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\begin{document}

\title{Warm and cool-edge populations reveal high resilience of seagrass
(Posidonia oceanica) to warming}



\author[1]{Scott Bennett}%
\author[2]{Teresa Alcoverro}%
\author[3]{Demetris Kletou}%
\author[3]{Charalampos Antoniou}%
\author[4]{Jordi Boada}%
\author[5]{Xavier Buñuel Moreno}%
\author[6]{Lidia Cucala}%
\author[7]{Gabriel Jorda}%
\author[3]{Periklis  Kleitou}%
\author[6]{Guillem Roca}%
\author[6]{Julia Santana-Garcon}%
\author[3]{Ioannis Savva}%
\author[8]{Adriana Verges}%
\author[9]{Núria Marbà}%
\affil[1]{University of Tasmania}%
\affil[2]{Spanish Institution for Scientific Research}%
\affil[3]{Marine \& Environmental Research MER Lab Ltd}%
\affil[4]{Centre d'Estudis Avançats de Blanes}%
\affil[5]{CEAB}%
\affil[6]{IMEDEA}%
\affil[7]{Instituto Espanol de Oceanografia Centro Oceanografico de Baleares}%
\affil[8]{School of Biological Sciences, University of New South Wales}%
\affil[9]{IMEDEA (CSIC-UIB), Institut Mediterrani d’Estudis Avançats}%


\vspace{-1em}



  
  \date{June 17, 2024}


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\selectlanguage{english}
\begin{abstract}
Disentangling spatial variation in climate change impacts is a pressing
challenge. Here we compared the performance of Posidonia oceanica
seagrass populations to temperature, throughout a year-long
translocation experiment across 2800 km in the Mediterranean Sea.
Transplants in central and warm-edge locations experienced temperatures
\textgreater{}29 \selectlanguage{ngerman}ºC during summer, representing thermal anomalies
\textgreater{}5ºC above long-term maxima for cool-edge populations,
1.5ºC for central and \textless{}1ºC for warm-edge populations. At the
onset of the experiment, a highly selective herbivory event removed 75\%
of cool-to-warm transplant biomass but left adjacent central and
warm-edge treatments intact. Despite big differences in thermal stress
and acute herbivory, cool-edge populations recovered and matched
warm-edge populations across all performance metrics. Central
populations displayed significantly lower growth and survivorship in
response to thermal stress. Our findings reveal that intraspecific
variation in thermal performance does not necessarily reflect thermal
geography and suggest greater resilience to warming for Posidonia
oceanica than previously recognised.%
\end{abstract}\selectlanguage{ngerman}%



\sloppy


\textbf{Introduction}

Ocean warming is having profound effects on benthic marine ecosystems
across the globe (Hoegh-Guldberg \& Bruno 2010). Foundation species such
as seaweeds, seagrasses and corals can be particularly susceptible to
warming and have undergone extensive thermal stress (Hughes \emph{et
al.} 2018), mortality (Marba \& Duarte 2010), and range contraction
(Wernberg \emph{et al.} 2016) over the past two decades. In conjunction
with direct physiological impacts, ocean warming is leading to changes
in species distributions resulting in new assemblages of species coming
together and competing for space and resources within an ever-changing
physical environment (Lenoir \emph{et al.} 2020). This combination of
factors creates a diverse and dynamic array of potential outcomes of
climate change on the structure and function of marine ecosystems.

Temperature effects on populations across a species geographical range
are not homogeneous and depend on the population's thermal range
position and a species evolutionary history and adaptive capacity
(Angilletta 2009; Somero 2010). For example, if individuals from
different populations display similar optimal and upper thermal limits,
then cool-edge populations will have lower sensitivity to warming than
warm-edge populations where ambient temperatures are already close to
upper thermal limits (Bennett \emph{et al.} 2019). This is an inherent
assumption of many species distribution models that rely on the realised
distribution of species to predict the likelihood of extirpation or
range shifts under climate change (Ara\selectlanguage{ngerman}újo \& Peterson 2012). If, on the
other hand, thermal performance differs between populations-- through
local adaptation or acclimatization -- then the thermal safety margin
between ambient conditions and an individual's upper thermal limit may
remain relatively constant between populations across a species range
(Bennett \emph{et al.} 2015). This is an inherent assumption used to
estimate the thermal sensitivity of reef building corals, for which the
magnitude and the duration of local thermal anomalies are strong
predictors of thermal bleaching (McClanahan \emph{et al.} 2019).
Determining whether thermal performance varies between populations and
how this relates to local climatology or a species distribution is
critically important for our capacity to anticipate and manage climate
change impacts. For the vast majority of species, within-species
patterns of thermal sensitivity remain unknown, highlighting a
fundamental knowledge gap in climate change ecology.

