Response of Coastal Fishes to the Gulf of Mexico Oil
F. Joel Fodrie
*, Kenneth L. Heck Jr.
1 Institute of Marine Sciences and Department of Marine Sciences, University of North Carolina at Chapel Hill, Morehead City, North Carolina, United States of America,
2 Dauphin Island Sea Lab and Department of Marine Sciences, University of South Alabama, Dauphin Island, Alabama, United States of America
The ecosystem-level impacts of the Deepwater Horizon disaster have been largely unpredictable due to the unique setting
and magnitude of this spill. We used a five-year (2006–2010) data set within the oil-affected region to explore acute
consequences for early-stage survival of fish species inhabiting seagrass nursery habitat. Although many of these species
spawned during spring-summer, and produced larvae vulnerable to oil-polluted water, overall and species-by-species catch
rates were high in 2010 after the spill (1,9896220 fishes km-towed
[m 6 1SE]) relative to the previous four years
(1,080643 fishes km-towed
). Also, several exploited species were characterized by notably higher juvenile catch rates
during 2010 following large-scale fisheries closures in the northern Gulf, although overall statistical results for the effects of
fishery closures on assemblage-wide CPUE data were ambiguous. We conclude that immediate, catastrophic losses of 2010
cohorts were largely avoided, and that no shifts in species composition occurred following the spill. The potential long-term
impacts facing fishes as a result of chronic exposure and delayed, indirect effects now require attention.
Citation: Fodrie FJ, Heck KL Jr (2011) Response of Coastal Fishes to the Gulf of Mexico Oil Disaster. PLoS ONE 6(7): e21609. doi:10.1371/journal.pone.0021609
Editor: Steven J. Bograd, National Oceanic and Atmospheric Administration/National Marine Fisheries Service/Southwest Fisheries Science Center, United States
Received March 7, 2011; Accepted June 2, 2011; Published July 6, 2011
Copyright: ß 2011 Fodrie, Heck. This is an open-access article distributed under the terms of the Creative Commons Attribution License, which permits
unrestricted use, distribution, and reproduction in any medium, provided the original author and source are credited.
Funding: The authors acknowledge support from the National Marine Fisheries Service, National Oceanic and Atmospheric Administration Marine Fisheries
Initiative and Northern Gulf Institute. The funders had no role in study design, data collection and analyses, decision to publish, or preparation of the manuscript.
Competing Interests: The authors have declared that no competing interests exist.
* E-mail: jfodrie @unc.edu
Prevailing models of ecological impacts resulting from oil
pollution are being revised after the April 2010 release of ,4.4
million barrels  of oil into the northern Gulf of Mexico (GOM).
In part, this is a legacy of the Exxon Valdez accident as a
watershed environmental catastrophe, and the extensive research
on acute and chronic impacts of the resulting inshore oil pollution
. Unlike the 0.25–0.5 million barrels released by the Valdez ,
however, the Deepwater Horizon (DH) disaster hemorrhaged oil
into the open ocean at 1500 m depth over a protracted 84-day
period . As a critical step toward new model development
applicable for detecting impacts of the DH spill, rigorous
observational data at organismal through community levels are
needed to guide ecosystem-based toxicology.
We have already learned that a significant fraction of the oil
released into the GOM from the Macondo well did not rise to the
surface, and this has implications for the ecosystem-level responses
we should anticipate. Rather, oil was emulsified at the well head
due to turbulent mixing, reduced buoyancy at depth, and addition
of Corexit 9500 dispersant. Subsequently, mid-water hydrocarbon
plumes  have been observed with stimulation of petroleum-
degrading bacteria . With this now understood, we revisit some
early concerns regarding impacts for nearshore fisheries .
During the DH spill, near-surface waters lacked any reliable
refuge from oil pollution, as slicks/sheens occurred at the
immediate surface and oil was emulsified throughout the water
column. For many fishes, including commercially valuable
snappers (Lutjanidae) and groupers (Serranidae), spawning occurs
during the spring or summer (table S1), and eggs, larvae and post-
larvae would have relied upon near-surface waters overlaying the
continental shelf during the DH spill [6–7]. Furthermore, eggs/
larvae and oil can be transported by the same hydrodynamic and
atmospheric processes, enhancing the probability of oil encounters
for many species. Because the population ecology of marine
species with bipartite life histories is disproportionately affected by
the health and survival of early life stages , understanding how
eggs, larvae and newly-settled juveniles coped with the DH spill is
essential for quantifying ecosystem responses.
We hypothesized that the strength of juvenile cohorts spawned
on the northern GOM continental shelf during May–September
2010 in the northern GOM would be negatively affected by egg/
larval-oil interactions. Oiled seawater contains toxic compounds
such as polycyclic aromatic hydrocarbons (PAHs) which, even
after weathering, can result in genetic damage, physical deformi-
ties and altered developmental timing for fish eggs/larvae [9–10].
