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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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http://bioscience.oxfordjournals.org September 2016 / Vol. 66 No. 9 BioScience 763
Ecological Consequences
of Shoreline Hardening:
A Meta-Analysis
RACHEL K. GITTMAN, STEVEN B. SCYPHERS, CARTER S. SMITH, ISABELLE P. NEYLAN,
AND JONATHAN H. GRABOWSKI
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.
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 (http://creativecommons.org/licenses/
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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764 BioScience September 2016 / Vol. 66 No. 9 http://bioscience.oxfordjournals.org
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.
Methods
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.,
epibiota).
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
2010).
Results
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
Soft
Sediment Biogenic Flora
Benthic
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,
–1.50
–1.00
–0.50
0.00
0.50
1.00
1.50
Log Response Ratio
–1.50
–1.00
–0.50
0.00
0.50
1.00
1.50
Log Response Ratio
Seawall Riprap Breakwater
b
11 (20) 8 (14) 5 (11)
22 (67) 7 (22)
8 (36)
a
Biodiversity
Abundance
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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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,
respectively).
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,
respectively).
Conclusions
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
–3
–2
–1
0
1
2
3
–3
–2
–1
0
1
2
3
Log Response Ratio
–3
–2
–1
0
1
2
3
Flora Ben. Infauna Birds Epibiota Nekton
a
b
c
NA
NA NA NA
Seawalls
Riprap
Breakwaters
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
–3
–2
–1
0
1
2
3
–3
–2
–1
0
1
2
3
Log Response Ratio
–3
–2
–1
0
1
2
3
Flora Ben. Infauna Birds Epibiota Nekton
a
b
c
NA NA
NA
Seawalls
Riprap
Breakwaters
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, 2016http://bioscience.oxfordjournals.org/Downloaded from
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770 BioScience September 2016 / Vol. 66 No. 9 http://bioscience.oxfordjournals.org
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
communities.
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
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http://bioscience.oxfordjournals.org 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
approaches.
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.
Acknowledgments
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://
bioscience.oxfordjournals.org/lookup/suppl/doi:10.1093/biosci/
biw091/-/DC1.
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Rachel K. Gittman (r.gittman@neu.edu), Steven B. Scyphers, and Jonathan
H. Grabowski are affiliated with the Marine Science Center at Northeastern
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... Studies have shown that armor reduces the mean abundance of key biological components of shoreline ecosystemsessentially driving down biomass on modified beaches (Dethier et al., 2016;Gittman et al., 2016). Decreases in the coverage and depth of beach wrack, the number of beached logs, and the density of supratidal invertebrates are just some of the effects of shoreline armoring and maintenance regimes (Dethier et al., 2017(Dethier et al., , 2016Heerhartz et al., 2014;Lee et al., 2018;Schooler et al., 2019;Toft et al., 2021). ...
... Researchers and practitioners are beginning to document and acknowledge the negative impacts of shoreline modification on ecological variability (Aguilera et al., 2014;Lawrence et al., 2021) and work in other systems has shown that anthropogenically modified ecosystems can be less variable across time and space (Buyantuyev and Wu, 2009;Gittman et al., 2016;Groffman et al., 2014;McKinney, 2006;Wu et al., 2011). For example, primary productivity is less variable through time in urbanized areas of Phoenix, AZ compared to areas with natural land cover (Buyantuyev and Wu, 2009). ...
... For example, primary productivity is less variable through time in urbanized areas of Phoenix, AZ compared to areas with natural land cover (Buyantuyev and Wu, 2009). Numerous studies have cited the spatial homogenization of biological communities across urban and modified landscapes (Groffman et al., 2014;McKinney, 2006), including shorelines (Gittman et al., 2016). Spatial variation in beach wrack, logs, and invertebrate communities is common across natural coastlines and can be driven by interacting differences in shore slope, shore type, aspect, landcover, latitude, and sediment type, among other characteristics (Dethier et al., 2016;Heerhartz et al., 2014;Reimer et al., 2018;Romanuk and Levings, 2003). ...
