Atlantic Ecology Division. Harwell, M. A. and J. H. Gentile. 2006. Ecological significance of residual exposures and effects from the Exxon Valdez oil spill. Integrated Environmental Assessment Management 2:204-246. Hatlen, K., C. A. Sloan, D. G. Burrows, T. K. Collier, N. L. Scholz, and J. P. Incardona. 2010. Natural sunlight and residual fuel oils are an acutely lethal combination for fish embryos. Aquatic Toxicology 99:56-64. Heintz, R., J. W. Short, and S. D. Rice. 1999. Sensitivity of fish embryos to weathered crude oil: Part II Increased mortality of pink salmon (Onchorhynchus gorbuscha) embryos incubating downstream from weathered Exxon Valdez crude oil. Environmental Toxicology and Chemistry 18:494-503. Heintz, R. A., S. D. Rice, M. G. Carls, and J. W. Short. 2012. The authors' second reply. Environmental Toxicology and Chemistry 31:475-476.

Hempel, G. 1979. Early life history of marine fish: the egg stage. Washington Sea Grant: distributed by University of Washington Press. Hicken, C. E., T. L. Linbo, D. H. Baldwin, M. L. Willis, M. S. Myers, L. Holland, M. Larsen, M. S. Stekoll, S. D. Rice, and T. K. Collier. 2011. Sublethal exposure to crude oil during embryonic development alters cardiac morphology and reduces aerobic capacity in adult fish. Proceedings of the National Academy of Sciences 108:7086-7090. Holdway, D. A. 2002. The acute and chronic effects of wastes associated with offshore oil and gas production on temperate and tropical marine ecological processes. Marine Pollution Bulletin 44:185-203. Holm, J., V. Palace, K. Wautier, R. Evans, C. Baron, C. Podemski, P. Siwik, and G. Sterling. 2003. An assessment of the development and survival of wild rainbow trout (Oncorhynchus mykiss) and brook trout (Salvelinus fontinalis) exposed to elevated selenium in an area of active coal mining. Pages 257-273 in Proceedings of the 26th Annual Larval Fish Conference. Institute of Marine Research Bergen (NO). Hose, J. E. 1994. Large-scale Genotoxicity Assessments in the Marine Environment. Environmental Health Perspectives 102:29-32. Hose, J. E. and E. D. Brown. 1998. Field applications of the piscine anaphase aberration test: lessons from the Exxon Valdez oil spill. Mutation Research/Fundamental and Molecular Mechanisms of Mutagenesis 399:167-178. Hose, J. E., M. D. McGurk, G. D. Marty, D. E. Hinton, E. D. Brown, and T. T. Baker. 1996. Sublethal effects of the (Exxon Valdez) oil spill on herring embryos and larvae: morphological, cytogenetic, and histopathological assessments, 1989 1991. Canadian Journal of Fisheries and Aquatic Sciences 53:2355-2365. Incardona, J. P., M. G. Carls, H. Teraoka, C. A. Sloan, T. K. Collier, and N. L. Scholz. 2005. Aryl hydrocarbon receptor-independent toxicity of weathered crude oil during fish development. Environmental Health Perspectives 113:1755-1762. Incardona, J. P., T. K. Collier, and N. L. Scholz. 2004. Defects in cardiac function precede morphological abnormalities in fish embryos exposed to polycyclic aromatic hydrocarbons. Toxicology and Applied Pharmacology 196:191-205. Incardona, J. P., L. D. Gardner, T. L. Linbo, T. L. Brown, A. J. Esbaugh, E. M. Mager, J. D. Stieglitz, B. L. French, J. S. Labenia, and C. A. Laetz. 2014. Deepwater Horizon crude oil impacts the developing hearts of large predatory pelagic fish. Proceedings of the National Academy of Sciences: 201320950. Incardona, J. P., T. L. Swarts, R. C. Edmunds, T. L. Linbo, A. Aquilina-Beck, C. A. Sloan, L. D. Gardner, B. A. Block, and N. L. Scholz. 2013. Exxon Valdez to Deepwater Horizon: comparable toxicity of both crude oils to fish early life stages. Aquatic Toxicology 142:303-316. Iverson, S. A. and D. Esler. 2010. Harlequin Duck population injury and recovery dynamics following the 1989 Exxon Valdez oil spill. Ecological Applications 20:1993-2006. Jiang, Z., Y. Huang, X. Xu, Y. Liao, L. Shou, J. Liu, Q. Chen, and J. Zeng. 2010. Advance in the toxic effects of petroleum water accommodated fraction on marine plankton. Acta Ecologica Sinica 30:8-15. Jung, J.-H., C. E. Hicken, D. Boyd, B. F. Anulacion, M. G. Carls, W. J. Shim, and J. P. Incardona. 2013. Geologically distinct crude oils cause a common cardiotoxicity syndrome in developing zebrafish. Chemosphere 91:1146-1155.

