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Protecting coastal communities has become increasingly important as their populations grow, resulting in increased demand for engineered shore protection and hardening of over 50% of many urban shorelines. Shoreline hardening is recognized to reduce ecosystem services that coastal populations rely on, but the amount of hardened coastline continues to grow in many ecologically important coastal regions. Therefore, to inform future management decisions, we conducted a meta-analysis of studies comparing the ecosystem services of biodiversity (richness or diversity) and habitat provisioning (organism abundance) along shorelines with versus without engineered-shore structures. Seawalls supported 23% lower biodiversity and 45% fewer organisms than natural shorelines. In contrast, biodiversity and abundance supported by riprap or breakwater shorelines were not different from natural shorelines; however, effect sizes were highly heterogeneous across organism groups and studies. As coastal development increases, the type and location of shoreline hardening could greatly affect the habitat value and functioning of nearshore ecosystems.
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Overview Articles September 2016 / Vol. 66 No. 9 BioScience 763
Ecological Consequences
of Shoreline Hardening:
A Meta-Analysis
Protecting coastal communities has become increasingly important as their populations grow, resulting in increased demand for engineered shore
protection and hardening of over 50% of many urban shorelines. Shoreline hardening is recognized to reduce ecosystem services that coastal
populations rely on, but the amount of hardened coastline continues to grow in many ecologically important coastal regions. Therefore, to inform
future management decisions, we conducted a meta-analysis of studies comparing the ecosystem services of biodiversity (richness or diversity)
and habitat provisioning (organism abundance) along shorelines with versus without engineered-shore structures. Seawalls supported 23%
lower biodiversity and 45% fewer organisms than natural shorelines. In contrast, biodiversity and abundance supported by riprap or breakwater
shorelines were not different from natural shorelines; however, effect sizes were highly heterogeneous across organism groups and studies. As
coastal development increases, the type and location of shoreline hardening could greatly affect the habitat value and functioning of nearshore
Keywords: biodiversity, bulkhead, ecosystem function, seawall, shoreline hardening
Over the last two centuries, humans have rapidly and
dramatically altered the global landscape, causing
many to refer to this period as the Anthropocene epoch
(Steffen et al. 2007). Some of the strongest examples of
anthropogenic change can be found along coastlines. With
roughly one-third of human populations living within 100
kilometers of a coastline and continued migration toward
coastal areas expected to increase this proportion to one-
half by 2030 (Small and Nicholls 2003, MEA 2005), coastal
ecosystems are among the most modified and threatened
globally (Adger et al. 2005). In efforts to protect people,
property, and critical infrastructure from coastal hazards
(e.g., erosive waves, storms, and flooding), as well as achieve
other human aspirations (e.g., maritime docking, and navi-
gation), coastal societies have historically armored or hard-
ened shorelines with a variety of engineering structures
(Dugan et al. 2011). Shoreline hardening, defined as the
installation of engineered-shore structures to (a) stabilize
sediment and prevent erosion and/or (b) provide flood
protection, is a common practice worldwide, with over
22,000 kilometers (roughly 14%) of shoreline hardened in
the United States alone (Gittman etal. 2015). Major coastal
cities such as New York, Sydney, and Hong Kong have 50%
or more of their shorelines hardened (Chapman and Bulleri
2003, Lam etal. 2009, Gittman etal. 2015). Given the current
levels of shoreline hardening and the projected growth of
coastal populations, understanding the ecological effects of
these structures is crucial for developing sustainable coastal
management and climate-adaptation strategies (Titus etal.
1998, Gittman etal. 2015). Specifically, understanding how
shoreline hardening affects biodiversity and ecosystem func-
tioning is necessary for evaluating the consequences of these
activities on associated ecosystem services, such as fisheries
production, property protection, and water quality benefits,
to coastal communities (Arkema etal. 2015, Scyphers etal.
2015, Gittman etal. 2016).
Although conservation and restoration practitioners have
been advocating for the implementation of “living shore-
lines” or “nature-based” strategies in lieu of traditional
“hard” approaches, such as seawalls or bulkheads, over
the last three decades (see Broome et al. 1988, Currin
et al. 2007), the science on the ecological consequences
of various shore-protection structures has lagged behind
(NRC 2007). Recent narrative reviews have identified many
of the impacts of engineered-shore structures on coastal
ecosystems and have recommended ways to minimize these
BioScience 66: 763–773. © The Author(s) 2016. Published by Oxford University Press on behalf of the American Institute of Biological Sciences. This is an
Open Access article distributed under the terms of the Creative Commons Attribution Non-Commercial License (
by-nc/4.0/), which permits non-commercial re-use, distribution, and reproduction in any medium, provided the original work is properly cited. For
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doi:10.1093/biosci/biw091 Advance Access publication 10 August 2016
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impacts (Chapman and Underwood 2011, Dugan etal. 2011,
Perkins etal. 2015); however, a comparative and quantita-
tive synthesis of the effects of engineered-shore structures
on coastal ecosystem services has yet to be conducted. The
purpose of this systematic review and meta-analysis was to
synthesize, quantify, and compare the effects of commonly
used engineered-shore structures on the coastal ecosystem
services of biodiversity and habitat provision. Moreover,
such a synthesis can help inform the development of effec-
tive coastal conservation policies and management actions.
To evaluate the biodiversity and habitat provision effects
of different engineered-shore structures, we conducted a
systematic review of all studies comparing the biodiversity
or abundance of organisms on shorelines with engineered
structures versus unmodified shorelines. Three catego-
ries of engineered-shore structures were considered: (1)
seawalls and bulkheads (figure 1a); (2) riprap revetments
(figure 1b); and (3) breakwaters and sills (figure 1c). For
the purposes of this review, all vertical walls constructed
parallel to shore in or above the high intertidal zone are
termed seawalls (figure 1a). Shore-parallel, sloped structures
constructed of unconsolidated rock or rubble in or above
the high intertidal zone are referred to as riprap revetments
(figure 1b). Structures constructed within the low intertidal
or subtidal zones are referred to as breakwaters (figure 1c).
We have elected to use the term breakwater in lieu of sill in
accordance with the terminology used by the United States
Army Corps of Engineer (USACE) in their guidance docu-
ment Low Cost Shore Protection (2001). The materials used
to construct the structures evaluated in the selected studies
vary and include concrete, granite or sandstone rock, marl,
wood, and vinyl sheeting. We defined natural shorelines as
rocky, soft-sediment, or biogenic (e.g., marshes, mangroves,
oyster reefs, or coral reefs present) shorelines without any
engineered-shore structures or modifications (figure 1d–f).
Peer-reviewed literature search. Using the Web of Science data-
base and the Google Scholar search engine, we searched the
literature with the following search terms: structure type
(seawall OR bulkhead OR riprap, OR breakwater OR sill)
AND response metric (richness OR diversity OR abundance
OR density OR cover OR growth OR fitness OR “ecosystem
service” OR habitat) AND shoreline hardening indicators
(“shore hard” OR “shore armor” OR “shore stabiliza-
tion” OR “shore protection”) to account for all literature
available by 5 November 2015. A total of 121 studies were
selected after reviewing the title, keywords, and abstract to
determine whether each study evaluated the effects of engi-
neered-shore structures on one or more ecological response
variables (e.g., species richness, and abundance). Of those
Figure 1. Example of engineered-shore structures: (a) a seawall; (b) riprap revetment; (c) breakwater; and natural
shorelines compared in this study: (d) rocky shoreline (granite platforms); (e) soft-sediment shoreline (sand beach); and (f)
biogenic shoreline (salt marsh). Rocky shorelines consist of consolidated rocky platforms and/or cobbles and boulders.
Soft-sediment shorelines consist of unconsolidated sediments (sands, muds, silts, clays) without intertidal vegetation.
Biogenic shorelines can include intertidal and shallow subtidal marsh, mangrove, bivalve or coral reef, or seagrass.
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studies, we only included those that compared the eco-
logical effects of one or more engineered-shore structures
with those of natural shorelines (e.g., unmodified rocky,
soft-sediment, or biogenic shores; figure 1d–f). Studies that
evaluated the ecological effects of biogenic methods of shore
stabilization (e.g., oyster or marsh restoration) alone were
not included because they could also be considered biogenic
habitat restoration. However, if the study compared the
effects of biogenic habitat restoration, such as marsh plant-
ing, combined with construction of an engineered-shore
structure (e.g., a rock breakwater) with those of a natural
shoreline, then the study was included in the analysis. The
evaluation of biogenic habitat restoration effectiveness in
restoring, enhancing, or sustaining ecosystems functions
has been covered elsewhere (e.g., Peterson and Lipcius
2003, Benayas etal. 2009, Shepard etal. 2011, Baggett etal.
