PHOTO-INDUCED TOXICITY OF DEEPWATER HORIZON SPILL OIL TO FOUR NATIVE

GULF OF MEXICO SPECIES

Matthew M. Alloy

Dissertation Prepared for the Degree of

DOCTOR OF PHILOSOPHY

UNIVERSITY OF NORTH TEXAS

December 2015

Approved By:

Aaron Roberts, Major Professor

Samuel Atkinson, Committee Member

Barney Venables, Committee Member

James Oris, Committee Member

James Stoeckel, Committee Member

Alloy, Matthew M. Photo-Induced Toxicity of Deepwater Horizon Spill Oil to Four Native

Gulf of Mexico Species. Doctor of Philosophy (Environmental Science), December 2015, 88 pp.,

references, 212 titles.

The 2010 Deepwater Horizon oil spill resulted in the accidental release of millions of

barrels of crude oil into the Gulf of Mexico (GoM). Photo-induced toxicity following co-exposure

to ultraviolet (UV) radiation is one mechanism by which polycyclic aromatic hydrocarbons

(PAHs) from oil spills may exert toxicity. Blue crab (Callinectes sapidus) are an important

commercial and ecological resource in the Gulf of Mexico and their largely transparent larvae

may make them sensitive to PAH photo-induced toxicity. Mahi-mahi (Coryphaena hippurus), an

important fishery resource, have positively buoyant, transparent eggs. These characteristics

may result in mahi-mahi embryos being at particular risk from photo-induced toxicity. Red

drum (Sciaenops ocellatus) and speckled seatrout (Cynoscion nebulosus) are both important

fishery resources in the GoM. They spawn near-shore and produce positively buoyant embryos

that hatch into larvae in about 24 h. The goal of this body of work was to determine whether

exposure to UV as natural sunlight enhances the toxicity of crude oil to early lifestage GoM

species. Larval and embryonic organisms were exposed to several dilutions of water

accommodated fractions (WAF) from several different oils collected in the field under chain of

custody during the 2010 spill and two to three gradations of natural sunlight in a factorial

design. Here, we report that co-exposure to natural sunlight and oil significantly reduced larval

survival and embryo hatch compared to exposure to oil alone.

ii

Copyright 2015

by

Matthew M. Alloy

iii

TABLE OF CONTENTS

Chapters

1. INTRODUCTION 1

Polycyclic Aromatic Hydrocarbons 1

Photo-induced Toxicity 2

Factors Influencing Photo-induced Toxicity 5

Photo-induced Toxicity and Oil 6

Deepwater Horizon Spill 8

References 11

2. PHOTO–INDUCED TOXICITY OF DEEPWATER HORIZON SLICK OIL TO BLUE CRAB

(CALLINECTES SAPIDUS) LARVAE 18

Introduction 19

Methods 19

Test Organism 19

Test Solutions 19

Toxicity Tests 21

Phototoxic Units 22

Statistical Analyses 23

Results 24

Toxicity Associated with WAF Preparation 24

Toxicity Associated with Oil Source 25

Discussion 30

Acknowledgements 33

References 34

3. ULTRAVIOLET RADIATION ENHANCES THE TOXICITY OF DEEPWATER HORIZON OIL TO

MAHI MAHI (CORYPHAENA HIPPURUS) EMBRYOS 37

Introduction 37

Materials and Methods 38

iv

Test Organism 38

Test Solutions 39

Toxicity Tests 40

Phototoxic Units 41

Statistical Analyses 42

Results 42

Discussion 47

Toxicity Associated with WAF Preparation Method 48

Toxicity Associated with Oil Source 49

Acknowledgements 52

References 53

4. SENSITIVITY OF EARLY LIFESTAGE RED DRUM (SCIAENOPS OCELLATUS) AND SPECKLED

SEATROUT (CYNOSCION NEBULOSUS) TO PHOTO–INDUCED TOXICITY OF NATURALLY

WEATHERED DEEPWATER HORIZON OIL 57

Introduction 57

Materials and Methods 59

Test Organisms 59

Test Solutions 59

Toxicity Tests 60

Phototoxic Units 61

Statistical Analyses 62

Results 62

Discussion 69

Acknowledgements 71

References 72

5. DISCUSSION 77

v

Influences on Polycyclic Aromatic Hydrocarbon Tissue Concentrations 77

Ultraviolet Wavelength Absorbing Pigmentation 80

Factors Relating to Species Sensitivity 82

Possible Effects on Recruitment 84

Acknowledgements 86

References 87

1

INTRODUCTION

Polycyclic Aromatic Hydrocarbons

Polycyclic aromatic hydrocarbons (PAHs) are a broad classification of environmental

contaminants. They consist of organic molecules with multiple ring structures and vary widely

in their chemical and toxicological properties [1, 2]. PAHs are rarely found in the environment

as a single chemical species. Typically, they are a complex mixture with a wide variety of

environmental sources including: geological seepage, industrial processing, combustion engine

operation and lubrication, volcanic activity, natural and anthropogenic fires, accidental release

from oil and coal extraction and transport [3].

The individual properties of PAHs vary widely. The two-ring PAH napthalene has a

molecular weight of 128.2 g/mol and a logkow of 3.37, while the five-ring PAH benzo[a]pyrene

has a molecular weight of 252 g/mol and a logkow of 6.04 [4]. The variety of characteristics of

PAHs means that they cover a fairly broad range of chemical and biological actions. They range

from volatile to relatively non-volatile and moderately hydrophobic to highly hydrophobic [4].

In a mixed PAH release, such as crude oil seep or spills, not all constituents of the mixture will

share the same environmental fate. PAHs associated with surface water can volatilize, oxidize,

photolyze, or biodegrade [5]. On the surface the volatiles will tend to diffuse to the gaseous

phase according to Henry’s Law [5, 6] leaving the less volatile fraction to undergo physical

weathering and possible dissolution into the water. They can also partition from the aqueous

phase to associate with sediments or be absorbed into biota [7]. Uptake of PAHs from

sediments or water into the organism differs among species. Exposure routes leading to uptake

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take the form of ingestion or diffusion based on the difference in hydrophobicity between the

organism and the aqueous environment [7]. Ingestion of PAHs can be from the food item’s

body burden, or incidental as with the ingestion of sediment particles during the course of

feeding. Diffusion is thought to be primarily across the gill surfaces during respiratory activity,

but it may also occur across the organism’s integument [8, 9]. PAHs above the surface of the

water or above the UV extinction depth will also be subject to UV photolysis [4, 10]. This can

create various breakdown products by itself [5], and bacterial metabolism can create others

[11].

Despite the variety in PAH characteristics and distributions, PAHs hold some modes of

toxicity in common, most notably by narcosis in aquatic organisms [12]. Narcosis is caused by

the partitioning of hydrophobic PAHs into the cell membranes of neuron tissues, depressing

central nervous system activity [13, 14]. Any PAH without functional groups and two to four

rings can have a narcotic effect on fish [15]. PAHs have been known to cause developmental

defects in larval organisms, specifically; cardiac malformation [16-20], craniofacial defects,

spinal-axis defects [21] ocular malformations [23], and neurodevelopmental defects [24]. It has

also been observed that fish immune system function can be depressed by PAHs [25, 26]. The

current model for the teratogenesis of PAHs is the activation of the Aryl Hydrocarbon receptor

by PAHs during key stages of development interferes with normal developmental regulation

[27-30].

Photo-induced Toxicity

Photo-induced toxicity is the phenomenon of a co-exposure of light increasing the

toxicity of a compound. Sometimes this increase can be more than a thousand fold [31-33]. It

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can occur broadly across kingdoms, in plants [34-36], and in animals [18, 37-51]. Compounds

that exhibit photo-induced toxicity vary across classifications, from pesticides, to endogenous

phytochemicals, and PAHs [33].

Photo-induced toxicity can refer to two classes of mechanisms, photosensitizers, and

photomodified toxicants. A photosensitizer is a molecule that, after cellular uptake, can cause

damage to cell membranes, enzymes, organelles, and DNA by generation of reactive oxygen

species (ROS) in the presence of UV light of a specific wavelength range. A photomodified

compound is one that has been altered in structure by UV light, enhancing its toxicity relative to

the parent compound. Often these are quinones. Modification can occur immediately prior to,

or even a relatively long time before exposure to the target organism. Pre-irradiation of a PAH

subject to photomodification prior to exposure to organisms will increase observed toxicity [52]

while photosensitizers require simultaneous exposure to UV and PAH. For example,

phenanthrene can photooxidize to 9,10-phenanthrenequinone, anthracene can become

anthraquinone or dihydroxyanthraquinone, and benzo[a]pyrene can become

phenanthrenequione, or benzo[a]pyrenequinone [10]. In the case of a metal bound to an

organic compound, photoreaction can release the metal, making it bioavailable which could

result in toxic effects [53, 54].

The mechanism of action for photosensitizers is the production of ROS by the

absorption of certain wavelengths of light [55]. All wavelengths of light travel at the same rate

of propagation based upon the media traveled through. Velocity is held constant, variation in

wavelength is then determined by the energy of the photons. Therefore, the shorter the

wavelength, the more energy the photon has. UV radiation consists of wavelengths between

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100nm and 400nm. The wavelengths of UV that are primarily responsible for photosensitization

reactions are between 320nm and 400nm, the absorbance range for UV for most PAHs [55].

Photosensitization occurs when a photodynamic PAH is absorbed by an organism which

is then subsequently exposed to ultraviolet radiation. This radiation must penetrate the cell to

reach the PAH and be of a wavelength that, once absorbed by the PAH molecule, excites an

electron to jump the gap from the PAH’s highest occupied orbital to the lowest unoccupied

orbital (HOMO-LUMO gap) [56]. This excited state is then often resolved by reaction with

dissolved oxygen to form an oxygen radical. Damage to cell structures most often results from

this oxygen radical starting a lipid peroxidation chain reaction. This action is dependent on the

wavelength/intensity of the UV exposure matching the HOMO-LUMO gap of the

photosensitizer in question. The wider the HOMO-LUMO gap, the less often an excited electron

will be able to achieve sufficient excitation to initiate a Type I or II reaction [33]. HOMO-LUMO

gap width has been used successfully to group PAHs into potentially phototoxic and non-

phototoxic [2, 57]. Once in the singlet excited state, the photosensitizer can react with

molecular oxygen or another receptive molecule or the electron can lose the gained energy by

thermal emission, fluorescence, or conversion to a triplet state (Figure 1). If the excited electron

is passed to a biological molecule it is known as a Type I reaction. If the singlet or triplet state

reacts with molecular oxygen it is known as a Type II reaction [58]. In the case of a Type I

reaction, the recipient molecule typically responsible for observed cellular damage is a lipid,

forming lipid peroxides that can result in cellular membrane damage leading to cell death [59].

In the case of a Type II reaction, singlet oxygen becomes an intermediary step towards reaction

with biomolecules to ultimately produce the same ROS damaging effects on organelles and

5

membranes [60, 61]. This damage on a tissue level can result in reduced function of an organ,

such as increased oxygen diffusion distance and reduced osmoregularity in the gill [62, 63].

Damage by an intermediary oxidation step or by direct oxidative damage to DNA can lead to

DNA adduct formation, and single strand breaks [2, 64]. It has also been demonstrated in the

literature that established PAH carcinogenicity can be enhanced by photo-activation [2]. Whole

organism effects that have been reported in the literature include: reduction in organism

fecundity [43], photo-avoidance behaviors [65], and feeding inhibition [66].

Figure 1. Photo-excitation leading to either a singlet state and reaction or a triplet state and a reaction.

Factors Influencing Photo-induced Toxicity

Incident solar radiation of any given location is dependent on several variables. The

latitude and season determine the base intensity of solar radiation prior to atmospheric

attenuation [67, 68]. Latitude also affects atmospheric attenuation. The angle of incidence close

to the equator presents less atmospheric gases to attenuate UV. At higher latitudes the angle of

6

incidence can almost double the effective thickness of the atmosphere, resulting in greater

attenuation of UV prior to reaching surface waters [69]. The thickness of the stratospheric

ozone layer is a strong determinant of small wavelength UV radiation. The majority of high

energy UVC (100nm to 280nm) is removed by the ozone layer. Lower energy ranges, UVB

(280nm to 315nm) and UVA (315nm to 400nm), are attenuated to a lesser degree before

reaching surface waters [70]. Many factors determine the depth of UV extinction including

intensity of incident UV at the surface and water column characteristics such as suspended

solids, dissolved organic carbon (DOC), and salinity [70]. The composition of the dissolved and

suspended components in the water can affect which wavelengths are most efficiently

screened out and what rate per unit depth [44, 70-73]. This wavelength specific screening can

shift potential for photosensitized toxicity by reducing the available energy needed for

photomodification or for its HOMO-LUMO gap. The degree to which dissolved organic carbon

has been photobleached also influences the UV in the system. Photobleached DOC fails to

attenuate as much UV as the same concentration of unbleached DOC [73]. Some forms of DOC

can influence photo-induced toxicity in other ways; Oris et al. [44] showed that humic acids can

reduce the bioaccumulation factor of PAHs in addition to UV attenuation. Salinity in systems

has been linked to greater UV penetrance in systems [70, 71]. Weinstein [74] reported that

salinity influenced PAH bioaccumulation and photo-induced toxicity in one estuarine species,

but not another. Weinstein concluded that the organism’s osmoregulation strategy was what

determined whether salinity could influence PAH bioaccumulation.