An additional challenge for predicting the impacts of climate change on
marine ecosystems is the influence of different ecological processes
such as biotic interactions, on the outcomes of warming (Araújo \& Luoto
2007). As species redistribute around the world, novel interactions are
emerging that may either accelerate or buffer ecosystems from change
under different environmental contexts (Gilman \emph{et al.} 2010;
Lurgi\emph{et al.} 2012). Incorporating this complexity into models is
challenging based on current approaches (Pagès \emph{et al.} 2018). For
example, thermal performance studies are routinely conducted in
controlled laboratory environments which are invaluable for
understanding complex physiological and evolutionary processes, but
often miss key ecological processes and trade-offs that may determine an
organism's success in the wild (Buñuel \emph{et al.} 2020). While the
ecological realism of experimental systems is improving (Ullah \emph{et
al.} 2018), calibrating laboratory-based experiments with real world
multi-species ecological interactions remains a challenge. At the other
end of the spectrum, direct impacts of thermal stress on natural
ecosystems from marine heatwaves and long-term warming offer invaluable
insights into the whole-of-ecosystem response to warming (Smale \emph{et
al.} 2019). However, such events are irregularly distributed and
historically uncommon (Oliver \emph{et al.} 2018). In addition, the
ecological consequences of marine heatwaves can be catastrophic,
undermining the health of the ecosystem for which the information is
needed (Wernberg \emph{et al.} 2016).

In this study we took an intermediate approach and conducted a 12-month
translocation experiment of the seagrass \emph{Posidonia oceanica}across
its geographical range. Translocation experiments have the benefit of
directly comparing the performance of geographically distant populations
under common, relatively natural ecological settings. We conducted our
study across a steep 5ºC gradient in average annual temperatures from
the western to eastern Mediterranean Sea. The Mediterranean has
experienced rapid warming between 0.25-0.65 ºC dec\textsuperscript{-1}
over the past three decades, 2-3 times faster than the average global
ocean (Marbà \emph{et al.} 2015). Contemporary climate change has
impacted central populations of \emph{P. oceanica} in recent decades
(Marba \& Duarte 2010) and previous studies have predicted that the
species could face functional extinction by 2050 (Jordà \emph{et al.}
2012; Chefaoui \emph{et al.} 2018). High thermal sensitivity in central
populations raises questions about the sensitivity of warm-edge
populations, which regularly face summer temperatures that are similar
to heatwave conditions attributed to losses in the western basin. The
aim of this experiment was to compare the eco-physiological performance
of \emph{P. oceanica} across its thermal distribution to quantify the
thermal sensitivity of different populations to climate warming. We test
the hypothesis that thermal performance reflects the thermal geography
of a species either through local adaptation or thermal niche
conservatism across its distribution.

\textbf{Methods}

\emph{Experiment locations and climate}

Trans-Mediterranean translocation of \emph{Posidonia oceanica} fragments
took place between Catalunya (Spain), Mallorca (Spain) and Cyprus in
July 2018 and were monitored until July 2019 (Fig. 1). Sea surface
temperature data for each transplant site were based on daily SST maps
with a spatial resolution of 1/4°, obtained from the National Center for
Environmental Information (NCEI,\emph{https://www.ncdc.noaa.gov/oisst}
(Reynolds \emph{et al.}2007). These maps have been generated through the
optimal interpolation of Advanced Very High Resolution Radiometer
(AVHRR) data for the period 1981-2019. Underwater temperature loggers
(ONSET Hobo pro v2 Data logger) were deployed at the transplant sites in
Catalunya, Mallorca and Cyprus and recorded hourly temperatures
throughout the duration of the experiment (one year). In order to obtain
an extended time series of temperature at transplant sites, a
calibration procedure was performed comparing logger data with sea
surface temperature from the nearest point on SST maps. In particular,
SST data were linearly fitted to logger data for the common period.
Then, the calibration coefficients were applied to the whole SST time
series to obtain corrected-SST data and reconstruct daily habitat
temperatures from 1981-2019. Local climate data was also compared to the
global thermal distribution of \emph{P. oceanica} to assess how
representative experimental sites were of the thermal distribution of
the species (Supplementary materials). Collectively, seawater
temperatures from the three locations span the 16\textsuperscript{th} -
99\textsuperscript{th} percentile of temperatures observed across the
global thermal distribution of \emph{P. oceanica.} As such Catalunya,
Mallorca and Cyprus are herein considered to represent the cool-edge,
centre and warm-edge of \emph{P. oceanica}distribution,
respectively\emph{.}

Transplantation took place toward warmer climates and procedural
controls were conducted within each source location, resulting in six
source-to-recipient treatments (Fig. 1). Initial collection of \emph{P.
oceanica} , handling and transplantation was carried out simultaneously
by coordinated teams in July 2018 (Table S1). Each recipient location
was subsequently resampled four times throughout the experiment, in
August/September 2018 (T1), October 2018 (T2), April 2019 (T3) and
May/June 2019 (T4, Table S1).

