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 less than 25% over the next three generations, or by 2050. It is listed as Least Concern.

This species is found in the Red Sea (Terraneo et al. 2017), the Gulf of Aden, Indian Ocean, East Africa, the Persian Gulf, the central Indo-Pacific, Australia, Japan, the East China Sea, the oceanic West Pacific, the central and eastern Pacific, Kingman Atoll (Brainard et al. 2005), the Hawaiian Islands and Johnston Atoll. It has also been confirmed from eastern Thailand, Taiwan (Huang et al. 2015) and the Lakshadweep Islands (DeVantier and Turak 2017).

It has a widespread distribution in the Eastern Tropical Pacific (Alzate et al. 2014): México: Puerto Escondido to Huatulco (Reyes-Bonilla and López-Pérez 1998, Reyes-Bonilla 2003); Costa Rica: Peninsula de Osa, Isla del Caño and Cocos Island (Cortés and Guzmán 1998); Panamá: Contadora Island, Saboga Island, Mogo Mogo Island, Pacheca Island, Iguana Island, Unnamed Island (Secas Island) and Uva Island (Glynn and Mate 1997); Colombia (Reyes 2000); Ecuador: Salango Island, Los Frailes, Sucre Island, La Plata Island and Galapagos Archipelago (Glynn 2003).

The depth range is 2-70 m, but the species primarily occurs from 5-20 m (Muir and Pichon 2019, Turak and DeVantier 2019, Montgomery et al. 2019, 2021; L. DeVantier pers. comm. 2024).

This species is common (DeVantier and Turak 2017). According to Glynn et al. (1996), the slow rates of sexual recruitment of agariciid corals critically affect recovery of eastern Pacific coral communities that have experienced severe disturbances. However, according to Guzmán pers. comm. (2008) recruits of this species in the Gulfs of Chiriqui and Panama are common on rocky substrates. 

In terms of relative abundance within the Eastern Tropical Pacific, this species has been categorized as common: Caño Island, Costa Rica; uncommon: Cocos Island, Costa Rica; Panamá; Colombia (including Malpelo Island); rare: Mexico; mainland Costa Rica, and Ecuador (including the Galápagos Archipelago) (Glynn and Ault 2000, Guzmán and Cortés 2001, Reyes-Bonilla 2003).

According to Glynn (1997), prior to the 1982-83 ENSO event approximately 100 colonies were present in Costa Rica (Caño Island) and 26 colonies in Galápagos (central southern islands). After that ENSO event, only 10 colonies survived in Costa Rica, and two colonies in the central southern islands of the Galapagos Archipelago (Glynn 1997). In 1998, only one single colony was known in the Galapagos central southern islands (at Punta Estrada, Santa Cruz), with a combined total of approx. 550 cm² live tissue. This colony was totally bleached on 11 May 1998 (Glynn 1998), and has not been found since (Glynn 2001), although another colony in the region on Champion Island is now known (G. Edgar pers. comm. 2008).

According to Guzmán and Cortés (2001), the number of colonies in Costa Rica (including Cocos Island) and Panamá are increasing (Guzmán and Cortés 2001). In addition, Guzmán et al. (2004), reported it as a very common species in the Coiba Archipelago, found in 50 to 75% of the sites; and, according to H. Guzmán pers. comm. (2008). This species is common at the Gulf of Chiriqui (46% of the sites), as well as in Las Perlas Archipelago (35% of the sites). The species is also moderately common at Malpelo, Colombia (G. Edgar pers. comm. 2008).

In Galapagos, after a major population decline following 1982-83 ENSO event; a few isolated patches have persisted and are possibly increasing in number at Darwin and Wolf islands (Chiriboga, C. Hickman and G. Edgar pers. comm. 2008).

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 1°C of adaptation is lower given this species is primarily restricted to depths shallower than 30 m and is moderately 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 less than 25% by the year 2050, 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 2-70 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 low susceptibility to crown-of-thorns starfish, is moderately susceptible to bleaching, has a relatively higher 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 is found in shallow, tropical reef environments. Colonies are massive to encrusting, sometimes with laminar margins. It usually occurs on walls or under overhangs in clear water. It can be found in most reef habitats (Wood 1983). Sparse colonies of this species are found from 9-15 m in the South China Sea and Gulf of Siam (Titlyanov and Titlyanova 2002).