Photo-induced Toxicity and Oil

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Photo-induced toxicity caused by PAHs has been reported in the literature in both the

laboratory and in situ. Calfee et al. [40] reported phototoxic effects in Daphina when exposed

to the water accommodated fraction (WAF) of oil collected from an abandoned offshore oil

field. Cleveland et al. [41] reported phototoxic effects to Mysid shrimp exposed to WAF from a

Guadalupe oil field. Ireland et al. [75] conducted a test between an upstream reference site and

a PAH contaminated section of river; the authors reported lethal photo-induced toxicity in

Daphnia in their field study as well as in a chronic laboratory assay. Little et al. [76] tested

underground free plume oil acquired under an abandoned oil field for photo-induced toxicity

using indoor solar simulators to tidewater silverside fish, and reported an order of magnitude

lower LC50s for their UV treatment compared to treatments lacking UV irradiation. Monson et

al. [77] investigated phototoxic effects on nematodes in PAH contaminated sediments, both in

the field and in control environments. The authors reported significant mortality in all

treatments where the test organisms were co-exposed to UV light and PAHs, and less than

twenty percent mortality all PAH treatments without UV light exposure. Peachey and Crosby

[78] performed PAH photo-induced toxicity assays on five different phyla of marine organisms

which inhabit reef flats known to have intense UV exposure. Among PAH exposures ranging

from 0 µg /L to 48 µg/L they reported four species which displayed no adverse effect with co-

exposure to UV and PAH, five species that displayed various toxic response within the tested

range and two species which displayed great sensitivity to co-exposure.

Part of the complexity of PAH toxicity is the tendency for PAH mixtures to change in

composition once released into aquatic systems. Photolysis, and bacterial metabolism can alter

toxicity of individual PAH compounds. This weathering effect has been tested and reported in

8

the literature to have increased phototoxic effect. Maki et al. [79] biodegraded Arabian crude

oil for 21 days, then exposed the degraded oil to solar irradiation for up to twenty-eight days.

The authors used this biodegraded and photo degraded oil to investigate photo-induced

toxicity to artemia napulii. They reported photo-induced toxicity to artemia exposed to WAF for

twenty-four hours increased as the solar weathering pretreatment period increased, exhibiting

the greatest toxicity at twenty-five days of irradiation. Environmental factors such as

temperature, pH, and dissolved oxygen can influence the rate of PAH photolysis [80]. The same

study concluded that photolysis rates in PAHs with four rings or more are influenced by pH and

dissolved oxygen. PAHs with fewer than four rings have photolysis rates independent of pH and

dissolved oxygen. Simulated solar weathering of Brazilian crude oil for sixty-four hours resulted

in decreases in the aliphatic and aromatic fractions, and an increase in the polar fraction of the

oil mixture [81]. Composition of PAHs in various crude oil sources is not homogenous, and oils

from different sources may weather differently [82].

The PAHs observed to be in Exxon-Valdez spill water included known photosensitizers

[15]. Ultraviolet light has been shown to penetrate as deeply as 25 meters in Port Valdez waters

under good conditions, and 6 meters even under turbid conditions [15]. Duesterloh et al. [83]

tested the photo-induced toxicity of weathered North Slope Alaskan crude oil to two species of

calanoid copepods. The authors exposed their test organisms to oil concentrations of 2 µg/L,

and two UVA treatments, high UVA (14,253 µW/cm2 total dose), low UVA (6,202 µW/cm2 total

dose), and a no-UVA treatment. They reported significant affects to both test organisms in all

UVA + oil co-exposures relative to oil only exposures.

Deepwater Horizon Spill

9

Beginning on the 20th of April, 2010, the Mobile Offshore Drilling Unit Deepwater

Horizon experienced a series of events that resulted in the sinking of the vessel and subsequent

oil release from the wellhead. On the 15th of July the oil release was controlled, and the

wellhead declared safe on the 19th of September, 2010. Over the period of uncontrolled oil

release –87 days in length– millions of barrels of oil was released [84].

Subsurface oil monitoring was conducted by various groups to monitor the amount of

oil release that was not reaching surface waters. PAH analysis of water samples taken at known

depth and locations indicated several subsurface plumes existed, some for months without

significant degradation, the largest of which was found at approximately 1,100 meters in depth

and to be in a continuous layer for 35 kilometers with a benzene, toluene, ethylbenzene, and

xylenes (BTEX) concentrations as high as 78 µg/L [85].

Approximately 25.5 thousand barrels of dispersants were utilized over the course of the

87 day oil spill [86]. The dispersants were not totally successful in the goal of preventing oil

from reaching the surface (Figure 2). Kujawinski et al. [87] reported that biota encountering

plumes within 10 kilometers of the wellhead might have been exposed to dispersant

concentrations within the range of 10 µg/L to 100 µg/L. They also reported measuring dioctyl

sodium sulfosuccinate (DOSS) –a component in both Corexit 9527 and 9500A formulations used

on the spill– in nanogram per liter concentrations in some sample sites as far as 300 kilometers

from the well head.

10

Figure 2. Maximum extent of the Deepwater Horizon surface slick.

11

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18

CHAPTER 2

PHOTO–INDUCED TOXICITY OF DEEPWATER HORIZON SLICK OIL TO BLUE CRAB (CALLINECTES

SAPIDUS) LARVAE1

Introduction

Polycyclic aromatic hydrocarbons (PAHs) are organic molecules with multiple carbon

ring structures that vary widely in their chemical and toxicological properties [1]. PAHs, as a

group, are lipophilic and readily accumulate in the tissues of aquatic organisms [2]. PAHs have

been shown to result in toxicity through a variety of mechanisms including photo-induced or

photo-enhanced toxicity [3,4]. Photo-induced toxicity is the phenomenon where a compound

exhibits increased toxicity in the presence of certain wavelengths of light [5]. Effects of PAH

photo-induced toxicity include increased mortality [6-10], reduced fecundity [3], increased

photo-avoidance behavior [11], and feeding inhibition [12]. The photo-induced toxicity of PAHs

to a wide variety of aquatic biota including daphnids [3, 13], fish [5, 6, 11, 14], bivalves [8],

marine corals [9], marine diatoms [15], and mysid shrimp [8, 10] is well documented.

Beginning on the 20th of April, 2010, the Mobile Offshore Drilling Unit Deepwater

Horizon (DWH) experienced a series of events that resulted in the sinking of the vessel and

subsequent oil release from the wellhead until it was sealed on July 15th, 2010. An estimated

700 million liters of oil was released, approximately 659 million liters more than the Exxon-

Valdez spill twenty-one years prior [16]. PAHs, including those with photodynamic activity, can

comprise a large fraction of crude oil [17].

1 This entire chapter is reproduced from Alloy, M. M., Boube, I., Griffitt, R. J., Oris, J. T., & Roberts, A. P. (2015).

Photo‐induced toxicity of Deepwater Horizon slick oil to blue crab (Callinectes sapidus) larvae. Environ. Toxicol. Chem., 34(9), 2061-2066, with permission from John Wiley and Sons.

19

The blue crab (Callinectes sapidus) ranges along the Western Atlantic coast as far north

as Nova Scotia, and as far south as Argentina [18]. It is found along the entire coastline of the

Gulf of Mexico and is a commercially and ecologically important species [18]. The mean yearly

blue crab hardshell harvest from the Gulf of Mexico from 2007 to 2011 was 75,846 metric tons,

and the mean blue crab softshell harvest for that time period was 847 metric tons [19]. In the

course of normal reproduction, blue crab release larva, known as zoea, into the water column

[20]. For the duration of their early life stages they remain in near-surface (<3 m) water where

they are likely to encounter UV light [20]. The goal of the current study was to determine the

sensitivity of blue crab zoea to photo-induced toxicity following exposure to DWH oil. To

accomplish this goal, zoea were exposed to a range of dilutions of DWH oil and gradations of

UV (natural sunlight filtered by plastics and mesh screening). Data from this study may be used

as a component of the Natural Resource Damage Assessment (NRDA) for blue crab following

the DWH oil spill.

Methods

Test organism

Organisms were obtained from two sources, the University of Southern Mississippi Gulf

Coast Research Lab (GCRL), and the University of Maryland Center for Environmental Science.

The source laboratories held their own adult brood stock and provided early lifestage Z3/Z4

zoea for testing. Prior to use in bioassays, zoea were held overnight (25 °C, 28-30 ppt salinity)

and fed rotifers obtained from the same in-house culture as the laboratory that provided the

zoea.

Test Solutions

20

Two field-collected oil samples were used in these experiments. Slick A is a surface slick

oil collected on July 29, 2010 from the hold of barge number CTC02404, which was receiving

surface slick oil from various skimmer vessels near the Macondo Well. Slick B is a more

weathered surface slick oil collected closer to shore on July 19, 2010 by the skimmer

vessel USCGC Juniper. Each of these oil samples are routinely used in testing as part of Natural

Resource Damage Assessment (NRDA). Synthetic seawater used as control/dilution water was

made by mixing BioSea Marine Mix (AquaCraft, Hayward, CA) with Milli-Q water to 28-30 ppt

salinity, 46 mS conductivity, and 8.0-8.4 pH.

Stocks of high energy water accommodated fraction (HEWAF) of Slick A or Slick B were

prepared by mixing a known volume of water with a mass of slick oil in a Waring C15 blender

(Stamford, Connecticut), on low power, for thirty seconds. The mixture was transferred to a

separatory funnel and allowed to settle in darkness for one hour before the lower portion was

drained for PAH analysis and utilization in test dilutions. Stocks of chemically enhanced water

accommodated fractions (CEWAF) using Slick A were prepared by known volume of water with

a mass of slick oil and dispersant (Corexit 9500) in a 10:1 ratio. The solution was stirred at a rate

sufficient to produce a vortex equal to 25% of solution height by a Teflon bar in a 4 L aspirator

bottle for at least 18 h prior to settling for one hour in darkness. The lower portion of the

solution was then drained for PAH analysis and utilization in test dilutions. Test dilutions were

prepared by dilution using synthetic seawater to a nominal percentage of HEWAF or CEWAF.

Samples of stocks and dilutions were taken with every preparation and shipped (4°C) to

ALS Environmental (Kelso, Washington) for analysis. PAH analytes were quantified using gas

chromatography mass spectroscopy in single ion monitoring mode (GC/MS-SIM), based on EPA

21

method 8270D [21]. The sum concentrations of 50 PAH analytes are reported and hereafter

referred to as tPAH50. Results of tPAH50 analyses for each test stock/renewal are reported in

Table 1.

Toxicity Tests

Zoea were exposed to a range of PAH concentrations and UV intensities in a factorial

design for 48 h. Exposures were conducted in 250 mL borosilicate glass crystallizing dishes

containing 10 zoea per dish. Two pairs of bioassays were conducted; (1) a Slick A HEWAF test

and a Slick A CEWAF test were conducted in parallel using Mississippi blue crab to examine

differences in toxicity related to WAF preparation, and (2) a Slick A HEWAF test and a Slick B

HEWAF test were conducted in parallel using Maryland blue crab (Mississippi crab were not

available) to examine differences in toxicity related to source oil. The exposure period for the

HEWAF/CEWAF pairing (Test 1) included an approximately 17 h equilibration period (i.e.,

exposure to WAF in darkness) followed by a morning WAF renewal, a 7 h solar exposure, an

overnight recovery period (17 h) followed by a second WAF renewal, a second 7 h solar

exposure, and an overnight recovery period (17 h). At the end of this 48h total exposure

period, mortality was assessed. The exposure regime for the Slick A/B HEWAF pairing (Test 2)

was similar, except that test media was renewed only once (at the end of the first solar

exposure) and mortality was assessed immediately after the second solar exposure, forgoing

the second overnight recovery period. Each bioassay had five PAH treatments with three to five

replicate dishes per PAH treatment. Tests conducted with Mississippi blue crab were carried

out using two UV treatments (10% and 100% ambient sunlight). Tests conducted with Maryland

blue crab were conducted using three UV treatments (10%, 50%, and 100% ambient sunlight).