These effects may be induced at very low (,1 ppb PAHs) levels of
exposure when persistent over days to weeks [11–12] - timescales
relevant for larval development and descriptive of the DH spill.
Additionally, emulsified oil droplets could mechanically damage
the feeding and breathing apparatus of relatively fragile larvae and
further decrease individual fitness. Unfortunately, observing egg/
larval mortality, growth or migration in situ is an enduring
challenge for biological oceanographers, as eggs/larvae are simply
too dilute and experience relatively high instantaneous mortality,
even in undisturbed systems .
In the absence of direct observations on eggs and larvae,
juvenile abundance data provide valuable indices of the acute,
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population-level responses of young fishes to the spill. Although
indirect evidence , early juvenile abundances are the
integrated products of early life-history processes such as
fertilization, larval growth/mortality, and settlement [6–8].
Therefore, effects of oil pollution on early life stages should be
detectable in time series data as shifts in the abundance of recently
settled juvenile fishes. We tested these predictions using 2006–
2010 survey data collected from the Chandeleur Islands, LA, to
Saint Joseph Bay, FL (Fig. 1), representing most of the nearshore
region directly impacted by oil. In contrast to the difficulties of
surveying marine larvae, quantitative measures of juvenile
abundances are tractable due to the tendency of settled fish to
aggregate in specialized nursery habitats . In the northern
GOM, many fish species, such as those in the drum (Sciaenidae),
snapper and grouper families have juveniles that are routinely
collected from shallow-water seagrass meadows they use as
primary nurseries .
Our dataset consisted of 853 individual trawl samples taken
between July 15 and October 31 of 2006–2010 within seagrass
meadows of the northern GOM (tables S2, S3). We collected
167,740 individual fishes representing 86 taxa, and examined
catch-per-unit-effort (CPUE) data for all species pooled together,
as well as separately for each of the 20 most abundant species. We
also tested for post-spill community-level shifts in seagrass-
associated fish assemblages using multivariate analyses . We
recognized that not all species were at equal risk for oil exposure
due to variation in spawning timing and larval distributions (tables
S1, S4). Furthermore, some species may have experienced release
from fishing pressure due to large-scale fishery closures  during
the spill (table S5), perhaps enhancing their larval production
during the summer spawning season. Therefore, we also
considered how these factors affected species-specific CPUEs
during 2010. In all analyses, comparisons among years were
considered as a proxy for the effects of oil disturbance (2006–2009
as undisturbed, 2010 as disturbed).
Within the oil-affected GOM, a five-year survey of seagrass-
associated fish communities did not indicate reductions in juvenile
abundances following the spill. Rather, of the twenty most
commonly collected fish species, twelve were characterized by
statistically higher catch rates in 2010 relative to 2006–20009
(a = 0.05; Table 1). Among the remaining eight taxa, pre- and
post-spill catch rates were statistically indistinguishable. Across our
entire study region, CPUE increased from 1,080643 fishes km-
(m 6 1SE) during 2006–2009 to 1,9896220 fishes km-
in 2010. When resolved among four geographical areas
(Chandeleur Islands, Gulf Islands, Grand Bay, Florida Bays;
Fig. 1), overall catch rates of juvenile fishes, as well as CPUE of the
most abundant species, pinfish (Lagodon rhomboides), were consis-
tently higher during 2010 than in 2008 or 2009, and in some areas
were higher in 2010 than all previous years (Fig. 2A–B; fig. S1, S2,
S3; table S6).
The species composition of juvenile fish assemblages was
unaltered in each sampling area during the months following the
DH disaster (Fig. 3). Prior to the spill, similarities among individual
trawl samples (SIMPER) ranged from 50.3% at the Chandeleur
Islands to 52.9% within Florida Bays (table S7). By comparison,
similarity percentages between pre- (2006–2009) and post-spill
Figure 1. Sampling region and study sites. Map of juvenile fish sampling stations, divided among four survey areas: Chandeleur Islands (blue
circles), Gulf Islands (green circles), Grand Bay (orange circles) and Florida Bays (red circles). 1. Chandeleur Is., LA; 2. Ship Is., MS; 3. Horn Is., MS; 4. Petit
Bois Is., MS; 5. Dauphin Is., AL; 6. Grand Bature Shoal, AL; 7. Point Aux Pines, AL, 8. Big Lagoon, FL; 9. Pensacola Bay, FL; 10. Choctawhatchee Bay, FL;
11. St. Andrew Sound, FL; 12. St. Joseph Bay, FL. The spread of surface oil during the 84-day spill is also shown (brown shading). Image at lower right
shows juvenile gray snapper (L. griseus), spotted seatrout (C. nebulosus) and pipefish (Syngnathus spp.).
