This species is comprised of a group of closely related yet clearly distinct genetic lineages representing at least eight newly defined species. The improved taxonomic scheme has helped to explain controversies over differences in life history strategies assumed for the species, such as reproductive mode or timing, however, evidence suggests that additional cryptic species remain to be identified (Schmidt-Roach et al. 2014).

This species is widespread and common. Global level, species-specific population data are limited; however, coral reefs have declined globally and are expected to continue rapidly declining due to increasing severe bleaching conditions under temperature stress caused by climate change as well as a variety of other threats. Our species-specific vulnerability traits analysis indicates this species is moderately susceptible to major threats related to coral reef degradation (e.g., disease and bleaching). We applied two analytical approaches involving two different global coral datasets and the species’ distribution map as proxies to infer population decline. Based on global coral cover monitoring data, this species experienced a suspected decline of less than 25% over the past three generations, or since 1989. Based on the projected onset of annual severe bleaching (ASB) conditions via both SSP2-4.5 and SSP5-8.5 scenarios of global climate model data, in combination with the species’ depth range, distribution and bleaching vulnerability, this species is suspected to decline by at least 61% over the next three generations, or by 2050. Since the species qualifies for a higher category under the projected decline, we therefore list it as Endangered A3ce. The change in status from the previous assessment reflects updated declines calculated from improved data on modeled coral cover loss and projected date of annual severe bleaching, along with improved knowledge of species traits.

This species occurs in the Red Sea, the Gulf of Aden, Pakistan (Ali et al. 2020), the southwest, northwest and central Indian Ocean, Persian Gulf, the central Indo-Pacific, tropical and subtropical Australia (Jones 2016), southern Japan, the South China Sea, Guam (Maynard et al. 2017), the oceanic west Pacific, the central Pacific, Austral Islands (Mayfield et al. 2015), the Hawaiian Islands, Johnston Atoll, Easter Island and the Eastern Pacific (Glynn 2003, Glynn et al. 2007). It is also confirmed from northern Vietnam, southern and southeastern China and Taiwan (Huang et al. 2015).

In the Eastern Tropical Pacific, the species has been reported from: Mexico: Baja California Sur, Nayarit, Jalisco, Colima, Michoacán, Guerrero and Oaxaca (Reyes-Bonilla 1998, Reyes-Bonilla and Lopez-Perez 1998, Reyes-Bonilla 2003, Calderon-Aguilar 2005, Reyes-Bonilla et al. 2005, Glynn et al. 2007); El Salvador: Del Amor beach, Los Cóbanos (Reyes-Bonilla and Barraza 2003); Costa Rica: Islas Murcielago Archipelago, Bahia Culebra, Bahia Brasilito, Samara, Cabo Blanco, Punta Leona, Herradura, Manuel Antonio, Punta Uvita, Peninsula de Osa, Golfo Dulce, Caño Island, and Cocos Island (Cortés and Guzmán 1998, Glynn et al. 2007, Guzmán and Cortés 2007); Panama: throughout the Gulfs of Chiriqui and Panama (Maté 2003, Glynn 1997, Glynn et al. 2007); Colombia: Gorgona Island, Ensenada de Utría and Tebada (Zapata and Vargas-Ángel 2003, Glynn et al. 2007); Ecuador: Salango Island, Los Frailes, Sucre Island and La Plata Island, and throughout the Galápagos Archipelago (Glynn et al. 2001, Glynn 2003, Hickman et al. 2005, Glynn et al. 2007).

The depth range is 1-110 m, but the species primarily occurs from 1-30 m (Muir et al. 2018, Muir and Pichon 2019, Turak and DeVantier 2019, L. DeVantier pers. comm. 2024).