Between 60-100 fragments were collected for each treatment. A fragment
was defined as a section of \emph{P. oceanica} containing one apical
shoot connected with approximately five vertical shoots and 10-15 cm of
rhizome with intact roots. Collection occurred at two sites within each
location, separated by approximately 1 km. Within sites, collections
were conducted between 4 -- 5 m depth and were spaced across the meadow
to minimise the dominance of a single clone and damage to the meadow.
Upon collection, fragments were transported for up to one hour back to
the nearest laboratory in shaded seawater.

\emph{Handling methods}

In the laboratory, fragments were placed into holding tanks with aerated
seawater, at ambient temperature and a 14:10 light-dark cycle. All
shoots were clipped to 25 cm length (from meristem to the tip of the
longest leaves), to standardise initial conditions and reduce biomass
for transportation. For transport by plane or ferry between locations,
fragments were packed in layers within cool-boxes. Each layer was
separated by frozen cool-packs wrapped in wet tea towels (rinsed in sea
water). All fragments spent 12 hrs inside a cool-box irrespective of
their recipient destination, including procedural controls (i.e.
cool-cool, centre-centre and warm-warm) to simulate the transit times of
the plants travelling furthest from their source location (Fig. 1a). On
arrival at the destination, fragments were placed in holding tanks with
aerated seawater at ambient temperature as described above in their
recipient location for 48 hrs, prior to field transplantation.

\emph{Measurement methods}

One day prior to transplantation, fragments were tagged with a unique
number and attached to a U-shaped peg. Morphological traits for each
fragment were measured and included: 1) length of the longest apical
leaf, width and number of leaves 2) total number of bite marks on leaves
of three vertical shoots per fragment, 3) number of vertical shoots, 4)
leaf count of three vertical shoots per fragment and 5) overall
horizontal rhizome length. A subset (n=10) of fragments per treatment
were marked prior transplantation to measure shoot growth. To do this,
all shoots within a single fragment were pierced using a hypodermic
needle. Two holes were pierced side-by-side at the base of the leaf/top
of the meristem.

\emph{Transplant methods}

All transplant sites were located in 4 -- 5 m depth in area of open
dead-matte, surrounded by \emph{P. oceanica} meadow. In Mallorca and
Cyprus, fragments were distributed between two sites, separated by
approximately 1 km. In Catalunya, lack of suitable dead matte habitat
meant that all fragments were placed in one site. Fragments were planted
along parallel transects at 50 cm intervals and with a 50 cm gap between
parallel transects (Fig. S1). Different treatments were mixed and
deployed haphazardly along each transect.

\emph{Resampling methods and herbivory}

On day 10 of the experiment, a severe herbivory event was recorded at
both warm-edge translocation sites. Scaled photos of all fragments were
taken at this time to record the effects of herbivory on transplants. At
the end of each main sampling period (T0 -- T1, T1-T2 and T3 -- T4), all
pierced fragments were collected and taken back to the laboratory to
measure shoot growth. At T1, T2 and T3, additional sets of fragments (n
= 10 per treatment) were marked using the piercing method to record
growth in the subsequent time period. In addition, at T1 and T3, n = 20
shoots within the natural meadow at each site were marked to compare
growth rates between the native meadow and transplants. Underwater shoot
counts and a scaled photo was taken to record fragment survivorship,
shoot mortality, bite marks, and shoot length among all remaining
fragments within each site and sampling time.

In the laboratory, morphological measurements (described above) were
repeated on the collected fragments and growth of transplant and natural
meadow shoots was measured. Growth (shoot elongation, cm
d\textsuperscript{-1}) of the marked shoots was obtained by measuring
the length from the base of meristem to marked holes of each leaf (new
growth) of the shoot and dividing the leaf elongation per shoot by the
marking period (in days). For each shoot, total leaf length (cm
shoot\textsuperscript{-1}) and the number of new leaves was recorded.
The rate of new leaf production (new leaves
shoot\textsuperscript{-1}d\textsuperscript{-1}) was estimated dividing
the number of new leaves produced per shoot and the marking period. New
growth was dried at 60 ºC for 48 hrs to determine carbon and nitrogen
content of the leaves, and carbon to nitrogen (C:N) ratios. Carbon and
nitrogen concentrations in the new growth leaf tissue was measured at
the beginning of the experiment and each subsequent time point for each
treatment. Nutrient analyses were conducted at Unidade de Técnicas
Instrumentais de Análise (University of Coruña, Spain) with an elemental
analyser FlashEA112 (ThermoFinnigan).