Response of Fishes to GOM Oil Spill
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(2010) samples were correspondingly high, ranging from 43.4%
within Grand Bay to 50.8% at the Chandeleur Islands.
Furthermore, pinfish, silver perch (Bairdiella chyrsoura), mojarras
(Eucinostomus spp.), pigfish (Orthopristis chrysoptera) and spotted
seatrout (Cynoscion nebulosus) drove similarity patterns both before
and after the spill (table S7). Species richness (S, up 15%,
p,0.001), diversity (ES
, up 11%, p = 0.006; H9, up 18%,
p,0.001) and evenness (J9, up 11%, p = 0.003) among trawl
samples were all slightly elevated during 2010 relative to 2006–
2009 averages (table S8; fig. S4), indicating that high CPUEs in
2010 were broad based.
When averaged across species, there was little statistical
evidence that either exposure risk or release from fishing pressure
significantly affected CPUEs during 2010. When comparing 2010
CPUE data against pre-spill catch rates, we did observe that fishes
characterized by moderate (spring spawning, nearshore larvae) or
high risk (spring-summer spawning, larvae distributed across the
continental shelf) exhibited decreases in CPUE following the spill
at the Chandeleur Islands and Grand Bay (Fig. 4A). However, no
statistically significant differences were found as a function of egg/
larval risk (F
= 1.4 10, p = 0.242) or sampling areas (F
= 0.999 ,
p = 0.440; table S9). Similarly, release from fishing pressure on
spawning fishes could be implicated, although not proven, as a
determinant of post-spill CPUEs. Along the Chandeleur and Gulf
Islands, increases in catch rates during 2010 relative to 2006–2009
were 800% and 950% higher, respectively (Fig. 4B), for species
identified in state and federal management plans than for species not
harvested by fishermen (table S5). No similar patterns were recorded
within Grand Bay or Florida Bays, however, and effects of fishing
=1.510, p = 0.223) and area (F
p = 0.225) on CPUE responses were not significant.
Collectively, no significant, acute impacts on the strength of
juvenile cohorts within seagrass habitats were detected following
the DH disaster. This was true for all species examined, bolstering
our confidence in the conclusion that ecosystem-level injuries were
not severe for this community of fishes. Unfortunately, our
assessment cannot be compared to the most analogous spill, the
IXTOC 1 blowout , due to a paucity of formal scientific
investigation following that accident (The 1979 IXTOC I blowout
at 3600 m depth, 80-km north of the Yucata´n Peninsula, was a
,3.5-million-barrel spill.). The most parsimonious explanation for
our data is that these fishes were resilient to the spill, possibly due
to the retention of a large proportion of spilled oil at depth. As
such, these data add to a developing literature [3–4] in which the
acute impacts of the spill may be concentrated in the deep ocean
rather than shallow-water, coastal ecosystems that were the focus
of early concern . For instance, gray snapper (Lutjanus griseus)
larvae were abundant in surface waters (0–25-m deep) over the
continental shelf from July through September , and were
among the most likely individuals to have contacted oil-polluted
water. Still, catch rates of gray snapper juveniles following the spill
Table 1. Relative frequencies and CPUE data for abundant fishes collected during sampling in seagrass meadows of the northern
(df = 851) Trend
Risk of oil
Lagodon rhomboides 60.22 644.86 1379.32 ,0.001
Eucinostomus spp. 72.67 119.94 60.21 0.086 NC Low Potential Bycatch
Bairdiella chrysoura 82.66 123.10 163.77 0.117 NC Moderate-Low No
Orthopristis chrysoptera 89.90 80.31 118.73 0.007
Moderate Potential Bycatch
Lutjanus griseus 91.64 23.63 43.02 0.003
Stephanolepis hispidus 93.29 11.95 70.61 ,0.001
Lutjanus synagris 94.68 14.79 19.18 0.171 NC High Yes
Cynoscion nebulosus 95.74 13.41 36.51 ,0.001
Syngnathus spp. 96.63 11.57 20.05 0.057 NC Moderate-Low No
Chilomycterus schoepfi 97.46 7.37 18.56 ,0.001
Leiostomus xanthurus 97.78 4.63 2.56 0.533 NC Moderate-Low Potential Bycatch
Opsanus beta 98.04 2.73 6.63 ,0.001
Arius felis 98.29 2.62 10.14 0.021
Low Potential Bycatch
Nicholsina usta 98.52 2.11 6.86 0.003
Sphoeroides spp. 98.70 2.26 2.24 0.974 NC Low No
Blenniidae 98.86 2.06 5.27 0.002
Mycteroperca microlepis 99.01 1.96 1.75 0.773 NC High Yes
Paralichthys spp. 99.16 1.97 2.90 0.133 NC Moderate Yes/ Potential
Archosargus probatocephalus 99.31 1.58 5.95 ,0.001
Lactophrys quadricorn is 99.43 1.47 3.22 0.036
Trend symbols indicate no change (NC) or statistically significant increase (q) in catch rates during 2010 relative to 2006–2009. Risk of oil encounters determined by
spawning season and across-shelf larval distribution for each species. Potential release from fishing pressure guided by state and federal management plans, as well as
shrimp-trawl bycatch data (SI).