Article
Humans have drastically modified marine nearshore ecosystems through shoreline armoring. Armor, in the form of seawalls and bulkheads, reduces the mean abundance of key ecological features of shoreline ecosystems, such as the amount of beach wrack, the number of beached logs, and the density of supratidal invertebrates. Armor also affects the physical and biological composition and diversity of these important ecological responses – altering the makeup of beach wrack and invertebrate species, for example. Less is known, however, about changes in variability – both over time and space – of ecological responses across natural, restored, and armored shores. Temporal and spatial variation in physical and biological variables can themselves be indicators of ecosystem health and effectiveness of restoration. Working alongside community (citizen) scientists, we found that beach wrack (a nutrient and habitat resource), logs (an element of habitat structure), and supratidal invertebrates (part of the consumer community) often increased following restoration. Further, not only were wrack, logs, and invertebrates on average more abundant and diverse at natural (never armored) shore types compared to armored shore types, but they also frequently had higher variance. In many cases, variance of ecological responses in restored shore types were more similar to natural shore types than armored shore types, indicating a positive effect of restoration. We found that differences among sample sites, rather than across sample years, explained more of the variation in ecological responses across all shore types. Because shoreline armoring is a pervasive human activity, public perception of this variability is key to the social context of restoration success. Participation in data collection through community science endeavors is one way to encourage an appreciation for natural variability within and across landscapes. We implore that shoreline monitoring efforts should evaluate and communicate ecosystem variability as a key indicator of restoration success.
... Roughly 14% of the United States shoreline is hardened, and a growing body of work addresses the consequences of implementing human-engineered structures on near natural shoreline (Gittman et al., 2016). Though studies show overwhelming negative impacts on ecosystems surrounding hardened shorelines, particularly in the case of seawall construction, research has focused on ecological changes more acutely than on physical geographic changes (Gittman et al., 2016). ...
... Roughly 14% of the United States shoreline is hardened, and a growing body of work addresses the consequences of implementing human-engineered structures on near natural shoreline (Gittman et al., 2016). Though studies show overwhelming negative impacts on ecosystems surrounding hardened shorelines, particularly in the case of seawall construction, research has focused on ecological changes more acutely than on physical geographic changes (Gittman et al., 2016). Most research concerning coastal sediment transport and the impacts of shoreline modification deal with ocean systems. ...
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From 2013 to 2020, water levels in Lake Michigan rose from an all-time low to the highest levels observed in nearly four decades, causing shoreline changes throughout the Great Lakes. These changes are particularly noticeable at North Avenue Beach in Chicago, an artificial beach where federal, state, and city agencies have mitigated rising waters by importing sand and constructing groins. Using spring season aerial imagery from 2012 to 2020, we calculated beach area for each year and compared sand cover loss between years to water level change per the USGS station located just south of the study area. Analysis reveals year-to-year loss in sand cover since 2013, with the largest single-year change occurring between 2018 and 2019. An inverse relationship with a slight lag exists between water level and these beach area changes. We calculated sand-groin distances from 2000 to 2020 to identify north–south effects. Of the six beach cells separated by groins, the northernmost two cells failed over the study period, and experienced the largest individual sand-groin distance losses. We modeled inundation to investigate whether the sand loss was explainable by lake level change alone, with particular attention given to hardened shoreline implemented north of the beach in 2015. Observed sand cover loss markedly exceeded predictions from the inundation modeling. In addition to water level changes, a local response to shoreline modification and obstruction of sediment transport at this site may influence sand cover.
... In particular, the coastal protection infrastructures are considerably growing and will continue to do this according to the increase of populations and related intensification of hazards (Scyphers et al., 2011;Hinkel et al., 2014). However, the downsides of common coastal armouring strategies (e.g., seawalls, revetments, groins) also need to be considered: habitat loss (Titus, 1998), lower floral and faunal biodiversity (Gittman et al., 2016), and depressed socio-economic resilience (Smith et al., 2020), especially due to their expensive maintenance cost. In this view, ecosystem-friendly alternatives to traditional coastal defence structures are emerging, such as natural and nature-based infrastructure (Sutton-Grier et al., 2018), nature-based solutions (Nesshöver et al., 2017), hybrid infrastructure (Sutton-Grier et al., 2015), ecosystem-based coastal defence (Temmerman et al., 2013), soft ecological engineering (Strain et al., 2019), and living shorelines. ...