This species is common (Veron et al. 2016, DeVantier and Turak 2017). Early life stages can regulate the population dynamics of this species (Bramanti et al. 2015). Populations of this species are highly differentiated both within the Western Indian Ocean and the tropical Southwestern Pacific, suggesting restricted gene flow at different spatial scales (marine province, ecoregions, islands/regions), leading to diverging lineages (Gélin et al. 2018). Recent genetic studies suggest that this species is relatively rare in the tropical west Pacific, but may be more common in the higher-latitude, colder water environments in the region (Torres and Ravago-Gotanco 2018). Gene flow is limited in Tropical Eastern Pacific populations, particularly among regions, but even over metre scales within populations (Combosch and Vollmer 2011).

The relative abundance of Pocillopora damicornis in the Eastern Tropical Pacific region has been categorized as follows:

Abundant: from Nayarit to Oaxaca, Mexico (Reyes-Bonilla 2003); Panama, Costa Rica and Colombia (Cortes and Guzmán, pers. comm.; Glynn and Ault 2000). Highly recovering in Caño Island, Costa Rica (Guzmán and Cortes 2001) and in Panama (Guzmán et al. in prep.).

Common: Gulf of California (Glynn and Ault 2000, Reyes-Bonilla 2003).

Uncommon: Revillagigedo Islands, Mexico (Reyes-Bonilla 2003), and Ecuador including the Galápagos Archipelago (Glynn and Ault 2000, Glynn 2003).

Rare: Del Amor beach, El Salvador (Reyes-Bonilla and Barraza 2003).

Glynn et al. (1988) report high rates of pocilloporid coral mortality across the eastern Pacific following the 1982/83 El Niño, ranging from 51% at Caño Island to 76-85% in Panama and 97-100% in the Galápagos Islands (Glynn et al. 1988). However, following this high coral mortality, numerous pocilloporid recruits have been observed at some sites in Costa Rica, Panama and the Galápagos Islands (Glynn unpublished data in Glynn et al. 1991). According to Cortés and H. Guzmán (pers. comm. 2008), Pocillopora damicornis seems to be very abundant and recovering in Panama, Costa Rica and Colombia, despite severe coral mortality after the ENSO events (1982-83 and 1997-98).

In the Galápagos Islands, pocilloporid communities were well developed off northeastern San Cristobal, Espanola and Floreana Island until the 1980s (Glynn 1994; 2003), but disappeared following the 1982-83 ENSO event, with minimal coral recruitment since (Glynn 2003). No live Pocillopora damicornis has been seen in the once coral-filled lava rock pools at Punta Espinosa, Fernandina, Galápagos Islands since 1983 (Glynn 2003).

In mainland Ecuador at Sucre Island, Machalilla, the predominant frame-building species before 1983 were Pocillopora elegans and Pocillopora damicornis, with the two species generating a rigid framework that covered over 1 ha of bottom (Glynn 2003). However, by 1991 the reef had declined to dispersed colonies of P. damicornis and P. elegans (Glynn 2003).

Species-specific, global level population information is limited. However, coral reefs are experiencing severe global level declines due to increasing water temperatures caused by climate change (Hoegh-Guldberg et al. 2017, Hughes et al. 2018, Donovan et al. 2021). For the purposes of this Red List assessment, we used species-specific vulnerability traits and two analytical approaches based on two global coral datasets to infer past (GCRMN 2021) and future (UNEP 2020) population trends.     

Approach 1: Future population trend
The projected onset of annual severe bleaching (ASB) was applied as a proxy to estimate global level population decline. ASB represents the date at which a coral reef will likely experience severe bleaching conditions annually, and beyond which the species will experience a greater than 80% decline as it is not expected to recover (van Hooidonk et al. 2014). ASB is defined as at least eight Degree Heating Weeks (DHW) occurring over a three-month period within a year, and where a DHW occurs when the sea surface temperature is at least 1°C above the maximum monthly mean (van Hooidonk et al. 2014; 2015). We defined the onset of ASB as corresponding to 80% or more decline, however, this is conservative as other studies have found that coral populations may experience near complete mortality and are unlikely to recover with just two incidences of ASB per decade (Obura et al. 2022).