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were high relative to the four previous years (up 82%, Fig. 2C;
area * pre/post spill context interaction p,0.001, table S6).
When averaged across species - and in some cases across species
with vastly different life histories - there were no statistically
significant differences in the response of fished or unfished species
to the spill (or their responses to subsequent management actions;
i.e., fishery closures). Still, there were notable patterns suggesting
that certain species may have been released from harvest pressure
during 2010, subsequently enhancing spawning activity and post-
spill cohort sizes despite any potentially negative oil impacts. For
example, spotted seatrout spawn during summer , but many
mature individuals are typically removed by recreational fishers
before reproducing. Following the fishery closures in 2010, we
recorded order-of-magnitude higher juvenile abundances of
spotted seatrout at the Chandeleur and Gulf Islands, as well as
elevated catch rates throughout our survey region (Fig. 2D; area,
pre/post spill context and 2-way interaction p,0.001, table S6).
Consistent with the patterns observed in the species-by-species
catch data and analyses of ‘risk’’ or ‘fishing’’ effects, there were no
significant post-spill shifts in community composition and
structure, nor were there changes in any of several biodiversity
measures. While natural recruitment variability can make it
difficult to detect population-level impacts for any one species
following large-scale disturbance , our whole-community
analyses and results are likely robust against these concerns.
Several other factors could have contributed to the high catch
rates of seagrass-associated fishes in 2010 despite large-scale oil
pollution. For instance, fishes may be uniquely buffered against oil
pollution due to their mobility or foraging ecology [21–22]. Also,
the major predators of fish eggs/larvae (e.g., gelatinous zooplank-
ton) may have been impacted by the spill, thereby reducing
natural mortality rates for coastal fishes . Regardless of the
mechanism(s) involved, thus far the potential for 2010 cohorts to
support regional fisheries over the next several years has persisted
despite the spill. This information is critical for projecting the
mode and tempo of ecological and economic recovery in the oil-
affected GOM, as well as guiding future conservation/restoration
activities to mitigate oil-spill injuries.
While these data provide reason for early optimism, attention
should now turn to possible chronic effects of oil exposure on fishes
as well as delayed indirect effects cascading through the post-spill
GOM. Fish may suffer growth, survival or reproductive penalties
years after exposure to oil , or experience altered migratory
behaviors . Oil sequestered in sediments may also affect
species laying benthic eggs for several years . More broadly,
ecosystems experiencing large-scale disturbance can carry or build
Figure 2. Catch rates of juvenile fishes, 2006–2010. Catch rates among years and sampling areas (Chandeleur Islands, Gulf Islands, Grand Bay
and Florida Bays) for: (A) all fishes pooled; (B) pinfish (L. rhomboides), (C) gray snapper (L. griseus), and (D) spotted seatrout (C. nebulosus). CPUE data in
panels B–D are presented on a log scale, and all data are shown as means of trawl samples (m + 1SE).
Response of Fishes to GOM Oil Spill
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instabilities over protracted periods that can eventually result in
delayed collapses of fisheries stocks .
Improved threat assessment for energy exploration  and
process-oriented studies of ecosystem responses will be long-term
initiatives resulting from the DH spill. In the short term, however,
observational data collected over relevant spatial and temporal
scales are invaluable for guiding and evaluating targeted studies of
oil toxicology . For fish species experiencing multiple stressors
such as habitat degradation  harvest pressure , climate
change , and now oil pollution, rigorous baseline survey data
and new syntheses are needed to enact effective ecosystem-based
Materials and Methods
We analyzed changes in northern Gulf of Mexico (GOM)
seagrass-associated fish communities during the last 5 years by
comparing survey data obtained either prior to (2006–2009) or
immediately following the Deepwater Horizon disaster (2010).
The survey region extended approximately 340 km along the
coastline, including a significant portion of the inshore area most
affected by oil (Fig. 1.). Each year, we made repeated sampling
trips to 12 sites, extending from the Chandeleur Islands, LA, to St.
