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Oyster Reefs as Natural Breakwaters Mitigate Shoreline
Loss and Facilitate Fisheries
Steven B. Scyphers
1,2
*, Sean P. Powers
1,2
, Kenneth L. Heck Jr.
1,2
, Dorothy Byron
2
1Department of Marine Sciences, University of South Alabama, Mobile, Alabama, United States of America, 2Dauphin Island Sea Lab, Dauphin Island, Alabama, United
States of America
Abstract
Shorelines at the interface of marine, estuarine and terrestrial biomes are among the most degraded and threatened
habitats in the coastal zone because of their sensitivity to sea level rise, storms and increased human utilization. Previous
efforts to protect shorelines have largely involved constructing bulkheads and seawalls which can detrimentally affect
nearshore habitats. Recently, efforts have shifted towards ‘‘living shoreline’’ approaches that include biogenic breakwater
reefs. Our study experimentally tested the efficacy of breakwater reefs constructed of oyster shell for protecting eroding
coastal shorelines and their effect on nearshore fish and shellfish communities. Along two different stretches of eroding
shoreline, we created replicated pairs of subtidal breakwater reefs and established unaltered reference areas as controls. At
both sites we measured shoreline and bathymetric change and quantified oyster recruitment, fish and mobile macro-
invertebrate abundances. Breakwater reef treatments mitigated shoreline retreat by more than 40% at one site, but overall
vegetation retreat and erosion rates were high across all treatments and at both sites. Oyster settlement and subsequent
survival were observed at both sites, with mean adult densities reaching more than eighty oysters m
22
at one site. We
found the corridor between intertidal marsh and oyster reef breakwaters supported higher abundances and different
communities of fishes than control plots without oyster reef habitat. Among the fishes and mobile invertebrates that
appeared to be strongly enhanced were several economically-important species. Blue crabs (Callinectes sapidus) were the
most clearly enhanced (+297%) by the presence of breakwater reefs, while red drum (Sciaenops ocellatus)(+108%), spotted
seatrout (Cynoscion nebulosus)(+88%) and flounder (Paralichthys sp.) (+79%) also benefited. Although the vertical relief of
the breakwater reefs was reduced over the course of our study and this compromised the shoreline protection capacity, the
observed habitat value demonstrates ecological justification for future, more robust shoreline protection projects.
Citation: Scyphers SB, Powers SP, Heck KL Jr, Byron D (2011) Oyster Reefs as Natural Breakwaters Mitigate Shoreline Loss and Facilitate Fisheries. PLoS ONE 6(8):
e22396. doi:10.1371/journal.pone.0022396
Editor: Howard Browman, Institute of Marine Research, Norway
Received February 28, 2011; Accepted June 21, 2011; Published August 5, 2011
Copyright: ß2011 Scyphers et al. 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: Funding for this project was provided by the National Oceanic and Atmospheric Administration Fisheries Office of Habitat Restoration through the
University of South Alabama’s Oyster Restoration Program and the Northern Gulf Institute. The funders had no role in study design, data collection and analysis,
decision to publish, or preparation of the manuscript.
Competing Interests: The authors have declared that no competing interests exist.
* E-mail: sscyphers@disl.org
Introduction
Nearshore, biogenic habitats of estuaries support a broad
spectrum of marine life and serve as nursery grounds for
economically-important fishes and shellfish [1–4]. Estuarine and
vegetated nearshore habitats comprise only 0.7% of global biomes,
yet the value of their ecosystem services has been estimated at $7.9
trillion dollars annually, or 23.7% of total global ecosystem
services [5]. Nearshore ecosystem services include disturbance
resistance, nutrient cycling, habitat, food production, and
recreation. Unfortunately, coastal and estuarine shorelines are
among the most degraded and threatened habitats in the world
because of their sensitivity to sea level rise, storms and increased
utilization by man [6,7]. Many previous efforts to protect
shorelines have involved the introduction of hardened structures,
such as seawalls, rocks or bulkheads to dampen or reflect wave
energy [8–10]. Although such structures may adequately mitigate
shoreline retreat, the ecological damages that result from their
presence can be great [8,10,11]. The cumulative effects of habitat
alteration and losses in the nearshore have had substantial
economic and ecological consequences [12,13] and threaten the
sustainability of many ecosystem services. Efforts to combat
degradation and loss of nearshore, biogenic habitats have
increased over the last decade [7,14,15]. Unfortunately, many
shoreline protection approaches still value engineering over
ecology in determining mitigation and restoration efficacy.
The ‘‘engineering first’’ approaches, including vertical bulk-
heads, concrete and granite rip-rap revetments and seawalls, are
often used by coastal engineers because they are viewed as
permanent and non-retreating structures. Unfortunately, insuffi-
cient concern may have been given to the ecological, aesthetic or
socioeconomic impacts of these hardened structures. A major
concern in implementing bulkheads and seawalls for coastal
property protection is that erosive wave energies are reflected back
into the water body, instead of being absorbed or dampened [10].
This subjects adjacent shorelines to even greater wave energy and
can cause vertical erosion down the barrier with subsequent loss of
intertidal habitats [10,16].
The benthic setting adjacent to many armored shores is
generally absent of complex, structured habitats [16]. Most
structurally complex, natural habitats are thought to function as
nurseries for many finfish and shellfish species because of their
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elevated faunal densities, enhanced growth or survival rates, or
higher contribution of individuals that emigrate offshore to adult
habitats [1,4]. Biogenic, three-dimensional structure can reduce
water velocities, increase sedimentation rates and enhance
propagule settlement and retention, indirectly creating a more
suitable environment for many species [17–20]. Despite the known
lack of ecological benefits, shoreline hardening has continued to
increase for decades primarily due to a lack of practical and
ecologically valuable alternatives. However, a growing initiative
for sustainable shoreline protection has focused on balancing
effective protection and habitat creation by a variety of new
methodologies collectively termed ‘‘living shorelines’’ [8].