22

Replicate dishes were suspended in an outdoor, flow-through water bath to maintain

temperature. Dishes were floated in polystyrene foam insulation board with holes cut to hold

replicate dishes in contact with the temperature bath water across the underside, and the

majority of the sidewall. Sunlight was used as the source of UV radiation. Screening materials

OP-4 (Professional Plastics, Fullerton, CA) plastic sheet transparent to UV (>90%) was used for a

full intensity (100% ambient) UV treatment. A metal mesh screen was added over the top of the

OP-4 plastic as an additional filter to achieve an approximately 50% ambient UV treatment. A

different formulation of acrylic, OP-3 (Professional Plastics, Fullerton, CA), plastic sheet allowed

transmittance of less than 10% of ambient UV. UV was measured continuously during the

exposures using a Biospherical PUV2500 radiometer (BioSpherical Instruments, San Diego, CA).

For each bioassay, organisms were loaded into dishes containing one of five test

dilutions and allowed to acclimate overnight in the absence of UV. After this period, dishes

were placed in the outdoor testing apparatus and exposed to approximately 7 h of sunlight,

referred to hereafter as the first solar exposure. At the end of 7 hours, dishes were taken to

shaded shelter where test solutions were renewed and survival recorded. Replicates remained

in the shelter for an overnight dark period. The following morning, test replicates were taken to

the outdoor testing apparatus for a second solar exposure. Survival was again assessed after

this second solar exposure.

Phototoxic Units

All tests were performed outdoors using ambient UV. Tests not performed in parallel

received different UV doses as a result. To account for differences in UV co-exposure, a

23

phototoxic unit was calculated using methods similar to Oris and Giesy [5]. Fourteen known

phototoxic PAHs present in the WAF preparations were used in the calculations: anthracene,

benzo[a]anthracene, benzo[e]pyrene, benzo[g,h,i]perylene, chrysene, fluoranthene, fluorene

(as well as C1 and 2), phenanthrene (as well as C1, 2, and 3), and pyrene. The aqueous

concentration of each PAH was calculated as a molar value and multiplied by its relative

photodynamic activity (RPA) compared to anthracene. The sum of each PAH’s molar

concentration multiplied by its RPA was used to calculate the sum equivalent molar

concentration of anthrancene. UV irradiances are reported as the integration of the test

duration at a resolution of one second, expressed as mW∙s/cm2. The anthracene equivalent

concentration is multiplied by the integration of the UV irradiance at λ=380nm.

Phototoxic dose =

Where ‘mW∙s’ only refers to the integration of irradiance at λ=380nm. Phototoxic units

are expressed as µM/L∙mW∙s/cm2.

Statistical Analyses

For each bioassay, a 2-factor analysis of variance (ANOVA) with a Dunnet’s post-hoc test

was used to determine differences in survival (percent survival data was arcsine-transformed to

meet ANOVA assumptions) between treatments in the statistical software JMP (version 11 SAS

Institute, Cary, NC). In each of the four ANOVAs, the factors were UV treatment, and tPAH50

concentration. Median lethal concentrations (LC50) were calculated using the drc package [22]

% WAF Initial tPAH50 1st Renewal 2nd Renewal

24

(µg/L) tPAH50 (µg/L) tPAH50 (µg/L)

Slick A HEWAF

Mississippi zoea 0% 0.03 0.45 0.16

0.08% 1.90 1.32 1.03

0.4% 8.42 6.02 4.58

2% 44.02 16.40 18.20

10% 208.95 88.48 68.75

Slick A CEWAF

Mississippi zoea 0% 0.03 0.45 0.16

1% 2.46 1.08 0.40

5% 3.74 4.71 1.94

15% 72.23 16.82 5.58

20% 76.54 28.61 8.37

Slick A HEWAF Maryland

zoea 0% 0.02 0.01 --

0.75% 14.71 13.56 --

1.50% 28.34 28.88 --

3% 54.09 58.17 --

6% 127.91 115.64 --

Slick B HEWAF Maryland

zoea 0% 0.02 0.02 --

3.50% 9.65 10.00 --

7.50% 19.03 16.11 --

15% 37.98 34.93 --

30% 71.81 66.89 --

Table 1. The nominal test dilution in percent of stock WAF (1g/L oil) and the measured tPAH50 concentration of the initial exposure and renewal solutions. “–“ indicates that there was no second renewal.

Results

Toxicity Associated with WAF Preparation

Parallel tests with Mississippi blue crab were designed to examine toxicity associated

with differences in WAF preparation (HEWAF vs CEWAF). In both the HEWAF and CEWAF

assays, organisms were exposed to two gradations of UV (100% and 10% ambient sunlight).

Organisms in the 100% ambient UV treatment were exposed to a mean intensity (± 1 SD) of

0.035 ± 0.024 mW/cm2/s and a cumulative, integrated dose of 908.2 mW·s/cm2 for the first

25

solar exposure and a mean intensity of 0.051 ± 0.026 mW/cm2/s and a cumulative dose of

1570.3 mW·s/cm2 for the second solar exposure. Organisms in the 10% ambient UV treatment

were exposed to a mean intensity of 0.004 ± 0.002 mW/cm2/s and a cumulative, integrated

dose of 90.8 mW·s/cm2 for the first solar exposure and a mean intensity of 0.005 ± 0.003

mW/cm2/s and a cumulative dose of 157 mW·s/cm2 for the second solar exposure. Both assays

also consisted of five PAH exposure concentrations. Measured tPAH50 concentrations in the

HEWAF exposure solutions were 0.03, 1.90, 8.42, 44.02, and 208.95 µg/L. Measured tPAH50

concentrations in CEWAF exposure solutions were 0.03, 2.46, 3.74, 72.23, and 76.54 µg/L.

In the HEWAF test, 100% ambient UV treatment, there were significant decreases in

survival at 44.02 µg /L compared to controls including 100% mortality in the 208.95 µg/L tPAH50

treatment (p<0.01) (Figure 1A). The phototoxic dose LC50 was 20.6 µM/L·mW∙s/cm2 (9.7-31.5

µM/L·mW∙s/cm2) in the HEWAF test. There were no significant differences in survival among

any PAH or UV treatment in the CEWAF test (p=0.07) (Figure 1B). There was insufficient

mortality to calculate a phototoxic LC50 from the CEWAF test.

Toxicity Associated with Oil Source

Parallel tests with Maryland blue crab were designed to examine differences associated

with oil samples (i.e. Slick A vs. Slick B). Because no toxicity was observed in CEWAF

preparations at the tPAH50 concentrations tested, these assays were run only with HEWAF

preparations. In both toxicity assays, organisms were exposed to one of three gradations of UV

(10%, 50%, and 100% ambient sunlight). Organisms in the 100% ambient UV treatment were

exposed to a mean intensity of 0.076 ± 0.009 mW/cm2/s and a cumulative, integrated dose of

1904.4 mW·s/cm2 for the first solar exposure, and a mean intensity of 0.044 ± 0.028 mW/cm2/s

26

and a cumulative dose of 1058 mW·s/cm2 for the second solar exposure. Organisms in the 50%

ambient UV treatment were exposed to a mean intensity of 0.038 ± 0.005 mW/cm2/s and a

cumulative, integrated dose of 952.2 mW·s/cm2 for the first solar exposure, and a mean

intensity of 0.022 ± 0.014 mW/cm2/s and a cumulative dose of 529 mW·s/cm2 for the second

solar exposure. Organisms in the 10% ambient UV treatment were exposed to a mean intensity

of 0.008 ± 0.001 mW/cm2/s and a cumulative, integrated dose of 190.4 mW·s/cm2 for the first

solar exposure, and a mean intensity of 0.004 ± 0.003 mW/cm2/s and a cumulative dose of

105.8 mW·s/cm2 for the second solar exposure. Both assays also consisted of five PAH exposure

concentrations. Measured Slick A tPAH50 concentrations in the exposure solutions were 0.02,

14.71, 28.34, 54.09, and 127.91 µg/L. Measured Slick B tPAH50 concentrations were 0.02, 9.65,

19.03, 37.98, and 71.81 µg/L.

In the Slick A test, survival was reduced by 74%, relative to controls, in the 127.91 µg/L

tPAH50 concentration in the 10% ambient UV treatment relative to controls (p=0.02). In the 50%

ambient UV treatment, significant decreases in survival were observed at tPAH50

concentrations ≥28.34 µg/L resulting in less than 10% survival (p<0.01). In the 100% ambient

UV treatment, significant decreases in survival (<20% survival) were observed in all tested

tPAH50 concentrations except the 0.02 µg/L controls (p<0.01) (Figure 2A). The phototoxic dose

LC50 was 9.5 µM/L·mW∙s/cm2 (8.1-10.8 µM/L·mW∙s/cm2) in the Slick A test.

In the Slick B test, there was no significant reduction in survival relative to controls in

the 10% ambient UV treatment. In the 50% ambient UV treatment, survival at tPAH50

concentrations ≥19.03 µg/L was less than 10%, and significantly reduced compared to control

survival (p<0.01). In the 100% ambient UV treatment, survival was significantly reduced to less

27

than 10% at all tPAH50 concentrations except the 0.02 µg/L controls (p<0.01) (Figure 2B). The

phototoxic dose LC50 was 9.9 µM/L·mW∙s/cm2 (9.4-10.5 µM/L·mW∙s/cm2) in the Slick B test.

28

Figure 1. Mean percent survival (± 1 SE) in Mississippi blue crab zoea exposed to [A] Slick A HEWAF or [B] Slick A CEWAF after two 7-h solar exposures. Asterisks indicate treatments with mean percent survival significantly different from controls within that UV treatment. Percent survival of each replicate by the log10 phototoxic dose + 1 of Mississippi blue crab zoea exposed to [C] Slick A HEWAF or [D] Slick A CEWAF.

29

Figure 2. [A] Mean percent survival (± 1 SE) of Maryland zoea exposed to [A] Slick A HEWAF or [B] Slick B HEWAF after two 7 h solar exposures. Asterisks mark treatments with mean percent survival significantly different from controls within that UV treatment. Percent survival of each replicate by the log10 phototoxic dose + 1 of Mayland blue crab zoea exposed to [C] Slick A HEWAF or [D] Slick B HEWAF.

30

Discussion

Standard laboratory-based toxicity testing underestimates acute PAH toxicity by not

accounting for interactions with natural stressors such as UV. In general, this study

demonstrates that low UV exposures required ten-fold more PAH compound to induce median

mortality compared to high UV exposures. The values generated in this study are comparable to

those reported by other investigators in studies of PAH photo-induced toxicity in marine

invertebrates [23, 24]. Concentrations of PAH which elicited significant acute mortality in the

presence of UV (0.3-12.9 µg/L) fall within the range of PAH concentrations reported in surface

waters during the Deepwater Horizon oil spill (0-84.8 µg/L) [25]. Furthermore, the measured

tPAH50 values are based on samples taken immediately after WAF preparation and do not

account for loss of PAH over time in the exposure chambers before renewal. Thus, the toxicity

values reported here are conservative and may underestimate toxicity. This demonstrates that

a UV component should be included in studies assessing the toxicity of PAHs to aquatic

organisms that may be exposed to UV in their habitats or injury to these organisms may be

greatly underestimated.

WAF preparation method had a significant effect on the occurrence of photo-induced

toxicity in blue crab larvae. The measured tPAH50 concentrations in the higher dilutions of

CEWAF fall within the range of values observed to result in significant toxicity in HEWAF

preparations. However, no significant toxicity, photo-induced or otherwise, was observed in

CEWAF preparations. This implies that the use of dispersants could reduce phototoxic hazard to

blue crab zoea. Previous investigations into dispersed oil toxicity have reported combinations of

oil and dispersant to be as toxic as oil alone [26], to reduce toxicity compared to oil-only

31

exposures [27], or to exhibit enhanced toxicity [28], as well as greater potential for

bioaccumulation [29, 30]. These previous investigations did not examine photo-induced

toxicity, and many of the test concentrations were tenfold higher or more than what was

utilized in the present study’s CEWAF assay. The highest Corexit exposure and its PAH

concentration are well below the reported NOECs in the studies cited above. This would imply

that while the addition of dispersant mediates photo-induced toxicity, the exposures in the

present study were below the threshold for non-UV acute toxicity.

The mechanism by which dispersant mediates photo-induced toxicity is unknown. Some

published models of photo-induced toxicity use PAH body burden as a predictor variable and

not waterborne PAH concentration [31]. Body burdens of PAH in larval blue crab were not

determined largely due to logistics and size and availability limitations of the test organisms.

However, a recent investigation has reported that in small bodied organisms (i.e. embryos or

larva with high surface area to body volume ratios), tissue concentrations did not significantly

improve model accuracy over water borne concentrations [31]. It is possible that dispersant

alters the bioavailability of PAH, reducing the rate of uptake [29]. Previous studies have

suggested it could increase bioaccumulation [30]. HEWAF preparations may contain more

bioavailable PAH through an as yet undetermined mechanism. It is also possible that dispersant

may absorb specific wavelengths of UV radiation, thus reducing the number of photochemical

reactions at the tissue level. Or, dispersion may enhance photolysis and loss of PAH over time in

the exposure systems. Further investigation into the mechanism by which dispersants mediate

photo-induced toxicity is warranted as dispersants were widely utilized in the Deepwater

Horizon incident, as well as many other accidental oil releases.