Joseph Bay, FL (29.68–30.72uN, 85.30–89.10uW). Sampling
occurred within mixed seagrass meadows that serve as primary
nursery habitat for many juvenile fishes that have recently settled
from the water column following a 5–45 day larval period [6,16].
Our samples were collected from seagrass mosaics that included
turtle grass (Thalassia testudinum), shoal grass (Halodule wrightii),
widgeon grass (Ruppia maritima), and manatee grass (Syringodium
filiforme), along with scattered unvegetated patches (table S3).
During each year, trawls were conducted from July 15 through
October 31 in order to record the abundances and composition of
fishes during the period when seagrass meadows are utilized as
primary nurseries by recently settled juveniles (refer to table S1 for
reproductive seasons of common fishes). Fishes were collected
using a 5-m otter trawl (2.0-cm body mesh; 0.6-cm bag mesh;
0.360.7-m doors) with conventional 4-seam balloon design
including float and lead lines but without tickler chains. Trawls
consisted of 3.960.1 (m 6 1SE) minute tows behind small (,7m)
research vessels traveling at 3.3+0.1 kilometers hour
853 samples were taken (table S2), and the trawl covered a linear
distance of 184.7 kilometers during our sampling. These trawls
occurred in depths of 0.5–2.5-m, and were conducted during
daylight hours. During our surveys, species were enumerated in
the field unless species-level identifications were not easily made.
Unidentified specimens were transported to the lab where
meristics were used by at least two different technicians to identify
each individual. In cases in which species could not be identified,
specimens were classified to the lowest taxonomic level possible.
Typically, fishes were 20–100 mm long (standard length),
indicative of recently-spawned juveniles. However, we did not
record individual sizes for all species, and, for pipefishes (Syngnathus
Figure 3. Community composition of seagrass-associated fish communities, 2006–2010. Multi-dimensional scaling plots for seagrass-
associated fish communities prior to (2006–2009; colored symbols) and following (2010; black circles) the DH spill. Data for (A) Chandeleur Islands, (B)
Gulf Islands, (C) Grand Bay and (D) Florida Bays are presented separately. Each datum represents a single trawl sample.
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spp.) and hard-headed catfish (Arius felis), we did observe that a
small proportion of our catch included reproductive adults. For
four species: gray snapper (50.560.8 mm [m 6 SE]), lane snapper
(Lutjanus synagris; 55.760.7 mm), spotted seatrout (60.861.1 mm)
and gag grouper (Mycteroperca microlepis; 157.563.2 mm); we did
record the sizes of all individuals throughout our surveys. Based on
our own otolith analyses (Fodrie unpublished) and published
reports of first-year growth among these four species (age-1 sizes:
gray snapper ,109 mm; lane snapper ,140 mm; spotted seatrout
,127 mm; gag grouper .198 mm), we calculated that .96% of
individuals were captured in the same year as they were spawned
Once enumerated, fishes were entered in to an Excel database,
and abundance data were converted into catch-per-unit-effort
(CPUE) data based on the linear distance over with each trawl
occurred. All statistical analyses were applied to these CPUE data.
Our complete CPUE dataset is included as a separate appendix in
our supporting information (dataset S1). This study was carried
out in strict accordance with the recommendations in the Guide
for the Care and Use of Laboratory Animals of the National
Institutes of Health. Our sampling protocol was approved by the
Committee on the Ethics of Animal Experiments of the University
of North Carolina at Chapel Hill (Permit Number: 10-114.0).
We investigated differences in the catch rates of seagrass-
associated fishes (all species pooled as well as the 20 most abundant
species individually) by unpaired t-tests comparing pre- (2006–2010)
and post-spill (2010) data (Table 1), as well as 2-way ANOVAs in
which sampling area (Chandeleur Islands, Gulf Islands, Grand Bay,
Florida Bays) and context (pre- versus post-spill) were fixed factors
(table S6). Regions were defined by basic geomorphology and
location, local water clarity, local salinity, and local seagrass
composition . Because variances were stable among groups,
no data transformations were required prior to analyses.
We analyzed similarities and differences in fish communities
among years (2006–2009 versus 2010) within each sampling area
(each area considered separately) using non-metric multidimen-
sional scaling (MDS), based on Bray-Curtis similarity indices
among all individual trawl samples (4
Pairwise comparisons between trawl samples across years were
conducted with analysis of similarity (ANOSIM) and similarity (or
dissimilarity) percentages (SIMPER) using PRIMER 5.2.2 soft-
ware (PRIMER-E Ltd; ).
We also examined patterns of species diversity among regions
and years by computing the following measures for each trawl
sample: S, number of species collected; ES
, species richness
rarefied to 20 individuals; H9, Shannon-Weiner diversity index
); and J9, Pielou’s evenness measure (PRIMER 5.2.2 software).
