Cladocora caespitosa (Linnaeus, 1767) is a colonial zooxanthellate scleractinian, endemic to the Mediterranean Sea, where it is considered the most important reef-building species (Kružić and Benković, 2008), with a carbonate production up to 1.1–1.7 Kg CaCO₃/m²/y (Peirano et al., 2001). This species can grow in a wide range of environmental conditions, from sheltered to exposed bottoms, and from the infralittoral plain, where it successfully competes with macroalgae, to the circalittoral one, where zooxanthellae can survive using indirect and diffuse light (Peirano et al., 1998; Peirano et al., 1999; Hoogenboom et al., 2010; Kersting et al., 2017a). Local hydrodynamic conditions, together with the typology and morphology of the substrate, influence colonies’ shape and growth patterns (Abel, 1959; Kersting and Linares, 2012). According to Peirano et al. (1998), C. caespitosa can be present as solitary colonies, beds (i.e., numerous and separate globose colonies) or banks (i.e., large frameworks of colonies reaching several square meters of extension), but also in mixed distributions, and it is also often found associated with rhodolits (Kersting et al., 2017b).

C. caespitosa is a long-living, slow growing species, with a very low recruitment rate and a high juvenile mortality (Kersting et al., 2014a). The longevity of this species (with a mean age of 50 years, up to 300 years for some colonies in the Columbretes Islands, Spain; Kersting and Linares, 2012), coupled with the production of a calcium carbonate skeleton, make C. caespitosa a suitable proxy to reconstruct past climate fluctuations on a decadal and centennial scale (Peirano and Kružić, 2004; Silenzi et al., 2005). At the same time, these same features make C. caespitosa extremely vulnerable to a variety of threats, like invasive algae (Kružić et al., 2008; Kersting et al., 2014b; Kersting et al., 2015), eutrophication (Kružić and Požar-Domac, 2007), pollution (El Kateb et al., 2016), trawling, dredging and related increase in sedimentation (Casado-Amezúa et al., 2015), limiting the distribution of extensive banks only to the Tunisian coast, the Gulf of Atalanta (Aegean Sea), the Ligurian Sea, the Columbretes Islands and Croatia (Laborel, 1961; Zibrowius, 1980; Peirano et al., 1998; Kersting and Linares, 2012). Despite these human pressures, the main reason for the ongoing regression of this coral over the last decades is the occurrence of mortality events, caused by the rapid increase of average sea temperatures and the increased frequency of marine heat waves (MHWs) (Rodolfo-Metalpa et al., 2005; Kružić et al., 2012; Jiménez et al., 2016; Garrabou et al., 2019; Garrabou et al., 2022). C. caespitosa is, in fact, well adapted to the seasonal temperature oscillations typical of the Mediterranean, but the extreme temperatures related to climate change (Lejeusne et al., 2010) affect its growth and health, causing bleaching and leading to colonies’ death (Jiménez et al., 2016). Considering its crucial role as ecosystem engineer (Jones et al., 1994) and its rapid decline and low recovery capability (Kersting et al., 2014a), C. caespitosa has been listed as Endangered in the IUCN Red List at Mediterranean regional level (Casado-Amezúa et al., 2015), and it is included in the Appendix II of the CITES, Annex II of the Protocol SPA/BD and in the Annex B of the EU Regulation Trade wild fauna and flora species (Otero et al., 2017). On the other hand, this coral has historically and recently shown unexpected rejuvenescence after warming-induced mortality events (i.e., the formation of a new polyp within an almost totally necrotic one) (Fedorowski, 1980; Kersting and Linares, 2019), suggesting an unexpected resilience and a high potential for successful translocation and restoration activities.