Living shoreline projects often involve the planting or
restoration of naturally-occurring biogenic habitats that have
numerous ecological benefits, in addition to providing a buffer for
wave action. In their natural setting, oyster reefs are often found
seaward of salt marshes and can attenuate erosive wave energies,
stabilize sediments and reduce marsh retreat, thereby making
them an attractive living shoreline approach [19,21,22]. Beyond
the targeted shoreline protection, living oyster reefs may provide
many ecosystem services including seston filtration, benthic-
pelagic coupling, refuge from predation and abundant prey
resources [2,18]. Given adequate recruitment and survival, oyster
reefs could be self-sustaining elements of coastal protection [21,22]
that enhance other habitats of the natural landscape, although few
studies have examined the premise of restoration through
facilitation [17,23].
Located on the northern Gulf of Mexico, Mobile Bay is one of
the best examples of a classic estuary [3] and, like many other
coastal areas, is highly developed with a large and increasing
proportion of its shorelines armored by bulkheads and seawalls
[10] (Figure 1). At last analysis in 1997, Douglass and Pickel
estimated that over 30% of the bay’s available coastline was
armored with over 10–20 acres of intertidal habitat lost, a high
percentage in this microtidal bay (,0.5 m tidal amplitude). The
historical armoring and marsh-edge losses have already had
negative fisheries consequences, with projections of further
reductions of blue crab harvest if armoring continues [24].
In this study, we experimentally examined the ecological effects
of constructing subtidal breakwater oyster reefs for coastal and
estuarine shoreline protection. In addition to documenting
changes in the physical setting near breakwaters and unaltered
control treatments, we quantified the habitat value for oysters,
fishes and mobile invertebrates. We focus particular attention on
the potential impacts on economically-important species, as this
provides insight into the economic implications of different
shoreline protection alternatives. We hypothesized that the
addition of breakwater reefs of oyster shell would: 1) mitigate
shoreline retreat, (2) provide substrate for recruitment and survival
of oysters, (3) support higher densities of small fishes, mobile
macro-invertebrates and larger and transient fishes and (4)
promote higher species richness and a different community
structure than unaltered control areas.
Methods
Ethics Statement
This study was conducted in accordance with the laws of the
State of Alabama and under IACUC protocols (Permit #05047-
FSH) approved by the University of South Alabama.
Study Setting and Site Selection
To determine the ecological and physical effects of created
breakwater oyster reefs, we conducted a manipulative field
experiment at two sites in coastal Alabama that contained
stretches of rapidly eroding coastlines. Study sites were selected
within regions known to have adequate larval supply of oysters
[25] and moderate wave climates [26]. At each site, we
constructed two breakwater reefs of loose oyster shell and
designated non-restored plots as controls in a randomized, paired
design (Figure 2). The first site, known locally as Point aux Pins,
received breakwater reefs in May 2007. The treatments at Point
aux Pins (site center point: 30.370098,288.308578) were located
along the southern extent of a peninsula of eroding salt marsh
habitat, largely comprised of fringing cordgrass (Spartina alterniflora)
and black needlerush (Juncus roemerianus). Remnants of oysters
(Crassostrea virginica) are found throughout the marsh and buried in
the subtidal sediments. The second site, Alabama Port (site center
point: 30.347917,288.121338), is located along the southwestern
shore of Mobile Bay, just north of the Dauphin Island bridge. The
treatments at Alabama Port were located along a two kilometer
stretch of eroding shoreline that has been encroached by armoring
at its northern and southern extents. Small patches of Spartina
alterniflora can be found at Alabama Port, but the most abundant
vegetation is Phragmites sp., which is largely present in the upper
intertidal zone. Both sites were selected within regions of high
oyster spat settlement (40–180 spat m
22
day
21
) [25].
Breakwater Reef Dimensions
The experimental oyster reefs were designed as subtidal wave-
attenuating breakwaters, a common coastal engineering approach
[8]. Each reef complex was comprised of three 5 m625 m
rectangular-trapezoid sections (Figure 3B). Each section consisted
of loose oyster shell, purchased from a local seafood processing
plant, placed on a geo-textile fabric to prevent subsidence and
secured by a plastic mesh covering (with 1 cm
2
openings) that was
anchored by rebar. The purpose of the mesh covering was to help
maintain the vertical relief of breakwaters until adequate
recruitment of oysters cemented the loose shell in place. The
initial height of each reef was slightly above MLLW (,1 m), under
the assumption that the loose oyster shell would settle below that
level and eventually become subtidal. The subtidal design of the
reefs allowed for maximum exposure for oyster settlement and
increased available substrate for foraging by transient and larger
Figure 1. Population Growth and Shoreline Armoring in Mobile
Bay, Alabama. Adapted with permission from Douglass and Pickel
1999, this figure depicts the rate and extent of shoreline armoring in
Mobile Bay. The vertical bars in the main graph show the proportion of
armoring while the line depicts the increasing population levels for
Mobile and Baldwin Counties.
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resident fishes, while maximizing potential capacity for wave
attenuation.
Hydrographic Environment
Mean surface water temperatures, recorded by electronic
thermometer, and salinity, measured by a refractometer, were
recorded during each sampling event. To observe longer term
patterns in salinity, we utilized publicly available data recorded by
hydrographic monitoring stations located at Cedar Point and
Dauphin Island, AL. The Cedar Point station is approximately
17.5 km from Point aux Pins and 4.0 km from Alabama Port site
center points. The Dauphin Island station is approximately
25.5 km from Point aux Pins and 11.0 km from Alabama Port
site center points. The Cedar Point station has been active since
2008 and the Dauphin Island station since 2003. To consider the
effects of wave climate and dominant wind direction and
magnitude on our study setting, we reviewed historical and
recently published coastal engineering studies [26,27].
Shoreline and Bathymetry Change
Vegetation retreat and changes in nearshore depth profiles were
monitored to evaluate the effect of the breakwaters on the
nearshore setting. Bathymetry surveys were conducted at both sites
during preliminary site selection and yearly following construction
at Point aux Pins. Bathymetric data was collected using a
Ceeducer Pro DGPS system with an integrated depth sounder
mounted to a 1 m62 m platform on pontoons. We surveyed each
site manually by walking the pontoon through multiple parallel
transects of the reef and control treatments. At each reef
treatment, the breakwater reef footprint was delineated using the
Ceeducer DGPS to measure reef spreading and consequential
reduction in reef height. The width of each reef section was also
measured by transect tape at reef construction and the end of the
study to measure changes in reef footprint. The data collected by
the Ceeducer unit was imported in ESRI’s ArcView, corrected for
tidal amplitude, and maps depicting depth at mean low water
(MLW) were created. To measure the shoreward retreat of
emergent vegetation, permanent rebar stakes were installed at
25 m intervals along the 100 m stretch of shoreline at each
replicate treatment. Each 6 m rebar stake was driven into the
marsh edge so that 1 m remained visible. These shoreline stakes
were installed shortly after breakwater construction at both Point
aux Pins and Alabama Port and were monitored periodically
thereafter. During each survey, marsh retreat was measured as the
distance from the rebar stake to the living vegetation line. Mean
differences between vegetation retreat rates adjacent to breakwa-
ters and controls were analyzed by repeated-measures analysis of
variance (ANOVA). Because of differences in reef creation dates
and sampling period, Point aux Pins and Alabama Port were
analyzed separately.