We investigated differences in community diversity via 2-way
ANOVAs in which sampling area (Chandeleur Islands, Gulf
Islands, Grand Bay, Florida Bays) and context (pre- versus post-
spill) were fixed factors. Because variances were stable among
groups, no data transformations were required prior to analyses.
Figure 4. Larval risk and fishery closure impacts. Effects of (A) egg/larval vulnerability and (B) harvest pressure on the responses of fishes to the
DH spill. Response of individual species calculated as the ratio of 2010 versus 2006–2009 CPUE data. Data are presented on a log scale as group
means (m + 1SE), with ratios .1 indicating that 2010 catch rates were elevated relative to 2006–2010 data.
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These approaches are proscribed in earlier syntheses for
detecting environmental impacts . Critiques of employing
parametric testing to detect ecosystem injury exist due to
interannual variability and reduced statistical power , although
those concerns have focused on analyses involving single species.
We determined the relative probability for oil-larvae encounters
(‘risk’) for the twenty most commonly captured fishes, and used
these data to explore how individual species responded differently to
large-scale oil pollution in the northern GOM. Information on the
seasonal timing of spawning and distribution of larvae from shore to
the outer margin of the continental shelf was collected from the
literature (tables S1 and S4), and used to define 4 levels of risk (in
addition to an ‘unknown’ [n = 4] category containing species for
which no data were available). ‘Low’ risk species (n = 6) included
those in which larvae remained inside estuaries, either in the
plankton or as benthic egg masses, regardless of spawning season.
‘Moderate-Low’ risk species (n = 4) were defined by having either 1)
larvae distributed in estuaries as well as across the continental shelf,
or 2) larvae distributed across the continental shelf, but not likely
during the spill period (i.e., April–September). Only two ‘Moderate’
risk species were identified: pigfish (Orthopristis chrysoptera) spawn
throughout summer, and have larvae distributed within nearshore
waters; while flounder (Paralichthys spp.) have larvae distributed
across the continental shelf, with a protracted spawning that extends
into April or May (i.e., some overlap with the oil spill). ‘High’ risk
species (n = 4) spawn offshore and have larvae distributed across the
continental shelf. Furthermore, spawning data indicates that these
species would have produced larvae sometime during the DH spill
(April–Sept in our classification scheme). Based on these risk guilds,
we examined the response of fishes to the spill by calculating the
ratio of 2010 CPUE data (averaged) to 2006–2009 CPUE data
(averaged) for each species. Following these calculations, ratios .1
indicate that average 2010 catch rates were higher than during the
previous 4 years, while ratios ,1 indicate that average 2010 catch
rates were lower than during the previous 4 years. Using each
species as a replicate measure, we used ‘risk’ and region
(Chandeleur Islands, Gulf Islands, Grand Bay, Florida Bays) as
fixed factors in a 2-way ANOVA that compared 2010 CPUE:
2006–2009 CPUE trends. Because variances were stable among
groups, no data transformations were required prior to analyses.
Similarly, we determined whether species were likely to have
experienced significant release from harvest pressure following
large-scale closures in the northern GOM, and examined how this
may have affected CPUE data in 2010. For each of the twenty
most commonly caught fish, we designated species as ‘fished’ if
they were included in any state or federal management plan as of
Dec 31, 2010 (table S5), or identified as ,1% (by biomass) of
bycatch in shrimp trawl fisheries within the northern GOM (table
S5). As before, we examined the response of fishes to the spill by
calculating the ratio of 2010 CPUE data (averaged) to 2006–2009
CPUE data (averaged) for each species. Using each species as a
replicate measure, we used ‘fishing pressure’ (with fished species
including species that are targeted or captured as incidental
bycatch at significant levels) and region (Chandeleur Islands, Gulf
Islands, Grand Bay, Florida Bays) as fixed factors in a 2-way
ANOVA that compared 2010 CPUE: 2006–2009 CPUE trends.
Because variances were stable among groups, no data transfor-
mations were required prior to analyses.
All univariate tests were conducted using StatView 5.0.1 software
(SAS Institute Inc.). Because each statistical analysis applied to
separate and easily distinguishable hypotheses, we made no
corrections to experiment-wise alpha for any of the univariate (total
fishes CPUE, individual fishes CPUE, risk guilds, harvest guilds,
diversity) or multivariate (ANOSIM) tests we conducted .
Figure S1 Catch rates of all fishes, pooled together, among
sampling areas prior to (2006–2009) and following (2010) the
Deepwater Horizon disaster.