Underwater photos of shoots were analysed using ImageJ software
(https://imagej.net). Maximum leaf length on each shoot in warm-edge
transplant sites (cool-warm, centre-warm and warm-warm) were recorded
for the initial (day 10) herbivore impact, T1, T2 and T3 and related to
transplant nutrient concentrations. Herbivore impact was estimated as
the proportional change in length of the longest leaf relative to
initial length at T0.

\emph{Thermal stress}

Long term maximum temperatures were recorded as the average of annual
maximum daily temperatures in each transplant site, averaged between
years from 1981-2019. Maximum thermal anomalies were calculated as the
difference between daily temperatures in a recipient site over the
course of the experiment and the long-term maximum temperature in the
source site for each corresponding population. `Heat stress' and
`recovery' growth periods of the experiment were defined as T0 -T2
(July-October) and T2-T4 (November-June), respectively, corresponding to
periods of positive and negative maximum thermal anomalies. Thermal
anomalies experienced by the different transplant treatments were
plotted using the `geom\_flame', function in the `HeatwavesR' package
(Schlegel \& Smit 2018) of R (version 3.6.1, 2019) .

\emph{Statistical analyses}

Differences in fragment survivorship between source-recipient treatments
were analysed with a chi-squared test at the end of the heat stress
period (T2) and end of the recovery period (T4). Differences in 1) Shoot
mortality per fragment, 2) herbivory impact, 3) C:N of leaves, 4)
growth, 5) new leaf production and 6) total leaf length were each
compared between treatments (n = 6) within discrete time periods using
Analysis of variance (ANOVA). Normality was tested using Shapiro-Wilks
test and Q-Q plots, and homogeneity of variance was tested with Levene's
test and residuals plot. No transformations were applied. Tukey-HSD
analysis was used to examine differences between groups for significant
main effects. To assess the relationship between thermal performance,
source location and 1) maximum absolute realised temperature and 2)
maximum thermal stress anomalies, linear mixed effects models were used
using the lme4 package in R (Bates \emph{et al.} 2015). Source location
and temperature metric were treated as fixed effects and recipient
location' and `time point' were treated as random effects, to account
for differences in performance across locations and seasonality.
Separate models were conducted for each temperature metric (i.e. max
temperature or heat stress anomaly) and performance metric (i.e. growth,
new leaf production and total leaf length) combination. We started with
a full model based on \emph{a priori} hypotheses about the inclusion of
terms and higher order interactions. We identified the minimum adequate
model by a stepwise removal of non-significant terms using likelihood
ratio tests of the model with the effect in question against the model
without the effect in question. Visual inspection of residual plots did
not reveal any obvious deviations from homoscedasticity or normality and
therefore no transformations were conducted. P-values were obtained by
likelihood ratio tests of the full model with the effect in question
against the model without the effect in question. All analyses were
conducted in R (version 4.0.3, 10/10/2020).

\textbf{Results}

\emph{Realised thermal anomalies}

The western Mediterranean Sea experienced marine heatwave conditions
during 2018 summer, coinciding with the translocation experiment.
Maximum daily cool-edge temperatures were 24.9 ºC, 1.3 ºC above long
term summer maximum (averaged 1981-2019). Maximum daily temperatures in
central sites were 29.1ºC, also 1.3 ºC above long term summer maximum
and the second highest temperature on record. Maximum daily temperatures
at \emph{P. oceanica's} warm-edge in the eastern Mediterranean were more
typical at 29.3 ºC and 0.18 ºC above long term maximum.

As a result of the marine heatwave in the western Mediterranean, thermal
anomalies experienced by cool-edge and central transplants were similar
in central and warm-edge locations. The highest anomalies were
experienced by cool-centre and cool-warm fragments, which experienced
maximum anomalies of 5.4 ºC and 5.7 ºC above long-term maximum
temperatures in the cool-edge, respectively. Positive thermal stress
anomalies lasted for 130 days for cool-warm fragments and 102 days for
cool-centre fragments over the course of the experiment (Fig. 2, Table
S1). Centre-warm and centre-centre fragments experienced maximum
anomalies of 1.53 ºC and 1.3 ºC, respectively for a duration of 37 and
33 days above long-term summer maxima in \emph{P. oceanica's} range
centre. Warm-warm transplants experienced equally high temperatures to
cool-warm and centre-warm transplants, but the smallest relative anomaly
(0.18 ºC for 2 days) above long-term summer maxima. The coolest daily
temperatures recorded throughout the course of the experiment ranged
from 10.5 ºC in the cool-edge, 13.9 ºC in the centre and 16.2 ºC in the
warm-edge.