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In the last decades, climate change and the rapid urbanization due to the development of the coastal economy have led to biodiversity loss and the fragmentation of habitat in many coastal zones. The presence of protected areas cannot prevent the progress of land degradation. However, these areas are very important because they provide significant ecosystem services and affect local tourism. With regard to increasing adaptation strategies to human pressures and climate change, the present study proposes a detailed monitoring activity and an ecological restoration plan which could improve the resilience of a protected coastal zone in the Pantano forest of Policoro, located on the Ionian coast (southern Italy). In this area, continuous phenomena of intensive deforestation, hydraulic reclamation actions, and fires have reduced the native species of particular naturalistic value, favouring the advancement of desertification, coastal erosion, and saltwater intrusion. The proposed actions are derived from a preliminary analysis on maps, UAV-images, climate data and from meetings with the local community. The operative process detailed in this article could be applied to other protected areas which are subjected to the same phenomena and problems.
... Artificialisation of Mayotte's shoreline is a reality, but remains limited nowadays (the urbanised shoreline increased from 1% in 1950 to 6% in 2016, Fig. 3 Gairin et al., 2021;Giraud-Renard et al., 2022) and worldwide (Adger et al., 2005;Airoldi and Beck, 2007;Dugan et al., 2011;Gittman et al., 2016;Matić-Skoko et al., 2020). As an example, more than 50% of the land in coastal areas is urbanised in several European countries (Airoldi and Beck, 2007). ...
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The shoreline is often at the interface of a combination of physical, ecological, and socio-economic forcing agents. Monitoring the shoreline changes across time is crucial to understand the causes of its evolution and put in place management measures. The analysis of aerial photographs from 1950 to 2016 at Mayotte Island (Indian Ocean) showed that the shoreline urbanisation is still low (6%) compared to the worldwide trend. However, a faster increase happened recently (from 3% in 1989 to 6% in 2016) owing to a strong demographic growth and socio-economic development. A multidisciplinary index was developed to assess the vulnerability of four study sites – Bandrélé, M’tsamboro, N’gouja, and Sakouli – (representative sites of beaches with fringing reefs throughout Mayotte with varying levels of urbanisation). The vulnerability of Bandrélé was lower than that of the other sites due to the presence of a mangrove at the back of the beach which plays a key role of buffer between the land and sea. M’tsamboro was the site with the highest anthropogenic pressure and highest vulnerability. Overall, as most of the shoreline is still natural at Mayotte, a sound management advice would be to put in place conservation measures to preserve natural coastal habitats, such as beaches, mangroves, seagrass beds, and coral reefs. The multidisciplinary vulnerability index developed in this study can be a useful tool to help coastal managers in the decision-making and prioritisation of actions to undertake on the shore.
... Intrinsic ecological functions and the role of natural dynamics were often overlooked and damaged. Consequently, coastal habitats with their characteristic species and system dynamics, often disappeared and biodiversity declined (Gittman et al. 2016, Dethier et al. 2017, Powell et al. 2019. Hard barriers and urban development also caused the natural coastal zones to become narrower (which is termed 'coastal squeeze'), with the consequence that natural buffer zones lost much of their function and sometimes even disappeared. ...
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The paper introduces nature‐based solutions (NBS) and their application in coastal adaptation management. NBS seek to make use of local natural elements and processes in coastal ecosystems, as much as possible, to harness forces of nature for the benefit of society. We focus on soft sedimentary coasts, like beaches and dunes, salt marshes, seagrass beds and mangroves. By shifting coastal management from conventional ‘Building in Nature' to ‘Building with Nature', NBS can be seen as a valuable alternative to the traditional approach, which is based on hydraulic, civil engineered designs. NBS can be applied in diverse situations and at various scales, from small‐scale (ecosystem elements, a small pond) to large‐scale (entire coastal stretches). The practice of NBS is also valuable for climate change adaptation, when forces of nature will increase. NBS requires a governance setting that makes use of an integrated approach with disciplines of ecology, economy and society working together. But integration is not yet common practise in many countries. We conclude that NBS are a promising alternative to the traditional approach. Because the practise still is relatively young, more field and laboratory projects should be executed, in particular under extreme weather conditions. The future challenge is to build up more stakeholder acceptance and (local) trust in the concept.