Figure S2 Catch rates of individual species, among sampling
areas prior to (2006–2009) and following (2010) the Deepwater
Horizon disaster. Data are presented for the 20 most abundant
Figure S3 Catch rates among sampling areas and years for the
20 most abundant species collected during trawl surveys.
Figure S4 Diversity measures for seagrass-associated fish
communities within sampling areas affected by the Deepwater
Table S1 Summary table for CPUE data (fish kilometer-
) of fishes prior to (2006–2009) and following (2010) the
Table S2 Distribution of trawl samples among sampling areas
(Chandeleur Islands, Gulf Islands, Grand Bay, Florida Bays) and
Table S3 Quantitative description of seagrass habitats sampled
throughout the northern Gulf of Mexico during 2006–2010.
Table S4 Information used to determine the likelihood of larvae
contacting oiled water during the summer of 2010.
Table S5 Summary table for the management status of the 20
most abundant fishes collected during our survey program.
Table S6 Summary table of the effects of sampling area and
year (context: pre- versus post-spill) on the catch rates of the 20
most abundant fishes collected during surveys in northern Gulf of
Mexico seagrass meadows.
Table S7 Comparisons of community structure between catch
data prior to (2006–2009) or immediately following (2010) the
Deepwater Horizon disaster (ANOSIM and SIMPER).
Table S8 Summary table of the effects of sampling area and
year (context: pre- versus post-spill) on the diversity (S, ES
and J9) of trawl samples collected within northern Gulf of Mexico
Table S9 Summary table of the effects of sampling area, larval
risk and harvest pressure on the change in catch rates of individual
species for pre- (2006–2009) and post-spill (2010) data.
Dataset S1 Complete CPUE data obtained for 2006–2009 trawl
surveys within seagrass meadows of the northern Gulf of Mexico.
Response of Fishes to GOM Oil Spill
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We are extremely grateful to the students and technicians who aided in the
field, especially C. Baillie, M. Brodeur, J. Myers, O. Rhoades and S.
Williams. B. Raines supplied the image of juvenile fishes in Fig. 1.
Constructive comments and support were provided by S. Powers, C.
Peterson, J. Kenworthy, and 2 anonymous reviewers.
Conceived and designed the experiments: FJF KLH. Performed the
experiments: FJF KLH. Analyzed the data: FJF. Wrote the paper: FJF
1. Crone TJ, Tolstoy M (2010) Magnitude of the 2010 Gulf of Mexico Oil Leak.
Science 330: 634–643.
2. Peterson CH, Rice SD, Short JW, Esler D, Bodkin JL, et al. (2003) Long-term
ecosystem response to the Exxon Valdez Oil Spill. Science 302: 2082–2086.
3. Camilli R, Reddy CM, Yoerger DR, Van Mooy BAS, Jukuba MV, et al. (2010)
Tracking hydrocarbon plume transport and biodegradation at Deepwater
Horizon. Science 330: 201–204.
4. Hanzen TC, Dubinsky EA, DeSantic TZ, Andersen GL, Piceno YM, et al.
(2010) Deep-sea oil plume enriches indigenous oil-degrading bacteria. Science
5. Kerr R, Kintisch E, Stokstad E (2010) Will Deepwater Horizon set a new
standard for catastrophe? Science 328: 674–675.
6. Able KW, Fahay MP Ecology of Estuarine Fishes, Temperate Waters of the
Western North Atlantic (The Johns Hopkins University Press). 566 p.
7. Miller BS, Kendall AW Early Life History of Marine Fishes (University of
California Press). 376 p.
8. Hjort J (1914) Fluctuation in the great fisheries of northern Europe reviewed in
the light of biological research. Rappo rts et Proce`s-Verbaux des Re´unions du
Conseil International pour l’Exploration de la Mer 20: 1–228.
9. Kocan RM, Hose JE, Brown ED, Baker TT (1996) Injury to the early life history
stages of Pacific herring in Prince William Sound after the Exxon Valdez oil spill.
Can J Fish Aquat Sci 53: 2366–2375.
10. Tuvikene A (1995) Responses of fish to polycyclic aromatic hydrocarbons
(PAHs). Ann Zool Fennici 32: 295–309.
11. Carls MG, Rice SD, Hose JE (1999) Sensitivity of fish embryos to weathered
crude oil: Part I. Low-level exposure during incubation causes malformations,
genetic damage, and mortality in larval pacific herring (Clupea pallasi). Environ
Toxicol Chem 18: 481–493.
12. Brown ED, Baker TT, Hose JW, Kocan RM, Marty GD, et al. (1996) Injury to
the early life history stages of Pacific herring in Prince William Sound after the
Exxon Valdez oil spill. Am Fish Soc Sym 18: 448–462.
13. Levin LA (2006) Recent progress in understanding larval dispersal: new
directions and digressions. Integr Comp Biol 46: 282–297.