The endangered status of this species and its fundamental role in Mediterranean ecosystems advocate its inclusion in restoration actions, since conservation measures alone may be not sufficient (Swan et al., 2016; Fariñas-Franco et al., 2018). In fact, restoration, defined as “the process of assisting the recovery of an ecosystem that has been degraded, damaged, or destroyed” (SER, 2023), has been acknowledged as complementary to conservation and necessary to recover ecosystems’ functioning (Normile, 2009). Tropical scleractinians are among the marine organisms most involved in restoration projects, together with other invertebrates (mainly anthozoans and bivalves), vertebrates (fish and birds), algae and plants (mangroves, salt-marsh plants, seagrasses) (Rinkevich, 1995; Epstein et al., 2001; Rinkevich, 2005; Bayraktarov et al., 2016; Destro et al., 2018; Boström-Einarsson et al., 2020; Dehnert et al., 2022; Dehnert et al., 2023). One of the most implemented methods in tropical reef restoration is the translocation or transplantation of coral colonies (Kotb, 2016; Ferse et al., 2021), a technique that does not only benefit the translocated individuals, but also provides tangible conservation benefits at population, species or ecosystems levels (IUCN Species Survival Commission, 2012). These characteristics make restoration a powerful tool to preserve specimens of locally rare and threatened species when the primary habitat is destroyed (Butt et al., 2021).

To design proper management and restoration measures, it is crucial to know life history traits, the ecological optima, and the potential threats compromising the survival of the target species. In this sense, the hermatypic coral C. caespitosa was here tested as a model species: considering its symbiosis with zooxanthellae, the species could be more sensitive to the effect of climate warming, as already demonstrated in tropical corals (e.g., Sammarco and Strychar, 2009; Brown and Cossins, 2011; Jones and Berkelmans, 2012; Pettay et al., 2015; Lough et al., 2018). Neverthless, autotrophic symbionts may also buffer the presence of multiple stressors (e.g., thermal stress and low food availability), as hypothesized for the temperate sea whip Eunicella singularis (Esper, 1791) (Fava et al., 2010).

Here we present the first attempt to apply the translocation technique, with restoration purposes, on C. caespitosa. A long-term monitoring on the transplantation site was conducted to assess the survival and status of relocated colonies. To complement and support the relocation activity, additional fragments of C. caespitosa were reared in aquaria as a potential source of corallites and colonies to supply future restoration actions. Additionally, considering the increasing impact of climate change on Mediterranean habitats, a MHW was simulated to test the resistance of C. caespitosa fragments to rising temperatures.

Ecosystem restoration is essential to recover degraded habitats worldwide impacted by a synergy of anthropogenic and natural stressors (SER, 2023). In this sense, the United Nations has declared the Ecosystem Restoration Decade (2021-2030) to boost global-scale initiatives engaging not just the academic or scientific environments but also non-governmental organizations, managers and stakeholders (UNEP, 2021). However, the scarcity of knowledge on species ecology and their response to environmental and anthropogenic stressors might impair the success of any restoration attempt.

Cladocora caespitosa is a well-studied species that, over the last years, experienced mass mortality events at Mediterranean scale (Kersting et al., 2013; Garrabou et al., 2019; Garrabou et al., 2022). Even though it is one of the most important reef-building species of the basin (Kružić and Benković, 2008), no restoration effort has been made so far. In this study, several colonies of C. caespitosa were translocated to avoid their damage due to demolition activities, following their fate for 4 consecutive years. This study represents one of the few records of long-term monitoring, since most of the studies usually report less than 18 months of survey (Boström-Einarsson et al., 2020), and one of the few focusing on a temperate coral (Steinberg et al., 2020).