\emph{Selective herbivory in transplants with higher nutritional
quality}

Ten days after initial deployment, transplants were surveyed revealing
selective overgrazing on cool-warm \emph{P. oceanica} (Fig. 2). Shoot
lengths of cool-warm transplants were reduced by 74.5 ± 4\% across both
transplant sites in Cyprus, from 22.3 ± 3 cm to 5.7 ± 9 cm. In contrast,
centre-warm and warm-warm plants that were interspersed with cool-warm
plants, lost significantly less biomass, from 25 ± 3\% and 27.9 ± 5 \%
of initial length, respectively (ANOVA, F\textsubscript{(2,194)} =
37.64, p\textless{}0.001, Fig. 2A). Selective grazing patterns were
consistent between two warm-edge sites separated by 1.6 km. No other
transplants in cool-edge or central locations were over-grazed by
herbivores. Selectivity patterns in grazing rates reflected differences
in nutritional quality between seagrass populations and regional
differences in herbivore assemblages. C:N of cool-edge \emph{P.
oceanica} leaves were 18.1 ± 1.4 (mean ± SE), significantly lower (i.e.
higher nutritional quality) at the start of the experiment than C:N
ratios for central 34.7 ± 1.4 or warm-edge 52.3 ± 1.8 populations
(ANOVA, F\textsubscript{(2,8)} = 97.76, p \textless{}0.001, Fig. 2B).

By week six, significant differences in shoot length were still observed
between treatments (ANOVA, F\textsubscript{(2,245)} = 87.69,
p\textless{}0.001), but difference between treatments was reduced (Fig.
2A). Similarly, C:N in leaf tissue of the transplants started to
equilibrate with local conditions. By the recovery period, no
significant differences in C:N (p = 0.83) and no differences in
herbivory impacts (i.e. shoot length, p = 0.79) were detectable between
source populations in warm-edge sites (Fig. 2). The morphology of the
bite scars observed on the consumed seagrass in Cyprus suggested that
herbivorous fishes (\emph{Siganus rivulatus, S. luridus, Sparisoma
cretense} ) and potentially sea turtles (\emph{Chelonia mydas} ) were
responsible for the intense grazing rates, although relative
contributions of each species could not be quantified (Fig. S2).

\emph{Survivorship}

Despite experiencing extreme thermal anomalies and grazing rates at the
start of the experiment, cool-warm transplants displayed high
survivorship (number of surviving rhizome fragments) with no significant
differences observed between treatments at the end of the thermal stress
period (\selectlanguage{greek}χ\selectlanguage{english}\textsuperscript{2} = 10.65, df = 5, p = 0.06, Fig. 3).
Survivorship then declined in all treatments over winter months (between
October and April), potentially due to dislodgement from storms. Notably
however, both centre-centre and centre-warm treatments displayed
significantly lower survivorship than all cool-edge and warm-edge
populations by the end of the experiment (\selectlanguage{greek}χ\selectlanguage{english}\textsuperscript{2}= 58.07,
df = 5, p\textless{}0.001, Fig. 3A).

The number of living shoots per fragment of rhizome was relatively
stable over time, with slight declines over the period of thermal stress
(Fig. 3B). Warm-warm and cool-centre treatments displayed significantly
lower rates of shoot mortality than central and other cool-edge
treatments at the end of the thermal stress period (ANOVA,
F\textsubscript{(5,388)} = 8.118, p \textless{}0.001). By the end of the
recovery period, changes in shoot numbers on the surviving fragments had
stabilised and cool-centre fragments had recovered to display no net
change from the beginning of the experiment.

\emph{Growth patterns}

The number of leaves per shoot remained relatively stable through time
and retained an average of 4.8 \selectlanguage{ngerman}± 0.13 leaves in cool-cool treatments,
4.3 ± 0.75 in centre-centre treatments and 4.3 ± 0.16 leaves per shoot
in warm-warm treatments by the end of the experiment. Centre-warm
treatments were the exception and declined from 4.0 ± 0.29 leaves
shoot\textsuperscript{-1} at the beginning, to 2.5 ± 0.79 leaves
shoot\textsuperscript{-1} by the end of the experiment. Underlying this
pattern of stability, however, were marked differences in leaf
production and turnover between treatments. New leaf production
displayed a positive relationship with maximum temperature and
significant differences between source populations, characterised by a
higher rate of new leaf production in treatments sourced from cool-edge
population, than centre or warm-edge populations (Table S2, Fig. 4A). In
relation to heat stress anomalies, new leaf production displayed a
strong positive relationship and no difference between source
populations (Table S2, Fig. 4B), suggesting local thermal adaptation or
acclimatization within source populations. \emph{P. oceanica} growth
rates (shoot elongation) and total leaf length per shoot displayed
non-significant relationships with maximum temperatures and heat stress
anomalies (p\textgreater{} 0.05, Fig. 4D, E, G, H). Rather, source
populations were the best predictor of growth and total leaf length over
the course of the experiment (Table S2). At the end of the thermal
stress period, growth rates were significantly different among
transplant treatments (ANOVA, F\textsubscript{(5,112)} = 3.8, p = 0.003,
Fig. 4F). Cool-cool and warm-warm treatments displayed the highest
growth rates despite experiencing up to 6ºC difference in temperature,
and centre-centre and centre-warm treatments displayed significantly
lower growth rates (Fig. 4F). By the end of the recovery period, growth
rates had increased in all treatments except centre-warm shoots which
remained significantly lower than cool-edge (Tukeys-HSD, p = 0.005) and
warm-edge controls (Tukeys-HSD, p = 0.018). For total leaf length,
shoots recorded their shortest length at the end of the heat stress
period and recovered to be significantly longer by end of the recovery
period (Fig. 4I). Moreover, at the end of the thermal stress period,
cool-edge procedural controls were significantly longer than all other
treatments and warm-edge procedural controls were significantly longer
than cool-warm or central-warm treatments (Table S2). These relative
differences held to the end of the recovery period (Fig. 4I).

