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Interacting effects of habitat structure and seeding with oysters on the intertidal biodiversity of seawalls

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The construction of artificial structures, such as seawalls, is increasing globally, resulting in loss of habitat complexity and native species biodiversity. There is increasing interest in mitigating this biodiversity loss by adding topographic habitat to these structures, and/or seeding them with habitat-forming species. Settlement tile experiments, comparing colonisation of species to more and less complex habitats, have been used to inform eco-engineering interventions prior to their large-scale implementation. Most studies have focused on applying one type of intervention (either adding habitat structure or seeding with native organisms), so it is unclear whether there are greater benefits to biodiversity when multiple interventions are combined. Using a fully orthogonal experiment, we assessed the independent and interactive effects of habitat structure (flat vs. crevice/ridges) and seeding with native oysters (unseeded vs. seeded) on the biodiversity of four different functional groups (sessile and mobile taxa, cryptobenthic and pelagic fishes). Concrete tiles (flat unseeded, flat seeded, complex unseeded and complex seeded) were deployed at two sites in Sydney Harbour and monitored over 12 months, for the survival and colonisation of oysters and the species density and abundances of the four functional groups. The survival of seeded oysters was greater on the complex than flat tiles, at one of the two sites, due to the protective role of crevices. Despite this, after 12 months, the species density of sessile invertebrates and the percentage cover of seeded and colonising oysters did not differ between complex and seeded tiles each of which supported more of these variables than the flat unseeded tiles. In contrast, the species density of mobile invertebrates and cryptobenthic fishes and the MaxN of pelagic fishes, at 1 month, were only positively influenced by seeding with oysters, which provided food as well as habitat. Within the complex seeded and unseeded tiles, there was a greater species density of sessile taxa, survival and percentage cover of oysters in the crevices, which were more humid and darker at month 12, had lower high temperature extremes at months 1 and 12, than on the ridges or flat tiles. Our results suggest that eco-engineering projects which seek to maximise the biodiversity of multiple functional groups on seawalls, should apply a variety of different microhabitats and habitat-forming species, to alter the environmental conditions available to organisms.
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RESEARCH ARTICLE
Interacting effects of habitat structure and
seeding with oysters on the intertidal
biodiversity of seawalls
Elisabeth Marijke Anne StrainID
1,2¤a
*, Vivian Ruth Cumbo
3‡
, Rebecca Louise Morris
4¤b‡
,
Peter David Steinberg
1,2‡
, Melanie Jane Bishop
3
1Sydney Institute of Marine Science, Mosman, New South Wales, Australia, 2Centre for Marine Science
and Innovation, School of Biological, Earth and Environmental Sciences University of New South Wales,
Sydney, New South Wales, Australia, 3Department of Biological Sciences, Macquarie University, Macquarie
Park, New South Wales, Australia, 4Centre for Research on Ecological Impacts of Coastal Cities, School of
Life and Environmental Sciences, The University of Sydney, Sydney, New South Wales, Australia
These authors contributed equally to this work.
¤a Current address: Institute for Antarctic and Marine Science, University of Tasmania, Tasmania, Australia
¤b Current address: National Centre for Coasts and Climate, School of Biosciences, University of Melbourne,
Victoria, Australia
‡ These authors also contributed equally to this work.
*strain.beth@gmail.com
Abstract
The construction of artificial structures, such as seawalls, is increasing globally, resulting in
loss of habitat complexity and native species biodiversity. There is increasing interest in miti-
gating this biodiversity loss by adding topographic habitat to these structures, and/or seed-
ing them with habitat-forming species. Settlement tile experiments, comparing colonisation
of species to more and less complex habitats, have been used to inform eco-engineering
interventions prior to their large-scale implementation. Most studies have focused on apply-
ing one type of intervention (either adding habitat structure or seeding with native organ-
isms), so it is unclear whether there are greater benefits to biodiversity when multiple
interventions are combined. Using a fully orthogonal experiment, we assessed the indepen-
dent and interactive effects of habitat structure (flat vs. crevice/ridges) and seeding with
native oysters (unseeded vs. seeded) on the biodiversity of four different functional groups
(sessile and mobile taxa, cryptobenthic and pelagic fishes). Concrete tiles (flat unseeded,
flat seeded, complex unseeded and complex seeded) were deployed at two sites in Sydney
Harbour and monitored over 12 months, for the survival and colonisation of oysters and the
species density and abundances of the four functional groups. The survival of seeded oys-
ters was greater on the complex than flat tiles, at one of the two sites, due to the protective
role of crevices. Despite this, after 12 months, the species density of sessile invertebrates
and the percentage cover of seeded and colonising oysters did not differ between complex
and seeded tiles each of which supported more of these variables than the flat unseeded
tiles. In contrast, the species density of mobile invertebrates and cryptobenthic fishes and
the MaxN of pelagic fishes, at 1 month, were only positively influenced by seeding with oys-
ters, which provided food as well as habitat. Within the complex seeded and unseeded tiles,
there was a greater species density of sessile taxa, survival and percentage cover of oysters
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PLOS ONE | https://doi.org/10.1371/journal.pone.0230807 July 16, 2020 1 / 21
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OPEN ACCESS
Citation: Strain EMA, Cumbo VR, Morris RL,
Steinberg PD, Bishop MJ (2020) Interacting effects
of habitat structure and seeding with oysters on the
intertidal biodiversity of seawalls. PLoS ONE 15(7):
e0230807. https://doi.org/10.1371/journal.
pone.0230807
Editor: Christopher W. McKindsey, Maurice
Lamontagne Institute, CANADA
Received: August 5, 2019
Accepted: May 20, 2020
Published: July 16, 2020
Copyright: ©2020 Strain et al. This is an open
access article distributed under the terms of the
Creative Commons Attribution License, which
permits unrestricted use, distribution, and
reproduction in any medium, provided the original
author and source are credited.
Data Availability Statement: All relevant data are
within the manuscript and its Supporting
Information files.
Funding: We thank The Ian Potter Foundation (PS),
Harding Miller Foundation (PS, ES, MB) and The
New South Wales Government Office of Science
and Research (PS, ES, MB) for their financial
support. We also thank Reef Design Lab Pty Ltd for
designing and producing the tiles used during the
study. This study was part of the World Harbour
Project. The funders provided support in the form
in the crevices, which were more humid and darker at month 12, had lower high temperature
extremes at months 1 and 12, than on the ridges or flat tiles. Our results suggest that eco-
engineering projects which seek to maximise the biodiversity of multiple functional groups
on seawalls, should apply a variety of different microhabitats and habitat-forming species, to
alter the environmental conditions available to organisms.
Introduction
Habitat complexity, the physical structure of an environment, is one of the key drivers of spe-
cies biodiversity [13]. At small scales, habitat complexity is frequently correlated with an
increased species diversity, species density and abundance of organisms [4,5]. Complex habi-
tats can promote species coexistence by providing a wider range of niches that reduce niche
overlap [6,7] and by increasing the area available for organisms to occupy [1]. Additionally,
habitat complexity may influence species diversity, species density and abundances by weaken-
ing predator-prey interactions [8], by dampening the effects of disturbance [9], by determining
the range of organismal body sizes that can be supported by an environment [4,10], and by
intercepting and retaining organic matter [11].
The critical role of habitat complexity in determining species diversity, species density and
abundance is of concern, given the ongoing habitat loss and homogenisation [2]. In urban
marine environments, natural habitats are being replaced by artificial structures (e.g. seawalls,
groynes, breakwaters, wharves) with reduced complexity [1214]. The smooth, relatively
homogenous, surface of these structures typically support a reduced species diversity and spe-
cies density as compared to the complex natural habitats they replace [6,15]. There is increas-
ing interest in how complexity might be incorporated into the design of urban marine
structures so as to enhance their ecological value [1619]. Both the addition of topographic
habitat structure (e.g. grooves, pits, ledges, water retaining structures; hereafter referred to as
habitat structure) and habitat-forming species (e.g., corals, bivalves and seaweeds) have been
proposed as mechanisms by which the biodiversity of these structures might be enhanced [19].
The type of complexity added and whether it is structural or organismal (i.e. habitat-form-
ing species) may influence its impact on the species density and abundances of colonising and
associated organisms. As animals of different body size utilise space differently, certain scales
of habitat structure may disproportionately benefit particular species that fit within their inter-
stices [10,20], providing them with protection from predators and/or abiotic stressors. Addi-
tionally, organisms vary in their niche requirements, so depending on the type of
microhabitats that the complexity creates, different taxa may be promoted [19]. For example,
on intertidal marine hard substrates, whereas algae and sessile invertebrates are each limited in
abundance by the availability of space for attachment and grazing, algae have the additional
requirement of a moist microhabitat, that prevents desiccation, and adequate light for photo-
synthesis. The addition of habitat structure that promotes water retention in well-lit environ-
ments may disproportionately enhance algae [2123]. Sessile invertebrates in contrast, could
benefit from microhabitats that provide protection from predators, but that are shaded, so that
algal competitors cannot survive [2426].
Whether complexity is provided by structural features or habitat-forming species may
influence the longevity of effects on the species density and/or abundances of colonising
organisms. In contrast to abiotic habitat structure, the complexity provided by habitat-forming
species may be more dynamic, varying as a function of habitat-former growth, and mortality
[27]. Nevertheless, even where habitat-forming species are short-lived, they may have
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of salaries for authors [ES, VC], but did not have
any additional role in the study design, data
collection and analysis, decision to publish, or
preparation of the manuscript. The specific roles of
these authors are articulated in the ‘author
contributions’ section.