2015) and is beyond the scope of this review. Finally, only
studies evaluating the effects of engineered-shore structures
on coastal shorelines (including open coast, estuarine, bay,
lagoon, and tidally influenced riverine shorelines) were
included. Studies of nontidal riverine or lake shorelines,
such as shorelines along the Great Lakes, United States,
were not included. Applying these criteria yielded 54 studies
for further review and analysis (supplemental appendix
S1). In 52 of the 54 studies considered, a control-impact
(CI) approach was used to compare hardened shorelines to
natural shorelines, whereas only two studies sampled hard-
ened and natural shorelines before and after hard shoreline
structures were installed (BACI design; e.g., Gittman etal.
2016). Studies that converted or experimentally manipu-
lated the configuration or substrate of hardened shorelines
(e.g., Bulleri 2005) were beyond the scope of this review and
therefore not included.
Data extraction. We extracted the means, standard deviations,
and sample sizes of community (e.g., taxonomic richness
and diversity) and individual taxa metrics (e.g., abundance,
density, percent cover, and biomass) for hardened and natural
shorelines from 32 of the 54 studies (table 1). The remain-
ing 22 studies either did not report the means, standard
deviations, or sample sizes for community and individual
taxa metrics in an extractable format or had no replication
(n = 1) at the level of shoreline type (e.g., seawall or natural;
supplemental table S1). Data were extracted from the text,
tables, and figures, with data extracted from figures using the
software program Data Thief (Tummers 2006). Data for each
metric were extracted for each structure or natural-shore
comparison (figure 1), with means averaged and standard
deviations calculated across replicate sites, time, and species
within a phylum or subphylum, but separately by shore zone
sampled (e.g., high intertidal, low intertidal, and subtidal)
and habitat-use group (flora, benthic infauna, birds, epibiota,
and nekton) when reported. Responses for flora included
marsh plants, mangroves, and upland shore plants; benthic
infauna included organisms living within soft sediments
(e.g., bivalves, amphipods, and polychaetes); birds included
shorebirds, gulls, and other waterfowl; epibiota include
both sessile and mobile organisms living on the surface of
the shoreline substrate (e.g., algae, bivalves, barnacles, and
gastropods); and nekton included fishes and free-swimming
crustaceans. Organisms were grouped into these categories
on the basis of their habitat use (e.g., benthic infauna versus
nekton) and groupings commonly used in the studies (e.g.,
Statistical analyses. We calculated effect sizes and correspond-
ing sampling variances for community and individual taxa
metrics as log response ratios; the proportional difference
between the means for three types of hardened shorelines
(seawalls, riprap revetments, and breakwaters) and natural
shorelines (Hedges etal. 1999). On the log scale, an effect
size of zero means no difference, whereas a negative value
means that the hardened shorelines had lower commu-
nity and individual metrics than natural shorelines. We fit
meta-analytic random-effects models to pooled community
(i.e., overall and organism-group biodiversity) and pooled
individual (i.e., overall and organism-group abundance)
effect sizes separately for seawall, riprap revetment, and
breakwater comparisons to natural shorelines. Because not
all response variables were restricted to specific shore zones
(e.g., subtidal or intertidal) and shore zonation classifica-
tions were not consistent across studies, we did not compare
effect sizes across shore zones. The total amount of residual
heterogeneity (τ2) was calculated for each model using
restricted maximum-likelihood estimation to account for
covariance among responses (Viechtbauer 2010). Residual
heterogeneity is variability among the true effects that is not
accounted for by the model (Viechtbauer 2010). Differences
in the functional responses of organism groups measured,
the study ecosystems, and study methods could all contrib-
ute to heterogeneity in effect sizes. To allow for consider-
ation of the potential sources of heterogeneity, we explored
the effect sizes across organism groups, as well as across
studies, through forest plots. Forest plots are recommended
for visually assessing the number and precision of the studies
included in the meta-analysis and the heterogeneity across
effect sizes (Vetter et al. 2013). We included several effect
sizes from the same publication, which are not independent;
therefore, we included a publication-level random effect to
account for the interdependency among multiple within-
study observations. To determine the percent difference in
biodiversity and abundance between hardened and natural
shorelines, we back-transformed the log response ratios and
then converted the back-transformed value to a percentage.
The potential for biases in favor of “significant effects” in
the published literature (the file drawer problem) is a con-
cern when conducting meta-analyses (Gillman and Wright
2010). To test for “file drawer” bias, we constructed funnel
plots for each random effects model and evaluated funnel
plot asymmetry using a regression test (Egger etal. 1997).
A funnel plot assumes that studies with smaller sample sizes
and higher sampling variances are more likely to be skewed
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and are less likely to be published (Duval and Tweedie 2000).
Therefore, asymmetry in the published data can be detected
by testing whether the observed effects are related to their
sampling sizes. In addition, when a significant effect size
was observed, we calculated Rosenthal’s fail-safe number,
which is the number of unpublished studies, with a mean
effect of zero, required to eliminate a significant overall effect
size (Rosenthal 1979, Møller and Jennions 2001). A fail-safe
number of 5K + 10 or higher, with K being the number of
number of studies included in the analysis, is considered to
be evidence of a robust average effect size (Rosenthal 1991).
All analyses were carried out using R 3.0.1 (R Development
Core Team 2016) with the R package metaphor (Viechtbauer
Of the 32 studies included in the analyses, 78% evaluated
seawalls, 28% evaluated riprap revetments, and 25% evalu-
ated breakwaters (table 1). More studies compared the eco-
system function of hardened shorelines with that of biogenic
shorelines (n = 16) than with that of rocky (n = 12) or soft-
sediment shorelines (n = 8). Most studies were conducted
along the Atlantic and Gulf of Mexico coasts of the United
States. All of the studies were published since the year 2000,
with nearly half of the studies published since 2010.
Seawalls. The overall mean log response ratio (LRR) between
seawalls and natural shorelines for biodiversity was 0.26
(95% CI: 0.40, 0.12); therefore, we found that biodiversity
Table 1. Studies included in the meta-analyses.
Authors Year Seawall Riprap Breakwater Rocky
Sediment Biogenic Flora
Infauna Birds Epibiota Nekton
Bilkovic and Mitchell 2013 X X X A A AB
Bilkovic and Roggero 2008 X X X B
Bozek and Burdick 2005 X X A
Bulleri and Chapman 2004 X X X A
Bulleri etal. 2004 X X A
Bulleri etal. 2005 X X A
Burt etal. 2009 X X AB AB
Chapman 2003 X X B
Chapman 2005 X X AB
Currin etal. 2007 X X A AB
Diaz-Agras etal. 2010 X X A
Drexler at al. 2013 X X A
Dugan and Hubbard 2006 X X AB
Dugan etal. 2008 X X AB AB
Gittman etal. 2016 X X X A A AB
Glasby etal. 2007 X X B
Harris and Strayer 2014 X X X B
Heatherington and
Bishop 2012 X X A
Hendon etal. 2000 X X X A
Jackson etal. 2015 X X A A
Lam etal. 2009 X X A
Lawless and Seitz 2014 X X X AB A
Lee and Li 2013 X X X A
Long etal. 2011 X X X AB A
Moreira etal. 2006 X X A
Morley etal. 2012 X X A AB A
O’Conner etal. 2010 X X A AB
Peters etal. 2015 X X A AB
Peterson etal. 2000 X X X A
Seitz etal. 2006 X X X AB AB
Sobocinski etal. 2010 X X X AB B
Strayer etal. 2012 X X X X B A AB A
Note: “A” indicates abundance data and “B” indicates biodiversity data.
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was 23% (CI: 33, 11) lower along shorelines with seawalls
when compared with that of natural shorelines (z = 3.67,
df = 10, p <.001, figure 2a and supplemental figure S1a).
The mean LRRs with 95% CIs between seawalls and natural
shorelines for biodiversity were negative across all natural
shoreline types (biogenic: 0.22 [0.41, 0.03], rocky: 0.12,
[0.24, 0.01], and soft sediment: 0.52, [0.78, 0.26]) and
all organism groups except epibiota (figure 3a). Biodiversity
was significantly lower for flora (66%; [80, 41]), benthic
infauna (20%; [33, 4]), birds (52%; [66, 34]), and
nekton (24%; [37, 10]; supplemental figure S2a).
The LRR for organism abundance between seawalls and
natural shorelines was 0.61 (0.98, 0.23), corresponding to
45% (62, 21) lower abundances of organisms along shore-
lines with seawalls when compared with those along natural
shorelines (z = 3.20, df = 21, p = .001, figures 2b and S1b).
The mean LRRs with 95% CIs for abundance were negative
for biogenic (0.74, [1.25, 0.22]) and soft sediment (1.11,
[1.72, 0.51]), but not rocky (0.64, [1.43, 0.16]) shoreline
comparisons. All organism groups except flora and epibiota
had negative mean LRRs with 95% CIs for
abundance between seawalls and natural
shorelines (figure 4a). The abundance of
benthic infauna, birds, and nekton were
66% (88, 8), 71% (86, 41), and 56%
(79, 9) lower, respectively, along shore-
lines with seawalls when compared with
along natural shorelines (supplemental
figure S3a).