32

In addition to the concentration of phototoxic PAHs, toxicity assessments will need to

consider the intensity of UV radiation and its penetration in marine surface waters. In the

present study, UV radiation intensity was varied to artificially simulate surface radiation with

minimal attenuation as well as depths where UV intensity is reduced to half, and to one tenth

of surface intensities. UV radiation attenuates with depth, and the rate of attenuation is

dependent on a variety of factors (dissolved organic carbon, suspended particulates, plankton

density, etc.) that vary on a site by site basis [32]. In open water areas of the Gulf of Mexico,

UV-A Z10 (the depth at which only 10% of surface UV-A irradiance remains) can be as deep as 37

m [32]. In coastal areas, UV attenuation varies greatly, and Z10 values as deep as 5 m and as

shallow as <1 m have been reported [32]. The high rates of UV attenuation in coastal waters are

primarily caused by non-photobleached dissolved organic carbon influx and suspended

particulates from freshwater inputs [33]. However, these are subject to great stochastic

variation. In order to extrapolate the results of the toxicity tests presented here to

invertebrates in the field, assessments of field UV exposure should be considered.

We have shown that co-exposure to slick oil from the Deepwater Horizon spill and

sunlight results in photo-induced toxicity in blue crab larvae. Effects were observed well within

the range of tPAH50 concentrations reported during the spill [25] and during relatively short

exposure periods (<48 h). Toxicity was most prevalent in HEWAF oil preparations, with no

toxicity observed in CEWAF preparations at the concentrations tested. The data from this study

combined with literature reports on UV penetration in the Gulf of Mexico demonstrate that

photo-induced toxicity of PAH should be considered in natural resource damage assessments of

the Deepwater Horizon oil spill. However, predicting actual PAH exposure and UV intensities

33

experienced by blue crab zoea in the Gulf of Mexico during the spill is outside the scope of this

study. The toxicity data presented here should be an important consideration for potential

effects of the Deepwater Horizon oil spill and future oil spills.

Acknowledgements

Data presented here are a subset of a larger toxicological database that is being

generated as part of the Deepwater Horizon Natural Resource Damage Assessment, therefore,

these data will be subject to additional analysis and interpretation which may include

interpretation in the context of additional data not presented here.

This work was supported as part of the Deepwater Horizon Natural Resource Damage

Assessment (NRDA). Data presented here are a subset of a larger toxicological database that is

being generated as part of the Deepwater Horizon NRDA, therefore, these data will be subject

to additional analysis and interpretation which may include interpretation in the context of

additional data not presented in this publication. The authors thank Stratus Consulting,

particularly Jeff Morris, Heather Forth, and Claire Lay, for their contributions. Thanks to

University of Southern Mississippi Gulf Coast Research Lab, and the University of Maryland for

providing test organisms.

34

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37

CHAPTER 3

ULTRAVIOLET RADIATION ENHANCES THE TOXICITY OF DEEPWATER HORIZON OIL TO MAHI-

MAHI (CORYPHAENA HIPPURUS) EMBRYOS

Introduction

Polycyclic aromatic hydrocarbons (PAHs) are a class of organic molecules composed of

multiple benzene rings that vary widely in their chemical and toxicological properties [1]. As a

group, PAHs are lipophilic and readily bioaccumulate [2]. PAHs exert toxicity through several

mechanisms including photo-induced toxicity [3-5]. Photo-induced toxicity is a phenomenon in

which a compound exhibits increased toxicity in the presence of certain wavelengths of light

[6]. Photodynamic PAHs, including those found in crude oil [7], have been shown to result in

photo-induced toxicity in aquatic organisms [5, 8-10]. Effects of PAH photo-induced toxicity

include increased mortality [1, 8, 11-14], reduced fecundity [3], increased photo-avoidance

behavior [15], and feeding inhibition [16].

Beginning on the 20th of April, 2010, the Mobile Offshore Drilling Unit Deepwater

Horizon experienced a series of events that resulted in the sinking of the vessel and subsequent

release of oil from the wellhead until it was sealed on July 15th, 2010 [17]. We have previously

shown that oil from the Deepwater Horizon spill is phototoxic to early lifestage blue crab at PAH

concentrations within the range of those that occurred during the spill [10]. The timespan and

scale of the oil release presented a hazard to both near-shore species such as the blue crab and

open water species like the mahi-mahi. While photo-induced toxicity of PAHs to fishes has been

reported [5, 11, 15, 18-20], little is known with respect to commercially and ecologically

38

relevant pelagic fish species that likely resided in the Gulf of Mexico at the time of the spill [21,

22].

The mahi-mahi (Coryphaena hippurus) occurs in almost every ocean body lower than 30

degrees in latitude with local distributions sometimes ranging farther north and south with

warmer currents [23]. It is found throughout the Gulf of Mexico and is an important

recreational resource as well as a commercially fished species [24]. The mean combined annual

mahi-mahi commercial and recreational harvest from the Gulf of Mexico from 2000 to 2012

was 980 metric tons [24]. In the course of a single reproductive event, mahi-mahi females

release as many as 1.5 million buoyant embryos in offshore waters [25]. For the duration of

their embryonic life stage the eggs remain buoyant in surface water where they are likely to

encounter UV light [25]. Due to the relative transparency, and positive buoyancy of their

embryos during early development, we hypothesized that mahi-mahi embryos would

specifically be at risk to photo-induced toxicity following exposure to PAHs. To test this

hypothesis, mahi-mahi embryos were exposed to a range of dilutions of water accommodated

fractions (WAF) of artificially weathered source oil, or to slick oil. Exposures were carried out

under gradations of UV achieved using natural sunlight filtered by plastics and mesh screening.

Data from this study may be used as a component of the Natural Resource Damage Assessment

(NRDA) following the DWH oil spill.

Materials and Methods Test organism

Organisms were obtained from a wild cohort of mahi-mahi broodstock maintained at

the University of Miami. Spawning occurred at dawn and embryos were collected in a purpose-

39

built egg collection vessel that was attached to a recirculating broodstock maturation system.

Embryos were prepared using methods described in Stieglitz et al., 2012 [26], whereby viable

embryos were treated with a formalin solution to remove parasites prior to use in toxicological

tests.

Test Solutions

Two source oils were used in these experiments, a field collected oil and an artificially

weathered oil. The field oil was obtained on July 29, 2010 from the hold of barge number

CTC02404, which was receiving surface slick oil from various skimmer vessels near the

Macondo Well. This slick oil is routinely used in NRDA testing for the Deepwater Horizon spill.

The artificially weathered oil was obtained from the MC252 riser. Artificial weathering was

accomplished using methods modified from Carls et al [24]. Oil was heated to approximately

98°C and stirred lightly until its mass was reduced by 33 to 38 percent. Filtered, UV-sterilized

Atlantic seawater was used in control/dilution preparations. Water physico-chemical

parameters were within the following ranges, 30-34 ppt salinity, 49-55 mS conductivity, and

7.9-8.4 pH.

Stocks of high energy water accommodated fraction (HEWAF) were prepared prior to

each test by mixing a known volume of water with a mass of slick oil in a Waring CB15 blender

(Stamford, Connecticut), on low power, for thirty seconds. The mixture was transferred to a

separatory funnel and allowed to settle in darkness for one hour before sampling for PAH

analysis and utilization in test dilutions. Stocks of chemically enhanced water accommodated

fractions (CEWAF) using were prepared by known volume of water with a mass of slick oil and

dispersant (Corexit 9500) in a 10:1 ratio. In darkness, the solution was stirred by a Teflon bar at

40

sufficient speed to achieve a 25% vortex in a 2 L aspirator bottle for 18-24 h prior to settling for

one hour. The solution was then sampled for PAH analysis and utilization in test dilutions. Test

dilutions were prepared by serial dilution in synthetic seawater to a nominal percentage of

HEWAF or CEWAF. All stocks were made to the same mass ratio of oil to water.

Samples of stocks and dilutions were taken with every preparation and shipped (4°C) to

Analytical Laboratory Services (Kelso, Washington) for analysis. Fifty specific PAH analytes were

quantified using gas chromatography mass spectroscopy in single ion monitoring mode

(GC/MS-SIM), based on EPA method 8270D [28]. The sum concentrations of these 50 PAH

analytes are hereafter referred to as tPAH50.

Toxicity Tests

The investigation consisted of four separate toxicity tests in which mahi-mahi embryos

were exposed to either CEWAF derived from slick oil, CEWAF derived from artificially

weathered source oil, HEWAF derived from slick oil, or HEWAF derived from artificially

weathered source oil. This approach allowed the investigation into the possible differences in

photo-induced toxicity brought about by dispersant based WAF preparation versus mechanical

WAF preparation, and possible differences from slick oil or artificially weathered oil used in

WAF preparations.

Mahi-mahi embryos were exposed to one of a range of PAH concentrations combined

with two or three UV intensities in a factorial design for 48 h. Exposures were conducted in 250

mL borosilicate glass crystallizing dishes containing 10 embryos per dish. The exposure period

included an approximately 17 h equilibration period, a 7 h solar exposure, an overnight

recovery period (17 h), a second 7 h solar exposure, and an overnight recovery period after

41

which percent hatch in each replicate was assessed. Each toxicity test had five to six PAH

treatments with five replicate dishes per PAH treatment.

Replicate dishes were suspended in an outdoor, flow-through water bath to maintain

temperature. Dishes were floated in polystyrene foam insulation board with holes cut to hold

replicate dishes in contact with the temperature bath water across the underside, and the

majority of the sidewall. Sunlight was used as the source of UV radiation. Screening materials

were suspended over replicate dishes to achieve gradations of UV (λ = 380 nm). A specially

formulated plastic sheet >90% transparent to UV was used for a full intensity (100% ambient)

UV treatment (KNF Corporation, Tamaqua, Pennsylvania, USA). A metal mesh screen was added

over the top of the full intensity plastic as an additional filter to achieve an approximately 50%

ambient UV treatment. A different formulation of plastic sheet allowing transmission of <10%

of ambient UV (Rosco Laboratories Inc, Stamford, Connecticut, USA) was used as a control. UV

was measured continuously during the exposures using a Biospherical radiometer (BioSpherical

Instruments, San Diego, CA).

Phototoxic Units

All tests were performed outdoors using ambient sunlight as a source of UV. Tests

performed on different days received different UV doses. To account for differences in UV co-

exposure, a phototoxic unit was calculated as described previously in Alloy et al. [10] using

methods similar to Newsted and Giesy [6]. Concentrations of fourteen known phototoxic PAHs

were used in the calculations: anthracene, benzo[a]anthracene, benzo[e]pyrene,

benzo[g,h,i]perylene, chrysene, fluoranthene, fluorene (as well as C1 and 2), phenanthrene (as

well as C1, 2, and 3), and pyrene. The aqueous concentration of each PAH was calculated as a

42

molar value and multiplied by its relative photodynamic activity compared to anthracene (RPA).

This produced the sum equivalent molar concentration of anthracene. This anthracene

equivalent concentration was multiplied by the integration of the UV irradiance (λ=380nm) to

produce the phototoxic dose. UV irradiances are reported as the integration of the test

duration at a resolution of one second, expressed as mW∙s/cm2. Phototoxic units are expressed

as µM/L·mW∙s/cm2.

Statistical Analyses

For each toxicity test, hatch data was arcsine transformed and a 2-factor analysis of

variance (ANOVA) with a Dunnett’s post-hoc test was used to determine differences in percent

hatch using the statistical software JMP (version 11, SAS Institute, Cary, NC). All statistical

comparisons were made using α= 0.05. UV treatment and tPAH50 concentration were used as

factors in each ANOVA. Median effect concentrations (EC50) for phototoxic dose were

calculated using the drc package in R (version 3.1.2) [29].

A table was generated of the anthracene equivalent molar concentrations required at a

given UV dose to meet three effect concentrations was generated using the mean UV intensity

of all tests, and the calculated effect concentrations for twenty percent, fifty percent, and

eighty percent. The mean UV intensity was multiplied by an exposure duration of 16 hours. This

UV dose was used as the 100% UV exposure for the table, and was reduced accordingly for

75%, 50%, and 25% UV calculations.

Results

Slick oil CEWAF dilutions contained 0.4, 1.5, 2.7, 5.1, and 10.0 µg/L tPAH50. The total UV

dose in each exposure was calculated by integrating the total irradiance received over the

43

timecourse of the test during both solar exposure periods. The slick oil CEWAF test utilized

three gradations of UV (100%, 50%, and 10%). Slick oil CEWAF 100% UV exposures received a

total integrated dose of 3276 mW·s/cm2 with a mean intensity (+/- 1SD) of 0.053 ± 0.023

mW/cm2/s. Significant toxicity was observed at exposures ≥ 2.7 µg/L tPAH50 in the 100% UV

treatment only (p < 0.001) (Figure 1A). To account for differences in UV exposure between

tests, EC50s were calculated in phototoxic dose units for each test. The slick oil CEWAF

phototoxic EC50 was 11.8 µM/L·mW∙s/cm2 (95% confidence interval = 7.84-15.7

µM/L·mW∙s/cm2) (Figure 1B).