Competing interests: The authors have read the
journal’s policies and the authors of this paper have
the following competing interests: Reef Design Lab
Pty Ltd. There are no patents, products in
development or marketed products associated with
this research to declare. This does not alter our
adherence to PLOS ONE policies on sharing data
and materials.
important ongoing effects by influencing the initial trajectory of species colonisation [28]. For
example, habitat-forming species that are dominant space occupants may inhibit some species
from colonising by pre-empting settlement space [29,30]. Alternatively, in ameliorating par-
ticular abiotic and biotic stressors, they may facilitate the survival of colonising species (e.g.
sessile or mobile taxa or cryptobenthic fishes) in otherwise unfavourable environments, and
enable them to recruit to size classes or developmental stages at which they can then persist,
even in the event of the demise of the habitat-forming species [31,32]
Additionally, where habitat-forming species provide the structure, their influence on bio-
logical communities may not only reflect their structural attributes but also their biological
functions [33,34]. For example, the facilitation of colonising organisms by habitat-forming
bivalves is not only a result of their structure, but also their filtration modifying patterns of set-
tlement [35] and producing pseudofaeces, which in turn can provide food and habitat to infau-
nal taxa [36]. Additionally, bivalves may serve as prey items for some fish and invertebrate
species, such that their influence on biodiversity is also trophic [37].
Small-scale experiments, investigating how different types of complexity enhance biodiver-
sity, can be informative in designing eco-engineering interventions that enhance the biodiver-
sity of artificial structures at larger spatial scales, and across a range of environmental settings.
These experiments may manipulate complexity by modifying abiotic habitat structure (e.g. by
adding pits, crevices or towers), or through the transplant of habitat forming species, such as
bivalves, corals and seaweeds [19]. The majority of research has assessed effects of abiotic habi-
tat structure and habitat-forming species on biodiversity independently of one another, but
there may be benefits of adding the two together, particularly if the two forms of complexity
interact synergistically [7,38]. The interactive effects of abiotic habitat structure and habitat-
forming species are largely untested on seawalls, particularly for cryptobenthic and pelagic
fishes that may use complexity for shelter or foraging [39]. Positive feedbacks between the two
forms of complexity may occur where habitat structure protects habitat-forming species from
predators and/or other environmental stressors (i.e. high maximum or variation in tempera-
tures, desiccation and wave exposure) allowing them to persist through time and support
dense and diverse assemblages [40,41]. However, in other instances there may be redundancy
in their combined effects [39] or they may act independently [42].
Recent studies of tiles attached to seawalls demonstrated that, cryptobenthic fishes inter-
acted more with structurally complex than flat tiles [39], while pelagic fishes interacted more
with tiles seeded with native bivalves and/or algae relative to unseeded tiles [7,39]. In general,
the effects of these small-scale manipulations were greater for cryptobenthic than pelagic fishes
[39]. This is because cryptobenthic fishes spend more time on complex habitats, as they are
better at avoiding predators in crevices than on flat surfaces [39]. Seeding may, however, also
influence patterns of habitat utilisation by pelagic fishes if the seeded organism represents a
prey item for them or enhances the availability of prey items relative to naturally fouled tiles
[39], or adjacent habitats [27]. These results suggest that adding habitat structure and seeding
with native bivalve on tiles can influence the behaviour of both cryptobenthic and pelagic
fishes [27,39] however the effects on species density and MaxN through time remains unclear.
Here we conduct a small-scale intertidal tile experiment in an urban context to assess how
manipulation of habitat structure, through the addition of cervices and ridges and seeding
with oysters to tiles, interact to influence the species density and abundances of sessile taxa,
mobile invertebrates, and cryptobenthic and pelagic fishes. The tiles are affixed to vertical sea-
walls to match the orientation of many artificial structures in the marine environment. We
hypothesise that in contrast to the temporally persistent complexity of crevices and ridges, the
habitat provided by transplanted oysters will be temporally dynamic due to processes such as
mortality and growth. Nevertheless, even in the event of considerable mortality of the oysters,
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we predict that initially seeded tiles will support distinct assemblages to those that are
unseeded, due to effects on species assembly that occur early in the colonisation process. Addi-
tionally, we expect that the presence of crevices will enhance oyster survivorship by providing
protection from environmental stressors (i.e. extreme temperatures and desiccation) and
reducing algal overgrowth through shading, such that there are non-additive effects of provid-
ing the two forms of complexity.
Despite the overall prediction of synergistic effects of crevices/ridges and seeding with oysters
on the species density and abundances, we expect that effects of the two factors will vary among
the four functional groups of organisms. We expect that sessile taxa, mobile invertebrates and
small cryptobenthic fishes (defined here as those associated with the substrate) will respond directly
to habitat structure and any amelioration of extreme temperatures, desiccation stress and preda-
tion it provides, so will display positive non-additive responses to abiotic habitat structure and oys-
ter seeding. Finally, we predict that pelagic fishes (defined here as those in the water adjacent to
tiles) that may utilise oysters as a food resource will spend more time at tiles seeded than unseeded
with oysters, resulting in greater species density and abundances associated with seeded tiles.
Methods
Permit
Permission to work on the seawalls was obtained directly through the North Sydney and
Leichhardt councils and permit P03/0029-4.1.
Study sites
The experiment was undertaken on sandstone vertical seawalls, which were located at two sites
(Morts Bay Park, Balmain, Site 1: -33.854, 151.185, Sawmillers Reserve, Waverton, Site 2:
-33.845, 151.200 GPS), within Sydney Harbour, New South Wales, Australia [7]. The sites (1
km apart) were each situated approximately 7 km from the mouth of the harbour, where the
semi-diurnal tides have a mean amplitude of 1.62 m. The experiment was conducted in the
mid-intertidal zone (i.e. MLWS + 0.7–0.9 m) on each of these seawalls, where low density pop-
ulations (relative to nearby natural rocky shores) of the native bivalve Sydney rock oyster, Sac-
costrea glomerata, naturally occurred [43].
Experimental design
We manipulated habitat structure on seawalls using 0.25 ×0.25 m concrete tiles that were
either flat (surface area = 0.0625 m
2
) or complex (surface area = 0.136 m
2
) with five 5 cm high
ridges, each separated by 1.5 to 5 cm wide crevices (Fig 1). The tiles were produced from
moulds designed by Reef Design Lab (Melbourne, Australia) using a marine concrete mix
commonly used for seawall construction. The flexible moulds allowed for manufacture of
complex geometry with undercuts and surface texture. We transplanted Saccostrea glomerata,
a habitat-forming oyster, onto a sub-set of tiles of each treatment (resulting in ‘seeded’ and
‘unseeded’ treatments), in a fully orthogonal design (Fig 1). In the Austral spring (November)
of 2015, we deployed five tiles of each of the four treatments in random order, approximately 1
m apart, at each of the two sites, to avoid any spatial auto-correlation in the mobile inverte-
brate observations (26).
For the seeded treatments, 52 juvenile S.glomerata (mean ±SE shell height: 23.7 ±3.9 mm)
were attached in clumps of 4–5 individuals using non-toxic epoxy glue (RMF—Vivacity
Marine Ltd Pty). On seeded flat tiles, the clumps of S.glomerata were placed haphazardly and
on the seeded complex tiles, half of the oysters (i.e. 26 individuals) were attached in clumps to
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the crest of ridges and the other half (26) were attached in clumps into the crevices. At each
site, the tiles were deployed in a single horizontal row. The complex tiles were positioned so
that the crevices and ridges were orientated horizontally.
Sessile and mobile taxa
In situ sampling. One, 6 and 12 months after the tiles were deployed, we assessed the
number of live seeded S.glomerata, the percentage cover of oysters and other sessile taxa, and
the abundances of mobile taxa on each tile, at low tide during the day. These time points were
selected to ensure that sampling was conducted at the beginning, middle and end of the experi-
ment, however month was considered a random factor in the analyses (see below) because we
had no a priori expectations about the effects of the treatments at these sampling points and
our goal was to assess the effects of the treatments over the entire duration of the experiment.
The percentage cover of oysters included both live and dead, transplanted and colonising oys-
ters summed together as it was not to distinguish between these categories in the photoqua-
drats. For each tile, we sampled the percentage cover of S.glomerata and other sessile taxa with
0.25 ×0.25 m photoquadrats, taken at a distance of 0.5 m from the tiles, using an Olympus
Tough style camera (TG-860, mode Automatic mode, 16 MP). Using Coral Point Count soft-
ware (CPCe) [44], we overlaid 100 randomly generated points on the flat tiles, or 50 points in
Fig 1. The four experimental treatments: a) flat unseeded tiles; b) complex unseeded tiles; c) flat seeded tiles; d)
complex seeded tiles. Complex tiles had ridges and crevices; seeding was with the Sydney rock oyster, Saccostrea
glomerata.
https://doi.org/10.1371/journal.pone.0230807.g001
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each of the crevices and ridge crests of the complex tiles. We sampled the same total number
of points (i.e. 100, 10 x zoom) on each of the flat and complex tiles to ensure sampling effort
was consistent between treatments. The S.glomerata and other sessile taxa directly under these
points (either primary colonisers, directly attached to the substrate, or secondary colonisers,
attached to another organism) were summed to calculate the percentage cover per tile (out of
100 points) and for each microhabitat (i.e. crevice or ridge: (number of points/50100)) on the
complex tiles. Simultaneously, in situ we counted the number of mobile invertebrate taxa (~
500μm diameter) on each tile by eye, noting their microhabitat (i.e. crevice or ridge) on the
complex tiles, as it was not possible to see these organisms in the photoquadrats.