Our meta-analyses for seawalls
included 20 biodiversity and 67 abun-
dance responses from 25 studies. The
total heterogeneity (τ2) in the true
effect sizes for biodiversity and abun-
dance were estimated to be 0.02 and
0.54, respectively. Fifty-four percent
of the total variability in biodiversity
effect sizes and 83% of the total vari-
ability in abundance effect sizes were
attributed to heterogeneity in the true
effects (I2). Although true effect size
estimates were heterogeneous for both
biodiversity (q = 26.25, df = 10, p = .003)
and abundance (q = 123.53, df = 21,
p < .001), nearly half of the studies
found significant, negative effects of
seawalls on the biodiversity and abun-
dance of organisms (supplemental fig-
ures S4 and S5). There was no evidence
of “file drawer bias” or asymmetry in
the published data for comparisons of
biodiversity or abundance of organisms
between seawalls and natural shorelines
(z = 1.51, p= .13 and z = 1.04, p = .30,
respectively). The Rosenthal fail-safe
number for the observed biodiversity
effect, or the number of unpublished studies with a mean
effect size of zero, needed to eliminate the overall effect
size at α = .05 is 122, which is greater than the 65 stud-
ies required for a robust effect size estimate. To eliminate
the overall observed abundance effect for seawall–natural
shoreline comparisons, 651 unpublished studies with a
mean effect size of zero would be needed, which is greater
than the number of studies required (n = 120) for the effect
size to be considered robust.
Riprap revetments. There was no difference in the biodiver-
sity or abundance of organisms found along shorelines with
riprap revetments and natural shorelines, with the mean
LRRs not being significantly different from zero (z = 1.82,
df = 7, p = .07 and z = 1.64, df = 6, p = .10, respectively,
figures 2 and S1). Mean biodiversity and abundance did not
differ between riprap and natural shorelines across organ-
ism groups (figures 3b, 4b, S2, and S3), with the exception
of a 39% (CI: 59, 9) reduction in flora biodiversity along
riprap shorelines (LRR = 0.49, 95% CI: 0.89, 0.09,
Log Response Ratio
Log Response Ratio
Seawall Riprap Breakwater
11 (20) 8 (14) 5 (11)
22 (67) 7 (22)
8 (36)
Figure 2. Overall log response ratios between engineered-shore structures
(seawall, riprap, breakwater) and natural shorelines for (a) biodiversity and
(b) abundance. The error bars represent 95% confidence intervals and data labels
show the number of studies and the total number of responses from the studies.
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768 BioScience September 2016 / Vol. 66 No. 9
figures 3b and S2). The total heterogeneity (τ2) in the true
effect sizes for biodiversity and abundance were estimated
to be 0.03 and 0.23 respectively. 78% of the total variability
in biodiversity effect sizes and 41% the total variability
in abundance effect sizes were attributed to heterogene-
ity in the true effects (I2). The true effect size estimates
were heterogeneous for both biodiversity (q = 29.37, df =
7, p < .001) and abundance (q = 0.07, df = 6, p < .001) and
varied considerably across studies (supplemental figures S6
and S7). Finally, we did not find evidence of “file drawer
bias” or asymmetry (z = 1.84, p = .07 and z = 0.78, p = .43,
Breakwaters. Similar to the results for
riprap revetments, there was no differ-
ence in the biodiversity or abundance of
organisms found along shorelines with
breakwaters when compared with those
along natural shorelines (figure 2). The
mean LRRs were not significantly dif-
ferent from zero (z = 0.46, df = 4, p = .65
and z = 0.97, df = 7, p = .33, respectively,
figures 2 and S1). The 95% CIs for
the mean biodiversity and abundance
LRRs encompassed zero for all organism
groups except for 39% (3, 88) greater
biodiversity of nekton (LRR = 0.33,
[0.03, 0.63], figures 3c, 4c, S2, and S3) on
shorelines with breakwaters compared
with that on natural shorelines. The total
heterogeneity (τ2) in the true effect sizes
for biodiversity and abundance were
estimated to be 0.32 and 0.22, respec-
tively. 96% of the total variability in
biodiversity effect sizes and 82% of the
total variability in abundance effect sizes
were attributed to heterogeneity in the
true effects (I2). The true effect size esti-
mates were heterogeneous for both bio-
diversity (q = 50.74, df = 4, p < .001) and
abundance (q = 42.86, df = 7, p < .001)
and varied considerably across stud-
ies (supplemental figures S8 and S9).
There was no evidence of “file drawer
bias” for published studies compar-
ing breakwaters and natural shorelines
(z = 0.81, p = .42 and z = 1.14, p = .25,
The design of engineered-shore struc-
tures and their functional similarity to
natural shorelines varies widely across
and within structure types (figure 1a–f;
Nordstrom 2014, Perkins et al. 2015).
Moreover, our analyses revealed some
clear distinctions in the quality of habitat
provided by the most common engineering alternatives to
natural shorelines. Most importantly, seawalls typically sup-
ported lower biodiversity and abundance of organisms than
did natural shorelines, indicating that these engineered-shore
structures are adversely affecting coastal ecosystems (figure
2). Biodiversity and abundance did not differ significantly
on riprap and breakwaters from natural shorelines; how-
ever, this lack of difference may reflect heterogeneity in the
effects of riprap or breakwaters across organism groups,
as well as a small number of studies (figures 2–4, S6–S9).
Studies included in this meta-analysis proposed that struc-
ture complexity and composition of substrate (Chapman and
Log Response Ratio
Flora Ben. Infauna Birds Epibiota Nekton
1 (2)
5 (5)
2 (2)
5 (9)
2 (2)
1 (2)
4 (4)
4 (6)
2 (2)
2 (3)
4 (8)
Figure 3. Overall log response ratios between engineered- shore structures (a)
seawall, (b) riprap, (c) breakwater and natural shorelines for biodiversity of
flora, benthic infauna, birds, epibiota, and nekton. The error bars represent
95% confidence intervals and data labels show the number of studies and the
total number of responses from the studies.
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Bulleri 2003, Seitz etal. 2006, Gittman etal. 2016), structure
placement within the intertidal or subtidal zones (Bozek
and Burdick 2005, Dugan etal. 2008, Bilkovic and Mitchell
2013), and associated wave and sediment dynamics (Bulleri
etal. 2004, Bulleri and Chapman 2004, Strayer etal. 2012),
may determine whether the biodiversity and abundance of
organisms differ between engineered and natural shorelines.
Therefore, we explored how reported differences in seawall,
riprap, and breakwater structure complexity, composition,
and placement related to the biodiversity and abundance of
different organism groups below.
Structure complexity and composi-
tion. Intertidal and shallow subtidal
habitats, particularly structurally complex
biogenic habitats (e.g., wetlands, man-
groves, and oyster reefs), provide refuge
for numerous small and juvenile nekton
species (e.g., Fundulus spp., Able et al.
2012; Penaeid shrimp, Boesch and Turner
1984); from abiotic stress (e.g., wave
energy, Möller et al. 2014); and from
predation (Peterson and Turner 1994).
Seawalls can alter the habitat available
to nekton by reducing the complexity
of intertidal and subtidal habitats (e.g.,
Chapman and Bulleri 2003, Bilkovic and
Roggero 2008). The vertical profile and
typically uniform surface of seawalls
(figure 1a) does not offer the same refuge
for nekton as boulders and camouflaging
sediment (Strayer et al. 2012), or dense
marsh vegetation (Hendon et al. 2000,
Peterson et al. 2000, Seitz et al. 2006,
Bilkovic and Roggero 2008, Gittman etal.
2016) characteristic of natural shorelines.
The lack of complexity along seawalls
likely explains why biodiversity and
abundance is lower for many organisms
than along natural shorelines.
The biodiversity and abundance of
nekton were similar between riprap and
natural shorelines (e.g., Seitz etal. 2006,
Bilkovic and Roggero 2008, Strayer etal.
2012) and potentially greater at shore-
lines with breakwaters when compared
with those at biogenic shorelines (only
type of natural shoreline evaluated, Burt
et al. 2009, Peters et al. 2015, Gittman
et al. 2016). Because both riprap and
breakwaters typically consist of piles of
unconsolidated rock and rubble of vary-
ing sizes and shapes (figures 1b and
1c; Nordstrom 2014), these structures
may provide nekton with equivalent or
greater refuge from predation or access
to food resources (e.g., epibiota, Clynick
et al. 2007; benthic infauna, discussed below) when com-
pared with less structurally complex natural shorelines.
In contrast to nekton, there were no differences in the
biodiversity or abundance of epibiota between seawalls
or breakwaters and natural shorelines (figures 3 and 4).
