This species is likely multiple species (Lewis pers comm. 2023).

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 more resilient 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, the Gulf of Aden, the southwest and northwest Indian Ocean, the Persian Gulf, the central Indian Ocean, the central Indo-Pacific, tropical Australia, southern Japan, the South China Sea, the oceanic West Pacific, the central Pacific, the Hawaiian Islands, Johnston Atoll, Pitcairn (Friedlander et al. 2014), and the far Eastern Pacific. It has also been confirmed from eastern Thailand, southern and southeastern China, and Taiwan (Huang et al. 2015).

In the Eastern Tropical Pacific, it occurs in: Mexico: Jaltemba Island to Punta Mita, and from Isabel Island to Marias Islands (Nayarit); Revillagigedo Islands; Zihuatanejo to Acapulco (Guerrero); and Puerto Escondido to Huatulco (Oaxaca) (Reyes-Bonilla 2003, Reyes-Bonilla et al. 2005, Calderon-Aguilar 2005); Costa Rica: Guanacaste, Bahía Culebra, Punta Mala, Manuel Antonio, Peninsula de Osa, Golfo Dulce, Caño Island, and Cocos Island (Guzman and Cortes 1992, Cortés and Guzmán 1998); Panama: throughout the Gulfs of Chiriqui and Panama (Glynn 1997, Maté 2003, Guzmán et al. 2004); Colombia (Glynn and Ault 2000, Reyes 2000, Zapata and Vargas-Ángel 2003): Ensenada de Utría, Tebada; Gorgona Island and Malpelo Island; Ecuador: Salango Island, Los Frailes, Sucre Island and La Plata Island, and throughout the Galápagos Archipelago (except for Fernandina and the west side of Isabela) (Glynn et al. 2001, Glynn 2003). 

The depth range is 0-110 m, but the species primarily occurs from 1-40 m (L. DeVantier pers. comm. 2024).

This species is very common (Veron et al. 2016, DeVantier and Turak 2017, Dietzel et al. 2021).

In the Eastern Tropical Pacific region, the relative abundance is as follows:

It is considered abundant at Clipperton Atoll (Glynn et al. 1996, Glynn and Ault 2000). According to Glynn et al. (1996), this species is dominant on the lower reef slopes, where it often covers more than 90% of the substratum. In addition, the lower zone limit of this species is unknown; colony abundance at some sites shows no signs of decreasing at 80 m depth (Glynn et al. 1996).

According to Maté (2003b), the relative abundance in the Gulf of Panama, varies from abundant to rare depending on site; and common to rare in the Gulf of Chiriquí. However, Guzman (pers. comm. 2008) disagrees with Maté (2003b), since Guzmán et al. (in prep. 2008) recorded it at 136 sites in the Gulf of Chiriquí and at 61 sites at las Perlas Archipelago. At Coiba Archipelago, Gulf of Chiriquí, this species was present in 75 to 100% of the sites studied by Guzmán et al. (2004).

This species was considered abundant at Caño Island, Costa Rica (Guzmán and Cortés 2001); common along mainland Colombia (Ensenada de Utría and Tebada), as well as in Cocos Island, Costa Rica (Glynn and Ault 2000, Guzman and Cortes 2006) and Malpelo Island, Colombia (G. Edgar pers. comm. 2008). Uncommon in mainland Costa Rica and the Galápagos Island, Ecuador (Glynn and Ault 2000, Glynn 2003). Rare in Mexico (from Nayarit to Oaxaca and the Revillagigedo Islands) and mainland Ecuador (Reyes-Bonilla 2003, Glynn 2003).

According to Glynn et al. (2000), population densities increased from ~1-4 colonies per 20 m² immediately following the 1982-83 ENSO to six to 12 colonies at Caño Island (Costa Rica) and Uva Island (Panama) after 10-yr. Survivorship to reproductive maturity at Caño Island, Uva Island and the southern Galapagos corresponded with recruitment rates of 0.1, 0.2 and 0 colonies/m², respectively, and recovered to pre-ENSO abundances at these sites (Glynn et al. 2000). Recruitment success at Uva Island was significantly related to maximum monthly positive sea surface temperature (SST) anomalies that occurred in the year preceding recruitment over the period 1982 to 1996; recruitment failed when SST anomalies exceeded 1.6 to 1.9°C during the severe ENSO events of 1982-83 and 1997-98 (Glynn et al. 2000).

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 not primarily restricted to depths shallower than 30 m and is more resilient 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 0-110 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 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 more resilient 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 more resilient to threats related to reef degradation. Therefore, past decline was inferred from the 20% percentile of estimated coral cover loss, resulting in a global level suspected decline of less than 25% since 1989, or over the past three generations.

This species occurs in most reef environments on shallow and deeper reefer slopes and on vertical walls. Generally, this species occurs broadly amongst coral reefs and coral communities on rocks and rubble substrata, but is absent from shallow platforms with high energy (Cortés and Guzmán 1998, Glynn et al. 2000); in some localities, this coral is found in cryptic habitats (Glynn et al. 2000). The maximum size is over 1 m. 

It is a broadcast-spawner, releasing masses of minute eggs and sperm (Glynn et al. 2000). Most colonies utilize an alternating periodic sequential hermaphrodite mode of reproduction (Glynn et al. 2000). Glynn et al. (2000) suggest year-round reproductive activity. In addition, spawning appears to increase during high temperature conditions (Glynn et al. 2000). According to Glynn et al. (2000), the species may reach reproductive maturity at seven years; while the minimum colony size at first reproduction was found to be 5 cm (about five years old). 

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.

This species is relatively resilient to bleaching (Winston et al. 2022).

El Niño Southern Oscillation (ENSO) events are the most important source of natural disturbance controlling coral communities (Glynn 1990). Pavona species have a high sensitivity to extreme elevated temperatures that interfere with reproduction and recruitment (Glynn et al. 2000). El Niño disturbance could have perilous consequences for small populations of eastern Pacific reef corals (Glynn 1988).

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 in the coast 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).

Other threats (Glynn et al. 2000): small colony size; slow skeletal growth; susceptibility to Acanthaster planci predation, and infrequent asexual fragmentation. Crown-of-thorns starfish (COTS) (Acanthaster planci) 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 the crown-of-thorns starfish 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 and increased susceptibility to disease, increased severity of ENSO events and storms, and ocean acidification.

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.

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.