To calculate ASB for each species we applied spatial data made publicly available via a United Nations Environment Programme report (UNEP 2020) that used the 2019 IPCC CMIP6 global climate models to estimate the projected onset of ASB for the years 2015-2100 on a 27 km x 27 km grid according to the 2018 WCMC-UNEP global coral reef distribution map, which has a resolution to 30 m depth. These data are available via two scenarios of Shared Socioeconomic Pathways (SSP), with SSP5-8.5 representing current global emissions and SSP2-4.5 representing a future reduction in emissions (UNEP 2020). We applied SSP5-8.5 since it follows the precautionary approach recommended by the IUCN Red List methodology and SSP2-4.5 since it represents a more moderate climate change scenario that better tracks current policy projections (Roelfsema et al. 2020, Obura et al. 2022). To acknowledge varying levels of coral adaptation to thermal stress, both of these spatial data layers are available for all quarter degree intervals between 0° and 2°C (UNEP 2020); however, coral adaptation in general is little understood and varies by species and locality (Bay et al. 2017, Matz et al. 2020, Logan et al. 2021). To account for adaptation, we calculated two estimates of ASB onset for both the SSP5-8.5 and the SSP2-4.5, where the first estimate assumes the species has no level of adaptation (0°C) and the second assumes a capacity for 1°C of adaptation. We clipped each of these four UNEP (2020) spatial data layers to the species’ distribution and calculated the average year of ASB onset across all overlapping grid cells. 

Based on this spatial analysis, the onset of ASB across this species’ range is projected to occur on average by the year 2035 for SSP5-8.5 and by 2039 for SSP2-4.5 assuming no level of adaptation and by the year 2062 for SSP5-8.5 and by 2070 for SSP2-4.5 assuming 1°C of adaptation. For species where the onset of ASB occurs within 3-generation lengths, the 3-generation reduction is calculated as 80% multiplied by two proportions: (i) the proportion of the species' depth range that is in 0-30 m range, and (ii) for widespread species, the proportion of cells within the species' range that are expected to experience ASB under SSP2-4.5 before 2050 (three generation lengths). We inferred that the uncertainty associated with the estimate of population decline based on no level of adaptation is lower given this species is primarily restricted to depths shallower than 30 m and is highly susceptible to bleaching. For widespread species, the final estimate of decline was further adjusted by excluding the proportion of cells within its range that were expected to experience ASB under SSP2-4.5 after 2050 (three generation lengths), in order to account for the potential resilience of species to the asynchronous variability of bleaching events that occur across the Indo-Pacific. The relative vulnerability to bleaching (i.e., highly susceptible, moderately susceptible, or more resilient) is primarily based on scientific species expert knowledge. The application of the species’ depth range as a vulnerability factor is based on the assumption that a coral species with shallow depth preferences is more frequently exposed to extreme temperatures and might decline at a faster rate in some places than species that also occur in deeper, cooler waters (Riegl and Piller 2003), although this is not always the case (e.g., Smith et al. 2016, Frade et al. 2018). Ocean acidification, which is measured by aragonite saturation, is also considered a major threat to corals due to the impacts of climate change, however, the impacts are expected to be more severe in cooler and/or deeper waters (Couce et al. 2013, van Hooidonk et al. 2014, Hoegh-Guldberg et al. 2017). Although the exact threshold of aragonite saturation that is expected to cause significant decline is not well-known, in the Pacific, changes in aragonite saturation are expected to be most severe in high-latitude reefs (van Hooidonk et al. 2014). Therefore, this species is suspected to experience a projected global level decline of at least 61% by the year 2050, or three generations in the future, regardless of the SSP2-4.5 or SSP5-8.5 scenario. 