Epibiota include both sessile organisms such as algae, oys-
ters, mussels, and barnacles, and mobile organisms such as
limpets, chitons, snails, and whelks that live on the surface
of hard substrates (e.g., shells, rocks, and plants). Despite
many studies reporting no differences in epibiota biodiver-
sity (e.g., species richness and Shannon diversity, Glasby
Log Response Ratio
Flora Ben. Infauna Birds Epibiota Nekton
2 (4) 6 (12) 3 (5)
11 (29)
8 (17)
7 (12)
3 (3)
5 (7)
4 (8)
1 (1)
5 (8)
4 (12)
Figure 4. Log response ratios between engineered-shore structures (a) seawall,
(b) riprap, (c) breakwater and natural shorelines for abundance of flora,
benthic infauna, birds, epibiota, and nekton. The error bars represent 95%
confidence intervals and data labels show the number of studies and the total
number of responses from the studies.
by guest on September 6, 2016 from
Overview Articles
770 BioScience September 2016 / Vol. 66 No. 9
etal. 2007) or overall abundance (e.g., Bulleri and Chapman
2004, Bulleri etal. 2005, Lam etal. 2009), several of these
same studies did report differences in community composi-
tion or dominance (via multivariate ordinations), particu-
larly for mobile grazers, such as limpets, snails, whelks, and
chitons (e.g., Bulleri and Chapman 2004, Lam etal. 2009).
Further research is needed to understand whether shore-
line structures that induce shifts in grazer communities
also affect the structure and function of nearshore marine
Of the few studies that reported differences in epibiota, a
majority compared engineered-shore structures with soft-
sediment or biogenic shorelines. Epibiota biodiversity and
abundance were lower along shorelines with seawalls and
riprap revetments when compared with those along soft-
sediment shores (Sobocinski etal. 2010, Strayer etal. 2012,
Harris et al. 2014). In contrast, studies reported higher
epibiota diversity and abundance on riprap and breakwaters
than on marsh and mangrove shorelines (O’Connor et al.
2010, Drexler etal. 2013, Peters etal. 2015, Gittman etal.
2016). Some shore-protection structures may serve as sur-
rogate habitats for native epibiota where natural hard sub-
strates, such as oyster reefs and mussel beds, have been lost
to overharvest, erosion, and poor water quality (Beck etal.
2011). However, the introduction of some types of hard sub-
strates into soft-sediment and biogenic shorelines may also
facilitate invasive species. Therefore, the location relative to
invasion pathways and substrate type should be carefully
considered (Ruiz etal. 1997).
Structure placement and associated wave and sediment
dynamics. Intertidal and shallow soft-sediment and biogenic
habitats provide refuge for benthic infauna—such as clams
(Seitz et al. 2006) and burrowing crustaceans (Dugan et al.
2008)—from predation (Lipcius et al. 2005) and are often
occupied by marine flora, such as marsh plants, mangroves,
and seagrasses. Larger nektonic predators (e.g., blue crabs) and
shorebirds (e.g., sand pipers, willets, and wading birds) forage
in soft-sediment and biogenic intertidal and shallow subtidal
habitats (Kneib 1982). Lower biodiversity and abundance of
benthic infauna and birds were associated with narrower soft-
sediment shores along seawalls (Dugan and Hubbard 2006,
Dugan etal. 2008), and lower abundances of benthic infauna
were also associated with coarser sediments (Sobocinski etal.
2010), leading us to conclude that seawalls reduced both the
quantity and quality of habitat available to these organisms.
Because they are typically placed in the high intertidal zone
(Titus et al. 1998), installation of a seawall and to a lesser
extent, a riprap revetment, can severe the connection between
upland and intertidal habitat, reflect wave energy and alter
sediment transport, and potentially increasing the depth of
the intertidal and nearshore subtidal zones reported in several
studies (Ruggiero and McDougal 2001, Peregrine 2003).
The loss or disruption of habitat suitable to upland flora
species by seawalls and riprap is likely the cause of the
reduced biodiversity observed by Strayer and colleagues
(2012) and the complete absence of high marsh at seawall
sites studied by Bozek and Burdick (2005). The loss of veg-
etated habitat can alter nutrient cycling in the intertidal (e.g.,
lower denitrification rates, O’Meara etal. 2015) and reduce
pollutant filtration (Reboreda and Cacador 2007), which
could have cascading effects via shifts in nutrient availabil-
ity and the bioaccumulation of toxins in benthic infauna,
epibiota, nekton, and birds (Franca et al. 2005). Unlike
the studies of seawalls and riprap, studies included in this
meta-analysis suggested that breakwaters can decrease the
depth of the shoreline via sediment deposition landward of
the breakwater, promoting the persistence of intertidal flora
such as marsh plants (Currin etal. 2007, Gittman etal. 2014,
2016). Flora abundance effects were only estimated from
one short-duration study on marsh dominated by Spartina
alterniflora and one short-duration study on the mangrove,
Avicennia marina. Both S. alterniflora and A. marina occupy
habitat seaward of typical seawall placement, leaving these
species vulnerable to loss from reflected, wave-induced ero-
sion or sea-level rise, often termed “coastal squeeze” (Pontee
2013), over longer (e.g., decadal) time scales (Titus et al.
1998), perhaps explaining why the above two short-term
studies did not find a difference between shorelines with
seawalls versus natural shorelines.
Study limitations. There was significant heterogeneity across
organism groups (figures 3, 4, S2, and S3) and studies
(figures S4–S9) for all structure types. However, seawalls had
a significant negative effect when compared with natural
shorelines for at least one metric (biodiversity or abundance)
for more than half of all studies and for all organism groups
except epibiota. Riprap and breakwater effects were more
heterogeneous than seawall effects in both magnitude and
direction across organism groups. There were fewer studies
on the ecological effects of riprap revetments (n = 9) and
breakwaters (n = 8) than seawalls (n = 25), which may have
increased heterogeneity in effect sizes and therefore limited
our ability (statistical power) to detect the effects of these
shore-protection structures relative to seawalls. However,
our results do suggest that some organism groups may be
adversely affected by riprap (e.g., flora and epibiota) or posi-
tively affected by breakwaters (e.g., nekton). Flora, such as
marsh plants, seagrasses, and mangroves, were represented
by only a single riprap study; however, a study by Patrick and
colleagues, which did not meet our criteria to include in the
analysis, showed a significant negative correlation between
seagrass percent cover and the percentage of riprap shore-
line in the estuary (2014). Therefore, research targeting the
effects of shore-protection structures on these organisms is
needed before more definitive conclusions can be drawn. In
general, additional studies examining the ecological effects
of riprap revetments and breakwaters are needed to inform
future decisions on the consequences of selecting these types
of structures.
A majority of studies occurred over a period of 1 year or
less and did not replicate their measurements or sampling
by guest on September 6, 2016 from
Overview Articles September 2016 / Vol. 66 No. 9 BioScience 771
through time. Observable changes to coastal habitats as a
result of shoreline hardening may only be detectable with
long-term measurements of multiple characteristics (e.g.,
Moody et al. 2013) or event-specific monitoring (e.g.,
storms, Gittman etal. 2014). Studies that track the effects
of different shore-protection structures on habitat-forming
organisms, such as marsh plants, seagrasses, mangroves,
and shellfish reefs, over multiple years to decades would
provide valuable insights on the stability and resilience
of these shoreline habitats and supported ecosystem ser-
vices. Using spatially and temporally replicated BACI or
beyond-BACI designs (Underwood 1994) may be par-
ticularly important for studies of habitat-forming species
if changes are a result of direct replacement of habitat with
a hard structure (e.g., Bozek and Burdick 2005) or if there
is high spatiotemporal variability in the physical environ-
ment (e.g., Bilkovic and Mitchell 2013). Finally, studies on
the effects of shore-protection structures on the broader
suite of ecosystem functions and services (e.g., nutrient
cycling, pollutant filtration, carbon sequestration, and sed-
iment stabilization) would allow coastal managers to bet-
ter compare the overall functionality of shore-protection
Implications for coastal conservation and management. Shoreline
protection will almost certainly continue to be a prior-
ity as coastal hazards, such as storms and sea level rise,
continute to threaten growing coastal populations and
infrastructure. We found that not all shore-protection
structures perform equally regarding their ecological
impacts on coastal ecosystems. Seawalls have clear nega-
tive consequences for coastal biodiversity and habitat
quality, and these ecological impacts should be considered
by coastal managers and decisionmakers when developing
coastal shoreline policies and permitting shoreline protec-
tion structures. In addition, a growing body of literature
suggests that natural alternatives, such as living or nature-
based shore protection or biogenic habitat restoration, can
reduce erosion while also enhancing other ecosystem ser-
vices (e.g., Meyer at al. 1997, Benayas etal. 2009, Scyphers
etal. 2011, Gittman etal. 2014). Policymakers and coastal
managers should consider the ecological effects of engi-
neered-shore structures when deciding how to best fulfill
the need to protect people, property, and infrastructure
while also conserving and sustaining coastal ecosystem
biodiversity and function.