Oyster Recruitment
To assess the value of the breakwater reef complexes for oysters
and other sessile invertebrates, we periodically collected quadrat
samples. Oyster settlement, growth and survival were quantified
using a 0.25 m
2
quadrat, which was haphazardly placed at three
locations on each reef section (n = 9 per replicate reef). The
exposed layer of shell within the quadrat was collected and placed
in a large container. Juvenile (#3 cm) and adult oysters (.3 cm)
were enumerated and measured in the field, and then returned to
the reef. Mortality was quantified by enumerating dead oysters,
which had both valves still articulated and were absent of fouling
organisms inside the shell. We sampled the breakwater reefs at
Point aux Pins in July, August and November 2007, May and
October 2008 and June 2009. We sampled the reefs at Alabama
Port in March, June and October 2008 and June 2009. For the
final sampling period of June 2009, six 0.25 m
2
quadrats were
sampled from each section (n = 18 per replicate reef) to account for
the reef spreading and to assure a similar proportion of reef surface
area was sampled.
We used univariate one-way ANOVA to test for differences in
densities of live juveniles, live adults and dead oysters among
Figure 2. Map of Study Sites in Mobile Bay and Mississippi Sound, Alabama. White triangles represent breakwater reef complexes and
white circles represent control treatments at the two restoration sites of (A) Point aux Pins and (B) Alabama Port. The locations of the (1) Cedar Point
and (2) Dauphin Island hydrographic monitoring stations are denoted by the numbered arrows.
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sampling events. Point aux Pins and Alabama Port recruitment
data were analyzed separately because the independent variable of
sampling date was different at each site. Density estimates
determined by individual quadrat samples (n = 3 all, except n = 6
for June 2009) for each reef section were averaged. The pooled
values from each of the three reef sections for each of the two
replicated treatments were used as replicates in a one-factor
ANOVA to test the effect of sampling date. These data were tested
for normality using the Kolmogorov-Smirnov test and homoge-
neity of variances using Bartlett’s test. To meet the assumptions of
ANOVA, all values were log transformed and retested. After
transformation, minor violations of normality and equal variances
were still present for live adults and dead oysters at both sites.
Because the violations from quadrat sampling are generally minor
and ANOVA is considered robust to such violation [28], we
proceeded with parametric ANOVA. When ANOVA results
showed significant differences, we used Tukey’s HSD post-hoc test
for multiple comparisons.
Fishes and Mobile Invertebrates
The response of fishes and mobile invertebrates was measured
using a combination of gear types to target small and large
individuals. Experimental gillnets (2 m630 m) were used to
capture larger species and individuals of coastal finfish species.
Sampling occurred twice each month for one year following
construction and monthly thereafter through all seasons, but was
reduced to every other month during winter. Gillnets were
deployed on adjacent sides of each reef or control treatment and
perpendicular to shore. Each net was comprised of two 15 m
panels (5 cm and 10 cm maximum opening) to broaden the size
Figure 3. Bathymetry Plots from the Western Experimental Breakwater Reef and Control Treatments at Point aux Pins. The top row
of 2006 plots was approximately one year prior to construction. The 2008 and 2009 plots are from one and two years post construction. Depth
gradients are shown in inset (A). A schematic of the initial reef shape is depicted in (B). The crest width of each reef was approximately 1-m at MLLW.
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range and body shape of animals captured. Gillnets were fished for
two hours starting one hour prior to sunrise. During winter
months, low tides prevented crepuscular sampling so nets were
fished for two hours starting one hour prior to sunset. Gillnets were
retrieved in the same order they were deployed, and soak time was
recorded as the time from when the net was first deployed until the
time retrieval began. All specimens captured were placed in
labeled bags and returned to the lab where they were identified,
measured and their biomass recorded.
To quantify smaller fishes and invertebrates, we seined adjacent
to each breakwater reef and control monthly, except every other
month during winter. At each treatment, a 6 m wide bag seine
with 6.25 mm mesh was towed three times between the treatment
and shore. All seine distances were 15 m and terminated into the
shore at Point aux Pins or a 4 m wide block net at Alabama Port.
All captured mobile invertebrates and fishes were placed in labeled
bags and returned to the laboratory where they were identified to
the lowest taxonomic level possible, measured and biomass
recorded.
To determine the effects of site and treatment on the
communities of fishes and invertebrates, we used multivariate
and univariate analyses. Differences in community structure
between reef and control treatments and between Alabama Port
and Point aux Pins sites were tested for each gear type using
permutational analysis of variance (PERMANOVA). Multivariate
PERMANOVA used Bray-Curtis similarity matrices of log (x+1)
transformed abundance data with 4,999 permutations [29].
Logarithmic transformations were applied to reduce the influence
of overwhelmingly abundant species. For univariate analyses on
gillnet data, PERMANOVA was used to test for site and treatment
effects on the total abundance, species richness and abundance of
demersal fishes in an approach similar to parametric ANOVA.
Univariate PERMANOVA tests were run on Euclidean distances
matrices with 4,999 permutations [30]. PERMANOVA was
chosen for univariate analyses because it allows for two-factor
designs, considers an interaction term and does not assume a
normal distribution of errors. The environmental classifications of
demersal, pelagic (including benthopelagic, pelagic, and pelagic-
neritic) and reef-associated fishes were acquired from FISHBASE
[31]. Seine data were analyzed identically to gillnet data analyses
as previously stated with the addition of a response variable
containing only decapod crustaceans. All multivariate tests and
univariate PERMANVOA were run in the software package
PRIMER-E v6 [32] with the PERMANOVA extension.