The survival rate recorded in this study after 19 months is higher (82.4%) than the average survival reported for tropical species (60-70%, Boström-Einarsson et al., 2020). However, in the long-term, we observed a decrease of survival percentage, reaching a value of 55.9%, still in line with previous short-term studies (Boström-Einarsson et al., 2020). This decline might be related with the extremely high temperatures recorded during summer 2022, which caused stress conditions and mortality events also in other cnidarians, such as the red coral Corallium rubrum (Linnaeus, 1758) (Roveta et al., 2023), and other gorgonian species, such as Paramuricea clavata (Risso, 1826) a and Eunicella cavolini (Kock, 1887) (Gómez-Gras et al., 2022). Despite the thermal anomalies experienced in the study area (Table S1, S2), more than half of the colonies survived, resisting these extreme conditions. This resilience might be related to the peculiar characteristics of the recipient site, shaded during most of the day, thus, reducing the direct exposure of the coral to sunlight, potentially limiting the damages of an excessive irradiance (Hoogenboom et al., 2006). As a matter of fact, a high irradiance coupled to an increased temperature might be detrimental for the fitness of the coral, affecting the photosynthetic rates and inducing bleaching (Ban et al., 2014). For future application of this translocation method, an accurate measurement of the irradiance should be performed both in the donor and recipient sites to ensure the survival of the transplanted colonies. Furthermore, it is recommended to apply respirometry and pulse amplitude modulation fluorimetry to monitor the photosynthetic potential of the transplanted colonies over time.

The percentage of detached colonies was lower (14.7%) than reported for tropical massive species (e.g., Diploastrea heliopora (Lamark, 1816)) after 6 months of transplantation (40%, Bongiorni et al., 2011), even though it increased during the following years of monitoring. Despite the selection of a sheltered recipient site, this higher detachment might be related to the abundant mucus production typical of the species (Herndl and Velimirov, 1986; Peirano et al., 1999), which has been demonstrated to be a factor favoring the observed phenomenon (Bongiorni et al., 2011). Moreover, the lack of a frequent cleaning and maintenance of the transplanted colonies could have been an additional factor promoting detachment (Bongiorni et al., 2011). Since the beginning, bleaching was observed in a single colony (2.9%), right after the summer of 2018. Nonetheless, this number increased at an alarming rate (14.7%) during our last monitoring in October 2022 (T4). The increase in duration of episodes of anomalously high SSTs in the western Mediterranean, as well as in the recipient site (up to 28.5°C; Table S2), might have induced a stronger stress in the translocated colonies. Additionally, a significant reduction of the total number of healthy corallites per colony was observed in T4, with an overall decreasing rate (-2.5 ± 0.4 corallites/month) higher than the increasing rate observed in the time interval between T0 and T3 (1.9 ± 0.7 corallites/month). Bleaching as well as tissue necrosis were the main causes of this trend, possibly related to the high temperatures (over 28°C) experienced in our site for more than 45 days (Table S2), as already demonstrated in controlled conditions by Rodolfo-Metalpa et al. (2006). A general reduction in colony size with the increase of temperature seems to be a common condition along the Italian coasts, being also observed in the Ligurian Sea since the ‘90s (Rodolfo-Metalpa et al., 2005; Azzola et al., 2022).

Compared to the field observations, a similar value of new corallites (1.3 ± 0.3 corallites/months) was obtained in the growth rate experiment, supporting the feasibility of aquaria as a potential ex-situ nursery also for this species (Ng et al., 2012; Merck et al., 2022). Budding of new polyps was observed mostly on fragments with multiple corallites both under light and dark, but with a lower rate in the last condition. Asexual reproduction through budding requires high energy expenditure and it is normally observed under favorable conditions and high availability of zooplankton (Rodolfo-Metalpa et al., 2008). Although fragments in dark conditions received the same amount of food, the scarce autotrophic contribution might explain this difference in the number of new polyps. This positive effect of light is further supported by the fact that in smaller fragments the production of new polyps was observed only under light conditions. Nonetheless, this imbalance between small and big fragments (both exposed to light) might be related: (i) on the one hand, to the experimental set-up, since following Leal et al. (2014), the quantity of food given was estimated per corallite, thus making the quantity of food available for 1-corallite fragments more diluted than for the ones with more corallites; (ii) on the other hand, to the fact that a fragment with more corallites presents a higher tentacle density, increasing the ‘encounter zone’ (i.e., by the extension of the polyps’ tentacles, in which prey are likely to be caught) (Madin, 1988), which likely renders the total energy available to be relatively higher in bigger colonies (Sakai et al., 2019).