Destructive sampling. After 12 months, the tiles were removed from the seawall, individ-
ually bagged and frozen at -5˚C until census. For each tile, we recorded the identity and per-
centage cover (both primary and secondary cover) of the sessile taxa (excluding S.glomerata)
separately from one pre-determined crevice (0.016 m
2
) and ridge (0.013 m
2
) on the complex
tiles (out of 50%) and from two areas of similar size on the flat tiles. We then removed all the
mobile taxa (>500 μm) from the same predefined areas on the complex and flat tiles, using
tweezers and by rinsing the tile area with seawater over a 500 μm sieve. All sessile and mobile
taxa were identified to species using a dissection microscope (Leica M125 x10 zoom) or, where
this was not possible, morphospecies. To ensure consistency, the same person identified and
enumerated the sessile and mobile taxa (listed in S1 Table), for all samples.
Fish
At the same time intervals (1, 6 and 12 months) as the sessile and mobile communities were
sampled, the fishes associated with the tiles were sampled during day-time high tides using
GoPro1cameras [7,38]. A camera was placed ~0.12 m above each tile and attached to the
seawall using stainless steel brackets and a camera housing [7,38]. The cameras faced down-
wards, so that the entire tile was in view [7,38]. The cameras were set to take photos every 2
seconds using the GoPro1setting ‘5 MP, wide’, and recorded images for 3 hours, during the
high tide [7,38]. The tiles were sampled haphazardly, over two days per site, using 10 cameras
per day. On each day, tiles at least 5 m apart were sampled to avoid any spatial auto-correlation
in the fish observations [7,38].
Using the video footage and relevant literature, we categorised the fish into two groups
‘cryptobenthic’ (associated with the substrate) or ‘pelagic’ (found in open water; listed in S1
Table). Throughout the study most of the pelagic fishes were adults (98%). Therefore, we did
not try to classify fish based on life stage. We quantified the number of species and abundance
of fish per frame for each of five 30 min periods of video footage (900 photos). The first 30 min
of video footage for each replicate tile were excluded from the data collection to avoid distur-
bance from the snorkelers. On the complex tiles we also noted whether the cryptobenthic fish
only, were situated in a crevice or on a ridge. The maximum number of species and abundance
recorded in any one single frame during each 30 min block was averaged across the five peri-
ods to give a single number (MaxN [45]) per tile and, for the cryptobenthic fish on the complex
tiles, microhabitat (i.e. crevice or ridge). The results presented therefore measure the difference
in the average maximum number of fishes per treatment, rather than the average number of
fishes (hereafter termed abundance) (37). MaxN is a commonly used measure of fish abun-
dance and ensures that no fish is counted twice within the sample period [45].
Environmental variables
At five periods (0–1, 2–3, 5–6, 8–9 and 11–12 months) after the tiles were deployed, we took
measurements of temperature. At each site, we deployed thirty DS1921G Themochron iButton
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data loggers (Thermodata Pty. Ltd. Warrnambool, Australia). We placed one iButton haphaz-
ardly on each flat tile (unseeded and seeded), and two iButtons (one in the crevice and one on
the crest of the ridge) haphazardly on each complex tile (unseeded and seeded), either inside
the oyster clumps (for seeded tiles) or outside the oyster clumps (for unseeded tiles). The iBut-
tons were waterproofed with Plastidip rubber coating (Plasti Dip International, Blaine, Minne-
sota, USA). The iButtons were attached to the tile with a cable tie, which was threaded through
a plastic straw, glued to the tile with non-toxic epoxy (RMF—Vivacity Marine Ltd Pty). This
enabled iButtons to be interchanged, so as to not exceed memory capacity, throughout the
experiment, without damaging the tiles or the colonising organisms. The iButtons were pro-
grammed to record temperatures at 20 min intervals, across a one-month period, with 0.5˚C
accuracy. These measurements were used to calculate the minimum, maximum, range and
standard deviation of temperature recorded across each month (16).
One, 6 and 12 months, after the tiles were deployed, we took measurements of humidity
with six DS1923 Hygrochron iButtons (Thermodata Pty. Ltd) and readings of light with the
external fibre optic sensor from a Diving PAM fluorometer (Walz Pty. Ltd.). As the Hygro-
chron iButtons could not get wet and the Diving PAM could only be used to make spot mea-
surements, readings of humidity and light were limited to low tide, with tiles sampled in
random order with respect to treatment over two consecutive days. For each of humidity and
light, we took one haphazardly positioned measurement from each flat tile and two measure-
ments from each complex tile (one in the crevice and one on the crest of the ridge). The
Hygrochron iButtons took measurements at 2 min intervals for 16 min, with the 8 measure-
ments per tile averaged to give a single value that was used in analyses.
Analyses
For each of the four functional groups (i.e. sessile taxa excluding S.glomerata, mobile inverte-
brates, cryptobenthic fishes, pelagic fishes), we calculated two univariate metrics: (1) species
density (i.e. the number of different species per tile, crevice or ridge), and (2) either cover (for
sessile organisms), counts (for mobile invertebrates) or MaxN (for fish). Species density and
abundance were calculated separately, at the level of tile and, for complex tiles, also separately
for crevices and ridges to assess how the two different microhabitats influenced the biodiver-
sity. The species density was calculated at the final time point, 12 months, via destructive sam-
pling for sessile and mobile organisms or in situ sampling for cryptobenthic or pelagic fishes,
while the abundance metrics were calculated separately at each time point using in situ sam-
pling and at month 12 using destructive sampling (for sessile and mobile taxa).
We used generalised linear mixed models (GLMMs) to test the effects of the initial treat-
ment conditions, adding habitat structure (fixed, 2 levels: flat, complex), seeding with oysters
(fixed, 2 levels: unseeded, seeded), site (random, 2 levels) and tile surface area (offset, m
2
), on
the species density of each of the four functional groups at 12 months. Analogous GLMMs
including these same factors as well as month (random repeated measures, 3 levels) were used
to test for the effects of adding habitat structure and seeding with oysters on the percentage
cover of S.glomerata (live and dead, transplants and colonising oysters) as well as the abun-
dance or MaxN, of the sessile and mobile taxa and cryptobenthic fishes and pelagic fishes. The
effect of adding habitat structure on the number of live seeded S.glomerata (transplants only)
was assessed using a GLMM that excluded the factor seeding.
To determine the effect of microhabitat identity on each of the same metrics for the sessile
and mobile taxa and cryptobenthic fish, a second set of GLMMs were run. For species density
data collected at 12 months, these had the factors: microhabitat (fixed 2 levels: crevice, ridge),
seeding with oysters (fixed, 2 levels: unseeded, seeded), site (random, 2 levels), area (offset m
2
)
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and tile (random, microhabitat nested in tile nested in seeding and site, 5 levels). Analyses of
cover, abundance and MaxN data also included the factor time (random repeated measure, 3
levels). As no pelagic fishes were recorded on the tiles, we did not test the effects of microhabi-
tat identity on the species density or MaxN of this functional group.
A third set of GLMMS tested for differences in the range of temperature (minimum, maxi-
mum and standard deviation), humidity and light levels among microhabitats of both the
complex and flat plates (2 fixed levels = crevice or ridge vs. flat) and between the two levels of
seeding with native oysters (2 fixed levels = unseeded vs. seeded). These analyses also had the
factors site, area and time (random repeated measures, 6 random levels for temperature, 3 lev-
els for humidity and 2 levels for light).
All statistical analyses were undertaken in R 3.5.0 [46]. The generalised linear mixed effects
models were conducted using the packages lme4 and figures were produced using ggplot 2
[47]. Likelihood ratio tests comparing models with and without the random effects were used
to obtain p-values for the random effects. In the models, species density, percentage cover and
counts (i.e. mobile invertebrate abundance and fish MaxN) were modelled using a Poisson dis-
tribution with a log link function. The average and standard deviation of temperature, average
humidity and light levels were modelled using a Gaussian distribution. All models were
checked for over-dispersion and spatial and temporal autocorrelation with plots and the resid-
uals were visually inspected for heteroscedasticity. Where appropriate, post hoc comparisons
were undertaken using eemeans to identify sources of treatment effects.
Results
A total of 80 taxa (excluding S.glomerata, the seeded species) were identified on tiles during
the study (S1 Table). These taxa were classified into four main groups: sessile algae and inverte-
brates (9 taxa), mobile invertebrates (39 taxa), cryptobenthic fishes (7 taxa) and pelagic fishes
(13 taxa) (S1 Table). At 12 months, the number of different species recorded across all tiles of
each treatment was: flat unseeded—24 taxa; flat seeded—43 taxa; complex unseeded—34 taxa;
and complex seeded—40 taxa.