To determine the effects of breakwater reefs on the most
common demersal fishes and decapods, we analyzed these taxa
separately as they include many economically-important coastal
species. We used Wilcoxon signed-rank tests to compare relative
abundances of each species ($1%) between the paired breakwater
reef and mudflat control treatments. This approach allowed us to
test for overall treatment effects, while controlling for date and site
variability through the paired experimental design but ignored
their interactive effects. Certain species that were closely related or
difficult to distinguish were analyzed as grouped taxa (e.g. Menidia
sp., Paralichthys sp.). For all tests, we considered results of p#0.05 to
be significant. The ANOVA and Wilcoxon tests were run using
the R Statistical Platform Version 10.1.1 [33].
Results
Hydrographic Environment
At Point aux Pins, mean surface water temperature over all
sampling events was 21.4uC(610.1 SD) measured by digital
thermometer, and salinity averaged 23.1 PSU (68.7 SD)
measured by refractometer. Mean water temperature at Alabama
Port was 21.8uC(67.8 SD) and salinity averaged 16.1 PSU (67.4
SD). Salinity data, shown as box and whisker plots, was acquired
from hydrographic monitoring stations at Cedar Point (Figure 4A)
and Dauphin Island (Figure 4B) to further investigate the salinity
regime over a longer time period. Cedar Point data shows 2008 to
have the highest salinity regime of the 2008–2010 years
(Figure 4A). The Dauphin Island station shows a similar pattern
with 2007 and 2008 having higher salinities than all other years
between 2003 and 2010. In addition to higher average salinity, the
outliers representing the lowest salinity measurements in 2007 and
2008 are substantially higher the other recent years indicating
fewer freshets.
Figure 4. Salinity Ranges Recorded by Hydrographic Monitoring Stations in Coastal Alabama. Box and whisker plots of salinity data
recorded by the hydrographic monitoring stations at (A) Cedar Point and (B) Dauphin Island. The Cedar Point Station has been active since 2008 and
the Dauphin Island Station since 2003.
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Most wind-driven wave energies along coastal Alabama
shorelines are generated by dominant south to southeasterly winds
from spring through early fall and north-oriented winds from late
fall throughout most of winter [26,27]. Fetch at Point aux Pins
averages approximately 15 km with a longest fetch of 32 km. The
erosion rate at this site has potentially increased in recent years
after Hurricane Katrina opened a one mile gap termed ‘‘Katrina
Cut’’ in Dauphin Island, a protective barrier island located due
south of Point aux Pins. The wave climate near Alabama Port is
strongly affected by prevalent southeast winds, as well as the wakes
of ships utilizing the Mobile shipping channel less than ten
kilometers to the East. At Alabama Port, average fetch is
approximately 21 km with a longest fetch of 34 km. For more
detailed discussion of wind and wave climates, erosion, sediment
sizes, Keddy exposure values and Knutson et al.’s vegetation
success scores, refer to Roland and Douglass (2005) [26].
Shoreline and Bathymetry Changes
Changes in the nearshore and shoreline environments of reef and
control sites were observed from measuring vegetation retreat and
bathymetric surveys. Bathymetric surveys at Point aux Pins found
that, in addition to a general trend of decreasing depth, areas
inshore of breakwater reefs appeared to gain more sediments than
areas inshore of control plots (Figure 3). The footprint of East and
West breakwaters expanded approximately 300% over the course of
the study, and reef crest height was reduced from approximately
1 m to 0.3 m. The living vegetation line at Point aux Pins retreated
nearly 6 m on average in slightly over two years (Figure 5A).
Repeated measures ANOVA found no differences in the vegetation
retreat rates between treatments, a strong effect of time and no
interaction between the two factors (Table S1). At Alabama Port,
breakwater reefs mitigated vegetation retreat by more than 40%
over two years (Figure 5B). Repeated measures ANOVA found a
marginally-significant treatment effect (p = 0.089) and a strong
effect of time with no interaction (Table S1).
Oyster Recruitment
Point aux Pins reefs were constructed in May 2007 and first
sampled for oyster recruitment the following July. Densities of
juvenile oysters continually increased until peaking at greater than
700 oysters m
22
in November 2007, but were much lower the
following year with ranges between 50 and 150 m
22
(F
5,30
= 28.15, p#0.001, Figure 6A). Adult oysters were found in
highest densities during November 2007 and May 2008 sampling
with approximately 35 oysters m
22
(F
5,30
= 38.29, p#0.001,
Figure 6B). The highest mortality was observed during the
October 2008 sampling event (F
5,30
= 22.492, p#0.001, Figure 6C)
and 88% of measured dead oysters were juveniles (#3 cm).
Alabama Port reefs were constructed in October 2007 and were
first sampled in March 2008. Live juveniles densities at Alabama
Port were between 70 and 140 m
22
in the last three sampling
events and higher than the first sampling event in March 2008
(F
3,20
= 47.40, p#0.001, Figure 6D). Adult oysters were observed
first and at a maximum in October 2008 (,75 oysters m
22
) and
found in lower densities in June 2009 (,20 oysters m
22
)
(F
3,20
= 18.82, p#0.001, Figure 6E), although October and June
were not significantly different. The first and highest mortality
(,70 oysters m
22
) was recorded in October 2008 (F
3,20
= 114.29,
p#0.001, Figure 6F), and juvenile oysters accounted for 80% of
the total dead.
Fishes and Mobile Invertebrates
Gillnet and seine sampling near breakwater reefs and controls
captured a diverse assemblage of fishes and mobile macro-
invertebrates. From the use of multiple gears, over 100 species of
fish and invertebrates were collected during the 30 month
sampling period. Gillnet sampling collected nearly 8,000 individ-
uals of 45 different species in 5 cm mesh panels while larger 10 cm
panels captured over 1,500 individuals of 44 different species.