Approach 2: Past population trend
Coral reef monitoring data were also applied as a proxy to estimate global level population decline. The Global Coral Reef Monitoring Network (GCRMN) compiled data related to the status and trends of coral reefs in 10 regions from 1978-2019 via the scientific monitoring observations of more than 300 network members located throughout the world. We applied the publicly available data on estimations of the percent of live hard coral cover loss at the 20%, 50% and 80% confidence intervals in the 37 subregions of the Indo-Pacific (GCRMN 2021) to estimate species population decline over the past three generations (1989-2019). The proportion of the species’ range that overlapped with each of the subregions was estimated using the Red List distribution map. The sum of the proportion of the subregional species distribution multiplied by the percent of coral cover loss in each subregion was then used to calculate the 20%, 50% and 80% estimates of coral loss across this species’ range.

To inform the choice of the best (i.e., lowest level of uncertainty) out of the three percentile declines, we considered 11 species-specific traits related to vulnerability to coral cover loss. Given this species’ depth range is 1-110 m and is predominately found at depths greater than 10 m, generalized abundance is considered common, overall population is not restricted or highly fragmented, does not occur off-reef, is highly susceptible to disease, does recover well from bleaching or disease, has a high susceptibility to crown-of-thorns starfish, is highly susceptible to bleaching, has a relatively lower susceptibility to the impacts of ocean acidification (Kornder et al. 2018), did not have >10,000 pieces exported annually in the aquarium trade between 2010-2019, it is overall suspected to be moderately susceptible to threats related to reef degradation. Therefore, past decline was inferred from the 50% percentile of estimated coral cover loss, resulting in a suspected global level decline of less than 25% since 1989, or over the past three generations.

This species occurs in all shallow water habitats from exposed reef fronts to mangrove swamps and wharf piles. It is found in mono-specific stands or multi-species reefs throughout its range from near the surface to a maximum depth of 20 m. It is commonly found from 1-15 m, rarely 18-20 m, in the South China Sea and Gulf of Siam (Titlyanov and Titlyanova 2002). This species is considered to be a main reef-framework builder and is found from 0.5-6 m of Panama (Sheppard 1982). This species is relatively tolerant of sedimentation and low salinity as long as there is adequate water motion. Colonies reproduce by fragmentation and by sexual reproduction (broadcast spawning) (Hodgson 1998). There is marked variability in reproductive strategies and genetic structure of this species throughout its geographic range (Adjeroud et al. 2014). Active feeding in this species plays a fundamental role in its metabolic dynamics and its susceptibility to bleaching (Lyndby et al. 2019; 2020).

In the Eastern Tropical Pacific, the species has not been reported from mangrove environments (Cortés and H. Guzmán pers. comm. 2008) but it is one of the major reef building species, forming intermeshing compact frameworks that can attain 2-3 m in relief (Glynn 2001). Amongst the reef building corals in the Eastern Tropical Pacific, pocilloporid species have the fastest growth rates (Guzmán and Cortés 1993). 

Pocilloporid corals, presumably including this species, are generally amongst the strongest coral competitors with relatively high rates of calcification (Glynn 2001). However, coral species exhibiting high rates of calcification usually have relatively high mortality rates (Glynn 2000). Pocilloporid corals also usually predominate at shallow depths (1-15m). Amongst the reef building corals in the Eastern Tropical Pacific region, pocilloporid species have the highest growth rates (Guzmán and Cortes 1993). They are the principal framework builders on Panamanian reefs (Glynn 2002).

This species is a broadcast spawner (Glynn et al. 1991) with the capacity to function as a simultaneous hermaphrodite (Glynn et al. 1991). According to Glynn et al. (1991), larval settlement in the Galápagos Islands presumably has been the predominant mode of recruitment, and the only observed form of recruitment in areas that experienced high mortality (97-100%) in 1983. Asexual reproduction by fragmentation has been reported as an important mechanism for reef recovery in Panama (Glynn et al. 1991). This species, like other pocilloporid species in the eastern Pacific, has low rates of recruitment (Glynn et al. 1991). Histological evidence indicates that spawning is likely to occur during a few days around the new moon (Glynn et al. 1991). Reproductive activity in the eastern Pacific its related to local thermal regimes, with a generally higher incidence of gravid corals at sites with stable, warm water conditions, or during warming periods in areas that experience significant seasonal variation (Glynn et al. 1991). Glynn et al. (1991) conclude that moderate El Nino warming can stimulate gametogenesis in Galápagos pocilloporid corals.