Growth rates in natural meadows displayed similar patterns to
transplants, albeit with higher rates over the same heat stress and
recovery periods (Fig. S2). During the thermal stress period, growth
rates were significantly higher in cool-edge meadows than warm-edge
meadows and both cool-edge and warm-edge meadows displayed significantly
higher growth rates than central meadows (ANOVA,
F\textsubscript{(2,66)}= 17.12, p\textless{}0.001). New leaf production
was highest in central meadows and significantly lower in cool-edge and
warm-edge meadows during the thermal stress period (ANOVA,
F\textsubscript{(2,66)} = 13.56, p\textless{}0.001).

\textbf{Discussion}

Improving predictions for how species and communities will be affected
by climate change and extreme events is an urgent challenge. Here we
examined how populations of the functionally important
seagrass\emph{Posidonia oceanica} could be affected by heat stress, by
conducting a translocation experiment across its range. Cool-edge,
central and warm-edge populations displayed clear differences in thermal
performance when grown under common conditions, but patterns contrasted
with \emph{a priori} expectations based on their thermal geography.
Cool-edge populations performed equally well to warm-edge populations
under common conditions and significantly better than central
populations in terms of growth rates and survivorship. The strong
performance of cool-edge populations was remarkable given they
experienced thermal anomalies of \textgreater{}5ºC above long term
summer maxima, compared to 1.5ºC for central populations and
\textless{}1ºC for warm-edge populations. Moreover, cool-edge
transplants suffered and subsequently recovered from severe overgrazing
at the warmest site that removed \textgreater{}75\% of aboveground
biomass, highlighting remarkable resilience of cool-edge populations to
combined effects of thermal stress and herbivory -- two pervasive
impacts of climate change in marine ecosystems.

\emph{Effects of thermal stress on seagrass populations}

Patterns of thermal performance in \emph{P. oceanica} varied among
populations and deviated from expectations with respect to range
position, highlighting the challenges of extrapolating climate change
vulnerability predictions from local observations. Among cool, central
and warm-edge populations, only measures of new leaf production
supported models of local-adaptation/acclimatization and no metrics
consistently supported expectations of thermal-niche conservatism
between populations. The positive relationship between new-leaf
production and thermal anomalies indicates higher rates of leaf turnover
(i.e., shorter lifespan) under warmer conditions, potentially reflecting
a coping strategy for thermal stress within fragments. Cool-edge and
warm-edge populations did display similar rates of growth and survival
under warm-edge (29ºC) conditions, suggestive of niche conservatism,
however central populations exhibited comparatively poor performance, in
contrast to expectations. Moreover, patterns of poor thermal performance
by central populations at the warm edge were consistent with results
among centre-centre procedural controls, which experienced similar
absolute temperatures to warm-edge treatments. Growth patterns from the
natural meadows over the same period further support these findings,
showing lower growth rates and higher new-leaf production in
centre-populations relative to cool and warm-edge meadows following the
thermal stress period, reinforcing that experimental findings were not
an experimental artefact.

Unlike centre-populations, the high resilience of cool-edge \emph{P.
oceanica} to thermal stress was consistent with documented responses of
congeneric \emph{Posidonia} species to a natural heatwave of similar
magnitude. In 2011, Western Australia experienced the most intense
marine heatwave on record (Hobday \emph{et al.} 2016) and\emph{Posidonia
australis} displayed relatively minor impacts despite experiencing
thermal anomalies in excess of 5ºC and absolute temperatures reaching
30ºC (Strydom \emph{et al.} 2020). In contrast to\emph{P. australis,}
other seagrasses (i.e. \emph{Amphibolis antarctica} ) experienced severe
losses (Fraser \emph{et al.} 2014) and kelp forests (\emph{Ecklonia
radiata} ) suffered a 100 km range contraction (Wernberg \emph{et al.}
2016) in response to the 2011 heatwave. While the response by \emph{P.
oceanica} to large anomalies in the current study was relative to
cool-edge (c.f. warm-edge in \emph{P. australis} ) conditions, the
similarity between species to survive temperatures up to 30ºC highlights
a remarkable latent capacity to withstand thermal stress in these edge
populations. Latent tolerance to extreme events potentially reflects the
warm evolutionary origins of the genus and the high resilience
\emph{Posidonia} has displayed to climatic upheaval over the past 60 mya
(Por 2009; Bianchi \emph{et al.} 2012).