Seines captured 71,640 individuals that represented 88 species or
grouped taxa. Demersal fishes appeared to be the most broadly
enhanced by the oyster reef structure when the overall percent
difference in CPUE between oyster reefs and mudflat controls was
calculated across both sites and all sampling events (Table S2). The
dominant pelagic and reef-associated species did not appear
strongly affected by oyster reef presence. Of the twelve species that
comprised at least 1% of the 5 cm gillnet catch, six were
categorized as demersal species. Four of these six demersal taxa
were more abundant on breakwater reefs than controls. Spotted
seatrout were 38% more abundant near breakwater reefs, and
displayed the strongest trend of enhancement among 5 cm
captured fishes. Twenty species comprised at least one percent
of the 10 cm gillnet catch, and eleven of these were demersal
fishes. Fourteen of the twenty species were captured more often
Figure 5. Shoreline Vegetation Retreat. Mean retreat (6SE) of living vegetation shoreward of each treatment at (A) Point aux Pins and (B)
Alabama Port.
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near breakwater reefs than controls. Nine species or grouped taxa
comprised at least one percent of seine catches, seven of which
were more frequently captured near breakwater reefs. Included in
these seven were three demersal fishes and three decapod
crustaceans.
We used multivariate PERMANOVA to test for differences in
the community structure between breakwater reef and control
treatments. PERMANOVA tests on 5 cm gillnet catches found
that site and the site-treatment interaction were both significant
factors (Table S3). There were no community-level differences
between our breakwater and control treatments with 5 cm
captured fishes. The communities of larger fishes captured by
10 cm gillnets differed significantly by site, treatment and the
interaction of the two factors (Table S3). The community structure
of smaller and juvenile fishes and mobile invertebrates captured by
seines were different between sites and treatments, with no
interaction between the two factors (Table S3).
We used univariate PERMANOVA tests on total abundance,
species richness and demersal and decapod abundances to detect
differences between breakwater reef and control treatments and
between Alabama Port and Point aux Pins. For 5 cm total
abundance, a significant interaction between site and treatment
was observed (Table S4) because total abundance was higher near
breakwaters at Point aux Pins but higher near controls at Alabama
Port (Figure 7A). For 10 cm gillnet catch, total abundance was
higher adjacent to oyster reefs than controls (Table S4). For both
10 cm and seine data, abundances were significantly higher at
Point aux Pins than Alabama Port, and no interaction was
observed between site and treatment (Table S4). The PERMA-
NOVA tests on species richness found significant differences
between sites across all gear types, between treatments only for
10 cm catches and no significant interactions (Table S4). For
10 cm catches, species richness was significantly higher near reefs
than controls (Table S4) and higher at Point aux Pins than
Alabama Port. Demersal fishes showed no differences between reef
and mudflat treatments for 5-cm catches (Table S3), but again
there was a significant interaction between site and
treatment(Figure 7B). For 10 cm, demersal fishes were more
abundant near breakwater oyster reefs (Figure 8A) and higher at
Point aux Pins. From seine catches, demersal fish abundance
showed no differences, but decapod crustacean abundance was
higher near reefs than mudflat controls (Table S4 and Figure 8B).
The relative abundance of each demersal fish and decapod
species ($1%) between breakwater and control treatments was
tested using paired Wilcoxon signed-rank tests. For 5 cm gillnet
samples, six demersal species contributed $1% of the total catch
Figure 6. Oyster Recruitment and Survival. Mean oyster densities (+SE) of live juvenile, live adult and dead oysters at Point aux Pins (A–C) and
Alabama Port (D–F). Different letters indicate statistical differences (p,0.05) from Tukey’s HSD post-hoc tests.
doi:10.1371/journal.pone.0022396.g006
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(Table S5). Of those, only sand seatrout abundance was
significantly enhanced by breakwater reefs (Figure 9A). Silver
perch, spotted seatrout and southern kingfish showed positive
trends of enhancement, but not statistically significant. Eleven
demersal fishes were analyzed from the 10 cm catches, seven of
which were significantly enhanced by reefs including sand
seatrout, spotted seatrout, red drum and black drum (Table S5
and Figure 9B). Only finetooth shark abundance in 10 cm gillnets
was significantly greater on controls than breakwater reef
treatments. Seine samples had nine species or taxa that comprised
$1% of the total catch, including three demersal fish species and
three decapods. Of the demersal fishes, which were silver perch,
Atlantic croaker and juvenile sciaenids, only silver perch showed a
significant difference and were more common near breakwater
reefs (Figure 9C). All three decapods, caridean shrimp, penaeid
shrimp and blue crabs were present in significantly higher densities
near breakwater reefs.
Discussion
Our study found that breakwater reefs constructed of loose
oyster shell provided substrate for oyster recruitment and harbored
a more diverse community of fishes and mobile invertebrates than
control areas without reefs. This habitat enhancement is
uncommon among shoreline protection schemes and could be a
vast improvement over traditional armoring techniques, many of
which have detrimental impacts on nearshore species [16]. While
our experimental breakwaters were an ‘‘ecology-first’’ approach
Figure 7. Relative Demersal Fish and Decapod Crustacean Abundance. Mean 61 SE CPUE of (A) demersal fishes separated by collection
method and (B) decapod crustaceans collected by seines near breakwater reefs and controls. Significant differences at P#0.05 from univariate
PERMANOA tests are indicated by asterisks.
doi:10.1371/journal.pone.0022396.g007
Figure 8. Total Abundance and Demersal Fish Abundance Separated by Site. Mean+1 SE catch per unit effort of (A) total fish and
invertebrate abundance and (B) demersal fish abundance collected by 5 cm gillnets. CPUE is presented as the total individuals captured for each hour
of soak time.
doi:10.1371/journal.pone.0022396.g008
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Figure 9. Relative Abundance of Dominant Demersal Fish and Decapod Taxa. Mean+1 SE CPUE of dominant demersal and decapod species
or grouped taxa between treatments. Significant differences at P#0.05 from Wilcoxon signed rank tests comparing paired breakwater reef and
control treatments.
doi:10.1371/journal.pone.0022396.g009
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and were successful in creating valuable habitat, they did not
provide the amount of protection that could be offered by well-
engineered methodologies. This shortcoming highlights the need
for coastal protection philosophies that balance ecology and
engineering. However, an approach similar to ours could serve as
an immediate solution to the habitat losses experienced along
many sheltered coasts. In these settings, breakwater oyster reefs
that were installed seaward of already armored shorelines could
mitigate losses of fish and shellfish habitat.