We thank the Handling Editor and three anonymous review-
ers for the thoughtful comments that greatly improved
this manuscript. This research was funded by a contract
to R. Gittman from the Pew Charitable Trusts (Award
Number12627) and supported by Northeastern University.
S. Scyphers was supported by a National Science Foundation
SEES Fellowship (OCE-1215825). C. Smith and I. Neylan
were supported by a North Carolina Coastal Recreational
Fishing License grant to C. Smith and C. Peterson and the
University of North Carolina at Chapel Hill.
Supplemental material
The supplemental material is available online at http://
References cited
Able KW, Vivian DN, Petruzzelli G, Hagan SM. 2012 Connectivity among
salt marsh subhabitats: Residency and movements of the mummichog
(Fundulus heteroclitus). Estuaries and Coasts 35: 743–753.
Adger WN, Hughes TP, Folke C, Carpenter SR, Rockström J. 2005. Social–
ecological resilience to coastal disasters. Science 309: 1036–1039.
Arkema KK, et al. 2015. Embedding ecosystem services in coastal plan-
ning leads to better outcomes for people and nature. Proceedings of the
National Academy of Sciences 112: 7390–7395.
Baggett LP, et al. 2015. Guidelines for evaluating performance of oyster
habitat restoration. Restoration Ecology 23: 737–745.
Beck MW, etal. 2011. Oyster reefs at risk and recommendations for conser-
vation, restoration, and management. BioScience 61: 107–116.
Benayas JMR, Newton AC, Diaz A, Bullock JM. 2009. Enhancement of
biodiversity and ecosystem services by ecological restoration: A meta-
analysis. Science 325: 1121–1124.
Bilkovic DM, Mitchell MM. 2013. Ecological tradeoffs of stabilized salt
marshes as a shoreline protection strategy: Effects of artificial structures
on macrobenthic assemblages. Ecological Engineering 61: 469–481.
Bilkovic DM, Roggero MM. 2008. Effects of coastal development on nearshore
estuarine nekton communities. Marine Ecology Progress Series 358: 27–39.
Boesch DF, Turner RE. 1984. Dependence of fishery species on salt marshes:
The role of food and refuge. Estuaries 7: 460–468.
Bozek CM, Burdick DM. 2005. Impacts of seawalls on saltmarsh plant
communities in the Great Bay Estuary, New Hampshire USA. Wetlands
Ecology and Management 13: 553–568.
Broome SW, Seneca ED, Woodhouse WW Jr. 1988. Tidal salt marsh restora-
tion. Aquatic Botany 32: 1–22.
Bulleri F. 2005. Experimental evaluation of early patterns of colonisation
of space on rocky shores and seawalls. Marine Environmental Research
60: 355–374.
Bulleri F, Chapman MG. 2004. Intertidal assemblages on artificial and natu-
ral habitats in marinas on the north-west coast of Italy. Marine Biology
145: 381–391.
Bulleri F, Chapman MG, Underwood AJ. 2004. Patterns of movement of
the limpet Cellana tramoserica on rocky shores and retaining seawalls.
Marine Ecology Progress Series 281: 121–129.
——. 2005. Intertidal assemblages on seawalls and vertical rocky shores in
Sydney Harbour, Australia. Austral Ecology 30: 655–667.
Burt J, Bartholomew A, Usseglio P, Bauman A, Sale PF. 2009. Are artificial
reefs surrogates of natural habitats for corals and fish in Dubai, United
Arab Emirates? Coral Reefs 28: 663–675.
Chapman MG, Bulleri F. 2003. Intertidal seawalls: New features of landscape
in intertidal environments. Landscape and Urban Planning 62: 159–172.
Chapman MG, Underwood AJ. 2011. Evaluation of ecological engineering
of “armoured” shorelines to improve their value as habitat. Journal of
Experimental Marine Biology and Ecology 400: 302–313.
Clynick BG, Chapman MG, Underwood AJ. 2007. Effects of epibiota on
assemblages of fish associated with urban structures. Marine Ecology
Progress Series 332: 201–210.
Currin CA, Delano PC, Valdes-Weaver LM. 2007. Utilization of a citizen
monitoring protocol to assess the structure and function of natural and
stabilized fringing salt marshes in North Carolina. Wetlands Ecology
and Management 16: 97–118.
Drexler M, Parker ML, Geiger SP, Arnold WS. 2014. Biological assessment
of eastern oysters (Crassostrea virginica) inhabiting reef, mangrove,
seawall, and restoration substrates. Estuaries and Coasts 37: 962–972.
by guest on September 6, 2016 from
Overview Articles
772 BioScience September 2016 / Vol. 66 No. 9
Dugan JE, Hubbard DM. 2006. Ecological responses to coastal armoring on
exposed sandy beaches. Shore and Beach 74: 10–16.
Dugan JE, Hubbard DM, Rodil IF, Revell DL, Schroeter S. 2008. Ecological
effects of coastal armoring on sandy beaches. Marine Ecology 29: 160–170.
Dugan JE, Airoldi L, Chapman MG, Walker SJ, Schlacher T. 2011. Pages
17–41 in Wolanski E, McLusky D, eds. Estuarine and Coastal Structures:
Environmental Effects, a Focus on Shore and Nearshore Structures.
Treatise on Estuarine and Coastal Science. Academic Press.
Egger M, Smith GD, Phillips AN. 1997. Meta-analysis: Principles and pro-
cedures. British Medical Journal 315: 1533–1537.
França S, Vinagre C, Caçador I, Cabral HN. 2005. Heavy metal concentra-
tions in sediment, benthic invertebrates and fish in three salt marsh
areas subjected to different pollution loads in the Tagus Estuary
(Portugal). Marine Pollution Bulletin 50: 998–1003.
Gillman LN, Wright SD. 2010. Mega mistakes in meta-analyses: Devil in the
detail. Ecology 91: 2550–2552.
Gittman RK, Fodrie FJ, Popowich AM, Keller DA, Bruno JF, Currin CA,
Peterson CH, Piehler MF. 2015. Engineering away our natural defenses:
An analysis of shoreline hardening in the US. Frontiers in Ecology and
the Environment 13: 301–307.
Gittman RK, Peterson CH, Currin CA, Fodrie FJ, Piehler MF, Bruno JF.
2016. Living shorelines can enhance the nursery role of threatened
estuarine habitats. Ecological Applications 26: 249–263.
Gittman RK, Popowich AM, Bruno JF, Peterson CH. 2014. Marshes with
and without sills protect estuarine shorelines from erosion better
than bulkheads during a category 1 hurricane. Ocean and Coastal
Management 102: 94–102.
Glasby TM, Connell SD, Holloway MG, Hewitt CL. 2007. Nonindigenous
biota on artificial structures: Could habitat creation facilitate biological
invasions? Marine Biology 151: 887–895.
Harris C, Strayer DL, Findlay S. 2014. The ecology of freshwater wrack
along natural and engineered Hudson River shorelines. Hydrobiologia
722: 233–245.
Hedges LV, Gurevitch J, Curtis PS. 1999. The meta-analysis of response
ratios in experimental ecology. Ecology 80: 1150–1156.
Hendon JR, Peterson MS, Comyns BH. 2000. Spatio–temporal distribu-
tion of larval Gobiosoma bosc in waters adjacent to natural and altered
marsh-edge habitats of Mississippi coastal waters. Bulletin of Marine
Science 66: 143–156.
Kneib RT. 1982. Habitat preference, predation, and the intertidal distri-
bution of gammaridean amphipods in a North Carolina salt marsh.
Journal of Experimental Marine Biology and Ecology 59: 219–230.
Lam NWY, Huang R, Chan BKK. 2009. Variations in Intertidal assem-
blages and zonation patterns between vertical artificial seawalls and
natural rocky shores: A case study from Victoria Harbour, Hong Kong.
Zoological Studies 48: 184–195.
Lipcius RN, Seitz RD, Seebo MS, Colón-Carrión D. 2005. Density, abun-
dance, and survival of the blue crab in seagrass and unstructured salt
marsh nurseries of Chesapeake Bay. Journal of Experimental Marine
Biology and Ecology 319: 69–80.
[MEA] Millennium Ecosystem Assessment. 2005. Ecosystems and Human
Well-Being. Island Press.
Meyer DL, Townsend EC, Thayer GW. 1997. Stabilization and erosion control
value of oyster cultch for intertidal marsh. Restoration Ecology 5: 93–99.
Møller AP, Jennions MD. 2001. Testing and adjusting for publication bias.
Trends in Ecology and Evolution 16: 580–586.
Möller I, etal. 2014.Wave attenuation over coastal salt marshes under storm
surge conditions. Nature Geoscience 7: 727–731.
Moody RM, Cebrian J, Heck KL Jr. 2013. Interannual recruitment dynamics
for resident and transient marsh species: Evidence for a lack of impact
by the Macondo oil spill. PLOS ONE 8 (art. e58376).