The most abundant taxa identified during the in-situ sampling were for each of the four
main groups: the barnacle Elminius modestus (sessile algae and invertebrates; cover,
mean = 19.650%, SE = +/- 1.947%), the conniwink Bembicium nanum (mobile invertebrates;
count per tile, mean = 2.308, SE = +/-1.342); the oyster blenny, Omobranchus anolius (crypto-
benthic fishes; MaxN, mean = 0.329, SE = +/- 0.053) and the yellowfin bream Acanthopagrus
australis (pelagic fishes; MaxN, mean = 0.428, SE = +/- 0.019). Destructive sampling at month
12 revealed the barnacle Elminius modestus (cover, mean = 16.996%, SE = +/- 2.358%) was the
most abundant sessile species and the limpet Patelloida mimula (count per tile, mean = 5.925,
SE = +/- 1.089) was the most abundant mobile invertebrate on the tiles. For each analysis,
except for the number of live seeded S.glomerata and the abundance of mobile taxa, there was
no significant difference in the effects of treatments between sites (S2S4 Tables), allowing
sites to be pooled together for the post-hoc tests.
Effect of habitat structure on the number of live seeded S.glomerata, and of
habitat structure and seeding on the percentage cover of oysters
The effect of habitat structure on the number of live seeded oysters differed between the two
study sites (Fig 2A and 2B,S2 Table). At site 1, where survivorship of the oysters was greater,
there were no detectable differences in the number of live S.glomerata between the complex
and the flat tiles, at any of the sampling times (Fig 2A,S2 Table). At site 2, where greater mor-
tality was observed, the number of live S.glomerata was across all sampling times significantly
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Effects of microhabitats and seeding with oysters on intertidal biodiversity
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greater on complex than flat tiles (Fig 2B,S2 Table). This difference was driven by the signifi-
cantly higher number of live seeded S.glomerata in the crevices than the ridges of the complex
tiles at Site 2, but not Site 1 (Fig 2B,S2 Table).
The effects of adding habitat structure and seeding with oysters on the percentage cover of
S.glomerata (which included both live and dead oysters and seeded as well as colonising indi-
viduals) differed through time (Fig 3A,Table 1,S3 Table). At 1 and 6 months, the percentage
cover of S.glomerata was greater on seeded than unseeded tiles, but was unaffected by habitat
structure (Fig 3A,S3 Table). By 12 months, the percentage cover of oysters was greater on the
seeded complex, unseeded complex and seeded flat tiles (each of which did not significantly
differ) than unseeded flat tiles (Fig 3A,S3 Table).
Similarly, within the complex tiles, whether seeding, microhabitat type, or an interaction
between the two influenced the percentage cover of S.glomerata, varied through time (Fig 3B,
Table 2,S3 Table). Initially, at 1 month, the percentage cover of S.glomerata was greater in seeded
than unseeded microhabitats, irrespective of whether they were a crevice or ridge (Fig 3B,S3
Table). At 6 months, the percentage cover of S.glomerata was greater in the crevices that were
seeded, than the ridges that were seeded, each of which had greater cover than the unseeded
microhabitats (which did not significantly differ; Fig 3B,S3 Table). At 12 months, the percentage
cover of S.glomerata was significantly greater in the crevices than in the ridges, irrespective of
seeding, with seeded ridges containing greater cover than unseeded ridges (Fig 3B,S3 Table).
Fig 2. Effect of habitat structure (flat vs. complex tiles; a, b) and microhabitat identity (crevice vs. ridge, on complex tiles; c, d) on the mean (+/-SE)
number of live seeded S.glomerata (transplants only) at each of two sites, 0, 1, 6 and 12 months following tile deployment, (n = 5). Tiles started with 52
oysters.
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Effects of microhabitats and seeding with oysters on intertidal biodiversity
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Effect of habitat structure and seeding with native oysters on the species
density, and abundances of the four functional groups
The effects of the treatments on the species density differed among the four functional groups
(Fig 4,Table 1,S4 Table). Whereas for sessile taxa, species density responded to the interacting
Fig 3. Effect of seeding with oysters (unseeded vs. seeded) and a) habitat structure (flat vs. complex) and b) microhabitat identity (crevice vs. ridge, on complex tiles) on
the mean (+/-SE) percentage cover of S.glomerata (both transplants and recruits) on experimental tiles 0, 1, 6 and 12 months following deployment, (n = 5). Data are
pooled across sites.
https://doi.org/10.1371/journal.pone.0230807.g003
Table 1. Summary of the results of general linear models testing for effects of habitat structure on the number of live seeded S.glomerata and the effects of habitat
structure and seeding with native oysters on the percentage cover of S.glomerata, the species density and abundances (percentage cover, counts or MaxN) of sessile
taxa, mobile taxa, cryptobenthic and pelagic fishes. Details of these analyses are given in Appendices S2-S5.
NA
Not applicable,
ns
p>0.05, p<0.05, ��p<0.01,
��p<0.001.
Factors Number of live
seeded S.
glomerata
Cover of S.
glomerata
Species
density of
sessile taxa
Cover of
sessile taxa
Species
density of
mobile taxa
Counts of
mobile taxa
Species density of
cryptobenthic fishes
MaxN of
cryptobenthic
fishes
Species
density of
pelagic fishes
MaxN of
pelagic
fishes
Habitat
ns
��� ��
ns ns ns ns ns ns ns
Seeding
NA
�� �
ns
��
ns
� ��
ns
Habitat x
Seeding
NA
��� �
ns ns
��
ns ns ns ns
Site ��
ns ns ns ns ns ns ns ns ns
Habitat x Site
ns ns ns ns
��
ns ns ns ns
Seeding x Site
NA ns ns ns ns
ns ns ns ns
Month ��� �� �� �� �� ��
Habitat x
Month
ns
��
ns ns ns ns
Seeding x
Month
NA
��
ns
��
ns
��
Site x Month ��
ns ns
��
ns ns
Habitat x Site
x Month
ns ns ns ns ns
Habitat x
Seeding x Site
NA ns ns ns ns ns ns ns ns ns
Habitat x
Seeding x
Month
NA
��
ns
��
ns ns
Seeding x Site
x Month
NA ns ns
ns ns
Habitat x
Seeding x Site
x Month
NA ns ns
��
ns ns
https://doi.org/10.1371/journal.pone.0230807.t001
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Effects of microhabitats and seeding with oysters on intertidal biodiversity
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Table 2. Summary of the results of general linear models testing the effects of microhabitats (crevices/ridges) on the number of live seeded S.glomerata, and the
effects of the microhabitats and seeding with native bivalves on the percentage cover of S.glomerata, the species density and abundances (percentage cover, counts
or MaxN) of sessile taxa, mobile taxa, benthic and pelagic fishes. Details of these analyses are given in supplementary material S4- S5.
NA
Not applicable,
ns
p>0.05,
p<0.05, ��p<0.01, ��p<0.001.
Factors Number of live
seeded S.
glomerata
Cover of S.
glomerata
Species
density of
sessile taxa
Cover of
sessile
taxa
Species
density of
mobile taxa
Counts of
mobile taxa
Species density of
cryptobenthic fishes
MaxN of
cryptobenthic
fishes
Tile: Microhabitats ��� ��
ns ns
��� ��
ns ns
Tile: Seeding
na
��
ns ns ns ns
� ��
Tile: Microhabitats ×Seeding
na
��
ns ns
�� �
ns ns
Tile: Site
ns ns ns ns ns ns ns ns
Month
ns ns
��� ��
ns
Tile: Site x Month
ns ns ns ns ns
Tile: Microhabitat x Tile: Site ���
ns ns ns ns ns ns ns
Tile: Seeding x Tile: Site
na ns ns ns ns ns ns ns
Tile: Microhabitat x Month
ns ns ns
��
ns
Tile: Seeding x Month
na
��
ns ns ns
Tile: Microhabitat x Month x
Site
��
ns ns ns ns
Tile: Seeding x Month x Tile:
Site
na ns ns ns ns
Tile: Microhabitat x Tile:
Seeding x Tile: Site
na ns ns ns ns ns ns ns
Tile: Microhabitat x Tile:
Seeding x Month
na
ns
��
ns
Tile: Microhabitat x Tile:
Seeding x Tile: Site x Month
na ns ns ns ns
https://doi.org/10.1371/journal.pone.0230807.t002
Fig 4. Effect of seeding with oysters (unseeded vs. seeded) and i) habitat structure (flat vs. complex tiles) or ii) microhabitat identity (crevice vs. ridge, on complex tiles)
on the mean (+/-SE) species density of a) sessile algae and invertebrates, b) mobile invertebrates, c) benthic fishes and d) pelagic fishes (n = 5). Invertebrates and algae
were sampled destructively, and fish were measured in situ at 12 months following tile deployment. Data are pooled across sites.
https://doi.org/10.1371/journal.pone.0230807.g004
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Effects of microhabitats and seeding with oysters on intertidal biodiversity
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effects of habitat structure and seeding with oysters, for mobile species and cryptobenthic fish,
species density was only influenced (positively) by seeding with oysters, and for the pelagic fish
there were no detectable effects of habitat structure or seeding with oysters on species density
(Fig 4ic–4D,Table 1,S4 Table). The species density of colonising sessile taxa was greater on
the flat seeded, complex unseeded and complex seeded tiles (each of which did not signifi-
cantly differ) than for the flat unseeded tiles (Fig 4ia,S4 Table). In contrast, the species density
of mobile invertebrates and cryptobenthic fishes was greater on the seeded flat and complex
tiles than the unseeded flat and complex tiles (Fig 4ib–4C,S4 Table).