An increase in the number of corallites was also recorded in some fragments exposed to the simulated MHW (both in light and dark conditions), suggesting temperature as a triggering factor for asexual reproduction through budding, a condition also common in other Mediterranean cnidarians (Di Camillo et al., 2021) as well as in tropical scleractinians (e.g., Pocillopora damicornis (Linnaeus, 1758), Jiang et al., 2017). Nonetheless, all experimental fragments showed clear signs of stress, displaying tissue retraction as well as a reduction in the tentacles size and reactivity under stimuli (Rodolfo-Metalpa et al., 2006; Bellis and Denver, 2017) (Figure 5).

The variation of the color of the tissue from E6-5 to E2 in the colonies kept in the dark was related to the lower content of zooxanthellae, always being lower in dark than in light conditions (Table 2) (Hill and Ralph, 2007), a result observed in both experiments. Surprisingly, in heat-treated fragments exposed to light, an increase in zooxanthellae density was recorded after 5 days compared to the control. Conversely, Rodolfo-Metalpa et al. (2006) encountered a reduction in their number after 34 days in nubbins exposed to 28°C. Considering the duration of our experiment, this observation might be a short-term response of the zooxanthellae to high temperatures, potentially decreasing in density after a mid-term exposure. Regardless of the exposure to light, our fragments presented at least 1 degree of magnitude higher concentrations of zooxanthellae compared not only to tropical scleractinians (Stimson et al., 2002), but also to other studies on C. caespitosa collected in different Mediterranean locations (Rodolfo-Metalpa et al., 2006; Hoogenboom et al., 2010; Rodolfo-Metalpa et al., 2010). This higher concentration is considered an adaptation of temperate corals to more shaded and cooler environments but might also represent a species adaptation to particular local conditions (Lasker, 2003; Eagleson et al., 2023). Future studies focused on the evaluation of the zooxanthellae content of Bagnoli transplanted colonies (constantly thriving in a shaded environment) will help to confirm this hypothesis.

The percentage of mitotic index recorded for C. caespitosa was comparable to the one recorded for other massive scleractinians, such as Siderastrea stellata Verril, 1878 from Brazil (Costa et al., 2013), and its increase with temperature was in line with Caroselli et al. (2015). While no difference in the mitotic index was recorded in the fragments exposed to light and higher temperature, the experimental fragments kept in the dark showed an increment of this parameter. However, this was not followed by an increase in zooxanthellae density (Table 2), a situation potentially explained by the capability of this coral to rely more on heterotrophic feeding, digesting its symbionts under thermal stress, as observed in various Acropora spp. (Fujise et al., 2014).

Overall, our results obtained in the field highlight the high capability of C. caespitosa to adapt to adverse environmental conditions, thus rendering this species a good candidate for future restoration actions. The aquaria experiments gave a first insight into the potential adaptability of this species at individual level. A higher number of genotypes is needed to assess intraspecific variability in the response to thermal stress and light exposure.

Despite the crucial roles played by this species as ecosystem engineer and in the Mediterranean shallow waters carbon budget (Rodolfo-Metalpa et al., 2010) and being protected by the European and international legislations, restoration efforts are still lacking. In this sense, our study represents the first insight into the species-based siting selection, highlighting for C. caespitosa the importance of considering the “exposure to light” criteria and the potential shading of colonies to ensure the survival and, later, the success of a restoration action of zooxanthellate species. Finally, considering the similar increase in corallites per month obtained in the field and in the laboratory, our study points out the suitability of aquaria rearing as a potential tool to supply coral fragments for restoration purposes.

The original contributions presented in the study are included in the article/Supplementary Material. Further inquiries can be directed to the corresponding author.

CC conceived the idea. CC and FT conducted the field activities. BC, MC, CR and SP performed the aquaria experiments. MC, CGC, TM, TPM and CR processed the data and run the statistical analyses. MC, CR, FT and LV wrote the original draft. All authors contributed to editing and reviewing the manuscript. All authors contributed to the article and approved the submitted version.