[NRC] National Research Council. 2007. Mitigating Shore Erosion along
Sheltered Coasts. National Academies Press.
O’Connor MI, Violin CR, Anton A, Ladwig LM, Piehler MF. 2011. Salt
marsh stabilization affects algal primary producers at the marsh edge.
Wetlands Ecology and Management 19: 131–140.
O’Meara T, Thompson SP, Piehler MF. 2015. Effects of shoreline hardening
on nitrogen processing in estuarine marshes of the US mid-Atlantic
coast. Wetlands Ecology and Management 23: 385–394.
Patrick CJ, Weller DE, Li X, Ryder M. 2014. Effects of shoreline alteration
and other stressors on submerged aquatic vegetation in subestuaries of
Chesapeake Bay and the mid-Atlantic coastal bays. Estuaries and Coasts
37: 1516–1531.
Peregrine DH. 2003. Water-wave impact on walls. Annual Review of Fluid
Mechanics 35: 23–43.
Perkins MJ, Ng TPT, Dudgeon D, Bonebrake TC, Leung KMY. 2015.
Conserving intertidal habitats: What is the potential of ecological engi-
neering to mitigate impacts of coastal structures? Estuarine, Coastal,
and Shelf Science 167B: 1–12.
Peters JR, Yeager LA, Layman CA. 2015. Comparison of fish assemblages
in restored and natural mangrove habitats along an urban shoreline.
Bulletin of Marine Science 91: 125–139.
Peterson CH, Lipcius RN. 2003. Conceptual progress towards predict-
ing quantitative ecosystem benefits of ecological restorations. Marine
Ecology Progress Series 264: 297–307.
Peterson GW, Turner RE. 1994. The value of salt marsh edge vs interior as
a habitat for fish and decapod crustaceans in a Louisiana tidal marsh.
Estuaries 17: 235–262.
Pontee N. 2013. Defining coastal squeeze: A discussion. Ocean and Coastal
Management 84: 204–207.
Reboreda R, Cacador I. 2007. Halophyte vegetation influences in salt marsh
retention capacity for heavy metals. Environmental Pollution 146:
Rosenthal R. 1979. The file drawer problem and tolerance for null results.
Psychological Bulletin 86: 638–641.
——. 1991. Meta-Analytic Procedures for Social Research. Sage.
Ruggiero P, Komar PD, McDougal WG, Marra JJ. 2001. Wave runup,
extreme water levels and the erosion of properties backing beaches.
Journal of Coastal Research 17: 407–419.
Ruiz GM, Carlton JT, Grosholz ED, Hines AH. 1997. Global invasions of
marine and estuarine habitats by non-indigenous species: Mechanisms,
extent, and consequences. Integrative and Comparative Biology 37:
Scyphers SB, Powers SP, Heck KL Jr., Byron D. 2011. Oyster reefs as natural
breakwaters mitigate shoreline loss and facilitate fisheries. PLOS ONE
6 (art. e22396).
Scyphers SB, Gouhier TC, Grabowski JH, Beck MW, Mareska J, Powers SP.
2015. Natural shorelines promote the stability of fish communities in an
urbanized coastal system. PLOS ONE 10 (art. e0118580).
Seitz R, Lipcius R, Olmstead N, Seebo M, Lambert D. 2006. Influence of
shallow-water habitats and shoreline development on abundance, bio-
mass, and diversity of benthic prey and predators in Chesapeake Bay.
Marine Ecology Progress Series 326: 11–27.
Shepard CC, Crain CM, Beck MW. 2011. The protective role of coastal
marshes: A systematic review and meta-analysis. PLOS ONE 6 (art.
Small C, Nicholls R. 2003. A global analysis of human settlement in coastal
zones. Journal of Coastal Research 19: 584–599.
Sobocinski KL, Cordell JR, Simenstad CA. 2010. Effects of shoreline
modifications on supratidal macroinvertebrate fauna on Puget Sound,
Washington, beaches. Estuaries and Coasts 33: 699–711.
Steffen W, Crutzen J, McNeill JR. 2007. The Anthropocene: Are
humans now overwhelming the great forces of Nature? Ambio 36:
Strayer DL, Findlay SEG, Miller D, Malcom HM, Fischer DT, Coote T. 2012.
Biodiversity in Hudson River shore zones: Influence of shoreline type
and physical structure. Aquatic Sciences: Research Across Boundaries
74: 597–610.
Titus J. 1998. Rising seas, coastal erosion, and the takings clause: How to
save wetlands and beaches without hurting property owners. Maryland
Law Review 57: 1279–1318.
Tummers B. 2006. DataThief III. (6 July 2016;
by guest on September 6, 2016 from
Overview Articles September 2016 / Vol. 66 No. 9 BioScience 773
Underwood AJ. 1994. On beyond BACI: Sampling designs that might
reliably detect environmental disturbances. Ecological Applications 4:
Unites States Army Corps of Engineers (USACE). 2001. Low Cost Shore
Protection. USACE. (6 July 2016;
Vetter D, Rücker G, Storch I. 2013. Meta-analysis: A need for well-
defined usage in ecology and conservation biology. Ecosphere 4:
Viechtbauer W. 2010. Conducting meta-analyses in R with the meta-
for package. Journal of Statistical Software 36: 1–48. (6 July 2016;
Rachel K. Gittman (, Steven B. Scyphers, and Jonathan
H. Grabowski are affiliated with the Marine Science Center at Northeastern
University, in Nahant, Massachusetts. Carter S. Smith and Isabelle P. Neylan
are affiliated with the Institute of Marine Sciences at the University of North
Carolina at Chapel Hill, in Morehead City.
by guest on September 6, 2016 from
... As sea waters begin to rise due to climate change, rock walls have been built along the coastal shoreline to prevent erosion, creating an unprecedented challenge for plants to reach land and germinate. 13 In this study, two of the three axes of rarity de ned previously were examined: habitat speci city and population size. To understand the habitat speci city of the two trees, the reproductive cycles of espesia populnea and Hibiscus tiliaceus were compared ; speci cally, relationships with potential oral visitors, dispersal abilities (fruit or seed drop rates, oating times, and abilities to move across seawalls), and establishment success (germination rates). ...
... (3) Thespesia populnea has a lower germination rate than Hibiscus tiliaceus. (4) e presence of man-made rock walls 13 impacts the distribution of espesia populnea more than the distribution of Hibiscus tiliaceus. 1). ...
... The response of the natural marine communities to anthropogenic disturbances has become a priority research subject in recent years. These perturbations, identified as any human activity that may cause stress to natural populations and assemblages , are rapidly and dramatically increasing (Gittman et al., 2016). One of the best examples can be found in the modification of coastlines (Bulleri & Chapman, 2010). ...
... This alteration of the environment is due to shoreline urbanisation combined with the increase in tourism, recreational and commercial activities among others (Bulleri & Chapman, 2010;Melero et al., 2017;Mosbahi et al., 2019). This results in the introduction and proliferation of artificial structures and in the replacement of natural substrates in marine coastal habitats worldwide (Sedano et al., 2019), the so-called shoreline hardening (Gittman et al., 2016). For instance, in various areas of Italy, France and Spain, shoreline hardening exceeds 45% of terrestrial cover (Bulleri & Chapman, 2010). ...
Full-text available
The aim of this investigation was to assess the differences in terms of α-diversity, species composition and trophic structure between peracarid assemblages inhabiting artificial hard substrates and those present on natural ones at two sites in the Western Mediterranean Sea. In each one, samples were collected from natural and artificial substrates, and the extracted peracarids were identified to the species level. Each sample was characterised by its Total Abundance, Species Richness, and 1-D Simpson, Shannon, Margalef and Equitability indices. A cluster plot based on the Bray-Curtis coefficient was conducted to detect the presence of distinct assemblages. In addition, the proportion of individuals belonging to each feeding guild was studied. The ecological indices revealed that assemblages from the artificial substrates at the Villaricos site showed lower values of diversity, while they were more diverse at the Calpe site. These results indicated that artificial hard substrates were not always inadequate ground for these animals. The species composition differed between artificial and natural hard substrates in the two studied sites and the trophic structure was also modified, although, in this case, the changes were similar in both sites. However, the difference in both respects seems to be higher between sites than between substrates, evidencing the importance of local scale processes.
... Aside from the uncertainty over the effectiveness of hard shorelines against erosion, environmental effects have not been well studied. From the studies that have been published, it appears as though hardened shorelines have a significantly adverse effect on the benthic organismal community and greatly reduce the amount of dry beach available for recreation (Gittman et al., 2016;Pilkey & Wright, 1988). This information, coupled with an increased awareness of the services of natural coastal habitats, has led to managers and property owners looking for alternative methods. ...