Roland and Douglass (2005) found that many stretches of
Alabama’s shoreline are faced with wave energies well above
critical limits where vegetation can naturally persist and proposed
breakwaters as a potential mechanism to reduce wave energies
[26]. The wave-attenuating capacity of the breakwaters in our
study was compromised because the loose shell reefs expanded
and flattened prior to the cementing together that could result
from oyster settlement and survival. The mesh covering used in
our study to maintain the breakwater reefs’ integrity was not rigid
enough to withstand the wave energy of our sites, but an
improvement in this aspect of the breakwater design could allow
for better shoreline protection and less disturbance of the reef. To
mitigate reef spreading and flattening, we suggest the introduc-
tion of a more rigid structure as a temporary backbone which
would deteriorate or could be removed after reef cementing
occurred.
At both Alabama Port and Point aux Pins, we documented
oyster recruitment and survival to reproductive size, but
substantial mortality limited reef cementing and success. The
high mortality recorded at both sites during October 2008
sampling appeared to be caused by predation or physical
disturbance, such as wave energy. During this sampling period,
very few exposed oysters were observed to be alive. In contrast,
nearly all live oysters observed were found sheltered inside of dead,
but still hinged oyster shells. This suggests that it is unlikely disease
was the cause of mortality, since structurally-protected oysters
would have no reprieve. Another factor that frequently affects
oyster survival is reef height as tall reefs escape the poor water
quality sometimes found near the sediment [34]. As previously
discussed, the vertical relief of our reefs did decline over time, but
again it is unlikely that sheltered oysters would survive if water
quality caused the observed mortality. Physical disturbance could
have caused many of the oyster shells that were on the surface and
available for settlement to be buried under other shells, also
explaining the lowered densities of live oysters. Predation is likely
the most plausible explanation for the differential mortality
between sheltered and exposed oysters. We frequently observed
black drum, southern oyster drills (Stramonita haemastoma) and
several species of crabs near the reefs. Stomach content analysis of
the black drum collected in gillnets usually found oyster shell
remains and dead oysters often showed signs of predation (S
Scyphers, Pers. Observ.). A recent mark and recapture study of
subtidal oyster reefs in coastal Alabama waters also documented
drills as the most prevalent cause of mortality due to visible
scarring on dead spat shells [35]. The high salinities and absence
of freshets observed during the drought conditions 2007 and 2008
were likely beneficial for the oyster drill predators which thrive in
higher salinity conditions [36,37].
The communities of fishes and mobile invertebrates that benefit
from oyster reefs have been well-described, but very few studies
have examined the enhancement from oyster reefs designed for
protecting shorelines. The elevated species richness and densities
that we observed during our study concur with most literature
describing oyster reef habitats [2]. From our seines, we found blue
crabs, penaeid and caridean shrimp, and juvenile silver perch were
more abundant near oyster reefs than mudflat controls. Higher
blue crab densities near reefs were likely due to the refuge value, as
their recruitment and survival is largely augmented by structured
habitats [38]. Blue crabs support an important commercial fishery
throughout Gulf and Atlantic estuaries and, along with caridean
and penaeid shrimp, are commonly found in the diets of several of
the larger fishes. From our 10 cm gillnet sampling, we found that
spotted seatrout, drum and flounder were substantially enhanced
by oyster reefs. The paradigm of abundance, biomass and species
richness being higher in structured areas and further increasing
with habitat complexity is a pattern observed in nearly all
nearshore ecosystems [20,39–41], but the relative importance of
food versus refuge within structured habitats remains unresolved
[42,43].
Landscape attributes, such as adjacent habitats or bathymetric
features, commonly influence community composition [44–46]
and are probably responsible for the interaction between site and
treatment for the total abundance and demersal abundance of
5 cm gillnet catches. The interaction was driven by demersal fishes
(Figure 7) and these catches were dominated by Atlantic croaker
and silver perch, both which are recognized to predominately feed
in non-structured habitats [47]. Geraldi et al. (2009) found very
little evidence of enhancement by oyster reefs restored in marsh
tidal creeks and concluded that the area was not limited by
complex structure and therefore the addition of oyster shell was
functionally redundant. Grabowski et al. 2005 concluded that
small or few reefs may not measurably enhance transient
predators. Interestingly, the broad enhancement we observed
occurred in a similar setting with each reef located near
structurally-complex saltmarsh habitat and of moderate size
(,225 m
2
).
It has proven quite challenging to predict the ecosystem services
to be expected from restoring reefs at different scales or in different
settings [34,41,42]. Ecosystem services provided by shallow marine
habitats have received considerable attention from natural and
social scientists seeking to quantify and predict potential benefits
from protection or restoration [5,48,49]. Historically, most of these
studies have focused on wetlands, seagrass meadows, coral reefs
and mangroves [5,50,51], all habitats that receive considerable
protection because of their productivity. Oyster reefs also provide
important ecosystem services [18], but are more challenging to
protect and manage because they are an exploited fishery [2]. A
long history of excessive and destructive harvesting coupled with
natural stressors like disease and storms have left shellfish
populations in global demise [52–54]. Most large or landscape
scale oyster reef restoration efforts have primarily targeted the re-
establishment of harvestable oysters, many of which failed to
achieve previous population levels. Some recent studies have
detailed shortcomings of oyster restoration and cast serious doubts
on the ability to achieve restoration success in subtidal and often
large-scale efforts [55]. However, other recent studies have
documented restored reefs that have persisted over decades [56]
and on unrivaled spatial scales [57]. Attempts to quantify the
economic benefits from restoring oyster reefs are very recent and
forthcoming and could provide more support for protecting and
restoring oyster reefs for the goods and services they provide
[58,59].
Awareness of the detrimental impacts of shoreline armoring
has increased in recent years, but movement towards more
ecologically-responsible methods has been limited by the lack of
cost-effective alternatives. ‘‘Living shoreline’’ approaches, includ-
ing breakwater reefs, that protect coastal uplands could provide a
more ecologically-responsible alternative to traditional armoring
and not only mitigate coastal erosion, but also enhance certain
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economically-valuable fish stocks. However, as our study
demonstrated, efforts to sustainably and responsibly protect
coastal shoreline habitats must balance both engineering and
ecology.
Supporting Information
Table S1 Results of Repeated-Measures ANOVA on
Vegetation Retreat.
(DOCX)
Table S2 Relative Abundance of the Most Abundant
Fishes and Mobile Invertebrates.