14. Hilborn R (1996) Detecting population impacts from oil spills: a comparison of
methodologies. Am Fish Soc Sym 18: 639–644.
15. Beck MW, Heck KL, Able KW, Childers DL, Eggleston DB, et al. (2001) The
identification, conservation, and management of estuarine and marine nurseries
for fish and invertebrates. Bioscience 51: 633–641.
16. Fodrie FJ, Heck KL, Powers SP, Graham WM, Robinson KL (2010) Climate-
related, decadal-scale assemblage changes of seagrass-associated fishes in the
northern Gulf of Mexico. Glob Change Biol 16: 48–59.
17. Underwood AJ in Detecting Ecological Impacts: Concepts and Applications in
Coastal Habitats, Schmitt RJ, Osenberg CW (Academic Press). pp 151–175.
18. NOAA Fisheries Service Deepwater Horizon Oil Spill Information. Available:
http://sero.nm fs.noaa.go v/deepwater_hor izon_oil_spill.htm. Accessed: 2011
19. D’Alessandro EK, Sponaugle S, Serafy JE (2010) Larval ecology of a suite of
snappers (family: Lutjanidae) in the Straits of Florida, west ern Atlantic Ocean.
Mar Ecol Prog Ser 410: 159–175.
20. Holt GJ, Holt SA (2000) Vertical distribution and the role of physical processes
in the feeding dynamics of two larval sciaenids Sciaenops ocellatus and Cynoscion
nebulosus. Mar Ecol Prog Ser 193: 181–190.
21. Peterson CH, Kennicutt MC, Green RH, Montagna P, Harper DE, et al. (1996)
Ecological consequences of environmental perturbations associated with offshore
hydrocarbon production: a perspective on long-term exposures in the Gulf of
Mexico. Can J Fish Aquat Sci 53: 2637–2654.
22. Edgar GJ, Marshall PA, Mooney P (2003) The effect of the Jessica grounding on
Gala´pagos reef fish communities adjacent to the wreck site. Mar Pollut Bul 47:
23. Graham WM, Condon RH, Carmichael RH, D’Ambra I, Patterson HK, et al.
(2010) Oil carbon entered the coastal planktonic food web during the Deepwater
Horizon oil spill. Environ Res Lett 5: 045301.
24. Heintz RA, Rice SD, Wertheimer AC, Bradshaw RF, Thrower FP, et al. (2000)
Delayed effects on growth and marine survival of pink salmon Oncorhynchus
gorbuscha after exposure to crude oil during embryonic development. Mar Ecol
Prog Ser 208: 205–216.
25. Wertheimer AC, Heintz RA, Thedinga JF, Maselko JM, Rice SD (2000)
Straying behavior of adult pink salmon (Oncorhynchus gorbuscha) exposed as
embryos to weathered Exxon Valdez crude oil. Trans Am Fish Soc 129:
26. Bue BG, Sharr S, Seeb JE (1998) Evidence of damage to pink salmon
populations inhabiting Prince William Sound, Alaska, two generations after the
Exxon Valdez Oil Spill. Trans Am Fish Soc 127: 35–43.
27. Pearson WH, Elston RA, Bienert RW, Drum AS, Antrim LD (1999) Why did
the Prince William Sound, Alaska, Pacific herring (Clupea pallasi) fisheries collapse
in 1993 and 1994. Can J Fish Aquat Sci 56: 711–737.
28. U. S. Department of the Interior, Minerals Management Service Environmental
Division, ‘‘Oil-Spill Risk Analysis: Gulf of Mexico Outer Continental Shelf
(OCS) Lease Sales, Eastern Planning Area, 2003–2007 and Gulfwide OCS
Program, 2003–2042’’ OCS Report MMS 2002–069.
29. Sagarin R, Pauchard A (2010) Observational approaches in ecology open new
ground in a changing world. Front Ecol Environ 8: 379–386.
30. Diaz RJ, Rosenberg R (2008) Spreading dead zones and consequences for
marine ecosystems. Science 321: 926–929.
31. Baum JK, Myers RA (2004) Shifting baselines and the decline of pelagic sharks
in the Gulf of Mexico. Ecol Lett 7: 135–145.
32. Byron D, Heck KL (2006) Hurricane effects on seagrasses along Alabama’s Gulf
Coast. Estuar Coast 29: 939–942.
33. Clark KR, Gorley RN (2001) PRIMER v5: User Manual/Tutorial. Plymouth:
Plymouth Marine L aboratory.
34. Moran MD (2003) Arguments for rejecting the sequential Bonferroni in
ecological studies. Oikos 100: 403–405.
Response of Fishes to GOM Oil Spill
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