Slick oil HEWAF dilutions contained 0, 0.1, 0.3, 0.9, 4.3, and 20.9 µg/L tPAH50. The slick

oil HEWAF test utilized two gradations of UV (100% and 10%) only. Slick oil HEWAF 100% UV

exposures received an integrated dose of 3022 mW·s/cm2 with a mean intensity of 0.044 ±

0.022 mW/cm2/s. Significant toxicity was observed to occur at exposures ≥ 5.0 µg/L tPAH50 in

the 100% UV treatment, and at the 20.9 µg/L tPAH50 exposures in the 10% UV treatment (p <

0.001) (Figure 1C). The slick oil HEWAF phototoxic EC50 was 6.77 µM/L·mW∙s/cm2 (5.91-7.64

µM/L·mW∙s /cm2) (Figure 1D).

Weathered source CEWAF dilutions contained 0, 0.2, 0.9, 3.2, 12.9, and 49.9 µg/L

tPAH50. Both weathered source oil tests utilized two UV gradations (100% and 10%). Weathered

source CEWAF 100% UV exposures received an integrated dose of 2436 mW∙s/cm2 with a mean

intensity of 0.040 ± 0.019 mW/cm2/s. Significant toxicity was observed to occur only in the 49.9

µg/L tPAH50 exposures in the 100% UV treatment (p < 0.001) (Figure 2A). The artificially

weathered source oil CEWAF phototoxic EC50 was 14.7 µM/L·mW∙s/cm2 (no solution to 95%

confidence interval) (Figure 2B).

44

Weathered source HEWAF dilutions contained 0, 2.0, 4.0, 6.7, 18.6, 67.9 µg/L tPAH50.

Weathered source HEWAF 100% UV exposures received an integrated dose of 1754 mW·s/cm2

with a mean intensity of 0.031 ± 0.014 mW/cm2/s. Significant toxicity was observed at

exposures ≥ 4.0 µg/L tPAH50 in the 100% UV treatment, and ≥ 18.6 µg/L tPAH50 in the 10% UV

treatment (p < 0.001) (Figure 2C). The artificially weathered source oil HEWAF phototoxic EC50

was 2.16 µM/L·mW∙s/cm2 (1.52-2.79 µM/L·mW∙s/cm2) (Figure 2D).

45

Figure 1. Hatching success in mahi-mahi embryos exposed to field collected slick oil (A) CEWAF by tPAH50 concentration and UV

treatment, (B) CEWAF by phototoxic dose. (C) HEWAF by tPAH50 concentration and UV treatment, and (D) HEWAF by phototoxic

dose. Errors bars represent 1SE. Asterisks indicate treatments with mean percent hatch significantly different from controls.

46

Figure 2. Hatching success in mahi-mahi embryos exposed to artificially weathered source oil (A) CEWAF by tPAH50 concentration

and UV treatment, (B) CEWAF by phototoxic dose. (C) HEWAF by tPAH50 concentration and UV treatment, and (D) HEWAF by

phototoxic dose. Errors bars represent 1SE. Asterisks indicate treatments with mean percent hatch significantly different from

controls.

47

Discussion

In all four toxicity tests, toxicity as decreased hatching success occurred in a PAH and UV

dependent manner. Normal hatching time for mahi-mahi at 25°C is approximately 40 hours

post fertilization [30]. This suggests that embryos that did not hatch within a normal timeframe

(up to 48hrs in this study) are developmentally delayed following co-exposure to PAH and solar

radiation. This likely affects their potential for survival and recruitment compared to control

embryos. Concentrations of tPAH50 observed to result in photo-induced toxicity in these

experiments are well within the range of concentrations (0-84.8 µg/L) reported in the spill area

during the active phase of the 2010 Deepwater Horizon spill [31]. Measured tPAH50 values in

this study are based on samples taken immediately after WAF preparation and do not account

for loss of PAH over time in the exposure chambers. Exposure solutions were held static for the

duration of the toxicity test. Thus, the toxicity values reported here are likely conservative and

time-integrated effect values are likely lower.

Concentrations of PAH observed to result in toxicity in this study are also in agreement

with other literature-reported values using oil and marine species. Barron and Ka’aihue [9]

report an LC50 of 30 µg/L total PAH in a 96 h UV and PAH co-exposure to pacific Herring

embryos. Sellin-Jeffries et al. [19] induced significant mortality after 48 h at 15 µg/L anthracene

in pacific herring young of the year in field UV exposure tests, and at 5 µg/L anthracene in

laboratory tests using larvae and artificial UV. Duesterloh et al. [32] report a LC20 of

approximately 2 µg/L PAH in a single 4 h UV and PAH co-exposure to a marine calanoid

copepod. Oil exposures without a UV component report much higher median effect values.

Singer et al. [33] reported a range of LC50s (16.3-40.2 mg/L) in a series of larval topsmelt

48

exposures to crude oil. Pollino et al. [34] reported a range of LC50s (1.28 mg/L WAF, 14.5 mg/L

dispersant only, and 1.37 mg/L WAF and dispersant) in a series of WAF, dispersant, and WAF

with dispersant exposures to day-of-hatch rainbowfish larvae. Interestingly, the range of PAH

concentrations observed in this study to be acutely phototoxic under natural sunlight to

embryonic mahi are similar to concentrations observed to result in sublethal effects in pelagic

marine fish without UV. Mager et al. [21] reported significantly reduced critical swimming

speed in juvenile mahi-mahi that were exposed once, as embryos, to concentrations as low as

1.8 µg/L tPAH50. Incardona et al. [22] published similar findings, reporting pericardial edema

EC50s as low as 0.8 µg/L in tuna and 12.4 µg/L in amberjack.

Toxicity Associated with WAF Preparation Method

There was a significant reduction in photo-induced toxicity comparing CEWAF to

HEWAF from the same source oil. Dilutions of CEWAF made using artificially weathered oil were

approximately sevenfold less phototoxic than HEWAF counterparts. Similarly, CEWAF made

from more weathered slick oil was approximately threefold less phototoxic than slick oil

HEWAF. This implies that the reduction of photo-induced toxicity in CEWAFs could be related to

the degree of weathering the oil has undergone prior to addition of dispersant.

We have reported a similar finding in a previous study conducted using blue crab zoea

co-exposed to UV and WAF [10]. In that study, photo-induced toxicity was also reduced in

CEWAF preparations compared to HEWAF preparations. Investigations into dispersed oil

toxicity have reported preparations of oil and dispersant to be as toxic as oil alone [35], to

reduce toxicity compared to oil-only exposures [36], or to exhibit greater toxicity [37]. Reports

of increased bioaccumulation of PAHs with dispersant use would also be expected to increase

49

photo-induced toxicity [38, 39]. However, none of the cited studies investigated photo-induced

toxicity specifically as a mechanism. Dispersants may mediate photo-induced toxicity by some

other mechanism than altered bioavailability or bioaccumulation. Dispersants may have some

unknown effect on irradiation by UV, thus reducing the number of photochemical reactions at

the tissue level. Alternatively, dispersion may enhance photolysis and loss of PAH over time in

the exposure systems. Further investigation into the mechanism by which dispersants mediate

photo-induced toxicity is warranted as dispersants were widely utilized in the Deepwater

Horizon incident, as well as many other accidental oil releases.

Toxicity Associated with Oil Source

Oil source significantly influenced toxicity in HEWAF preparations as indicated by lack of

overlap in the phototoxic EC50 95% confidence intervals. HEWAF prepared from artificially

weathered source oil was more phototoxic (twofold) than HEWAF prepared from Slick A source

oil. This could be explained by analysis of the photodynamic PAH content of each preparation.

HEWAF stock prepared with artificially weathered source oil was calculated to have more than

double the anthracene equivalent of HEWAF stock prepared with slick oil, resulting in increased

toxic effect (Table 1). Interestingly, this could not be used to explain the lack of difference in

CEWAF. CEWAF stock prepared with artificially weathered source oil contained more than

twenty times the concentration of photodynamic PAHs than the CEWAF stock prepared with

slick oil. However, there was no difference in toxicity observed between the two. Furthermore,

CEWAF reduced toxicity on a tPAH50 basis when compared to HEWAF preparations of the same

source oil. We have previously reported similar findings in photo-induced toxicity tests with

blue crab zoea [10]. Taken together, these findings suggest that the mechanism by which

50

dispersant mediates photo-induced toxicity may be independent of photodynamic PAH

concentration.

The positive buoyancy of mahi-mahi embryos removes significant UV attenuation by the

water column as a factor in possible exposure scenarios. However, a variety of weather factors

can reduce the amount of UV that reaches the surface. Table 2 illustrates how photochemical

reciprocity shifts the concentration of PAH needed to produce the same phototoxic dose as UV

is reduced. Even with the lowest UV given in the table –requiring fourfold more PAH– the

concentrations calculated to for the EC20s range from 1.16-19.2 nM/L anthracene equivalent.

The latter of which is slightly less than the photodynamic PAH concentration in the highest

artificially weathered source HEWAF exposure in the present study. Thus, even under relatively

low UV conditions significant hatch failure could occur at the higher PAH concentrations

reported during the spill.

This study demonstrates that a UV component is important to a complete

understanding of the toxicity of PAHs to aquatic organisms, or hazard may be greatly

underestimated. Mahi-mahi and other pelagic fish species with similar life history traits may be

particularly vulnerable to photo-induced toxic effects. Their embryos are buoyant and

transparent, and newly hatched larvae remain close to the surface. This places them at risk for

UV exposure during all early life stages. Toxicity tests using Deepwater Horizon oil, but without

a UV component, have reported sublethal effects at low PAH concentrations (0.8-18.2 µg/L) in

mahi-mahi and other pelagic predatory fishes [21,22]. Effect concentrations determined in the

absence of UV may underestimate the toxicity of oil. It is possible that co-exposure to UV may

induce or exacerbate those sublethal effects at even lower concentrations of PAH.

51

Here, we have shown that co-exposure to slick oil from the Deepwater Horizon spill and

sunlight results in photo-induced toxicity in mahi-mahi embryos. Effects were observed well

within the range of tPAH50 concentrations reported during the spill and during relatively short

exposure periods (48 h) using WAF preparations both with and without dispersants. This study

demonstrates that photo-induced PAH toxicity likely contributed to overall toxic effects from

the Deepwater Horizon oil spill on early lifestage fish.

WAF Prep Type Oil type µM/L ANT equivalent

HEWAF Slick 0.79

HEWAF Artificially weathered source 1.75

CEWAF Slick 0.05

CEWAF Artificially weathered source 1.28

Table 1. Calculated anthracene equivalent concentrations from the four stock WAF

preparations. Concentrations presented in micromoles of anthracene equivalent per liter.

Integrated UV Dose

mWs/cm2

EC20 (ANT Equivalents)

nM/L (95% CI)

EC50 (ANT Equivalents)

nM/L (95% CI)

EC80 (ANT Equivalents) nM/L (95% CI)

Slick A 2423 1.68 (1.28-2.07) 2.79 (2.44-3.15) 4.66 (3.49-5.83)

HEWAF 1812 2.24 (1.72-2.77) 3.74 (3.26-4.22) 6.23 (4.66-7.79)

1214 3.35 (2.56-4.13) 5.58 (4.87-6.29) 9.29 (6.96-11.6)

607 6.70 (5.12-8.27) 11.2 (9.74-12.6) 18.6 (13.9-23.3)

WS 2423 0.289 (0.130-0.448) 0.891 (0.627-1.15) 2.74 (1.54-3.94)

HEWAF 1812 0.387 (0.174-0.600) 1.19 (0.839-1.54) 3.67 (2.06-5.27)

1214 0.578 (0.259-0.895) 1.78 (1.25-2.30) 5.48 (3.08-7.87)

607 1.16 (0.519-1.79) 3.56 (2.50-4.60) 11.0 (6.16-15.7)

Slick A 2423 1.95 (1.25-2.64) 4.87 (3.24-6.48) 12.1 (3.88-20.3)

CEWAF 1812 2.60 (1.67-3.53) 6.51 (4.33-8.66) 16.2 (5.19-27.1)

1214 3.89 (2.50-5.27) 9.72 (6.46-12.9) 24.1 (7.75-40.5)

52

607 7.78 (4.99-10.5) 19.4 (12.9-25.9) 48.3 (15.5-81.0)

WS 2423 4.81 (2.09-7.51) 6.06 (NA -13.7) 7.63 (NA-22.6)

CEWAF 1812 6.43 (2.80-10.0) 8.10 (NA-18.3) 10.2 (NA-30.2)

1214 9.60 (4.18-15.0) 12.1 (NA-27.3) 15.2 (NA-45.1) 607 19.2 (8.35-30.0) 24.2 (NA-54.5) 30.5 (NA-90.3)

Table 2. Calculated anthracene equivalent concentrations at the EC20, EC50, and EC80 and their

95% confidence intervals for four UV doses.

