This species is widespread and uncommon. 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 about 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. Since the species does not qualify for a Near Threatened or threatened category under the projected decline, we defer to the estimate of past decline and list the species as Near Threatened A2bce. 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 is distributed in the South Pacific from Fanning Atoll, Kiribati Islands (Glynn and Ault 2000), American Samoa, New Caledonia, Fiji (D. Fenner pers. comm. 2008), and the Marshall Islands (Veron et al. 2016).

In the eastern tropical Pacific, it occurs in Mexico: Concepcion Bay to Coronados Islands; Carmen Island, Loreto and Agua Verde Bay; San Jose Island, La Paz and Cerralvo Island; Cabo Pulmo to Cabo San Luca (Baja California Sur) (Reyes-Bonilla and Caleron-Aguilera 2019); Jaltemba Island to Punta Mita, and from Isabel Island to Marias Islands (Nayari); Bandera Bay to Tenacatita Bay (Jalisco; Carpizo-Ituarte et al. 2011); Revillagigedo Islands; Manzanillo (Colima); Zihuatanejo to Acapulco (Guerrero); and Puerto Escondido to Huatulco (Oaxaca) (Reyes-Bonilla and Lopez-Perez 1998, Reyes-Bonilla 2003), El Salvador (Reyes-Bonilla 2002, Reyes-Bonilla and Barraza 2003), Costa Rica: Guanacaste, Bahia Culebra, Punta Mala, Manuel Antonio, Peninsula de Osa, Golfo Dulce, Caño Island, and Cocos Island (Cortés and Guzmán 1998), Panama: throughout the Gulfs of Chiriquí and Panama (Glynn 1997, Maté 2003, H. Guzmán pers. comm. 2008, Tortolero-Langarica et al. 2017), Colombia: Ensenada de Utría, Tebada, Gorgona Island and Malpelo Island (Reyes 2000, Zapata and Vargas-Ángel 2003), Ecuador: Salango Island, Los Frailes, Sucre Island and La Plata Island, and the Galapagos Archipelago  (Glynn et al. 2001, Glynn 2003, Hickman 2005).

The depth range is 1-30 m.

This species is uncommon (Veron et al. 2016).

According to H. Guzmán (pers. comm. 2008), this species is widespread and abundant within much of the region (Costa Rica, including Cocos Island, Panama, Colombia and throughout central and north Galápagos Archipelago), and in recent years, populations appear to be increasing, and new recruits and colonies are more frequently observed. As an illustration, Guzmán et al. (2004) reported this species in 75 to 100% of studied sites in Coiba Archipelago, Panama.

Reyes-Bonilla (2003) considered this species as abundant at the Gulf of California and nearby areas, and common from Nayarit to Oaxaca including the Revillagigedo Islands. According to Glynn and Ault (2000), it is common at Panama, mainland Colombia, and the Galápagos Islands (Ecuador), uncommon at Costa Rica (including Cocos Island) and mainland Ecuador, and rare at Clipperton Atoll.

Colonies of this species are preferred prey for Acanthaster planci (COTS) which sometimes affects the abundance of this species (Cortes and Guzman 1998).

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 2025 for SSP5-8.5 and by 2026 for SSP2-4.5 assuming no level of adaptation and by the year 2054 for SSP5-8.5 and by 2068 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 1-30 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 not 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 25% since 1989, or over the past three generations. Given that ASB is projected to occur beyond the three-generation length period, the final species decline was based on past coral cover loss.

This species occurs widely on coral reefs and coral communities on rocks, except on shallow platforms with high energy (Cortés and Guzmán 1998). In several localities, it can also grow in cryptic habitats (A. Chiriboga and H. Guzman pers. comm. 2008). This species, along with Pavona clavus and Porites lobata, can sometimes build reef frameworks or contribute to pocilloporid reef building (Glynn 2001). It can reach more than one metre high and may build patches of hundreds of square metres (Cortés and Guzmán 1998). According to Cortés and Guzmán (1993), growth rates of this species vary between 0.83 and 0.86 cm/yr.

Reproduction is mainly sexual and it may be a broadcast spawner (Glynn et al. 1996). It is an alternating periodic sequential hermaphrodite (Glynn et al. 1996, Glynn et al. 2000) and 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 11 years. This may slow the recovery of populations, especially in areas that have experienced catastrophic mortality. Sexual recruitment of this species was low after the El Niño of 1982-83, with only moderate recovery since 1983 (Glynn et al. 1996). Glynn et al. (1996) suggest that the factor which limits the number of corals reaching recruitment stages is probably mortality during planktonic development and settlement or shortly thereafter. According to Glynn et al. (1996), the low rates of sexual recruitment of agariciid corals critically slows recovery of eastern Pacific coral communities that have experienced severe disturbances. 

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 bleaching response of this species is variable (Khen et al. 2023).  
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). In general, after the 1982-83 El Niño event mean reef mortalities ranged between 50% and 75% at Caño Island, Costa Rica, and Gulf of Chiriquí, Panama, and between 85% and 97% at the Gulf of Panama, Panama, and the Galápagos Archipelago, Ecuador, respectively (Glynn 1990, Glynn 2001). Sexual recruitment of this species into eastern Pacific coral communities disturbed by ENSO 1982-83 has been low, with only moderate recovery evident since 1983 (Glynn et al. 1996). 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 include small colony size, slow skeletal growth, susceptibility to Acanthaster planci predation (COTS), and infrequent asexual fragmentation (Glynn et al. 2000).

Colonies of this species are preferred prey of Acanthaster planci (COTS) and this sometimes causes abundance declines (Cortes and Guzman 1998). 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.