Within the complex tiles, the effects of microhabitat identity and seeding on species density
at 12 months also varied among functional groups (Fig 4ii,Table 2,S5 Table). Whereas for the
sessile taxa, species density was greater in crevices than on ridges (Fig 4iia), irrespective of
seeding, for mobile invertebrates, species density was greatest in the seeded and unseeded crev-
ices, followed by the seeded ridges and then the unseeded ridges (Fig 4iib) and for crypto-
benthic fishes, species density was only influenced (positively) by seeding (Fig 4iic,Table 2,S5
Table).
Both the in situ and destructive sampling revealed that neither the addition of habitat struc-
ture nor seeding with oysters affected the cover of sessile taxa (Fig 5ia,Table 1,S5 and S6
Tables). In contrast, in situ sampling revealed that the abundance of mobile invertebrates was
significantly influenced by interacting effects of the two treatments, month and site (Fig 5iabc,
Table 1,S5 Table). At 1 month, and for both sites, there was no effect of adding habitat struc-
ture or seeding with oysters on the abundance of mobile invertebrates (Fig 5ib,Table 1,S5
Table). At 6 months, and for both sites, there were significantly higher numbers of mobile
invertebrates on the unseeded and seeded complex tiles than the unseeded and seeded flat tiles
(Fig 5ib,Table 1,S5 Table). At 12 months, and for both sites, there were significantly higher
numbers of mobile invertebrates on seeded complex tiles than on any other treatment (Fig 5ib,
S5 Table). The destructive sampling of mobile invertebrates after 12 months, however, revealed
a greater abundance on the complex or flat seeded tiles (which did not significantly differ),
Fig 5. Effect of seeding with oysters (unseeded vs. seeded) and i) habitat structure (flat vs. complex tiles) or ii) microhabitat identity (crevice vs. ridge, on complex tiles)
on the mean (+/-SE) a) percent cover of sessile algae and invertebrates, b) count of mobile invertebrates, c) MaxN of benthic fishes and d) MaxN of pelagic fishes
sampled in-situ 1, 6 and 12 months following tile deployment (n = 5). Data are pooled across sites.
https://doi.org/10.1371/journal.pone.0230807.g005
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than either the complex unseeded tiles or the flat unseeded tiles, with the latter treatment
recording the fewest mobile invertebrates (S6 Table). The MaxN of the cryptobenthic fish was
much higher adjacent to the seeded than the unseeded tiles through time (Fig 5ic,S5 Table)
while the MaxN of pelagic fishes was greater for seeded than unseeded tiles at 1 month,
although this effect subsequently disappeared at 6 and 12 months (Fig 5D,S5 Table).
Within the complex tiles, the percentage cover of sessile taxa was unaffected by microhabi-
tat type or seeding with native oysters, irrespective of whether sampling was in situ or destruc-
tive (Fig 5iia,Table 2,S5 and S6 Tables). In contrast, the abundance of mobile invertebrates
was significantly influenced by the interacting effects of these two treatments, the effects of
which varied through time (Fig 5iib,Table 2,S5 and S6 Tables). At 1 month, there was no
effect of microhabitat type or seeding with oysters on the abundance of mobile invertebrates.
At 6 months, there was a greater abundance of mobile invertebrates in the crevices than the
ridges and by 12 months, both the in-situ and the destructive sampling showed the number of
mobile invertebrates was significantly greater in the seeded crevices than the unseeded crev-
ices, followed by the seeded ridges, with the unseeded ridges containing the least (Fig 5iic,S5
and S6 Tables). Across all time points, the MaxN of the cryptobenthic fish was much higher
adjacent to the seeded than the unseeded microhabitats, irrespective of whether they were
crevices or ridges (Fig 5iid,Table 2,S5 Table).
Effects of habitat structure and seeding with native oysters on the
environmental variables
Of the five environmental variables measured (maximum, minimum and standard deviation
temperature, humidity and light), all but minimum temperature displayed significant treat-
ment effects (Figs 6and 7,S7 Table). Maximum and standard deviation temperature as well as
light displayed effects of microhabitat that were independent of seeding, but which varied
through time (Figs 6and 7,S7 Table). At all-time points (excluding 9 months), the maximum
temperature was significantly lower in the crevices than the flat tiles, but there were no detect-
able differences between the ridges and the flat tiles (Fig 6A,S7 Table). At 1, 3 and 12 months,
but not 6 and 9 months, the crevices displayed a significantly smaller standard deviation of
temperature as compared to flat tiles, but ridges displayed no difference to the flat tiles (Fig 6C,
S7 Table). At both time points (1, 12 months) there were significantly higher light readings on
flat tiles and on ridges, which displayed similar readings to one another, than in crevices (Fig
7B,S7 Table). In contrast, humidity displayed effects of microhabitat that were dependent on
seeding and on time (Fig 7A,S7 Table). At all the three time points a similar humidity was
Fig 6. Effect of microhabitat type (flat vs crevice vs ridge) and seeding with native oysters (unseeded vs seeded) on the mean (+/-SE) (a) maximum, b) minimum and
(b) standard deviation (SD) of temperature through time (months) since tile deployment (n = 5). Data are pooled across sites.
https://doi.org/10.1371/journal.pone.0230807.g006
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recorded on flat tiles as on ridges, irrespective of seeding (Fig 7A,S7 Table). The crevices and
flat microhabitats displayed no difference in humidity at 1 month and displayed differences
only when unseeded at 6 months, but by 12 months crevices were more humid than flat micro-
habitats, independent of seeding (Fig 7A,S7 Table).
Discussion
Our study provides the first experimental test of how habitat structure (crevices/ridges) and
seeding with habitat-forming species (oysters) interact to influence the species density and
abundances of four functional groups of organisms (sessile and mobile taxa and cryptobenthic
and pelagic fishes) on vertically orientated substrate, in a highly modified urban marine envi-
ronment. The two types of experimental manipulations generally had positive effects on the
species density and abundances. Nevertheless, whether the effects of habitat structure or seed-
ing with oysters were independent or interactive varied among functional groups.
The addition of habitat structure had a positive influence on the survival of seeded S.glo-
merata at the site with the highest morality and, at both sites, on the percentage cover of seeded
and colonising oysters. Consequently, by 12 months, we found a similarly high percentage
cover of oysters on seeded and unseeded complex tiles, although in the absence of habitat
structure, oyster cover was greater on seeded than unseeded tiles. The generally greater sur-
vival and colonisation of oysters on the complex tiles relative to the flat tiles, appeared to be
due to the protective role of the crevices, which reduced oyster mortality. The crevices may
have protected oysters against both predation by pelagic fishes, which were more abundant
around the seeded than unseeded tiles during the first 6 months, and summer low tide temper-
ature extremes, which can be lethal to juvenile oysters [7,48]. Maximum temperatures
recorded in the crevices were up to 6.1˚C cooler than on the ridges or flat plates in the first few
months of our study, which corresponded to the Austral summer. Irrespective of the mecha-
nism, these results highlight the importance of habitat structure in providing a refuge for juve-
nile oysters against predatory and environmental stressors.
Interestingly, there were no long-term effects of seeding with S.glomerata on the percentage
cover of oysters on the complex tiles. Each of our study sites contained low density populations
of wild oysters (relative to nearby rocky shores) [43] and oysters recruited in significant num-
bers to crevices, swamping the effects of seeding. Hence, at sites with a supply of S.glomerata
recruits, the benefits of seeding may for enhancing percentage cover of oysters, in the longer-
Fig 7. Effect of microhabitat (flat vs crevice vs ridge) and seeding (unseeded vs seeded) on the mean (+/-SE) (a) humidity (%) and (b) light (mmol m
-2
s
-1
), through time
(months), (n = 5). Data are pooled across sites.
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term, be minimal. Instead, seeding efforts may be better directed at sites where there are dis-
persal barriers to oyster settlement (perhaps introduced by coastal structures), but where envi-
ronmental conditions are otherwise suitable for their survivorship, growth and reproduction
[49].
The functional groups of organisms that responded most strongly to the manipulations of
habitat structure and seeding with oysters were those whose body size most closely matched
the dimensions of the interstices formed by the interventions. That is, our experimental
manipulations which created microhabitats at the scale of cms, were most effective for enhanc-
ing the species density and/or abundances of benthic algae and invertebrates and crypto-
benthic fishes but not pelagic fishes on seawalls [38,5052]. After 12 months, the species
density of sessile algae and invertebrates and the percentage cover of S.glomerata was greater
on the tiles for which one or both types of interventions were manipulated relative to the flat
unseeded tiles. In contrast, the species density and abundances (counts or MaxN) of mobile
taxa and cryptobenthic fishes on the tiles, were only influenced, positively, by seeding with oys-
ters. These results suggest that even through the habitat structure provided by the seeded S.glo-
merata varied through time due to morality and growth of the oysters, there may be lasting
effects of the initial seeding on mobile taxa and cryptobenthic fishes.