Short periods of subaerial exposure during extreme low tides are not lethal to this species, but negatively affects sexual reproduction, which might have deleterious effects at the population level. The periodic occurrence of extreme low tides in the tropical eastern Pacific may be one factor responsible for the high rate of asexual reproduction (e.g., fragmentation) in pocilloporid corals of this region (Castrillón-Cifuentes et al. 2017).

Pocillopora species are preyed on by at least nine groups of consumers. These vary in their consumption patterns, but include:

a) Species that bite off colony branch-tips: pufferfishes (Arothron), parrotfishes (Scaridae), filefishes (Monacanthidae) (Glynn 2002). 

b) Species that scrape skeletal surface: hermit crabs (Trizopagurus, Aniculus, and Calcinus) (Glynn 2002).

c) Species that remove tissues but leave the skeleton intact: gastropods (Jenneria pustulata and Quoyula sp. (Glynn 2002)), butterflyfishes, angelfishes, damselfish (Stegastes acapulcoensis), and Acanthaster planci (Glynn 2002).

d) Species that abrade tissues and skeleton: Eucidaris galapagensis (Glynn 2001).

Jenneria and Acanthaster can kill whole, relatively large (approx. 30 cm in diameter) colonies of Pocillopora (Glynn 2002). Pocilloporid species can have crab (Trapezia sp.) and alpheid shrimp as mutualistic symbionts that protect the coral from the attack of the crown-of-thorns sea star (Glynn 2001).

The age at first maturity of most reef-building corals is typically three to eight years (Wallace 1999). Based on this, we infer that the average age of mature individuals of this species is greater than eight years. Based on average sizes and growth rates, we also infer that the average length of one generation is 10 years. Longevity is not known, but is likely to be greater than 10 years. Therefore, any population decline rates estimated for the purposes of this Red List assessment are measured over a time period of 30 years.

The collection of this species for the aquarium trade may lead to overharvest and localised reductions in abundance, especially for populations of naturally rare species (Bruckner and Borneman 2006). However, the wild collection of corals is highly selective and considered low impact in the long-term relative to other activities such as coral mining and dynamite fishing (Green and Shirley 1999, Pratchett et al. 2020). 

This species exhibited variable bleaching (0-50%) and low mortality in the 1998 bleaching event in Palau (Bruno et al. 2001). The larval stage of this species is particularly sensitive to increases in sea surface temperature, as to ocean acidification (Putnam et al. 2013, Jiang et al. 2019); there are also trans-generational effects to subsequent offspring (Putnam and Gates 2015, Smith 2019). However, individuals of this species in highly variable environments may have increased resilience to some aspects of climate change (Jiang et al. 2020). Larvae have the capacity to reduce the symbiont cell density without a harmful effect on their survivorship under thermal stress (Haryanti et al. 2015). This species has been found susceptible to bleaching on reefs off Guam (Maynard et al. 2017). Thermal stress causes polyp bailout in this species (Fordyce et al. 2017). Excess symbionts increase the susceptibility of this species to bleaching (Cunning and Baker 2012), while growth of this species is lowered if the symbiont species changes from clade C to the more thermotolerant clade D (Cunning et al. 2015). Alteration of the microbiome of this species may lessen the effects of bleaching (Rosado et al. 2019). Colony morphology may be linked to bleaching susceptibility in different morphs of the Pocillopora damicornis species complex (Smith et al. 2017). Increased temperature can mitigate the adverse effects of acidification on the calcification of juveniles of this species, but at a substantial cost to asexual budding (Jiang et al. 2018). Heat stress events reduce the diversity, quantity and functional potential of biogenic volatile organic compounds emitted by this species, further compromising the healthy functioning of coral reef ecosystems (Lawson et al. 2021). Thermal stress reduces the resilience of this species to ocean acidification by impairing control over calcifying fluid chemistry (Guillermic et al. 2021).