The idea of \emph{P} . \emph{oceanica} as resilient to heat stress may
appear at odds with the prevailing literature, which has demonstrated
(Diaz-Almela \emph{et al.} 2009; Marba \& Duarte 2010) and projected
(Jordà \emph{et al.} 2012; Chefaoui \emph{et al.} 2018) high sensitivity
of \emph{P. oceanica} to warming. However, the results of the current
study are consistent with these previous findings and help place them in
a broader geographical context. Central populations of \emph{P.
oceanica} were severely affected by thermal anomalies up to 1.5ºC above
long term summer maxima in the current study. Anomalies of this
magnitude are slightly cooler than those experienced during the 2003
Mediterranean heatwave, which caused a steep increase in shoot mortality
in central populations (Marba \& Duarte 2010). Our findings therefore
support previous evidence that \emph{P. oceanica} in central populations
are living close to their thermal limits but suggest that populations
elsewhere may indeed have greater tolerance of high temperatures. The
reason why central populations of \emph{P. oceanica} are more severely
affected by high temperatures than cool and warm-edge populations is
unclear. The genetic structure of \emph{P. oceanica} is characterised by
strong separation between the eastern and western basins, consistent
with a vicariance event during the last glacial maxima
(Arnaud\selectlanguage{english}-Haond\emph{et al.} 2007). Within the western basin, genetic
connectivity is relatively low for both the cool-edge and central
experimental collection locations, but both these locations are directly
connected to the same central node within the regional metapopulation
(Rozenfeld\emph{et al.} 2008). Central and cool-edge populations,
therefore, share more genetic and climatological similarities over
geological timescales than cool-edge and warm-edge populations (Chefaoui
\emph{et al.} 2017), whereas central and warm-edge populations share
more contemporary similarities in terms of selection pressure on upper
thermal limits. Despite this, central populations exhibit the greatest
sensitivity to temperature, across the distribution of \emph{P.
oceanica} forcing a re-evaluation of the vulnerability and management of
this species in response to climate change.

While field-based evidence of performance could not separate cool-edge
and warm-edge populations, the maximum realised temperature in our
experiment was 29.3oC. Therefore, we cannot rule out the possibility
that differences in performance between populations occur at higher
temperatures. Indeed, previous laboratory-based experiments have
suggested that warm-edge populations can survive temperature up to 36oC,
whereas cool-edge populations and central populations of \emph{P.
oceanica} display upper thermal limits around 30oC (Bennett \emph{et
al.} 2021)\emph{.} Such experimental evidence would suggest that
cool-edge populations were close to their upper threshold during the
current experiment whereas warm-edge populations retained a larger
thermal buffer. Previous laboratory experiments, however, did not
document the subsequent recovery of \emph{P. oceanica} following thermal
stress. While cool-warm and cool-centre treatments clearly showed signs
of thermal stress in the current study, our results suggest that they
have greater capacity to endure and recover from heat stress than
central-populations.

\emph{Selectivity and recovery of seagrass from herbivory}

In addition to direct impacts of temperature on \emph{P. oceanica,} high
rates of grazing by herbivores on cool-warm transplants highlights the
importance of indirect impacts of warming on ecosystem function. The
selective patterns of grazing at warm-edge sites were consistent with
the contrasting patterns of nutritional quality of transplants from
different locations. These results are consistent with previous studies
that have demonstrated higher rates of herbivory on nutrient enriched
plants (Prado \& Heck Jr 2011; Campbell \emph{et al.} 2018).
Interestingly, increased grazing on nutrient-rich seagrass was only
observed in warm-edge, not central sites, where low nutrient meadows
were also present. This pattern, in conjunction with feeding scars
suggests that warm-affiliated fishes and sea turtles were likely
responsible for the observed over-grazing. Tropical range extending
fishes such as the rabbitfish \emph{Siganus rivulatus} and \emph{S.
luridus} are the dominant herbivores in Cyprus (Santana-Garcon \emph{et
al.} personal communication) and while currently absent from Spain, are
expected to establish in the coming decades (Daniel \emph{et al.} 2009).
Herbivore pressure by native cool-affiliated species is already high in
the western Mediterranean (Pages \emph{et al.} 2012) and therefore the
cumulative effects of cool and warm water herbivores on nutrient rich
seagrass in the western Mediterranean needs to be assessed and monitored
carefully.