The similar species density of sessile taxa recorded between tiles irrespective of whether
they contained habitat structure, seeded oysters or both suggests there was some redundancy
in the effects of the treatments, or, alternatively, because oyster cover did not differ between
the three treatments, that oyster cover was the driver of species density. Oysters may influence
species density by: enhancing the surface area for attachment and grazing; by providing micro-
habitats between their shells that are protected from extreme low-tide air temperatures, desic-
cation stress and/or predators; through their biological functions of filter feeding, which can
reduce the abundance of competitively dominant spores, or create currents that facilitate set-
tlement; and through biodeposition which provides organic material [5355]. The species den-
sity of mobile invertebrates and cryptobenthic fishes, may have responded more strongly to
the initial seeding treatment than sessile taxa as they can move onto tiles within hours of their
deployment to graze and utilise the microhabitat. In contrast, the colonisation of sessile taxa is
dependent on larval settlement and may take longer to establish. Yet despite the rapid response
of mobile species density to seeding with oysters, their abundances first displayed effects of
habitat structure at 6 months and seeding at 12 months, Differences in the species density of
sessile taxa among flat tiles and those with habitat structure may reflect the settlement choices
of the organisms for cooler, shaded, moist and potentially less wave exposed surfaces [56],
greater post-settlement mortality on unprotected surfaces, or reduced competition for space
where more hard substrate is available for settlement.
Our experimental manipulations of seeded oysters were predicted to influence the species
density and MaxN of pelagic fishes by enhancing prey availability and altering their foraging
behaviour. However, contrary to our expectations, there were no temporally persistent effects
of seeding on the species density or abundances of pelagic fishes. Previous research suggests
that the relatively short-lived effect of seeding on the MaxN of pelagic fish, which disappeared
after 6 months, was related to changing patterns of foraging on the tiles [7]. The most abun-
dant pelagic fishes at 1 and 6 months were the Yellowfin bream (Acanthopagrus australis) and
the Fanbelly leather jacket (Monacanthus chinensis). Both of these species are common around
artificial habitats [57,58] and are known predators of the Sydney rock oyster, S.glomerata
[5961], including those growing on the unprotected surfaces of seawalls [7,62]. The absence
of an effect of seeding on pelagic fishes after 6 months may reflect the loss of the oyster prey
resource due to growth of oysters out of size classes most vulnerable to predation [7], and/or
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the complete removal of unprotected individuals from exposed surfaces, leaving only those in
hard-to-get protected microhabitats.
On seawalls, the distribution and abundances of organisms are often limited by abiotic and
biotic stressors [15,6367]. There are increasing efforts to eco-engineer missing microhabitats
on seawalls to enhance the biodiversity and ecosystems functioning of their communities (e.g.
clearance rates, carbon sequestration or bioprotection), [23,38,42,6769]. Small-scale experi-
ments such as that conducted here can be useful in identifying appropriate eco-engineering
approaches before they are scaled up. Our study clearly demonstrates that crevices can
enhance shading and moisture retention and reduced the high temperature extremes at low
tides compared with ridges and flat tiles. They may also reduce fish predation on the colonising
taxa, although this was not directly quantified in this study.
Overall the presence of crevices increased the range of environmental conditions present
on complex tiles which in turn could lead to greater co-existence of sessile and mobile taxa
through portioning of resources [6]. In demonstrating that crevices support increased species
density of sessile taxa and species density and abundances of mobile taxa as compared to
ridges, our results support previous suggestions that the addition of crevices to homogenous,
flat surfaces such as seawalls might be a successful ecoengineering approach [19]. Nevertheless,
to fully assess the benefits, we would also need to compare the associated community of the
creviced tiles to those of featureless seawalls.
Our research suggests that in addition to cervices, seeding with oysters influences the biodi-
versity of the sessile taxa, mobile taxa and cryptobenthic fishes supported by artificial substrata
[70,71]. We demonstrated that the species density of sessile taxa benefited from adding either
habitat structure or seeding with oysters, while the species density and/or abundances of
mobile taxa and cryptobenthic and pelagic fishes were primarily influenced, positively, by
seeding. As different groups of organisms responded most strongly to different types of inter-
vention, eco-engineering projects aimed at maximising the biodiversity of multiple functional
groups might benefit from the creation of a variety of different types of microhabitats on any
given structure, that increase the breadth of niche space available to organisms [19]. Alterna-
tively, projects that focus on enhancing individual species (e.g. large predators or protected
species) may want to target particular interventions that have the greatest benefit for these [72,
73].
Our small-scale tile experiment tested the effects of adding habitat structure and seeding
with S.glomerata on the species density and abundances of four functional groups, at two sites
and through time. Using this design, it was not possible to disentangle the effects of the initial
treatment conditions from the effects of later colonising biota on the responses of the four
functional groups. Further research, which manipulates different size classes and numbers of
live and dead oysters, across multiple sites and locations, over which there is a greater range of
species, is needed to inform the design of eco-engineering interventions. Such research should
consider how site characteristics, the local species pool, and the scale of complexity relative to
the size of the organisms influences outcomes of habitat manipulation [74]. When coupled
with information on the costs and benefits of eco-engineering interventions, and the scales of
interventions that maximise ecological benefits, the results of such studies will enable manag-
ers to make informed decisions regarding the best measures, given goals and site
characteristics.
Supporting information
S1 Table. List of functional groups and species utilising each of the four experimental
treatments during the experiment. Species/morphospecies are marked as present (+) or
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Effects of microhabitats and seeding with oysters on intertidal biodiversity
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absent (-) based on sampling at month 12, whereby sessile and mobile species were censused
using destructive sampling and pelagic and cryptic fish were sampled using GoPros in situ. If
species were recorded at months 1 and/or 6, but not at month 12, they are listed but marked as
absent at 12 months.
(DOCX)
S2 Table. Results of generalised linear models testing the effects of habitat structure (flat
vs. complex tiles) and microhabitat identity (crevice vs. ridge, nested within the complex
tiles), on the number of live seeded S.glomerata sampled in-situ.The surface area of the
tiles or microhabitats (offset), site and month (repeated measure) were also included in the
model. Post hoc tests for significant factors of interest are shown. Tests significant at α= 0.05
are shown in bold.
(DOCX)
S3 Table. Results of generalised linear models testing the effects of habitat structure (flat
vs. complex) and microhabitat identity (crevice vs. ridge, nested within the complex tiles)
and seeding with oysters (unseeded [US] vs. seeded [S]) on the percentage cover of S.glo-
merata sampled in-situ through time. The surface area of the tiles or microhabitats (offset),
site and month (repeated measure) were also included in the model. Post hoc tests for signifi-
cant factors of interest are shown. Tests significant at α= 0.05 are shown in bold.
(DOCX)
S4 Table. Results of generalised linear models testing the effects of seeding with oysters
(unseeded vs. seeded) and habitat structure (flat vs. complex tiles) or microhabitat identity
(crevice vs. ridge, nested within complex tiles) on the species density sampled destructively
at month 12 for the sessile algae and invertebrates (sessile) and mobile invertebrates
(mobile) and in-situ at month 12 for cryptobenthic fishes and pelagic fishes. The surface
area of tiles or microhabitats (offset) and site were also included in the model. Post hoc tests
for significant factors of interest are shown. Tests significant at α= 0.05 are shown in bold.
(DOCX)
S5 Table. Results of generalised linear models testing the effects of seeding with oysters
(unseeded vs. seeded) and habitat structure (flat vs. complex tiles) or microhabitat identity
(crevice vs. ridge, nested within complex tiles) on the percentage cover of sessile algae and
invertebrates (sessile), the total abundance of mobile invertebrates (mobile) and the MaxN
of each of cryptobenthic fishes and pelagic fishes. Functional groups were sampled in-situ, 1
6- and 12-months following deployment of tiles. The surface area of tiles or microhabitats (off-
set), site and month (repeated measure) were also included in the model. Post hoc tests for sig-
nificant factors of interest are shown. Tests significant at α= 0.05 are shown in bold.
(DOCX)
S6 Table. Results of generalised linear models testing the effects of seeding with oysters
(unseeded vs. seeded) and habitat structure (flat vs. complex tiles) or microhabitat identity
(crevice vs. ridge, nested within complex tiles) on the percentage cover of the sessile algae
and invertebrates (sessile) and the total abundance of mobile invertebrates (mobile) sam-
pled destructively at month 12. The surface area of tiles or microhabitats (offset), site and
month (repeated measure) were also included in the model. Post hoc tests for significant fac-
tors of interest are shown. Tests significant at α= 0.05 are shown in bold.
(DOCX)
S7 Table. Results of generalised linear models testing the effects of seeding with oysters
(unseeded vs. seeded) and microhabitat identity (crevice or ridge each tested separately vs.
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Effects of microhabitats and seeding with oysters on intertidal biodiversity
PLOS ONE | https://doi.org/10.1371/journal.pone.0230807 July 16, 2020 17 / 21
flat) on the environmental variables (maximum temperature, minimum temperature,
standard deviation temperature, average humidity and average light). Site and month
(repeated measure) were also included in the model. Post hoc tests for significant factors of
interest are shown. Tests significant at α= 0.05 are shown in bold.
(DOCX)
Acknowledgments
We thank the many people that helped in deploying and monitoring the experiment, including
Emma Wilkie, Peter Simpson, Stephanie Bagala, Caleb Rankin, Dominic McAfee, Maria
Vozzo and the Port Stephens Fisheries Institute and the NSW Department of Primary Industry
for providing the oysters. This study was part of the World Harbour Project.
Author Contributions
Conceptualization: Elisabeth Marijke Anne Strain, Vivian Ruth Cumbo, Rebecca Louise Mor-
ris, Peter David Steinberg, Melanie Jane Bishop.
Data curation: Elisabeth Marijke Anne Strain, Vivian Ruth Cumbo, Rebecca Louise Morris,
Melanie Jane Bishop.