In general, species of this genus are highly susceptible to bleaching (McClanahan et al. 2007, Hughes et al. 2017, Khen et al. 2023), but also have high recovery potential (Darling et al. 2013). Pocilloporid species as well as other major reef building corals within the Eastern Tropical Pacific region (Porites, Pavona, Gardinoseris) catastrophically declined in the Galápagos Archipelago and Cocos Island after 1983. Recovery observed since that time was in large part nullified by the 1997-98 ENSO event (Glynn 2000). According to Glynn et al.(1988), pocilloporid coral mortality in the eastern Pacific was high, ranging from 51% at Caño Island to 76-85% in Panama and 97-100% in the Galápagos Islands (Glynn et al. 1988).

Glynn (1994) suggests that the sea urchin Eucidaris galapagensis (syn. E. thouarsii) provides important biotic control of pocilloporid reef development. This urchin is the most persistent corallivore in the Galapagos Islands, where it is often observed grazing on pocilloporid corals (Glynn 2001).

Overfishing is probably responsible for some ecological imbalance on coral reefs that could prolong recovery from other disturbances (Glynn 2001). Moreover, Edgar et al. (unpublished manuscript 2008) reported that over-exploitation of sea urchin predators (lobsters and fishes), along with ENSO, has a major effect in the condition and distribution of corals in the Galapagos Islands, by increasing the grazer and bioerosion pressure on corals.

Coral mortality associated with phytoplankton blooms has been reported from Caño Island, Costa Rica, and Uva Island, Panama, in 1985; where mortality of pocilloporid species (especially P. capitata and P. elegans) was in the order of 100% and 13% respectively at 3m depth (Guzmán et al. 1990).

According to Glynn (2001), pocilloporid coral harvesting is an important threat in the Eastern Tropical Pacific region, specially along the continental coast. This activity has virtually eliminated pocilloporid corals from Acapulco (Mexico), Bahia Culebra (Costa Rica), Taboga Island (Panama), and parts of the coast of Ecuador (Glynn 2001). Nevertheless, this activity is now largely excluded from Costa Rica and Panama (H. Guzmán pers. comm. 2008).

Bryant et al. (1998), based on four anthropogenic factors (coastal development; overexploitation and destructive fishing practice; inland pollution and erosion, and marine pollution), estimated a high threat to coral reefs along the coasts of Costa Rica, Panama and Colombia. High levels of siltation caused by accelerated coastal erosion have degraded coral reefs in Costa Rica, Colombia and Ecuador (Glynn 2001). Derelict fishing gear is a key driver in the fragmentation and loss of this species on the marginal rocky reefs of Ecuador (Figueroa-Pico et al. 2020). This species is very sensitive to microplastics (Tang et al. 2021). Tissue regeneration and coral health of this species were significantly reduced after a 96 h exposure to a crude oil / seawater mixture (May et al. 2020). Intensive fish farming produces effluents are detrimental to reproduction and growth of this species (Villanueva et al. 2006). This species is under threat from dredging in Eastern Australia (Jones et al. 2016). This species is sensitive to disease (Rodríguez-Villalobos and Reyes-Bonilla 2019). 

Other threats include: a) predation principally by Acanthaster and Jenneria (Glynn 2002, 1994, 2000), and b) harvesting for the curio trade, an activity that has virtually eliminated pocilloporid corals from Acapulco (Mexico), Bahia Culebra (Costa Rica), Taboga Island (Panama), and parts of the coast of Ecuador (Glynn 2001).