Kennedy, C., L. McDonald, R. Loveridge, and M. Strosher. 2000. The effect of bioaccumulated selenium on mortalities and deformities in the eggs, larvae, and fry of a wild population of cutthroat trout (Oncorhynchus clarki lewisi). Archives of Environmental Contamination and Toxicology 39:46-52. Kuhnhold, W. W. 1970. The influence of crude oils on fish fry. Food and Agriculture Organization of the United Nations. In Ruivo, M. (1972). Marine pollution and sea life. Langheinrich, U., G. Vacun, and T. Wagner. 2003. Zebrafish embryos express an orthologue of HERG and are sensitive toward a range of QT-prolonging drugs inducing severe arrhythmia. Toxicology and Applied Pharmacology 193:370-382. Liu, Z., J. Liu, Q. Zhu, and W. Wu. 2012. The weathering of oil after the Deepwater Horizon oil spill: insights from the chemical composition of the oil from the sea surface, salt marshes and sediments. Environmental Research Letters 7:035302. Lönning, S. and B. E. Hagström. 1976. Deleterious effects of corexit 9527 on fertilization and development. Marine Pollution Bulletin 7:124-127. Machlis, G. E. and M. K. McNutt. 2010. Scenario-Building for the Deepwater Horizon Oil Spill. Science 329:1018-1019. McIntosh, S., T. King, D. Wu, and P. V. Hodson. 2010. Toxicity of dispersed weathered crude oil to early life stages of Atlantic herring (Clupea harengus). Environmental toxicology and chemistry 29:1160-1167. McNutt, M., R. Camilli, G. Guthrie, P. Hsieh, V. Labson, B. Lehr, D. Maclay, A. Ratzel, and M. Sogge. 2011. Assessment of flow rate estimates for the Deepwater Horizon/Macondo well oil spill. Flow Rate Technical Group Report to The National Incident Command, Interagency Solutions Group:1-22. Middaugh, D., M. Hemmer, and E. Lores. 1988. Teratological effects of 2, 4-dinitrophenol," produced water" and naphthalene on embryos of the inland silverside Menidia beryllina. Diseases of Aquatic Organisms 4:53-65. Motta, P. J., M. Maslanka, R. E. Hueter, R. L. Davis, R. De la Parra, S. L. Mulvany, M. L. Habegger, J. A. Strother, K. R. Mara, and J. M. Gardiner. 2010. Feeding anatomy, filter- feeding rate, and diet of whale sharks Rhincodon typus during surface ram filter feeding off the Yucatan Peninsula, Mexico. Zoology 113:199-212. Muhling, B. A., M. A. Roffer, J. T. Lamkin, G. W. Ingram, Jr., M. A. Upton, G. Gawlikowski, F. Muller-Karger, S. Habtes, and W. J. Richards. 2012. Overlap between Atlantic bluefin tuna spawning grounds and observed Deepwater Horizon surface oil in the northern Gulf of Mexico. Marine Pollution Bulletin 64:679-687. Murie, D. and D. Parkyn. 2008. Age, Growth and Sex Maturity of Greater Amberjack (Seriola dumerili) in the Gulf of Mexico. MARFIN Final Report (NA05NMF4331071). NMFS (National Marine Fisheries Service). 2011. Compilation of Federal Regulations- Fisheries of the Caribbean, Gulf and South Atlantic. Department of Commerce NOAA National Marine Fisheries Service. NMFS (National Marine Fisheries Service). 2012. Recreational Fisheries Statistics Queries. http://www.st.nmfs.noaa.gov/recreational-fisheries/access-data/run-a-data-query/index OSAT 2 (Operational Science Advisory Team). 2011. Summary Report for Fate and Effects of Remnant Oil in the Beach Environment. New Orleans: Gulf Coast Incident Management Team. http://www.restorethegulf.gov/sites/default/files/u316/OSAT‐ 2%20Report%20no%20ltr.pdf