(DOCX)
Table S3 Results from Multivariate PERMANOVA
Tests.
(DOCX)
Table S4 Results from Univariate PERMANOVA Tests.
(DOCX)
Table S5 Results from Wilcoxon Signed-Rank Tests on
Single Species or Grouped Taxa.
(DOCX)
Acknowledgments
Advice from Loren Coen, Joel Fodrie, Marti Anderson, Matt Ajemian and
Michelle Brodeur and helpful comments from Howard Browman and two
anonymous reviewers greatly improved our study and manuscript. We are
grateful for the help provided by Carly Canion, numerous technicians,
interns and graduate students that assisted in the field and lab components
of this study. We thank Scott Douglass for permission to modify and
include Figure 1. Caitlin Bovery and Mallory Scyphers lent their expertise
on bathymetry plots and data presentation. We acknowledge the Dauphin
Island Sea Lab for the hydrographic data, and we appreciate the assistance
of Laure Carassou for analysis.
Author Contributions
Conceived and designed the experiments: SS SP KH DB. Performed the
experiments: SS SP KH DB. Analyzed the data: SS SP KH DB.
Contributed reagents/materials/analysis tools: SS SP KH DB. Wrote the
paper: SS SP KH DB.
References
1. Beck MW, Heck KLJ, Able KW, Childers DL, Eggleston DB, et al. (2001) The
identification, conservation, and managment of estuarine and marine nurseries
for fish and invertebrates. Bioscience 51: 633–641.
2. Coen LD, Luckenbach MW, Breitburg DL (1999) The role of oyster reefs as
essential fish habitat: a review of current knowledge and some new perspectives.
American Fisheries Society Symposium 22: 438–454.
3. Gunter G (1967) Some Relationships of Estuaries to the Fisheries of the Gulf of
Mexico. In: Lauff GH, ed. Estuaries. Washington, D.C.: American Association
for the Advancement of Science. pp 621–638.
4. Heck KL, Hays G, Orth RJ (2003) Critical evaluation of the nursery role
hypothesis for seagrass meadows. Marine Ecology Progress Series 253: 123–136.
5. Costanza R, d’Arge R, de Groot R, Farber S, Grasso M, et al. (1997) The value
of the world’s ecosystem services and natural capital. Nature 387: 253–261.
6. Syvitski JPM, Vorosmarty CJ, Kettner AJ, Green P (2005) Impact of humans on
the flux of terrestrial sediment to the global coastal ocean. Science 308: 376–380.
7. Vitousek PM, Mooney HA, Lubchenco J, Melillo JM (1997) Human domination
of Earth’s ecosystems. Science 277.
8. NRC (2007) Mitigating shore erosion along sheltered coasts. Council NR, ed.
Washington, D.C.: National Academies Press. 174 p.
9. Pilkey OH, Wright W (1988) Sea-walls vs. Beaches. Journal of Coastal Research.
pp 41–64.
10. Douglass SL, Pickel BH (1999) The tide doesn’t go out anymore - the effect of
bulkheads on urban shorelines. Shore and Beach 67: 19–25.
11. Kennish MJ (2002) Environmental threats and environmental future of estuaries.
Environmental Conservation 29: 78–107.
12. Chambers JR (1991) Coastal Degradation and Fish Population Losses. In:
Stroud R, ed. Stemming the Tide of Coastal Fish Habitat Loss: National
Coalition for Marine Conservation. pp 45–51.
13. Jackson JBC, Kirby MX, Berger WH, Bjorndal KA, Botsford LW, et al. (2001)
Historical overfishing and the recent collapse of coastal ecosystems. Science 293:
629–638.
14. Minton MD (1999) Coastal Wetland Restoration and Its Potential Impact on
Fishery Resources in the Northeastern United States. American Fisheries Society
Symposium 22: 405–420.
15. Thayer GW, McTigue TA, Bellmer RJ, Burrows FM, Merkey DH, et al. (2003)
Science-based restoration monitoring of coastal habitats, Volume One: A
framework for monitoring plans under the Estuaries and Clean Waters Act of
2000 (Public Law 160–457). NOAA Coastal Ocean Program Decision Analysis
Series No 23 : 35 plus appendices.
16. Bilkovic D, Roggero M (2008) Effects of coastal development on nearshore
estuarine nekton communities. Marine Ecology Progress Series 358: 27–39.
17. Bruno JF, Kennedy CW (2000) Reduction of wave-related disturbance in cobble
beach habitats by Spartina alterniflora: Importance of patch size. Oecologia 122:
98–108.
18. Coen LD, Brumbaugh RD, Bushek D, Grizzle R, Luckenbach MW, et al. (2007)
AS WE SEE IT: Ecosystem Services related to oyster restoration. Marine
Ecology Progress Series 341: 303–307.
19. Dame RF, Patten BC (1981) Analysis of energy flows in an intertidal oyster reef.
Marine Ecology Progress Series 5: 115–124.
20. Orth RJ (1977) The importance of sediment stability in seagrass communities.
In: Coull B, ed. Ecology of Marine Benthos. ColumbiaSC: University of South
Carolina. pp 281–300.
21. Meyer DL, Townsend EC, Thayer GW (1997) Stabilization and Erosion Control
Value of Oyster Cultch for Intertidal Marsh. Restoration Ecology 5: 93–99.
22. Piazza BP, Banks PD, La Peyre MK (2005) The potential for created oyster shell
reefs as a sustainable shoreline protection strategy in Louisiana. Restoration
Ecology 13: 499–506.
23. Bruno JF, Stachowicz JJ, Bertness MD (2003) Inclusion of facilitation into
ecological theory. Trends in Ecology and Evolution 18: 119–125.
24. Jordan SJ, Smith LM, Nestlerode JA (2009) Cumulative effects of coastal habitat
alterations on fishery resources: toward prediction at regional scales. Ecology
and Society 14: 16.
25. Kim C-K, Park K, Powers SP, Graham WM, Bayha KM (2010) Oyster larval
transport in coastal Alabama: Dominance of physical transport over biological
behavior in a shallow estuary. Journal of Geophysical Research 115: 1–16.
26. Roland RM, Douglass SL (2005) Estimating wave tolerance of Spartina
alterniflora in coastal Alabama. Journal of Coastal Research 21: 453–463.
