toxicity of oil from bp deepwater horizon blowout on the...
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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
By
Kelsey L. McEachern
July 2014
i
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.
ii
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.
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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
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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
v
LIST OF TABLES
Table 1. Comparison of Finfish Toxicity Tests
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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
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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
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would be dependent on the spatiotemporal severity of exposure particularly in relation to the
timing of natural density-dependent population regulation.
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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
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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,
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$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).
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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).
Figure 1. Conceptual diagram of ecological risk resulting from the Deepwater
Horizon well blowout.
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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
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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.
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
Figure 2. Satellite image of surface oil in northern Gulf of Mexico after
Deepwater Horizon well blowout.
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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
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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
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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
12
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
13
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
14
(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
15
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
16
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
17
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
18
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
19
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.
20
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
21
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
22
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
23
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
24
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
25
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.
26
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
Where:
= Score (0, 1, 2 or 3)
= the number of times the score, , was recorded
N = the number of larvae within that treatment
27
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.
28
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.
B C
D
A
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.
29
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%
0 1000 2000 3000 4000 5000
C3-NapthaleneC2-NapthaleneC1-Napthalene
Napthalene
PAH Concentration (ppb, g/L)
0 50 100 150 200 250
Benzo[e]pyreneC2-ChrysenesC1-Chrysenes
ChryseneBenzo[a]anthracene
C2-Fluoranthenes / PyrenesC1-Fluoranthenes / Pyrenes
FluoranthenePyrene
C3-DibenzothiopheneC2-DibenzothiopheneC1-Dibenzothiophene
DibenzothiopheneC3-Phenan/anthraC2-Phenan/anthraC1-Phenan/anthra
PhenanthreneAnthracene
C3-FluoreneC2-FluoreneC1-Fluorene
FluoreneBiphenyl
Figure 4. Average relative concentration and composition of major PAH constituents in 4 CEWAF
stock preparations. Note two different scales.
30
(Figure 5). The remaining 23 common constituents of PAHs analyzed contributed less than 1%
to the total TU.
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
Toxic Units (TU)
0 2 4 6 8 10
Benzo[e]pyreneC2-ChrysenesC1-Chrysenes
ChryseneBenzo[a]anthracene
C2-Fluoranthenes / PyrenesC1-Fluoranthenes / Pyrenes
FluoranthenePyrene
C3-DibenzothiopheneC2-DibenzothiopheneC1-Dibenzothiophene
DibenzothiopheneC3-Phenan/anthraC2-Phenan/anthraC1-Phenan/anthra
PhenanthreneAnthracene
C3-FluoreneC2-FluoreneC1-Fluorene
FluoreneBiphenyl
0 20 40 60 80 100
C3-NapthaleneC2-NapthaleneC1-Napthalene
Napthalene
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.
31
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.
32
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
Figure 6. Developmental abnormalities observed as a result of 24 hr acute exposure to MC252 CEWAF.
33
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.
34
Regression E3; R2 = 0.96 ; p = 0.0036
Log + 1 Concentration
0.0 0.2 0.4 0.6 0.8 1.0 1.2 1.4 1.6 1.8
Mean
Ran
k S
co
re
0.0
0.5
1.0
1.5
2.0
2.5
3.0
3.5
Exposure 1
Exposure 2
Exposure 3
Skeletal
Regression E1; R2 = 0.77 ; p = 0.0211Regression E2; R2 = 0.89 ; p = 0.0156Regression E3; R2 = 0.90 ; p = 0.0145
Log + 1 Concentration
0.0 0.2 0.4 0.6 0.8 1.0 1.2 1.4 1.6 1.8
Mean
Ran
k S
co
re
0.0
0.5
1.0
1.5
2.0
2.5
3.0
3.5
Exposure 1
Exposure 2
Exposure 3
Cardiac
Regression E1; R2 = 0.80 ; p = 0.0166
Log + 1 Concentration
0.0 0.2 0.4 0.6 0.8 1.0 1.2 1.4 1.6 1.8
Mean
Ran
k S
co
re
0.0
0.5
1.0
1.5
2.0
2.5
3.0
3.5
Exposure 1
Exposure 2
Exposure 3
Yolk
Regression E1; R2 = 0.72 ; p = 0.0331Regression E2; R2 = 0.77 ; p = 0.0489Regression E3; R2 = 0.81 ; p = 0.0390
Log + 1 Concentration
0.0 0.2 0.4 0.6 0.8 1.0 1.2 1.4 1.6 1.8
Mean
Ran
k S
co
re
0.0
0.5
1.0
1.5
2.0
2.5
3.0
3.5
Exposure 1
Exposure 2
Exposure 3
Finfold
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.
35
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).
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
Total Length
PAH (mg/L)
0 1 2 3 4 5
Len
gth
(m
m)
0.8
1.0
1.2
1.4
1.6
1.8
2.0
2.2
2.4
2.6
Exposure 1
Exposure 2
Exposure 3
Figure 8. Total length of post hatch yolk sac larvae as a function of exposure concentration.