Padrós, F. and S. Crespo. 1996. Ontogeny of the lymphoid organs in the turbot Scophthalmus maximus: a light and electron microscope study. Aquaculture 144:1-16. Page, D. S., J. M. Neff, P. F. Landrum, and P. M. Chapman. 2012. Sensitivity of pink salmon (Oncorhynchus gorbuscha) embryos to weathered crude oil. Environmental Toxicology and Chemistry 31:469-471. Peterson, C. H., S. D. Rice, J. W. Short, D. Esler, J. L. Bodkin, B. E. Ballachey, and D. B. Irons. 2003. Long-term ecosystem response to the Exxon Valdez oil spill. Science 302:2082- 2086. Ramachandran, S. D., P. V. Hodson, C. W. Khan, and K. Lee. 2004. Oil dispersant increases PAH uptake by fish exposed to crude oil. Ecotoxicology and Environmental Safety 59:300-308. Reddy, C. M., J. S. Arey, J. S. Seewald, S. P. Sylva, K. L. Lemkau, R. K. Nelson, C. A. Carmichael, C. P. McIntyre, J. Fenwick, G. T. Ventura, B. A. S. Van Mooy, and R. Camilli. 2011. Composition and fate of gas and oil released to the water column during the Deepwater Horizon oil spill. Proceedings of the National Academy of Sciences. 109(50), 20229-20234. Richards, W. J., M. F. McGowan, T. Leming, J. T. Lamkin, and S. Kelley. 1993. Larval fish assemblages at the loop current boundary in the Gulf of Mexico. Bulletin of Marine Science 53:475-537. Rico-Martínez, R., T. W. Snell, and T. L. Shearer. 2013. Synergistic toxicity of Macondo crude oil and dispersant Corexit 9500A to the Brachionus plicatilis species complex (Rotifera). Environmental Pollution 173:5-10. Rose, K. A., Cowan, J. H., Winemiller, K. O., Myers, R. A., & Hilborn, R. (2001). Compensatory density dependence in fish populations: importance, controversy, understanding and prognosis. Fish and Fisheries, 2(4), 293-327. Roy, G., R. Vuillemin, and J. Guyomarch. 2005. On-site determination of polynuclear aromatic hydrocarbons in seawater by stir bar sorptive extraction (SBSE) and thermal desorption GC–MS. Talanta 66:540-546. Ryerson, T. B., R. Camilli, J. D. Kessler, E. B. Kujawinski, C. M. Reddy, D. L. Valentine, E. Atlas, D. R. Blake, J. de Gouw, and S. Meinardi. 2012. Chemical data quantify Deepwater Horizon hydrocarbon flow rate and environmental distribution. Proceedings of the National Academy of Sciences 109:20246-20253. Saco-Alvarez, L., J. Bellas, O. Nieto, J. M. Bayona, J. Albaiges, and R. Beiras. 2008. Toxicity and phototoxicity of water-accommodated fraction obtained from Prestige fuel oil and Marine fuel oil evaluated by marine bioassays. Science of the Total Environment 394:275-282. Schofield, K., S. Marcy, and G. Suter. 2007. Unspecified Toxic Chemicals: Detailed Conceptual Diagram. CADDIS Volume 2: Sources, Stressors & Responses. U.S. Environmental Protection Agency. Schrope, M. 2011. Deep Wounds. Nature 472:152-154. Sharp, J., K. Fucik, and J. Neff. 1979. Physiological basis of differential sensitivity of fish embryonic stages to oil pollution. Marine Pollution: Functional Responses, Vernberg, W. B., A. Calabrese, F. P. Thurberg and F. J. Vernberg, Eds., Academic Press, New York p 85-108, 1979. 9 Fig, 32 Ref. Sinderman, C. J. 1994. Quantitative Effects of Pollution on Marine and Anadromous Fish Populations. U.S. Department of Commerce NOAA NMFS.