... In addition to putting property and infrastructure at risk, erosion can diminish coastal habitats, particularly in areas where human activities have limited natural sources of sediments or stabilizing vegetation. Furthermore, hard armoring such as seawalls, groins, and bulkheads, which are intended to protect coastal property, can actually worsen erosion locally and on adjacent lands (Gittman et al. 2016) Saltwater Intrusion. Saltwater intrusion into coastal freshwater systems -surface and groundwater -has become a growing problem for New Jersey's coastal communities and freshwater ecosystems. ...
... Following two decades of particularly destructive tropical storms and hurricanes, coastal communities are expanding their tools for keeping people safe and protecting property and infrastructure. Although hard armoring continues to expand along populated coastal areas across the country, communities are increasingly embracing natural infrastructure as part of the solution (Gittman et al. 2016). Approaches range from protection and restoration of natural systems and use of living shorelines to voluntary buyouts and protection of coastal open space. ...
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“Nature-based Solutions” (NbS) can play an important role in community adaptation and resilience by not only ameliorating climate-related risks but also through enhancing the quality of life for community residents. This guide delves into the opportunities for integrating NbS into community adaptation planning processes with a special focus on the U.S. Climate Resilience Toolkit's “Steps to Resilience” framework.
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Jamaica, like most Small Island Developing States, is at high risk from coastal hazards due to its exposure to tropical storms, high levels of coastal development, vulnerable coastal communities and the predicted impacts of climate change. Environmental degradation has been linked to increased vulnerability to disasters. Nature‐based solutions, although not formally present in the literature, are being implemented at various scales in Jamaica. This paper presents an overview of three marine and coastal nature‐based solutions being used in Jamaica ‐ protected area management (Special Fishery Conservation Areas), mangrove restoration and coral restoration ‐ presenting an overview of current applications in Jamaica, and argues that these conservation projects traditionally focussed on biodiversity have co‐benefits as nature‐based solutions for disaster resilience. The paper closes with several future research objectives to further the use of NbS for disaster resilience in Jamaica.
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Forage fish are schooling species commonly occurring in both offshore pelagic and nearshore coastal habitats. Beyond use by some species for spawning, the dynamics of nearshore habitat use are not well understood. The objective of our study was to evaluate the spring–summer dynamics of forage fish occurrence in nearshore habitats of the Strait of Juan de Fuca, Washington. We suspected that habitat changes resulting from removal of two large dams on the Elwha River (2009–2011) may have altered fish presence and abundance. Monthly beach seine sampling in four regions along 40 km of shoreline was conducted from April to September between 2006 and 2019. We caught nearly 600,000 fish, comprising 82 different species. Nine species of forage fish accounted for 81.7% of all fishes caught; most were classified as postlarvae and juveniles based on size. There were spatial differences in the forage fish assemblage between two of our sites but no discernable year effects and no obvious impact of dam removal on forage community composition. Three species represented 78.8% of the catch: Pacific Herring Clupea pallasii, Pacific Sand Lance Ammodytes hexapterus, and Surf Smelt Hypomesus pretiosus. We used a Bayesian generalized linear mixed model to evaluate spatial and temporal variability in the probability of occurrence of these species. Each species exhibited a unique pattern of intra‐annual, interannual, and regional fluctuations. Pacific Herring occurrence progressively increased monthly, Pacific Sand Lance occurrence decreased, and Surf Smelt probability of occurrence peaked in June. Temporal variations in distribution and abundance of these species are likely driven by life history differences and biological requirements. We speculate that specific characteristics of each region, including proximity to spawning areas, spawn timing, extant current patterns, and ecosystem processes, drove variations in distribution between species.
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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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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.
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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.
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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.