\emph{Conclusions}

The combination of warming temperatures, extreme events and species
redistributions has the potential to fundamentally change marine
ecosystems in the Anthropocene. Identifying general principles,
developing models and testing predictions about how these impacts will
unfold across species distributions is critical to anticipate and
adaptively manage ecosystems over coming decades. Here we demonstrate
that cool, central and warm-edge populations of seagrass differ in their
response to temperature, however these responses are not necessarily
linear or conform to \emph{a-priori} expectations. Moreover, the highly
selective effects of warm-affiliated herbivores on cool-edge
populations\emph{,} highlight how species redistributions and species
interactions can have unexpected and consequential effects on ecological
performance under climate change. Our findings make a strong case for
ongoing empirical comparisons of thermal performance for functionally
important species and highlight the complexity of inferring thermal
sensitivity from a single location. In doing so these findings provide
hope for greater resilience to warming in Mediterranean seagrasses than
previously recognised.

\textbf{Acknowledgements}

This study was funded by Fundacion BBVA (project Interbioclima). S.B.
received funding from the~European Union's Horizon 2020 research and
innovation programme under grant agreement No 659246, Juan de la Cierva
Formacion contract from the Spanish Ministry of Economy, Industry and
Competiveness and the Australian Research Council (DE200100900). N.M.
and S.B. received funding from the Spanish Ministry of Economy, Industry
and Competitiveness (MedShift, CGL2015-71809-P). J.B. received funding
from the Spanish Ministry of Science and Innovation, Juan de la Cierva
FJC18-035566-I. T.A. received funding from the Spanish Ministry of
Science and Innovation (Project: UMBRAL, CTM2017-86695-C3-1-R). Authors
are grateful to Anais Barrera and Marc Casas for their help during the
field and laboratory work in Catalunya.

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\textbf{Figures}

\textbf{Fig. 1} A) Map of source and recipient transplant sites across
the Mediterranean Sea. Colours indicate the source and direction of
transplants in Catalunya (blue), Mallorca (yellow and orange) and Cyprus
(red). B) Illustrated example of a transplant fragment, characterised by
an apical shoot and multiple vertical shoots. C) Daily temperature
anomalies experienced by \emph{P. oceanica} transplants, with respect to
the maximum annual temperature recorded in the site of origin (averaged
between 1981-2019). Dashed vertical line, highlights switch between
positive (stress) and negative (recovery) thermal anomalies. Maximum
annual temperatures recorded in Catalunya = 25oC, Mallorca = 29.1oC and
Cyprus = 29.3oC. Daily temperatures experienced by transplants relative
to their thermal regimes of origin in D) Catalunya, E) Mallorca and F)
Cyprus. Shaded areas represent daily temperatures above thermal regime
of origin. Different shades of blue, yellow, orange and red within D, E
and F, illustrate temperature profiles in the different recipient sites.
Grey line represents the long-term daily temperature, averaged between
1981-2019.

\textbf{Fig. 2} Patterns of shoot length and nutritional quality of
seagrasses over time in response to herbivory. A) Effects of a selective
feeding event in August 2018 in Cyprus where herbivores selectively
overgrazed cool-edge transplants (marked by asterisk). B)
Carbon:Nitrogen ratios of \emph{P. oceanica} leaves in the different
treatments throughout the course of the experiment. C:N ratios in
cool-cool plants (blue dashed line) were only measured in July 2018 and
April 2019. C) Differences in herbivory on cool-warm transplants,
compared to adjacent centre-warm and warm-warm transplants.

\textbf{Fig. 3} Patterns of transplant survivorship (A) and transplant
shoot survival (B) over the course of the experiment. Transplant
fragment survival is represented as the proportion of transplants
remaining at the end of each growth period, relative to the expected
number of transplants (i.e. initial conditions minus fragments removed
for growth measurements). Shoot survivorship is the change in the number
of shoots per transplant, relative to initial conditions. Blue points
and lines represent the fragments from Catalunya (cool edge), orange
lines and points represent fragments from Mallorca (central) and red
represents fragments from Cyprus (warm edge).

\textbf{Fig. 4} Thermal performance of \emph{Posidonia
oceanica}transplants across the Mediterranean Sea. Rows represent
performance metrics of new leaf production (top), shoot elongation
(middle) and total leaf length (bottom). Columns represent temperature
metrics of absolute maximum temperature (left), maximum thermal stress
anomaly (middle) and a comparison between heat stress period (July-Oct
2018, left, grey) and recovery periods (Nov -- June 2019, white, right).
Points represent mean +- SE rates per treatment. Shape of point
illustrates the time-period measurements were recorded in. Colour of
point represents the different transplant treatments and procedural
controls as per previous Figures.\selectlanguage{english}
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