36
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).
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).
SV:VT
PAH (mg/L)
0 1 2 3 4 5
SV
:VT
Len
gth
Ra
tio
0.0
0.2
0.4
0.6
0.8
1.0
1.2
1.4
1.6
Exposure 1
Exposure 2
Exposure 3
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
37
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.
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
Yolk Length:Diameter
PAH (mg/L)
0 1 2 3 4 5
Yo
lk L
en
gth
: D
iam
ete
r R
ati
o
0.0
0.5
1.0
1.5
2.0
Exposure 1
Exposure 2
Exposure 3
Figure 10. Ratio of yolk length to diameter versus exposure concentration. Regression E2; R2 = 0.66;
p = 0.0481
38
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.
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
Log + 1 Concentration
0.0 0.2 0.4 0.6 0.8 1.0 1.2 1.4 1.6 1.8 2.0
Perc
en
t N
on
-Via
ble
(%
)
0
20
40
60
80
100
Exposure 1
Exposure 2
Exposure 3
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: R
2 = 0.79, p = 0.0178 Regression E3: R
2 = 0.83,
p = 0.0110.
39
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).
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
Log + 1 Concentration
0.0 0.2 0.4 0.6 0.8 1.0 1.2 1.4 1.6 1.8 2.0
Pe
rce
nt
Mo
rta
lity
(%
)
0
20
40
60
80
100
Exposure 1
Exposure 3
Exposure 3
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
graphic presentation. Regression E2: R2 = 0.89, p = 0.0039; Regression E3: R
2 = 0.72, p = 0.0323
40
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
41
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
42
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).
43
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
44
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
45
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
46
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
47
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
48
(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,
49
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
50
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
51
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
52
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
53
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
54
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.
55
Table 1. Comparison of EC50s and LC50s from different types of oil and exposure methodologies.
Species Test Conditions Response Toxicant /OriginEC50/LC50
(ppm)Reference
WAF Bluefin Tuna embryos
(Thunnus thynnus)24 h static Cardiac Edema MC252 Weathered Source 0.0008 Incardona et al. 2014
Yellowfin Tuna embryos
(Thunnus albacares)24 h static Cardiac Edema MC252 Weathered Source 0.0023 Incardona et al. 2014
Greater Amberjack embryos
(Seriola dumerili)24 h static Cardiac Edema MC252 Weathered Source 0.0124 Incardona et al. 2014
Pacific Herring embryos
(Clupea pallasi )
16 d (checked
daily)Survival
Weathered Alaska North Slope
crude oil on gravel53.3 Carls et al. 1999
Pacific Herring embryos
(Clupea pallasi )
16 d (checked
daily)
Spinal
Abnormality
Weathered Alaska North Slope
crude oil on gravel
33.5 (LWO) -
3.6 (MWO)Carls et al. 1999
Killifish embryos
(Fundulus heteroclitus)Survival #2 Fuel Oil (API Reference Oil III) 1.5 Sharp et al. 1979
CEWAF Red Drum embryos
(Sciaenops ocellatus)24 h static Viability
MC252 Weathered Source &
Corexit 95000.38 - 1.63 Present Study
Red Drum embryos
(Sciaenops ocellatus)24 h static Mortality
MC252 Weathered Source &
Corexit 95000.48 - 2.43 Present Study
Red Drum embryos
(Sciaenops ocellatus )48 h static Survival
Western Gulf of Mexico Oil WAF &
Corexit 9527>100 Fucik et al. 1995
Red Drum embryos
(Sciaenops ocellatus )48 h static Survival
Central Gulf of Mexico Oil WAF &
Corexit 9527>100 Fucik et al. 1995
Atlantic Herring embryos
(Clupea harengus )24 h static
Blue Sac disease
(BSD)
Medium South American Crude
Oil & Corexit 95008.5 McIntosh et al. 2010
Turbot yolk-sac larvae
(Scophthalmus maximus)48 h static Survival
Fresh Kuwait Crude Oil & Corexit
95272 Clark et al. 2001
Turbot yolk-sac larvae
(Scophthalmus maximus)48 h static Survival
Fresh Forties Crude Oil & Corexit
95000.44 Clark et al. 2001
Inland Silverside juveniles
(Menidia beryllina )96 h static Survival
Fresh Kuwait Crude Oil & Corexit
95270.55 Clark et al. 2001
Inland Silverside juveniles
(Menidia beryllina )96 h static Survival
Weathered Kuwait Crude Oil &
Corexit 95271.09 Clark et al. 2001
Inland Silverside juveniles
(Menidia beryllina )96 h static Survival
Fresh Forties Crude Oil & Corexit
95000.49 Clark et al. 2001
Inland Silverside adults
(Menidia beryllina )96 h static Survival No. 2 Fuel Oil & Corexit 9500 2.61
EPA National Contingency Plan
Schedule Toxicity Summary
56
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
57
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
58
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
59
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
60
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
61
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
62
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.
63
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