27. Schroeder WW, Wiseman WJJ (1985) An Analysis of the Winds (1974–1984)
and Sea Level Elevations (1973–1983) in Coastal Alabama. Mississippi-Alabama
Sea Grant Consortium Publication MASGP-84-024. pp i-102.
28. Underwood AJ (1981) Techniques for analysis of variance in experimental
marine biology and ecology. Oceanography and Marine Biology: An Annual
Review 19: 513–603.
29. Anderson MJ (2001) A new method for non-parametric multivariate analysis of
variance. Austral Ecology 26: 32–46.
30. Anderson MJ (2001) Permutation tests for univariate or multivariate analysis of
variance and regression. Canadian Journal of Fisheries and Aquatic Sciences 58:
626–639.
31. Froese R, Pauly D (2010) Editors, Fishbase. World Wild Web Electronic
Publication. www.fishbase.org, version (11/2010).
32. Clarke KR, Gorely RN (2006) Primer v6: User Manual/Tutorial. PRIMER-E,
Plymouth.
33. CRAN (2010) R: A language and environment for statistical Computing. R
Foundation for Statstical Computing, Vienna, Austria ISBN 3-900051-07-0,
URL http://wwwR-projectorg.
34. Lenihan HS (1999) Physical-biological coupling on oyster reefs: how habitat
structure influences individual performance. Ecological Monographs 69:
251–275.
35. Gregalis KC, Powers SP, Heck KLJ (2008) Restoration of oyster reefs along a
bio-physical gradient in Mobile Bay, AL. Journal of Shellfish Research 27:
1163–1169.
36. Wells HW (1961) The fauna of oyster beds, with special reference to the salinity
factor. Ecological Monographs 31: 239–266.
37. Garton D, Stickle WB (1980) Effects of salinity and temperature on the
predation rate of Thais haemastoma on Crassostrea virginica spat. Biological
Bulletin 158: 49–57.
38. Moksnes P-O, Heck KL (2006) Relative importance of habitat selection and
predation for the distribution of blue crab megalopae and young juveniles.
Marine Ecology Progress Series 308: 165–181.
39. Goreau T, Goreau N, Goreau T (1979) Corals and coral reefs. Scientific
American 241: 124–136.
40. Rozas LP, Odum WE (1987) Fish and macrocrustacean use of submerged plant
beds in tidal freshwater marsh creeks. Marine Ecology Progress Series 38:
101–108.
Mitigating Shoreline Loss and Facilitating Fish
PLoS ONE | www.plosone.org 11 August 2011 | Volume 6 | Issue 8 | e22396
41. Wittman J (1985) Refuges, biological disturbance, and rocky subtidal community
structure. New England Ecological Monographs 55: 421–445.
42. Bostrom C, Mattila J (1999) The relative importance of food and shelter for
seagrass-associated invertebrates: a latitudinal comparison of habitat choice by
isopod grazers. Oecologia 120: 162–170.
43. Grabowski JH, Powers SP (2004) Habitat complexity mitigates trophic transfer
on oyster reefs. Marine Ecology Progress Series 277: 291–295.
44. Geraldi NR, Powers SP, Heck KL, Cebrian J (2009) Can habitat restoration be
redundant? Response of mobile fishes and crustaceans to oyster reef restoration
in marsh tidal creeks. Marine Ecology Progress Series 389: 171–180.
45. Grabowski JH, Hughes AR, Kimbro DL, Dolan MA (2005) How habitat setting
influences restored oyster reef communities. Ecology 86: 1926–1935.
46. Micheli F, Peterson CH (1999) Estuarine vegetated habitats as corridors for
predator movements. Conservation Biology 13: 869–881.
47. Peterson CH, Grabowski JH, Powers SP (2003) Estimated enhancement of fish
production resulting from restoring oyster reef habitat: quantitative valuation.
Marine Ecology Progress Series 264: 249–264.
48. Hobbs RJ, Harris LE (2001) Restoration ecology: repairing Earth’s ecosystems in
the new millenium. Restoration Ecology 9: 239–246.
49. Peterson CH, Lipcius RN (2003) Conceptual progress towards predicting
quantitative ecosystem benefits of ecological restorations. Marine Ecology
Progress Series 264: 297–307.
50. Moberg F, Folke C (1999) Ecological goods and services of coral reef ecosystems.
Ecological Economics 29: 215–233.
51. Ronnback P (1999) The ecological basis for economic value of seafood
production supported by mangrove ecosystems. Ecological Economics 29:
235–252.
52. Beck MW, Brumbaugh RD, Airoldi L, Carranza A, Coen LD, et al. (2011)
Oyster reefs at risk and recommendations for conservation, restoration, and
management. Bio Science 61: 107–116.
53. Kirby MX (2004) Fishing down the coast: historical expansion and collapse of
oyster fisheries along continental margins. Proceedings of the National Academy
of Sciences 101: 13096–13099.
54. Rothschild B, Ault J, Goulletquer P, Heral M (1994) Decline of Chesapeake Bay
oyster population: a century of habitat destruction and overfishing. Marine
Ecology Progress Series 111: 29–39.
55. Mann R, Powell EN (2007) Why oyster restoration goals in the Chesapeake Bay
are not and probaly cannot be acheived. Journal of Shellfish Research 26:
905–917.
56. Powers SP, Peterson CH, Grabowski JH, Lenihan HS (2009) Success of
constructed oyster reefs in no-harvest sanctuaries: implications for restoration.
Marine Ecology Progress Series 389: 159–170.
57. Schulte DM, Burke RP, Lipcius RN (2009) Unprecedented restoration of a
native oyster metapopulation. Science 325: 1124–1128.
58. Brumbaugh RD, Torropova C (2008) Economic valuation of ecosystem services:
A new impetus for shellfish restoration? Basins and Coasts 2: 8–15.
59. Grabowski JH, Peterson CH (2007) Restoring Oyster Reefs to Recover
Ecosystem Services. In: Cuddington K, Byers JE, Wilson W, Hastings A, eds.
Ecosystem Engineers, 4: Academic Press.
Mitigating Shoreline Loss and Facilitating Fish
PLoS ONE | www.plosone.org 12 August 2011 | Volume 6 | Issue 8 | e22396