Acknowledgements

Data presented here are a subset of a larger toxicological database that is being

generated as part of the Deepwater Horizon Natural Resource Damage Assessment, therefore,

these data will be subject to additional analysis and interpretation which may include

interpretation in the context of additional data not presented here.

This work was supported as part of the Deepwater Horizon Natural Resource Damage

Assessment (NRDA). Data presented here are a subset of a larger toxicological database that is

being generated as part of the Deepwater Horizon NRDA, therefore, these data will be subject

to additional analysis and interpretation which may include interpretation in the context of

additional data not presented in this publication. The authors thank Stratus Consulting,

particularly Jeff Morris and Claire Lay, for their contributions.

53

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57

CHAPTER 4

SENSITIVITY OF EARLY LIFESTAGE RED DRUM (SCIAENOPS OCELLATUS) AND SPECKLED

SEATROUT (CYNOSCION NEBULOSUS) TO PHOTO–INDUCED TOXICITY OF NATURALLY

WEATHERED DEEPWATER HORIZON OIL

Introduction

Polycyclic aromatic hydrocarbons (PAHs) are a class of organic molecules with multiple

carbon rings. As a class, PAHs vary widely in toxicological and chemical properties [1]. In

general, PAHs are lipophilic [2]. PAHs exert toxicity through several mechanisms, however,

photo-induced or photo-enhanced toxicity often occurs at concentrations below the threshold

of other modes of toxicity [3-5]. Photo-induced toxicity is the phenomenon in which a

compound exhibits increased toxicity in the presence of certain wavelengths of light [5]. Effects

of PAH photo-induced toxicity include increased mortality [6-11], reduced fecundity [3],

delayed hatching [12] increased photo-avoidance behaviors [13, 5], and feeding inhibition [14].

The photo-induced toxicity of PAHs to aquatic species is well documented in the literature [5, 6,

11, 15-17].

On the 20th of April, 2010, the mobile offshore drilling unit Deepwater Horizon

experienced a series of events that resulted in the sinking of the vessel itself and subsequent

release of millions of barrels of oil from the wellhead unit until it was sealed on July 15th, 2010

[18]. Crude oil contains PAHs including those with photodynamic activity [19]. The Exxon Valdez

and Prestige oil spills are two examples where releases of oil resulted in potential photo-

induced PAH toxicity to early lifestage finfish [11,16, 20], shellfish [9, 12, 21], and marine

zooplankton [21].

58

Red drum (Sciaenops ocellatus) are found along the southern Atlantic and Gulf of

Mexico coasts [22]. It is an important commercial and recreational species. Commercial fishing

of red drum peaked in 1986 with nearly 15 million pounds caught [23]. Concerns about stock

depletion led to responses at the state and federal level to protect red drum populations,

including Executive Order 13449 in 2007, limiting the catch of red drum in federal waters [24].

Recreational fishing of red drum has increased since 1986 from 5 million pounds to more than

15 million pounds harvested in 2008 [23]. The mean catch between 2002 through 2011 for Gulf

of Mexico red drum was 8.9 million fish [23].

Red drum spawn in the open water neritic zone beginning in August and continuing into

December, however, regional variation in timing may shift this window by several weeks [22].

The number of eggs released per spawning event is dependent on female body size, varying in

laboratory conditions from two hundred thousand to in excess of two million eggs per female

[22]. Over the course of a single spawning season, a female may spawn every 2-4 days [25]. Red

drum larvae migrate with the tide into estuaries and settle in the cover of submerged

seagrasses [22].

Similarly, speckled seatrout (Cynoscion nebulosus) are distributed as far north as Virginia

and found along the shores of all Gulf of Mexico states [26]. The mean catch, combining harvest

and release, between 2002 through 2011 for Gulf of Mexico speckled seatrout was 29.9 million

fish [26]. Overfishing led to management responses and bans on several fishing methods in

order to allow for stocks to recover. Speckled seatrout spawn from April through October but,

again, this window of time varies regionally based on local conditions [26]. Speckled seatrout

spawn in a variety of habitats; they have been observed to spawn over submerged seagrass,

59

among floating masses of detritus, and in tidal areas with little or no vegetation present [26].

Spawning typically occurs in relatively shallow waters (3-4.6 m) [26]. Although older, larger

females spawn more frequently, over the course of a single season, a female may spawn every

2-9 days and release >1,000 eggs/g bodyweight [26]. Speckled seatrout larvae typically hatch in

estuarine habitat very close to their spawning area [26]. As larvae, they remain associated with

structure such as seagrass until they reach the juvenile life stage [26].

In both red drum and speckled seatrout, the eggs are buoyant, and larvae hatch in near

the surface where they are likely to encounter UV light. The goal of this study was to determine

the sensitivity of red drum and speckled seatrout larvae to photo-induced toxicity following

exposure to DWH spill oil. To accomplish this goal, larvae of both species were exposed to a

range of dilutions of two different field-collected oils and three gradations of UV. Data from this

study may be used as a component of the Deepwater Horizon Natural Resource Damage

Assessment (NRDA).

Materials and Methods

Test Organisms

Organisms were obtained from in house cultures at the Texas Parks and Wildlife Sea

Center (Lake Jackson, Texas, USA). Two species were tested, red drum and speckled seatrout.

Prior to use in bioassays, embryos were held until hatch (approximately 18-24 hpf). All

organisms used in tests were less than 48 hpf at the commencement of testing.

Test Solutions

Two oil samples collected in the field under chain of custody were used in these

experiments. Slick A is a surface slick oil collected on July 29, 2010 from the hold of barge

60

number CTC02404, which was receiving surface slick oil from various skimmer vessels

responding to the spill. Slick B is a more weathered surface slick oil collected on July 19, 2010 by

the skimmer vessel USCGC Juniper. Both of these oil samples are routinely used in testing as

part of the DWH NRDA. Filtered seawater from the Sea Center Hatchery (collected from a gulf

channel) was used in all control solutions and in preparation of treatment dilutions.

Stocks of high energy water accommodated fraction (HEWAF) of Slick A or Slick B were

prepared by mixing a known volume of water with a mass of slick oil (1g/L) in a Waring CB15

blender (Stamford, Connecticut), on low power, for thirty seconds. The mixture was transferred

to a separatory funnel and allowed to settle for one hour in darkness before sampling for PAH

analysis and utilization in test dilutions as described in Alloy et al. [11, 12]. Test dilutions were

prepared by dilution in filtered seawater to a nominal percentage of HEWAF. Samples of stocks

and dilutions were taken with every preparation and shipped (4°C) to ALS Environmental (Kelso,

Washington) for analysis. Fifty PAH analytes being used for NRDA analyses were quantified

using gas chromatography mass spectroscopy in single ion monitoring mode (GC/MS-SIM),

based on EPA method 8270D [27]. The sum concentrations of these 50 PAH analytes are

hereafter referred to as tPAH50.

Toxicity Tests

Larvae were exposed to one of a range of PAH concentrations combined with one of a

range of UV intensities in a fully factorial design for approximately 20 h. Exposures were

conducted in 250 mL borosilicate glass crystallizing dishes containing 10 larvae per dish. The

exposure period included an approximately 8 h equilibration period (under indoor low intensity

lighting), followed by a 7-h solar exposure, followed by a 5-h no-UV period before mortality was

61

assessed. Each bioassay had five PAH treatments with five replicate dishes per PAH treatment.

Tests were conducted with three UV (λ= 380nm) treatments (nominally 10%, 50%, and 100%

ambient natural sunlight).

Replicate dishes were suspended in an outdoor, flow-through water bath to maintain a

target temperature of 28°C. Dishes were floated in polystyrene foam insulation board with

holes cut to hold replicate dishes in contact with the temperature bath water across the

underside, and the majority of the sidewall. Sunlight was used as the source of UV radiation.

Screening materials were suspended over selected dishes to achieve treatment gradations of

UV. A specially formulated plastic sheet transparent to UV was used for a full intensity (100%

ambient) UV treatment (Professional Plastics, Fullerton, CA, USA). A metal mesh screen was

added over the top of the full intensity plastic as an additional, neutral density filter to achieve

an approximately 50% ambient UV treatment. A different formulation of plastic sheet allowing

transmission of < 10% of ambient UV (Professional Plastics, Fullerton, CA, USA) was used as a

control. UV was measured continuously during the exposures using a Biospherical BIC2104R

radiometer (BioSpherical Instruments, San Diego, CA).

Phototoxic Units

All tests were performed outdoors using ambient UV. Tests not performed in parallel

received different UV doses as a result. To account for differences in UV exposure, a phototoxic

unit was calculated using methods similar to Oris and Giesy [5] and previously described in Alloy

et al. (2015) [12]. Briefly, fourteen known phototoxic PAHs present in the WAF preparations

were used in the calculations: anthracene, benzo[a]anthracene, benzo[e]pyrene,

benzo[g,h,i]perylene, chrysene, fluoranthene, fluorene (as well as C1 and 2), phenanthrene (as

62

well as C1, 2, and 3), and pyrene. The aqueous concentration of each PAH was calculated as a

molar value and multiplied by its relative photodynamic activity (RPA) compared to anthracene

[5]. The sum of each of these values was used to calculate the sum concentration of anthracene

equivalents. UV doses are reported as the integration of the test duration at a resolution of one

second, expressed as mW∙s/cm2. The anthracene equivalent concentration is multiplied by the

integration of the UV irradiance at λ=380nm to produce the phototoxic dose.

Phototoxic dose =

Where ‘mW∙s’ refers to the integration of irradiance only at λ=380nm. Phototoxic units

are expressed as µM/L·mW∙s/cm2.

Statistical Analyses

For each toxicity test, percent survival data was arcsine-transformed to meet ANOVA

assumptions. A 2-factor analysis of variance (ANOVA) with a Dunnett’s post-hoc test was used

to determine differences in survival between treatments using UV treatment and tPAH50

concentration as factors. ANOVAs were performed using the statistical software JMP (version

11 SAS Institute, Cary, NC). Statistical significance was determined using α= 0.05. Median lethal

concentrations (LC50) for whole-test phototoxic dose, and individual UV treatments within

toxicity tests were calculated using the drc package in R (version 3.1.2) [28].

A table was generated of the anthracene equivalent molar concentrations required at a

given UV dose to meet three effect concentrations was generated using the mean UV intensity

of all tests, and the calculated lethal concentrations for twenty percent, fifty percent, and

eighty percent. The mean UV intensity was multiplied by an exposure duration of 7 hours. This

63

UV dose was used as the 100% UV exposure for the table, and was reduced accordingly for

75%, 50%, and 25% UV calculations.

Results

Red drum larvae were exposed to Slick A HEWAF dilutions containing 0.00, 1.46, 3.13,

5.67, and 11.76 µg/L tPAH50. Red drum exposed to the 100% UV treatment received a total

integrated dose of 705.79 mW•s/cm2 with a mean intensity (± 1SD) of 0.038 ± 0.021

mW/cm2/s. Mean survival of all treatments is detailed in Figure 1A. Significant mortality

occurred in the 100% UV treatment following exposure to tPAH50 concentrations ≥3.13 µg/L

(ANOVA, F(14,59)=11.42, Dunnett’s post-hoc p< 0.01). Median mortality tPAH50 concentrations

(LC50s) were calculated separately by UV treatment. The tPAH50 LC50 for the 50% UV and the

100% UV treatments were 12.2 µg/L tPAH50 (95% confidence interval = 7.37-17.0 µg/L tPAH50)

and 3.42 µg/L tPAH50 (95% confidence interval = 2.47-4.37 µg/L tPAH50). A model using

phototoxic units was also used to calculate a phototoxic LC50. The phototoxic LC50 for Slick A

HEWAF to red drum larvae was 1.41 µM/L•mW•s/cm2 (95% confidence interval = 1.17-1.65

µM/L•mW•s/cm2) (Figure 1B).

Red drum larvae were exposed to Slick B dilutions containing 0.00, 2.27, 5.11, and 8.25

µg/L tPAH50. 100% UV exposures of red drum larvae exposed to Slick B HEWAF received a total

integrated dose of 846.9 mW•s/cm2 with a mean intensity of 0.032 ± 0.015 mW/cm2/s. Mean

survival in all treatments is detailed in Figure 2A. Significant mortality occurred in the 50% and

100% UV treatments following exposure to tPAH50 concentrations ≥5.11 µg/L (ANVOA,

F(11,48)=80.47, Dunnett’s post-hoc p< 0.01) and ≥2.27 µg/L, respectively (Dunnett’s post-hoc p<

0.01). The tPAH50 LC50 in the 50% and 100% UV treatments were 5.79 µg/L tPAH50 (95%

64

confidence interval = 5.45-6.14 µg/L tPAH50) and 1.99 µg/L tPAH50 (95% confidence interval =

1.78-2.19 µg/L tPAH50), respectively. The phototoxic LC50 for Slick B to red drum larvae was 1.06

µM/L•mW•s/cm2 (95% confidence interval = 0.993-1.12 µM/L•mW•s/cm2) (Figure 2B).