Formal analysis: Elisabeth Marijke Anne Strain.
Funding acquisition: Elisabeth Marijke Anne Strain, Peter David Steinberg, Melanie Jane
Bishop.
Investigation: Elisabeth Marijke Anne Strain.
Methodology: Elisabeth Marijke Anne Strain.
Project administration: Elisabeth Marijke Anne Strain.
Writing – original draft: Elisabeth Marijke Anne Strain.
Writing – review & editing: Elisabeth Marijke Anne Strain, Vivian Ruth Cumbo, Rebecca
Louise Morris, Peter David Steinberg, Melanie Jane Bishop.
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... Biofilms of diatoms and cyanobacteria, as well as invertebrates and macroalgae, settle, feed and grow on hard substrates [15]. These provide food and habitat to mobile invertebrates and fish [8,17]. As compared with natural hard substrates, urban structures typically support distinct ecological communities, often of reduced native biodiversity, enhanced non-native biodiversity and fewer ecological functions [15,18,19]. ...
... To date, enhancement of complexity through eco-engineering has predominantly been applied at small, experimental scales, to existing structures (but see [27] for a larger-scale example) and limited to adding a single type of microhabitat (e.g. water-retaining feature, pits, holes [6], but see [8,23] for examples simultaneously manipulating two types of complexity). Despite the positive effects of complexity that have been described at seascape and landscape scales [3], these interventions have produced effects on biodiversity ranging from highly positive to neutral [6,21,28] and on individual functional groups of organisms that range from positive to negative [21]. ...
... Based on ecological theory [1,3] and the results of small experimental-scale marine eco-engineering studies (e.g. [8,21]), we expected that the addition of complex habitat panels to otherwise relatively flat and featureless seawalls would enhance biodiversity at panel and site scales. General effects of complexity were, to the contrary, not found by this study. ...
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Urbanization is leading to biodiversity loss through habitat homogenization. The smooth, featureless surfaces of many marine urban structures support ecological communities, often of lower biodiversity, distinct from the complex natural habitats they replace. Eco-engineering (design for ecological co-benefits) seeks to enhance biodiversity and ecological functions on urban structures. We assessed the benefits to biodiversity of retrofitting four types of complex habitat panels to an intertidal seawall at patch (versus flat control panels) and site (versus unmodified control seawalls and reference rocky shores) scales. Two years after installation, patch-scale effects of complex panels on biodiversity ranged from neutral to positive, depending on the protective features they provided, though all but one design (honeycomb) supported unique species. Water-retaining features (rockpools) and crevices, which provided moisture retention and cooling, increased biodiversity and supported algae and invertebrates otherwise absent. At the site scale, biodiversity benefits ranged from neutral at the high- and mid-intertidal to positive at the low-intertidal elevation. The results highlight the importance of matching eco-engineering interventions to the niche of target species, and environmental conditions. While species richness was greatest on rockpool and crevice panels, the unique species supported by other panel designs highlights that to maximize biodiversity, habitat heterogeneity is essential. This article is part of the theme issue ‘Ecological complexity and the biosphere: the next 30 years’.
... Pre-cast concrete tiles with physical complexity such as crevices, grooves and/or pits have been attached to surfaces of coastal defence structures in intertidal areas to test their potential to increase species richness and abundance on these structures (see Strain et al., 2018;O'Shaughnessy et al., 2020O'Shaughnessy et al., , 2021. These studies have demonstrated that increasing physical complexity can result in greater diversity and abundances of colonising organisms relative to flat surfaces (Strain et al., 2020). Additionally, tiles seeded with native bivalves, thus increasing biotic complexity, can also provide more favourable microhabitats for colonising species than flat surfaces (McAfee et al., 2016;Strain et al., 2020). ...
... These studies have demonstrated that increasing physical complexity can result in greater diversity and abundances of colonising organisms relative to flat surfaces (Strain et al., 2020). Additionally, tiles seeded with native bivalves, thus increasing biotic complexity, can also provide more favourable microhabitats for colonising species than flat surfaces (McAfee et al., 2016;Strain et al., 2020). Seeding foundation species such as bivalves can enhance recruitment of colonising organisms and enable them to reach less vulnerable size classes or developmental stages, through reduction of abiotic stressors (Connell and Slatyer, 1977) and negative biological interactions (Gratwicke and Speight, 2005), and/or by producing faecal matter to serve as food or habitat for infauna (Commito and Boncavage, 1989). ...
... To date, most eco-engineering studies have tested the effects of adding physical and biological complexity separately (but see Bradford et al., 2020;Strain et al., 2020;Vozzo et al., 2021). The positive effects of physical and biological complexity for biodiversity may be enhanced by combining the two approaches (Strain et al., 2018;Bradford et al., 2020;Vozzo et al., 2021). ...
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Increasing human population, urbanisation, and climate change have resulted in the proliferation of hard coastal infrastructure such as seawalls and breakwaters. There is increasing impetus to create multifunctional coastal defence structures with the primary function of protecting people and property in addition to providing habitat for marine organisms through eco-engineering - a nature-based solutions approach. In this study, the independent and synergistic effects of physical complexity and seeding with native oysters in promoting diversity and abundances of sessile organisms were assessed at two locations on Penang Island, Malaysia. Concrete tiles with varying physical and biological complexity (flat, 2.5 cm ridges and crevices, and 5 cm ridges and crevices that were seeded or unseeded with oysters) were deployed and monitored over 12 months. The survival of the seeded oysters was not correlated with physical complexity. The addition of physical and biological complexity interacted to promote distinct community assemblages, but did not consistently increase the richness, diversity, or abundances of sessile organisms through time. These results indicate that complexity, whether physical or biological, is only one of many influences on biodiversity on coastal infrastructure. Eco-engineering interventions that have been reported to be effective in other regions may not work as effectively in others due to the highly dynamic conditions in coastal environment. Thus, it is important that other factors such as the local species pools, environmental setting (e.g., wave action), biological factors (e.g., predators), and anthropogenic stressors (e.g., pollution) should also be considered when designing habitat enhancements. Such factors acting individually or synergistically could potentially affect the outcomes of any planned eco-engineering interventions.
... Progress has also been made in the increased design suitability of artificial reefs and wave power foundations through increased complexity of refugia (Bender et al., 2020;Komyakova et al., 2019a;Langhamer and Wilhelmsson, 2009;Rouanet et al., 2015;Callaway et al., 2017) and through exact replication of the topographic complexity of natural reef surfaces (Evans et al., 2021;Erramilli and Genzer, 2019). Many of these approaches are relatively new, challenged by large data gaps, inconsistencies in the outcomes and generally occur on a relatively small-scale (in particular hard eco-engineering) (Bradford et al., 2020;Strain et al., 2020;Morris et al., 2017), highlighting the need for further research and longer-term monitoring of these innovations. Despite these challenges, a range of promising results have been reported and several are discussed here. ...
... In these highly modified environmentswhere artificial structures cannot be substituted with more natural optionseco-engineering solutions focus on improving the biodiversity or functionality of these coastal defences through the addition of topographic complexity. Key strategies include the introduction of micro-texture (Coombes et al., 2015), crevices and grooves (Martins et al., 2010;Borsje et al., 2011), artificial rock pools (Firth et al., 2014;Evans et al., 2015), precast habitat enhancement units and panels Browne and Chapman, 2014;Perkol-Finkel et al., 2018) and biotic complexity through seeding with habitat-forming species such as oysters and mussels (Bradford et al., 2020;Strain et al., 2020), corals (Ng et al., 2015) and canopy-forming algae (Perkol-Finkel et al., 2012). In addition, some studies have suggested the type of material (Dennis et al., 2018;McManus et al., 2018) and/or slope (Chapman and Underwood, 2011) of the habitat enhancement units and panels should be manipulated to increase their ecological benefits. ...
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... Over the past 70 years, about 3000 km 2 of tidal mudflats have been reclaimed using artificial hard structures along Jiangsu's coastline (Zhang et al., 2013). These artificial hard structures, including seawalls, breakwaters, ripraps and other structures, provide appropriate substrates for rocky intertidal species to settle upon and thereby facilitate their northward distribution shift (Bishop et al., 2017;Strain et al., 2019;Strain, Cumbo, et al., 2020;Wang et al., 2020). ...
... Structural complexity of the habitat could provide microhabitat for intertidal species as "thermal refugia," and micro-scale physiological thermal tolerance can play an important role in determining whether species could survive in the face of extreme events (Li, Tan, et al., 2021;Lima et al., 2016). The accumulation of microscale physiological variations could enhance the macro-scale thermal resilience of intertidal species Strain, Cumbo, et al., 2020), and thereby influence species distributional shifts affected by climate change (Bates et al., 2018;Li, Tan, et al., 2021;Liao et al., 2021;Schils & Wilson, 2006). At different time-scales, heat hardening can work synchronously with seasonal acclimatization to increase resistance of raze clam Sinonovacula constricta to high temperatures . ...