Technical Report
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North Carolina’s coast attracts millions of visitors from around the world every year. Home to a wide diversity of species, rich marine habitats, and scenic views, North Carolina’s coast is a hotspot for vacationers and marine scientists alike. Coastal recreation and tourism contributed an estimated $1.1 billion to the state economy in 2013, with over $225 million from the Outer Banks tourism alone. Fisheries and seafood industries also contributed a sizable $188 million to North Carolina’s economy, excluding any profits made from recreational fishing(Hallac, 2016; Harrison et al., 2017). With so much of the state relying on the ocean economy, North Carolina’s coast is especially important to protect. Coastal shorelines have long been stabilized with hardened structures, such as seawalls and bulkheads, in order to provide protection from coastal erosion and hazards. However, these engineered structures often decrease the ecological integrity of natural shorelines. In an attempt to mitigate human impacts on shorelines across the globe, coastal managers have recently emphasized “living shorelines,” a term used to describe structures that incorporate natural components, such as oysters and marsh, to provide both shoreline protection and valuable habitat. Other hybrid approaches include granite sills, which serve as a hard substrate for the recruitment of other living components, like oysters and fish. While living shorelines are becoming a popular concept, there is relatively little information that quantifies the ecological value and ecosystem services that they offer. Considering projected sea-level rise and increasing frequency of tropical storms, there is an urgent need for understanding how coastal communities can maximize their shoreline stabilization efforts while maintaining ecological integrity. In this capstone, the University of North Carolina at Chapel Hill Institute for the Environment (UNC-IE) Morehead City Field Site undergraduate students plan addressed the differences between four living shorelines. The four studied sites include the following: [1] the Maritime Museum discontinuous granite sill (DGS) in Beaufort, [2] the Duke University Marine Lab continuous granite sill (CGS) on the west side of Pivers Island, [3] the NOAA continuous oyster sill (COS) on the east side of Pivers Island, and [4] the UNC Institute of Marine Sciences(UNC-IMS) discontinuous oyster sill (DOS) in Morehead City (Figure 1).Wave attenuation was mapped and measured at each shoreline in order to contribute to our understanding of the physical differences between each living shoreline type. Nutrients, water quality, and biota around each site was also measured to better quantify the environmental impacts and quality of each shoreline in terms of habitat. The data collected on these factors will provide insight into the environmental implications that different types of living shorelines may have.
... As saltmarshes have declined, property owners have taken action to prevent coastal erosion by building hardened shorelines (Erdle et al., 2006;Swann, 2008;Gittman et al., 2016). Hardened shorelines include sea walls, jetties, revetments, breakwaters, and bulkheads. ...
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Living shorelines are a way of protecting the shoreline from erosion while supporting the productivity of coastal habitats to maintain their many desirable ecosystem services. We focused on evaluating hardened versus living shorelines for coastal protection and compared them to adjacent natural marsh shorelines suffering excess erosion rates. Living shorelines are a suite of shoreline conservation and restoration techniques, which usually involve a breakwater that dampens wave energy so vegetation can take root and stabilize the shoreline sediments. This synthesis looked at six different sites, all containing a natural, living, and hardened shoreline across two different energy groups (low and high) to see how hydrologic, geomorphic, and vegetative parameters affected shoreline processes. The erosion rate of the coastline and its geomorphic shape were significantly influenced in the two energy groups, with the high energy coastlines eroding more quickly and having a steeper slope. Hardened shorelines were found to have little to no erosion, while natural shorelines had the greatest amount of edge erosion over time. Living shorelines lessened this rate of erosion compared to the natural marsh. The natural and living shorelines were similar in slope and sediment parameters, while hardened shorelines had steep slopes and higher sand content. We found that coastlines with high turbidity, high erosion rates, large wave power, and relative exposure resulted in steeper slopes and a higher percentage of sand in the sediment, but reduced vegetation percent cover and percent of marsh dominant vegetation species. This research increases our knowledge about what environmental conditions may be most suitable for living shorelines, of importance as sea levels rise and increase coastal erosion rates.
... While numerous entities have aimed to create artificial reefs to promote life, few have integrated wave attenuation properties into the design for coastline protection. Similarly, not all coastline protection structures provide the complex habitats that the natural environment does (Gittman et al. 2016). ...
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An artificial spur-and-groove (SAG) reef formation designed to perform as a submerged breakwater was studied. Not only do natural SAG reefs effectively attenuate waves, but they also create important circulation within the system that encourages the success of corals and the overall ecosystem. Modular components were developed and imported into the CFD software FLOW-3D® HYDRO to test the wave-structure interaction while adjusting the geometry of the structure, wave conditions, and depth due to tidal changes in environmental conditions similar to Moloka’i, Hawaii. Results showed highest wave height reduction with larger wave heights as well as lower submergence during low tide conditions.
... Wind-blown sediment transport is a ubiquitous process on Earth and many other planetary surfaces such as Mars and Venus. In addition to land forming, it can also bring many aeolian hazards, such as soil nutrients removal (Wen and Zhen, 2020), infrastructure damage (Hassan and Guirguis, 2013;Mehdipour and Baniamerian, 2019), and human respiratory system impairment (Li and Sherman, 2015;Middleton, 2017), which will be even worse in the highly populated coastal areas, where onethird of the human population lives (Gittman et al., 2016;Currin, 2019). ...
Fences have been widely used in coastal protection engineering for their low cost, simple deployment, and easy integration with ecosystems. The effects of fence porosity and height on dune development have been investigated while not much attention has been paid to the effects of fence opening configurations, such as opening size and geometry, and porosity distributions. In this study, we deployed eight fences with same height and similar porosity, but different opening configurations on a sandy beach in Pingtan, Fujian Province. Results indicate that there is a similar two-dune-one-trough pattern for all fences at the beginning of dune development, and opening size, orientation, and geometry, and porosity distribution control the leeward dune peak locations. Fences with small openings and non-uniform porosity have high trapping efficiency, and upper denser porosity may be the optimal design as these fences have the highest trapping efficiency and capacity. The conclusions from this study can provide guidance on practical fence design.
... The land cover of local and regional catchments feeding into these estuaries also have experienced significant anthropogenic impacts, such as agriculture conversion and urban development (Greene et al., 2015a). Other stressors include habitat destruction, shoreline armoring, invasive species, and climate change impacts such as sea level rise, higher water temperature and ocean acidification (Gittman et al., 2016;Gleason et al., 2011). ...
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Estuaries provide essential ecosystem functions, including serving as nurseries for fish and aquatic invertebrates (i.e., nekton). Understanding how anthropogenic stressors affect the habitat quality of estuaries is a crucial component of fisheries conservation. A meaningful composite indicator for estuary stress on the nekton community at regional scales should encapsulate major anthropogenic effects over an assemblage of important species. This study introduces a composite stressor index, using meta-analysis to synthesize trawl and seine-based sampling data on 39 nekton species from 25 monitoring programs covering 38 Pacific Coast estuaries spanning four ecoregions. Candidate stressor variables include pollution sources, urban and agricultural land use, and stream impairments, considered at different spatial scales. For each of the 39 focal species, we screened these stressors using hierarchical statistical models that account for environmental variability (e.g., salinity) across sampling events and systematic biases due to monitoring program protocols and species-specific biogeographic ranges. Based on this screening, we identified a subset of stressors most strongly related to nekton presence/ absence. Then we used principal component analysis (PCA) to identify major thematic axes of variation across these stressors and the 38 estuaries. Hierarchical nekton models fit to these axes (i.e., thematic stressors) indicate that urban development is negatively related to most nekton species, while agriculture is often positively related to species presence, though with smaller estimated effects. These models were also used to create the estuary-level composite index of anthropogenic stress. Urban development is the major contributor to the index scores of most estuaries, except in the Oregon-Washington Coast ecoregion, where agriculture is dominant. The composite index indicates that estuaries in the Southern California Bight ecoregion are the most impacted, along with South San Francisco Bay.
... In the Salish Sea, WA, USA, 29% of shorelines are armored (MacLennan et al., 2017). Armor placement can have negative impacts on biodiversity and organism abundance both regionally (Dethier et al., 2016) and worldwide (Gittman et al., 2016b), and as a result, efforts to remove armor and restore natural processes are increasing. Studies in the Salish Sea have focused on measuring restoration effectiveness in upper intertidal-supratidal areas where impacts of armor (Dethier et al., 2016) and therefore restoration actions (Toft et al., 2021) have been most pronounced on beach characteristics such as wrack, logs, sediments, vegetation, and invertebrates. ...