In the Eastern Tropical Pacific region, this species usually occurs on coral reefs and coral communities on rocky substrata at depths of 3-30 m (H. Guzmán, Chiriboga and G. Edgar pers. comm. 2008); it also occurs on walls or under overhangs in clear water. Colonies are among the predominant framework builders in the Eastern Tropical Pacific region (Glynn 2000). The reported average growth rate is 10.4 and 13.2 mm per year in Costa Rica and Panama, respectively (Glynn 1985, Guzmán and Cortés 1989). Reproduction is mainly sexual and the species is a broadcast spawner with an alternating periodic sequential hermaphrodite mode of reproduction (Glynn et al. 1996, Glynn et al. 2000). However, the species has a high proportion of mixed sexual patterns with gonochoric colonies usually predominating over hermaphrodites (Glynn et al. 1996). This pattern of sexuality could promote outbreeding during periods of high population abundance, but still allow sexual reproduction by selfing at times of severe population decline (Glynn et al. 1996). According to Glynn et al. (1996), the earliest age of sexual reproduction is 20 years, which may retard the recovery of the population, especially in areas that have experienced catastrophic mortality.  

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.

Colonies of this species seem not to be as sensitive to high temperatures as other species of corals (Fong and Glynn 1998). After El Niño 1982-83, approximately 92% of the southern Galápagos population died (Glynn 1994, 1997), as well as approximately 90% of the known colonies at Caño Island (Glynn 1997). According to Glynn (1988), El Niño disturbance could have perilous consequences for eastern Pacific small populations reef corals, including this species. This species has revealed no sexual recruitment where seed populations are absent or rare (Galápagos Islands), and only low recruitment (Panamá) in areas with colonies that survived the ENSO disturbance (Glynn 1996).

Crown-of-thorns starfish (COTS), Acanthaster planci, can influence the population structure and distribution of this species (Fong and Glynn 1998, Glynn 2001), but no outbreaks resulting in mass coral mortality are known (Glynn 2001). At Uva Island, Panamá, corals that recruit are subject to significant mortality by COTS predation (Glynn 1996). Colonies of this species are preferred prey for COTS, which may affect its population structure and distribution (Fong and Glynn 1998, Glynn 2001). COTS are found throughout the Pacific and Indian Oceans, and the Red Sea. These starfish are voracious predators of reef-building corals, with a preference for branching and tabular corals such as Acropora species. Populations of COTS have greatly increased since the 1970s and have been known to wipe out large areas of coral reef habitat. Increased breakouts of COTS has become a major threat to some species, and have contributed to the overall decline and reef destruction in the Indo-Pacific region. The effects of such an outbreak include the reduction of abundance and surface cover of living coral, reduction of species diversity and composition, and overall reduction in habitat area.

In general, the major threat to corals is global climate change, in particular, temperature extremes leading to bleaching, increased severity of storms, and ocean acidification (Hughes et al. 2017, Hoegh-Guldberg et al. 2018). As of 2020, nearly all coral reefs globally have been impacted by at least one major bleaching event, with many reefs having experienced multiple severe thermal stress events (Heron et al. 2016, Sully et al. 2019). During the most recent, and first, multi-year global bleaching event 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 (Blunden et al. 2018, Vargas-Angel et al. 2019, Eakin et al. 2019). While coral populations can be resilient to coral bleaching and bounce back (e.g., Diaz-Pulido et al. 2009, Pisapia et al. 2016), more frequent bleaching events in the future are expected to prevent full reef recovery and cause local extinctions of some species (van Hooidonk et al. 2016, Sheppard et al. 2020).

Localized threats human threats to corals include coastal construction (Nguyen et al. 2013), overfishing and destructive fishing (Mumby et al. 2007, Albert et al. 2012), pollution from agriculture and industry (Bruno et al. 2003), domestic pollution (Cunning et al. 2019), tourism activities (Lamb et al. 2014) and invasive species (Hume et al. 2014). Some of these threats impact corals directly, such as being physically disturbed and smothered with sediment during a construction project (Erftemeijer et al. 2012), while others operate indirectly via ecosystem processes and linkages between corals and other reef organisms. Macroalgae is major competitor with corals that reduces growth, causes disease, prevents new coral recruitment and can tip the entire ecosystem into a less diverse and less productive ‘algal-dominated’ reef (Hughes 1994, Bellwood et al. 2004). Macroalgal levels are controlled by both bottom-up provision of nutrients (Fabricius 2005), and top-down herbivory by parrotfish (Mumby et al. 2007), hence, while the immediate threat to the coral is be the algae, the ultimate threat may be sewage, fertilizer from agriculture or overfishing of herbivorous fish. The complex nature of the coral reef ecosystem means that while the immediate threat may be obvious (e.g. macroalgae, disease, crown-of-thorns outbreak), the ultimate threat is often less clear (Nyström et al. 2008, Anthony et al. 2015).

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. The severity of these combined threats to the global population of each individual species is not known.

This is not a target taxon in either the live or ornamental coral trade (Wood et al. 2012).

All stony corals are listed on CITES Appendix II. Parts of the species’ range overlaps with Marine Protected Areas.

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.