In general, the major threat to corals is global climate change, in particular, temperature extremes leading to bleaching and increased susceptibility to disease, increased severity of ENSO events and storms, and ocean acidification. The most recent, and first, multi-year, global bleaching event (spanning hundreds of kilometres or more) was from 2014 to 2017. Nearly 30% of reefs suffered mortality level-stress, more than 50% of affected reef areas were impacted at least twice, and some locations saw almost complete coral cover loss (Eakin et al. 2019). The average interval between bleaching events is now more than 50% less than before, preventing full reef recovery (Hughes et al. 2018). Bleaching events, leading to coral mortality, are predicted to become more frequent and severe. 

Coral disease has emerged as a serious threat to coral reefs worldwide with increases in numbers of diseases, coral species affected, and geographic extent (Ward et al. 2004, Sutherland et al. 2004, Sokolow et al. 2009). Outbreaks of coral diseases have damaged coral reefs worldwide with the most widespread, virulent, and longest running coral disease outbreak currently occurring on the Florida Reef Tract and throughout the Caribbean. The disease, stony coral tissue loss disease, has been ongoing since 2014 (Precht et al. 2016) and has devastated affected reefs along Florida (Walton et al. 2018, Williams et al. 2021) and throughout the Caribbean (Alvarez-Filip et al. 2019, Kramer et al. 2019). Numerous disease outbreaks have also occurred in the Indo-Pacific (Willis et al. 2004, Aeby et al. 2011; 2016), Indian Ocean (Raj et al. 2016) and Persian Gulf (Howells et al. 2020). Escalating anthropogenic stressors combined with the threats associated with global climate change of increases in coral disease, frequency and duration of coral bleaching and ocean acidification place coral reefs in the Indo-Pacific at high risk of collapse.

Localized threats to corals include fisheries, human development (industry, settlement, tourism, and transportation), changes in native species dynamics (competitors, predators, pathogens and parasites), invasive species (competitors, predators, pathogens and parasites), dynamite fishing, chemical fishing, pollution from agriculture and industry, domestic pollution, sedimentation, and human recreation and tourism activities (Crabbe 2019). The severity of these combined threats to the global population of each individual species is not known.

This species is moderately traded for aquaria, with 1,000-10,000 pieces being collected from the wild and exported annually between 2010-2020 (CITES 2021).

All stony corals (Order: Scleractinia) are listed on CITES Appendix II, and under Annex B of the European Union Wildlife Trade Regulations. Moreover, several countries (e.g., India, Israel, Jordan, Djibouti, Fiji and the Philippines) at various stages have banned either the trade or the export of CITES II listed species, which includes all stony corals. Other countries such as Indonesia, trade maricultured corals, with quotas, production limit and regulations in place to ensure the trade is sustainable. Having timely access to national-level trade data from CITES is valuable for monitoring trends for this species. Consideration of the suitability of species for aquaria should also be included as part of fisheries management, and population surveys should be carried out to monitor the effects of harvesting alongside other population trends.

This species has been successfully transplanted in Cousin Island Special Marine Reserve in the Seychelles (Montoya-Maya et al. 2016). Growth and survival rates of transplanted coral fragments of this species suggest that successful and low-cost restoration of tropical eastern Pacific coral reefs is feasible using this species (Lizcano-Sandoval et al. 2018).

Recommended measures for conserving this species include research in taxonomy, population, abundance and trends, ecology and habitat status, threats and resilience to threats, restoration action; identification, establishment and management of new protected areas; expansion of protected areas; recovery management; and disease, pathogen and parasite management. Artificial propagation and techniques such as cryo-preservation of gametes may become important for conserving coral biodiversity.
The Convention on Biological Diversity adopted an updated Strategic Plan for Biodiversity 2011-2020, which now includes Aichi Biodiversity Target 11, calling for 10% of coastal and marine areas to be conserved by 2020. And in 2016, the IUCN World Conservation Congress agreed upon a target of >30% global marine protection by 2030.

It is crucial that global warming is constrained well below 2°C (the goals of the Paris Agreement).