Sinderman, C. J. 2006. Effects of Coastal Pollution on Yields Fish and Shellfish Resources. Pages 163-200 in M. J. Kennish, editor. Coastal Pollution Effects on Living Resources and Humans. Taylor & Francis. Singer, M., D. Aurand, G. Bragin, J. Clark, G. Coelho, M. Sowby, and R. Tjeerdema. 2000. Standardization of the preparation and quantitation of water-accommodated fractions of petroleum for toxicity testing. Marine Pollution Bulletin 40:1007-1016. Smith, R. L. and J. A. Cameron. 1979. Effect of water soluble fraction of Prudhoe Bay crude oil on embryonic development of Pacific herring. Transactions of the American Fisheries Society 108:70-75. Stene, A. and S. Lønning. 1984. Effects of 2-methylnaphthalene on eggs and larvae of six marine fish species. Sarsia 69:199-203. Sumaila, U. R., A. M. Cisneros-Montemayor, A. Dyck, L. Huang, W. Cheung, J. Jacquet, K. Kleisner, V. Lam, A. McCrea-Strub, W. Swartz, R. Watson, D. Zeller, D. Pauly, and T. Quinn. 2012. Impact of the Deepwater Horizon well blowout on the economics of US Gulf fisheries. Canadian Journal of Fisheries & Aquatic Sciences 69:499-510. Tunnell, J. W. J. 2011. An expert opinion of when the Gulf of Mexico will return to pre-spill harvest status following the BP Deepwater Horizon MC 252 oil spill. Gulf Coast Claims Facility; Harte Research Institute for Gulf of Mexico Studies, Washington, D.C. USCG (US Coast Guard) 2011. On scene coordinator report Deepwater Horizon oil spill: Submitted to the National Response Team. USEPA (US Environmental Protection Agency).1998. Guidelines for Ecological Risk Assessment. Washington DC: Risk Assessment Forum, USEPA. EPA/630/R095//002F. Westernhagen, H. v. D., V.; Cameron, P.; Berg, J.; Furstenberg G. 1988. Developmental defects in pelagic fish embryos from the western Baltic. Helgoländer Meeresuntersuchungen 43:45-60. White, P. A. 2002. The genotoxicity of priority polycyclic aromatic hydrocarbons in complex mixtures. Mutation Research/Genetic Toxicology and Environmental Mutagenesis 515:85-98. Whitehead, A., B. Dubansky, C. Bodinier, T. I. Garcia, S. Miles, C. Pilley, V. Raghunathan, J. L. Roach, N. Walker, and R. B. Walter. 2012. Genomic and physiological footprint of the Deepwater Horizon oil spill on resident marsh fishes. Proceedings of the National Academy of Sciences 109:20298-20302. Whitehouse, B. G. 1984. The effects of temperature and salinity on the aqueous solubility of polynuclear aromatic hydrocarbons. Marine Chemistry 14:319-332. Wurl, O. and J. P. Obbard. 2004. A review of pollutants in the sea-surface microlayer (SML): a unique habitat for marine organisms. Marine Pollution Bulletin 48:1016-1030.