Speckled seatrout larvae were exposed to Slick A dilutions which contained 0.00, 0.25,

0.43, 1.02, and 2.49 µg/L tPAH50. Speckled seatrout larvae in the 100% UV treatment exposed

to Slick A HEWAF received a total integrated dose of 1108.9 mW•s/cm2 with a mean intensity

of 0.042 ± 0.019 mW/cm2/s. Mean survival in all treatments is detailed in Figure 3A. Significant

mortality occurred at tPAH50 concentrations ≥2.40 µg/L in the 50% UV treatment (ANOVA,

F(14,60)=24.89, Dunnett’s post hoc p< 0.01), and ≥0.43 µg/L in the 100% UV treatment (Dunnett’s

post hoc p< 0.01). The tPAH50 LC50 in the 50% and 100% UV treatments was 2.41 µg/L tPAH50

(95% confidence interval = 0.741-4.07 µg/L tPAH50) and 0.827 µg/L tPAH50 (95% confidence

interval = 0.626-1.03 µg/L tPAH50), respectively. The phototoxic LC50 for Slick A to speckled

seatrout larvae was 0.516 µM/L•mW•s/cm2 (95% confidence interval = 0.478-0.553

µM/L•mW•s/cm2) (Figure 3B).

Speckled seatrout larvae were exposed to Slick B dilutions containing 0.00, 0.08, 0.18,

0.56, and 1.63 µg/L tPAH50. Speckled seatrout larvae in the 100% UV treatment exposed to

Slick B HEWAF received a total integrated dose of 1439.73 mW•s/cm2 with a mean intensity of

0.056 ± 0.019 mW/cm2/s. Mean survival in all treatments is detailed in Figure 4A. Significant

mortality occurred at 1.63 µg/L tPAH50 in the 50% UV treatment (ANOVA, F(14, 60)=26.53,

Dunnett’s post-hoc p< 0.01), and tPAH50 concentrations ≥0.18 µg/L in the 100% UV treatment

(Dunnett’s post-hoc p< 0.01). The tPAH50 LC50 in the 50% and 100% UV treatments were 1.01

µg/L tPAH50 (95% confidence interval = 0.802-1.21 µg/L tPAH50) and 0.182 ug/L tPAH50 (95%

65

confidence interval = 0.159-0.206 µg/L tPAH50), respectively. The phototoxic LC50 was 0.287

µM/L•mW•s/cm2 (95% confidence interval = 0.239-0.335 µM/L•mW•s/cm2) (Figure 4B).

66

Figure 1. [A] Mean percent mortality ±1 SE of larval red drum exposed to slick A HEWAF and one of three UV treatments. Asterisks indicate significant difference from control treatment of the same UV (α=0.05). [B] Phototoxic dose-response model of mortality in larval red drum exposed to slick A HEWAF and UV.

67

Figure 2. [A] Mean percent mortality ±1 SE of larval red drum exposed to slick B HEWAF and one of three UV treatments. Asterisks indicate significant difference from control treatment of the same UV (α=0.05). [B] Phototoxic dose-response model of mortality in larval red drum exposed to slick B HEWAF and UV.

68

Figure 3. [A] Mean percent mortality ±1 SE of larval speckled seatrout exposed to slick A HEWAF and one of three UV treatments. Asterisks indicate significant difference from control treatment of the same UV (α=0.05). [B] Phototoxic dose-response model of mortality in larval speckled seatrout exposed to slick A HEWAF and UV.

69

Figure 4. [A] Mean percent mortality ±1 SE of larval speckled seatrout exposed to slick B HEWAF and one of three UV treatments. Asterisks indicate significant difference from control treatment of the same UV (α=0.05). [B] Phototoxic dose-response model of mortality in larval speckled seatrout exposed to slick B HEWAF and UV.

70

Discussion

Co-exposure to natural sunlight increased the toxicity of DWH spill oil to larval red drum

and speckled seatrout in a tPAH50 and UV dose dependent manner. All calculated tPAH50 LC50s

(0.18-5.78 µg /L) in the present study are within the lower range of PAH concentrations (0-84.8

µg /L) reported in the DWH spill area [29]. Other photo-induced PAH toxicity tests, including

those using DWH spill oil, report total PAH EC50 or LC50 values that are within an order of

magnitude of those observed in the red drum and speckled seatrout tests [11, 12, 16, 20]. It is

worth noting that the tPAH50 values in this study are based on analysis of chemistry samples

taken immediately after WAF preparation and do not account for loss of PAH over time in the

actual exposures. Given that exposure solutions were not renewed for the duration of the

bioassay, the toxicity values reported here are likely higher than if we calculated them using

time-integrated doses due to the reduction in PAH concentration in the exposure solutions over

time.

The phototoxic LC50s (0.25-1.13 µM/L•mW•s/cm2) are substantially lower than those

reported in two similar studies on early lifestage mahi-mahi and blue crab [11, 12]. The

reported EC50s (phototoxic dose) for effects on hatching success in embryonic mahi-mahi were

between 6.8-11.8 µM/L•mW•s/cm2 (Slick A only). Median lethal concentrations (phototoxic

dose) reported for blue crab zoea were 9.5-20.6 µM/L•mW•s/cm2 (Slick A), and 9.9

µM/L•mW•s/cm2 (Slick B). While it is difficult to compare across taxa and lifestage, these data

suggest that red drum and speckled seatrout, both members of the drum family (Sciaenidae),

may be among the more sensitive test organisms to oil photo-induced toxicity. It is unlikely that

UV-sensitivity alone explains this result as survival in 10%, 50%, and 100% UV controls (no PAH)

71

was >90% in most cases. Regardless of mechanism, the sensitivity of larval red drum and

speckled seatrout lowers the range at which photo-induced PAH toxicity is observed among

Gulf of Mexico species.

Both red drum and speckled seatrout are estuarine species. This makes likely exposure

scenarios more complex than species that both spawn and live exclusively in open waters.

Immediately after hatching, red drum and speckled seatrout larvae at a similar age as those

used in this study can be found in both open and nearshore waters [25, 26]. In open water

areas of the Gulf of Mexico, UV-A Z10 (the depth at which only 10% of surface UV-A irradiance

remains) can be as deep as 37 m [31]. This suggests that, in open water areas, the likelihood for

co-exposure to PAH and UV is quite high. In coastal areas, UV attenuation varies greatly, and Z10

values as deep as 5 m and as shallow as <1 m have been reported [31]. Thus, the potential for

larvae to be exposed to UV in habitats where oil was trapped among seagrasses, salt marsh

vegetation, and sediments [30] may be more site- and season-specific. However, the sensitivity

of red drum and speckled seatrout larvae to oil photo-induced toxicity may result in significant

effects even when UV exposure is relatively low. For example, table 1 illustrates how

photochemical reciprocity shifts the concentration of PAH needed to produce the same

phototoxic dose as UV is reduced. The PAH doses required for median mortality under the

lowest UV given range from approximately 1.4-3.9 nM/L anthracene equivalent Thus, even

under relatively low UV conditions significant mortality could occur at the higher PAH

concentrations reported during the spill.

72

Co-exposure to UV as natural sunlight greatly increases the toxicity of DWH spill oil to

larval red drum and speckled seatrout. Effects were observed during short test durations (<24

h) and well within the range of tPAH50 concentrations reported during the spill. Data presented

here suggests that these two species may be among the most sensitive tested to date to oil

photo-induced toxicity. Significant toxicity was observed even under reduced (50% ambient) UV

intensity indicating that effects in the field may occur in more turbid, nearshore coastal waters

as well as highly transparent open water.

Integrated UV Dose LC20 (ANT Equivalents) LC50 (ANT Equivalents) LC80 (ANT Equivalents)

µWs/cm2 nM/L nM/L nM/L

RD 1065 0.279 (0.211-0.348) 0.988 (0.803-1.17) 3.49 (2.27-4.72) Slick A 800 0.372 (0.281-0.463) 1.31 (1.07-1.56) 4.65 (3.03-6.28)

533 0.558 (0.422-0.695) 1.97 (1.61-2.34) 6.98 (4.54-9.42) 266 1.12 (0.845-1.39) 3.96 (3.22-4.69) 14.0 (9.10-18.9)

RD 1065 0.622 (0.550-0.694) 0.951 (0.864-1.04) 1.45 (1.28-1.63) Slick B 800 0.828 (0.733-0.924) 1.27 (1.15-1.38) 1.93 (1.70-2.16)

533 1.24 (1.10-1.39) 1.90 (1.73-2.07) 2.90 (2.56-3.25) 266 2.49 (2.20-2.78) 3.81 (3.46-4.15) 5.82 (5.13-6.51)

SST 1065 0.412 (0.370-0.454) 0.626 (0.567-0.684) 0.950 (0.810-1.09) Slick A 800 0.549 (0.493-0.604) 0.833 (0.755-0.91) 1.26 (1.08-1.45)

533 0.823 (0.740-0.907) 1.25 (1.13-1.37) 1.90 (1.62-2.18) 266 1.65 (1.48-1.82) 2.50 (2.27-2.74) 3.80 (3.24-4.36)

SST 1065 0.166 (0.141-0.192) 0.349 (0.306-0.391) 0.730 (0.589-0.872) Slick B 800 0.221 (0.187-0.256) 0.464 (0.408-0.52) 0.972 (0.784-1.16)

533 0.332 (0.281-0.383) 0.696 (0.612-0.781) 1.46 (1.18-1.74) 266 0.666 (0.564-0.768) 1.40 (1.23-1.56) 2.92 (2.36-3.50)

Table 1. Calculated anthracene equivalent concentrations at the LC20, LC50, and LC80 and their

95% confidence intervals for four UV doses.

73

Acknowledgements

Data presented here are a subset of a larger toxicological database that is being

generated as part of the Deepwater Horizon Natural Resource Damage Assessment, therefore,

these data will be subject to additional analysis and interpretation which may include

interpretation in the context of additional data not presented here.

This work was supported as part of the Deepwater Horizon Natural Resource Damage

Assessment (NRDA). Data presented here are a subset of a larger toxicological database that is

being generated as part of the Deepwater Horizon NRDA, therefore, these data will be subject

to additional analysis and interpretation which may include interpretation in the context of

additional data not presented in this publication. Thanks to the Texas Parks and Wildlife’s Sea

Center for the use of their aquaculture facilities and providing test organisms.

74

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CHAPTER 5

DISCUSSION

Influences on Polycyclic Aromatic Hydrocarbon Tissue Concentrations

Co-exposure of early lifestage organisms to polycyclic aromatic hydrocarbons (PAH) and

ultraviolet light (UV) resulted in photo-induced toxicity in all tested species. All species tested in

this work exhibit type III survivorship and rely on very large cohorts of offspring to produce

adequate recruitment. Significant reductions in organism success from photo-induced toxicity

could be deleterious to populations following PAH exposure. This study demonstrates that a UV

component is important to a complete understanding of the toxicity of PAHs to aquatic

organisms, or its environmental hazard may be greatly underestimated.

For photosensitization reactions to occur the organism must carry a body burden of

photo-dynamic PAHs and UV light of the correct range of wavelengths has to reach the cells

where the PAHs have partitioned to [1]. The uptake of PAHs in a natural system is complex and

modulated by various factors. Bioconcentration of PAHs also varies between species and taxa.

The blue crab represented crustacea in this work. Crustaceans are known to have larger

bioconcentration factors than fishes because of relatively poor PAH metabolism and excretion

compared to fishes. All other factors being equal, it would be expected that blue crab zoea

would be more sensitive to photo-induced toxicity than the other tested species because of a

greater accumulation of PAHs when exposed to equal waterborne concentrations. This was not

the case as blue crab were the least sensitive species of the four tested in this work. By the

same reasoning, it might be expected that mahi-mahi would take up similar body burdens to

78

red drum and speckled seatrout and thus exhibit very similar sensitivity of photo-induced PAH

toxicity. This was indeed the case between red drum and speckled seatrout, but not so with

mahi-mahi. Both red drum and speckled seatrout were exposed to PAHs for a shorter duration

than mahi-mahi but exhibited greater sensitivity. This may be due to the difference in the

lifestage of exposure between the three species. Red drum and speckled seatrout were both

exposed to PAHs as hatched larvae while mahi-mahi embryos spent the majority of their

exposure to both UV and PAHs within the protection of the egg chorion. It may be that the

mahi-mahi chorion hampers the partitioning of PAHs into the developing organism. Since there

were no renewals of test solutions the PAH exposure to the mahi-mahi may have been further

hampered by bacterial degradation and photolysis prior to hatch.