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Coastal shorelines are a key interface between terrestrial and aquatic ecosystems and are vital for human livelihood. As a result, shorelines have experienced substantial human modifications worldwide. Shoreline “hardening” – the construction of armor including seawalls, bulkheads, or rip-rap – is a common modification that has substantial negative ecological effects. Currently, restoration involving the removal of armor and replacement with “living” shorelines is becoming an established practice. Still, the ecological response to armor removal is oftentimes unpredictable and site-specific. We hypothesized that the confluence of larger-scale geophysical features might strongly influence ecological restoration outcomes at particular locations. To measure the effectiveness of armor removal in the context of broad-scale geophysical features across the Salish Sea, WA, USA, we studied 26 paired restored and natural reference beaches of the same shoretype (feeder bluff, accretion shoreform, or pocket beach), as well as corresponding fetch, sub-basin, and percent of shoreline sediment drift cell armored. Sites were restored for an average of six years. We gauged restoration effectiveness based on levels of five ecological response variables: beach wrack (percent, depth), logs (count, width), sediments (percent sand), vegetation (percent overhanging, count of fallen trees), and insects (density, taxa richness). We found that armor removal often restored these variables to natural levels, but that restoration response was dependent on geophysical features such as shoretype and fetch. Natural beaches did have higher measurements of overhanging vegetation, fallen trees, and insect taxa richness, as these features likely take time to mature at restored beaches. Feeder bluffs had a higher proportion of surface sand and number of fallen trees than other shoretypes, coinciding with the erosion of bluff material, whereas natural pocket beaches within bordering rocky headlands had higher insect densities. Sites with a large fetch had higher input of deposited wrack and logs, whereas sites with a small fetch had higher input from localized terrestrial sources – fallen trees and eroding sand. By incorporating the effectiveness of restoration with landscape features such as shoretype and fetch, we can more effectively plan for future restoration actions and better predict their outcomes.
... Compiling these four intertidal surveys gives us insight into methodological and habitat effects on estimates of richness and composition at PCPP. Of the three habitats, cobble beach had the highest reported taxonomic richness and riprap the lowest (compared with Gittman et al. 2016). However, rarefaction and extrapolation estimates suggest that riprap richness might be similar to, and sediment richness might even exceed, that of cobble. ...
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The value of biodiversity and of documented biodiversity surveys is well established. Extracting historical biodiversity data and synthesizing them with current data can provide a more comprehensive estimate of total diversity and guide future monitoring. We demonstrate the utility of compiling historical and recent biodiversity data to better characterize taxon richness and composition. Our focus is an otherwise unmonitored habitat in an unmonitored British Columbia provincial park, in a heavily impacted region of the Salish Sea that was designated a United Nation Biosphere Reserve in 2021. We conducted surveys and compiled historical records that together spanned three intertidal habitats and 43 years. From these combined data we report a total of 99 taxa, an order of magnitude increase over the number listed in the park’s Master Plan. These include seven non-native species, of which four are newly reported here. Rarefaction, extrapolation, and multivariate dissimilarity analyses revealed the roles of methods and habitat types in contributing to differences in taxon richness and composition among surveys. This data compilation illustrates many of the challenges and opportunities in aligning and assembling independent space-time snapshots of alpha (i.e., local) diversity to better understand the gamma (i.e., regional) diversity of a marine protected area and provides the foundational data needed to design effective future monitoring at molecular to ecosystem scales.
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Marine phytoplankton may adapt to ocean change, such as acidification or warming, because of their large population sizes and short generation times. Long-term adaptation to novel environments is a dynamic process, and phenotypic change can take place thousands of generations after exposure to novel conditions. We conducted a long-term evolution experiment (4 years = 2100 generations), starting with a single clone of the abundant and widespread coccolithophore Emiliania huxleyi exposed to three different CO2 levels simulating ocean acidification (OA). Growth rates as a proxy for Darwinian fitness increased only moderately under both levels of OA [+3.4% and +4.8%, respectively, at 1100 and 2200 μatm partial pressure of CO2 (Pco2)] relative to control treatments (ambient CO2, 400 μatm). Long-term adaptation to OA was complex, and initial phenotypic responses of ecologically important traits were later reverted. The biogeochemically important trait of calcification, in particular, that had initially been restored within the first year of evolution was later reduced to levels lower than the performance of nonadapted populations under OA. Calcification was not constitutively lost but returned to control treatment levels when high CO2–adapted isolates were transferred back to present-day control CO2 conditions. Selection under elevated CO2 exacerbated a general decrease of cell sizes under long-term laboratory evolution. Our results show that phytoplankton may evolve complex phenotypic plasticity that can affect biogeochemically important traits, such as calcification. Adaptive evolution may play out over longer time scales (>1 year) in an unforeseen way under future ocean conditions that cannot be predicted from initial adaptation responses.
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Restoration of degraded ecosystems is an important societal goal, yet inadequate monitoring and the absence of clear performance metrics are common criticisms of many habitat restoration projects. Funding limitations can prevent adequate monitoring, but we suggest that the lack of accepted metrics to address the diversity of restoration objectives also presents a serious challenge to the monitoring of restoration projects. A working group with experience in designing and monitoring oyster reef projects was used to develop standardized monitoring metrics, units, and performance criteria that would allow for comparison among restoration sites and projects of various construction types. A set of four universal metrics (reef areal dimensions, reef height, oyster density, and oyster size–frequency distribution) and a set of three universal environmental variables (water temperature, salinity, and dissolved oxygen) are recommended to be monitored for all oyster habitat restoration projects regardless of their goal(s). In addition, restoration goal-based metrics specific to four commonly cited ecosystem service-based restoration goals are recommended, along with an optional set of seven supplemental ancillary metrics that could provide information useful to the interpretation of prerestoration and postrestoration monitoring data. Widespread adoption of a common set of metrics with standardized techniques and units to assess well-defined goals not only allows practitioners to gauge the performance of their own projects but also allows for comparison among projects, which is both essential to the advancement of the field of oyster restoration and can provide new knowledge about the structure and ecological function of oyster reef ecosystems.
Coastal ecosystems provide numerous services, such as nutrient cycling, climate change amelioration, and habitat provision for commercially valuable organisms. Ecosystem functions and processes are modified by human activities locally and globally, with degradation of coastal ecosystems by development and climate change occurring at unprecedented rates. The demand for coastal defense strategies against storms and sea-level rise has increased with human population growth and development along coastlines worldwide, even while that population growth has reduced natural buffering of shorelines. Shoreline hardening, a common coastal defense strategy that includes the use of seawalls and bulkheads (vertical walls constructed of concrete, wood, vinyl, or steel), is resulting in a “coastal squeeze” on estuarine habitats. In contrast to hardening, living shorelines, which range from vegetation plantings to a combination of hard structures and plantings, can be deployed to restore or enhance multiple ecosystem services normally delivered by naturally vegetated shores. Although hundreds of living shoreline projects have been implemented in the United States alone, few studies have evaluated their effectiveness in sustaining or enhancing ecosystem services relative to naturally vegetated shorelines and hardened shorelines. We quantified the effectiveness of (1) sills with landward marsh (a type of living shoreline that combines marsh plantings with an offshore low-profile breakwater), (2) natural salt marsh shorelines (control marshes), and (3) unvegetated bulkheaded shores in providing habitat for fish and crustaceans (nekton). Sills supported higher abundances and species diversity of fishes than unvegetated habitat adjacent to bulkheads, and even control marshes. Sills also supported higher cover of filter-feeding bivalves (a food resource and refuge habitat for nekton) than bulkheads or control marshes. These ecosystem-service enhancements were detected on shores with sills three or more years after construction, but not before. Sills provide added structure and may provide better refuges from predation and greater opportunity to use available food resources for nekton than unvegetated bulkheaded shores or control marshes. Our study shows that unlike shoreline hardening, living shorelines can enhance some ecosystem services provided by marshes, such as provision of nursery habitat.
Multiple stressors affect estuarine shorelines including erosion, sea level rise and impacts from human development of adjacent lands. Increasingly common features of coastal development are vertical shoreline stabilization structures such as bulkheads. Bulkheads are designed to prevent land loss and flooding through the construction of a vertical wall anchored to the land. However, they break the connection between land and water and are barriers to upland plant migration. This disconnect can affect hydrology, alter nutrient and sediment supplies, and lead to marsh loss. We measured the effects of bulkheads on sediment nitrogen fluxes, including denitrification (DEN), at three representative estuarine shoreline types: natural marsh (no bulkhead), bulkhead without marsh, and bulkheads with marshes of varying widths. Sediment cores were taken mid-marsh or, 2 m seaward of bulkhead in sites lacking marsh in northern, central and southern coastal regions of North Carolina. Concentrations of N2 and O2 were measured using a membrane inlet mass spectrometer. In addition, sediment organic matter and inorganic nitrogen concentrations were quantified. Average DEN rate was 93.1 ± 7.0 µmol N m−2 h−1 with the highest rates in the summer and lowest rates in the winter. Sediment oxygen demand was positively correlated with DEN rate (R2 = 0.43, p < 0.01), which suggests that DEN is affected by carbon lability. DEN was not affected by bulkhead presence (R2 = 0.01, p = 0.52), but marsh presence significantly affected yearly DEN rates (R2 = 0.13, p < 0.01). These data indicate that bulkheads do not directly affect nitrogen processing, but indirectly reduce cycling rates through marsh loss.