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Aim Understanding the formation and maintenance of biogeographical breaks is fundamental for developing analyses related to biodiversity and conservation. Biogeographical patterns along China's coast are changing dramatically in the face of climate change and alterations in land-use. In this paper, we sought to clarify the mechanisms responsible for the formation and maintenance of a biogeographical barrier on China's coast. Location Coastline of northern China. Methods We have reviewed literature of research related to biogeographical and phylogeographical patterns of intertidal macrobenthos along the coast of Jiangsu Province and adjacent areas, summarized the distribution patterns and biogeographical breakers. We have also reviewed literature about the processes and drivers on coastal biogeographical breaks, to clarify the mechanisms acting to the northward shift of the biogeographical break. Results The Yangtze (Changjiang) River Estuary Biogeographical Barrier (YREBB) at 30°–31°N, which serves as a coastal biogeographical boundary for the Cold Temperate Northwest Pacific Province and the Warm Temperate Northwest Pacific Province for marine species, has moved northward to ~33°–34°N due to the changes in habitat continuity, oceanographic circulation and climate factors. Consequently, a new biogeographic barrier for intertidal macrobenthos, the Subei Biogeographical Barrier (SBB) on the central coast of Jiangsu Province, has emerged. Main conclusions The formation and maintenance of the SBB are closely related to the larval dispersal potential, larval settlement success and post-settlement population establishment, all of which have been profoundly influenced by anthropogenic environmental changes. The northward shift of the YREBB and the appearance of the SBB provide an excellent model system for investigating the impacts of climate change and land-use change on coastal biogeographic patterning and for clarifying the mechanisms underlying the formation and maintenance of biogeographical barriers in the face of the unprecedented environmental changes.
... Substrates that are either complex by nature (e.g., oyster shell, limestone rock) or are fabricated to incorporate complexity (e.g., concrete blocks with cracks, surface texture, pools and holes; BESE units) can protect oyster recruits from predators as well as environmental stressors such as high temperatures that are being exacerbated by climate change (McAfee et al., 2018a;Strain et al., 2018). Successful oyster recruitment, in turn, leads to complex biogenic habitat, that supports associated invertebrates and juvenile fish (Strain et al., 2020), and can serve as a prey resource for fish and mobile crustaceans (Grabowski and Peterson, 2007;Grabowski et al., 2012). Additionally, complex substrates can facilitate biodiversity while oyster reefs are developing, and still acquiring their own biogenic complexity (Chapman and Blockley, 2009;Ido and Shimrit, 2015;Paalvast, 2015;Strain et al., 2020). ...
... Successful oyster recruitment, in turn, leads to complex biogenic habitat, that supports associated invertebrates and juvenile fish (Strain et al., 2020), and can serve as a prey resource for fish and mobile crustaceans (Grabowski and Peterson, 2007;Grabowski et al., 2012). Additionally, complex substrates can facilitate biodiversity while oyster reefs are developing, and still acquiring their own biogenic complexity (Chapman and Blockley, 2009;Ido and Shimrit, 2015;Paalvast, 2015;Strain et al., 2020). Consequently, utilization of complex substrates will be particularly important in environments where high post-settlement mortality limits oyster reef establishment, or where enhancement of native biodiversity is a key ecosystem service goal ( Table 1). ...
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... As such, so-called "Working with Nature" (WwN) techniques have gained traction over recent decades as fundamental tools in conservation, and significant progress has been made to seat the efforts to create novel or restored ecosystems within broader ecological theory (Palmer et al., 2016;Statton et al., 2018). For example, the modification of otherwise flat concrete surfaces of seawalls, greatly increasing the availability of ecological niches and substrate, has been used to encourage the settlement of native marine flora and fauna and increase biodiversity on seawalls (Perkol-Finkel et al., 2018;Strain et al., 2018Strain et al., , 2020. WwN approaches have also been used to control so-called "soft" or sediment habitats, such as the use of melaleuca fences to reduce wave energy and increase sediment accretion for mangrove restoration attempts (Van Cuong et al., 2015). ...
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... The same crevice and ridge were sampled across all tiles. When standardized by area, such subsampling provides estimates of abundance and richness that do not significantly differ from sampling the full tile (Strain et al., 2020). Counts of mobile organisms and percentage cover of sessile organisms were recorded. ...
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Over the last decade there has been a global effort to eco-engineer urban artificial shorelines with the aim of increasing their biodiversity and extending their conservation value. One of the most common and viable eco-engineering approaches on seawalls is to use enhancement features that increase habitat structural complexity, including concrete tiles molded with complex designs and precast “flowerpots” that create artificial rock pools. Increases in species diversity in pits and pools due to microhabitat conditions (water retention, shade, protection from waves, and/or biotic refugia) are often reported, but these results can be confounded by differences in the surface area sampled. In this study, we fabricated three tile types (n = 10): covered tile (grooved tile with a cover to retain water), uncovered tile (same grooved tile but without a cover) and granite control. We tested the effects of these tile types on species richness (S), total individual abundance (N), and community composition. All tiles were installed at 0.5 m above chart datum along seawalls surrounding two island sites (Pulau Hantu and Kusu Island) south of Singapore mainland. The colonizing assemblages were sampled after 8 months. Consistent with previous studies, mean S was significantly greater on covered tiles compared to the uncovered and granite tiles. While it is implied in much of the eco-engineering literature that this pattern results from greater niche availability allotted by microhabitat conditions, we further investigated whether there was an underlying species-individual relationship to determine whether increases in S could have simply resulted from covered tiles supporting greater N (i.e., increasing the probability of detecting more species despite a constant area). The species-individual relationship was positive, suggesting that multiple mechanisms are at play, and that biodiversity enhancements may in some instances operate simply by increasing the abundance of individuals, even when microhabitat availability is unchanged. This finding underscores the importance of testing mechanisms in eco-engineering studies and highlights ongoing mechanistic uncertainties that should be addressed to inform the design of more biodiverse seawalls and urban marine environments.
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Urban areas have broad ecological footprints with complex impacts on natural systems. In coastal areas, growing populations are advancing their urban footprint into the ocean through the construction of seawalls and other built infrastructure. While we have some understanding of how urbanisation might drive functional change in terrestrial ecosystems, coastal systems have been largely overlooked. This study is one of the first to directly assess how changes in diversity relate to changes in ecosystem properties and functions (e.g. productivity, filtration rates) of artificial and natural habitats in one of the largest urbanised estuaries in the world, Sydney Harbour. We complemented our surveys with an extensive literature search. We found large and important differences in the community structure and function between artificial and natural coastal habitats. However, differences in diversity and abundance of organisms do not necessarily match observed functional changes. The abundance and composition of important functional groups differed among habitats with rocky shores having 40% and 70% more grazers than seawalls or pilings, respectively. In contrast, scavengers were approximately 8 times more abundant on seawalls than on pilings or rocky shores and algae were more diverse on natural rocky shores and seawalls than on pilings. Our results confirm previous findings in the literature. Oysters were more abundant on pilings than on rocky shores, but were also smaller. Interestingly, these differences in oyster populations did not affect in situ filtration rates between habitats. Seawalls were the most invaded habitats while pilings supported greater secondary productivity than other habitats. This study highlights the complexity of the diversity-function relationship and responses to ocean sprawl in coastal systems. Importantly, we showed that functional properties should be considered independently from structural change if we are to design and manage artificial habitats in ways to maximise the services provided by urban coastal systems and minimise their ecological impacts.
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Ecosystem engineers that modify the thermal environment experienced by associated organisms might assist in the climate change adaptation of species. This depends on the ability of ecosystem engineers to persist and continue to ameliorate thermal stress under changing climatic conditions-traits that may display significant intraspecific variation. In the physically stressful intertidal, the complex three-dimensional structure of oysters provides shading and traps moisture during aerial exposure at low tide. We assessed variation in the capacity of a faster- and slower-growing population of the Sydney Rock Oyster, Saccostrea glomerata, to persist, form three-dimensional structure and provide a cool microhabitat to invertebrates under warmer conditions. The two populations of oysters were exposed to a temperature gradient in the field by attaching them to passively warmed white, grey and black stone pavers and their growth, survivorship and colonisation by invertebrates was monitored over a 12-month period. Oysters displayed a trade-off between fast growth and thermal tolerance. The growth advantage of the fast-growing population diminished with increasing substrate temperature, and at higher temperatures, the faster-growing oysters suffered greater mortality, formed less habitat, and were consequently less effective at ameliorating low-tide air temperature extremes than slower-growing oysters. The greater survivorship of slower-growing oysters, in turn, produced a cooler microclimate which fed back to further bolster oyster survivorship. Invertebrate recruitment increased with habitat cover and was greater among the slower than the faster-growing population. Our results show that the capacity of ecosystem engineers to serve as microhabitat refugia to associated organisms in a warming climate displays marked intraspecific variation. Our study also adds to growing evidence that fast growth may come at the expense of thermal tolerance.
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In many parts of the world, shorelines of estuaries are being modified because of increasing urban infrastructure. Consequently, habitats are being lost and replaced by other, artificial habitats, or are being severely modified in deleterious ways. Management of these changes requires recognition of the types of changes that are being made to estuarine shorelines, what sorts of impacts they cause and what sorts of managerial intervention are possible or desirable. Because experimentation in these managed habitats is in its early stages, it is important to provide sound ecological advice based on realistic hypotheses about consequent ecological changes. Wherever possible, management should be evaluated by well-planned experiments, involving analyses of changes in managed areas compared with changes in appropriate control areas. Where the goals of managerial action are the recovery of assemblages of species, what happens in managed areas should also be compared by equivalence tests with what happens in reference areas or relative to pre-defined ecological measures of what would constitute recovery. Here, different types of altered habitats, impacts, ecological changes and purposes of management are discussed. The needs for and nature of different types of hypotheses, experiments and analyses are reviewed with reference to relevant examples.