Another influence on PAH uptake and body burdens is the lipid fraction of the organism

in question. Lipids are the main partition compartment for PAHs in organisms [2, 3].

Composition of the organism is then important to its rate of uptake and final body burden at

equilibrium. Lipid content of early lifestage organisms can differ greatly from adults, especially

during yolk-dependent portions of development [4]. This can also vary throughout

development as some organisms hatch with little to no yolk left, e.g. blue crab zoea, while

others like mahi-mahi are dependent on it for longer durations after hatching. Prior to hatch, in

pelagic embryos, lipids can constitute as much as eighty percent of the embryo dry mass [4].

However, egg chorions often hamper movement of PAHs into the egg. This may reduce the

importance to lipid fraction to PAH uptake in pre-hatch fish exposure. In crab and a shrimp total

lipid constituted between two and one half percent up to ten percent of total wet mass [5, 6].

Fish larvae also vary in their total lipid composition by species. Rainuzzo et al. [7] reported lipid

79

composition at hatch to range from fourteen percent dry weight to twenty seven percent dry

weight among four fish species. So the lipid component of PAH partitioning into larvae would

be expected to be dependent on both the species and the developmental stage of the larvae in

question. This would imply that by relative expected lipid fractions the blue crab zoea would

have the lesser lipid fraction and thus body burden, making them less sensitive to photo-

induced PAH toxicity than fish larvae. This agrees with what was observed in testing between

blue crab and the red drum and speckled seatrout. However, by lipid fraction the mahi-mahi

embryos would likely have an equal or larger body burden of PAHs than the red drum and

speckled seatrout larvae as the larvae use up remaining yolk. In four fish species Rainuzzo et al.

[7] reported a reduction in lipid mass from the time of hatch to the time of first feeding,

although this reduction varied among the four species. As before, the difference between the

mahi-mahi embryonic tests and red drum and speckled seatrout larval tests may be attributable

once again to the relative impermeability of the chorion to PAHs. Additionally, lipid content is

more predicative of PAH body burdens for invertebrates than for fish species because fishes

readily metabolize PAHs [8]. However, recent studies have put forth that for very small bodied

organisms, such as those tested in this work, bioconcentration factors and lipid fraction may

not be as relevant as the waterborne exposure itself [2]. At these small masses the surface area

to volume ratio is such that direct diffusion into the organism is often the main form of gas

exchange and as a consequence PAH uptake. Following that viewpoint, post hatch exposure

might best be thought of as a combination of the dissolved concentration in the water and the

undissolved microdroplets and surface slick that comes into direct contact with the organism,

particularly gill surfaces.

80

Ultraviolet Wavelength Absorbing Pigmentation

The above assumes that differences in PAH partitioning and metabolism are the primary

factors influencing differences between species sensitivity to photo-induced PAH toxicity, which

may not be the case. UV exposure and sensitivity also determine a given species sensitivity. UV

protective environmental factors, organism behaviors and adaptations can have a great

influence on photo-induced toxic outcomes. However, the experimental design for this work

eliminated appreciable UV attenuation by water characteristics, and UV protective behaviors

such as vertical migration. This leaves only biochemical defenses to influence the absorbed UV

dose to the organism. Many of these biochemical defenses may also ameliorate reactive oxygen

species generation and damage.

Crustaceans accumulate carotenoids from their diet [9]. Carotenoids are known for

direct absorption of near UV wavelengths. These pigments are important to crustaceans in a

variety of metabolic functions and development [9]. As such the mother concentrates these

pigments into her developing eggs. Maternal transfer of these pigments into eggs gives the

hatched zoea a starting reserve prior to their first feeding event. Carotenoids are powerful

antioxidants and may offer some level of protection against ROS damage to zoea by that route.

This may have offered blue crab zoea both a level of UV protection that reduced PAH ROS

generation as well as protection against some amount of ROS that was generated despite UV

absorption. Because the zoea are entirely dependent on maternal transfer for these

carotenoids before their first feedings the variation between Slick A tests in blue crab might be

attributable to differences in maternal carotenoid supply.

81

Most marine fish species also accumulate dietary pigments as adults. However, the eggs

of the tested fish species are all transparent. Species such as red drum, speckled seatrout, and

the mahi-mahi disperse their eggs and provide no parental care. The primary defense against

predation for buoyant pelagic eggs is transparency throughout a very brief development period

before hatching. As a side effect, this permits high energy wavelengths of UV to pass freely

through the organism during development which could cause DNA damage. To manage this

many species pass on mycosporine-like amino acids (MAAs) and gadusol to their eggs [10].

These transparent compounds offer a range of protection against high energy, shorter

wavelength UV-B (peak absorbance varies by compound and typically ranges from 268 nm to

320 nm) while affording UV-A transparency as some predatory pelagic species’ vision extends

into the upper UV-A. These pigments do not absorb the UV-A range [11], which is most

responsible for photo-induced PAH toxicity. So that if mahi-mahi embyos were to absorb similar

PAH body burdens to blue crab zoea and then were exposed to similar UV conditions, it would

be expected that the mahi-mahi would be more sensitive than the blue crab because of the

photoprotective and antioxidant properties of their pigments over the MAAs and gadusol

available to the mahi-mahi. Something similar would be expected with red drum and speckled

seatrout larvae. Gadusols and MAA from development would remain in the developing larvae

but would still not afford significant protection against UV-A or ROS. In the absence of other

considerations this is consistent with the results of this work, blue crab were the least sensitive

larvae and speckled seatrout were the most sensitive.

Factors Relating to Species Sensitivity

82

Across all marine and freshwater species ultimate cause of mortality in PAH photo-

induced toxicity is the co-exposure of UV and photodynamic PAHs. However, the proximate

causes of mortality are loss of ion balance and asphyxiation. This is brought about by a chain of

events that ends with the destruction of the integumentary and gill surfaces of the organism.

Co-exposure of UV and photodynamic PAHs produces ROS, which lead to a lipid peroxidation

chain reaction in the cell membrane. This, in turn, leads to a loss of the cell’s osmoregulation

abilities which can result in apoptosis or necrosis. Tissues at the outer surface of the organism,

gill and integument, are subject to the most intense damage [12]. These are also the tissues

where most of the oxygen/carbon dioxide exchange occurs. In some small fish larvae the

majority of respiration happens by diffusion through the skin rather than across the gills.

Sufficient damage to these tissues removes the organisms’ ability to respire and asphyxiation

follows. In a photo-induced toxicity study focused on the effects on gills it has been observed

that several responses to damage contribute to the loss of ion balance and gas exchange

beyond direct destruction of gill tissue [12]. These were reported to include mucosal cell

hypertrophy, edema, and vasoconstriction [12]. These act to increase diffusion distances and to

hamper control of ion balance in both freshwater and marine organisms. In hyperosmotic

conditions this means elevated ion levels in the bloodstream and in hypoosmotic conditions

this means decreased ion levels in the bloodstream [13]. Organisms with great control over ion

balance can often be resistant to PAH toxicity by having more robust ion regulation capabilities.

The blue crab are known to have a greater range of hypoxia tolerance than mahi-mahi, the

former being able to survive in 16% oxygen saturation conditions while the latter has been

known to die >66% saturation conditions [14, 15]. There is a similar relationship with salinity

83

tolerance, mahi-mahi are able to tolerate lower salinities and fresh water for only brief periods

while the euryhaline blue crab can tolerate very low salinities and hypersaline conditions for

prolonged periods [16, 17]. This is not surprising as in their natural habitats only the blue crab

would normally see a wide range of salinities and low oxygen concentrations. The red drum and

speckled seatrout share similar habitat with the blue crab and also exhibit a wider range of

tolerance to hypoxic conditions and salinities than the mahi-mahi [18]. If the organisms in this

work were to be ordered by tolerance to hypoxia it would be expected that the blue crab would

be the most resistant to photo-induced PAH toxicity and the mahi-mahi the least. Yet the mahi-

mahi fall in the middle by phototoxic LC50. This could be explained by differences in lifestage

testing between embryos and larvae. Testing mahi-mahi larvae instead of embryos may yield

results more in line with the association of hypoxia tolerance as a predictor of resistance to

photo-induced PAH toxicity. Pigments in blue cab that absorb UV and act as antioxidants as well

as differing PAH metabolism in blue crab, as discussed above, may cloud the comparison. The

sheephead minnow (Cyprinodon variegatus) is an estuarine fish that was used in testing related

to this work. While not included in this body of work, testing on the sheepshead minnow

showed it to be very resistant to photo-induced PAH toxicity in comparison to the other tested

fishes. The sheephead minnow is an exceptional osmoregulator, known to be tolerant to

salinities ranging from 0 ppt to 142 ppt [19, 20]. Sheepshead were tested both as embryos and

as larvae. Larvae were shown to be more sensitive than embryos when exposed to overlapping

concentrations, however in both tests the phototoxic LC50 was an order of magnitude greater

than mahi-mahi embryos. Testing of more pelagic species and more estuarine species is

84

necessary to develop a more defined relationship between salinity and hypoxia tolerance and

resistance to photo-induced PAH toxicity.

Taking into account all of the discussed factors above it is not surprising that the blue

crab zoea were the least sensitive of the four tested species. The lower sensitivity of the mahi-

mahi compared to the red drum and speckled seatrout may be attributable to the protection of

the chorion through the majority of exposure. The origin of the differences in sensitivity

between the red drum and speckled seatrout is less clear at this time. Further testing might be

required to make the root causes clear. It is worth note that despite the protective factors in its

favor, the blue crab was still susceptible to photo-induced PAH toxicity at relatively low PAH

concentrations and with co-exposure to ambient sunlight.

Possible Effects on Recruitment

For the species tested in this work, and many type III survivorship marine organisms,

sufficient recruitment requires large spawns. Both eggs and young larvae are vulnerable to

predation, although predation pressure is thought to reduce as larvae grow. Hatched larvae are

also subject to starvation. A very generalized statement of pelagic spawning species is that

99.99% of fertilized embryos do not survive long enough to reach sexual maturity [21]. A

significant increase in mortality rate (M) from photo-induced PAH toxicity can have a

multiplicative reduction in overall recruitment [21]. For example, given a mahi-mahi spawn of

one million fertilized embryos the normal expectation would be that about one hundred would

reach sexual maturity. A five percent increase in the mortality rate would reduce one hundred

to sixty three. A ten percent increase in M would reduce the expected one hundred to forty. A

fifty percent increase in M would result in a one hundredfold reduction in expected

85

recruitment. However, this is an adjustment to M in absence of other factors already

contributing to M such as the availability of food. Early lifestage larvae such as those studied in

this work are dependent on zooplankton and their naupulii as a major food source. If photo-

induced PAH toxicity reduced the availability of these organisms to the fish and crab larvae

there would most likely be a greater increase to M than by considering direct mortality induced

by photo-induced PAH toxicity alone. This reduction in recruitment, even in the perhaps

unlikely absence of trophic cascade effects on adults, would adversely affect future

reproduction and recruitment. Just as larvae are dependent on zooplankton as a primary food

source, fingerling fish and pre-settlement megalopa crab depend on earlier lifestage organisms

of different species and even their own as prey. In this manner, significant reductions in larvae

hatching success and early survival could have deleterious effects to organisms not directly

affected by photo-induced PAH toxicity.

The hazard that photo-induced PAH toxicity presents to marine species is a complex

matter that varies by the physiological characteristics of the species, and their lifestage. Among

the species tested in this work sensitivity ranged across two orders of magnitude. Even the least

sensitive species, the blue crab, were shown to be vulnerable to relatively short duration

exposures. As representatives of the estuary, nearshore, and offshore habitats the tested

species demonstrate the possibility for photo-induced PAH toxicity to organisms that inhabit

these systems. Further testing on other representative species, particularly pelagics, is needed.

Factors that influence PAH concentration and UV penetration also must be taken into account

in an assessment of the environmental risk poised to species residing in these systems.

86

Acknowledgements

Data presented here are a subset of a larger toxicological database that is being

generated as part of the Deepwater Horizon Natural Resource Damage Assessment, therefore,

these data will be subject to additional analysis and interpretation which may include

interpretation in the context of additional data not presented here.

This work was supported as part of the Deepwater Horizon Natural Resource Damage

Assessment (NRDA). Data discussed here are a subset of a larger toxicological database that is

being generated as part of the Deepwater Horizon NRDA, therefore, these data will be subject

to additional analysis and interpretation which may include interpretation in the context of

additional data not presented in this work.

I thank my advisor, Dr. Aaron Roberts, for seeing me through two graduate degrees and

all the trials and tribulations that are inherent with the endeavor. I thank my committee for

their input, advice, guidance, and patience with the peculiarities involved with working

together with both government and consulting agencies. I thank my friends for their

understanding when I could only give them redacted explanations about my work these past

years. I thank my family for their unconditional support throughout my graduate career. I

thank NOAA and Stratus (Abt Associates) for this immense opportunity.

87

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