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 19% 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 70% 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 is distributed in Fiji (DeVantier and Turak 2017) and the far eastern Pacific. A record from Nha Trang, Vietnam (Latypov 2011) may need verification.

It occurs in the Eastern Tropical Pacific region from: Mexico: Baja California Sur, Nayarit, Jalisco, Colima, Michoacán, Guerrero and Oaxaca (Reyes-Bonilla and Lopez-Perez 1998, Reyes-Bonilla 2003, Calderon-Aguilar 2005, Reyes-Bonilla et al. 2005); El Salvador: Del Amor beach, Los Cobanos (Reyes-Bonilla and Barraza 2003); Costa Rica: Bahia Culebra, Samara, Manuel Antonio, Caño Island (Cortes and Guzmán 1998), and Cocos Island (Glynn and Ault 2000); Panama: Coiba Archipelago (Guzmán et al. 2004), Gulf of Chiriquí and Las Perlas Archipelago (Guzmán et al. in prep. 2008); Colombia: Ensenada de Utría; Gorgona Island and Malpelo Island (Zapata and Vargas-Angel 2003); Ecuador: Salango Island, Los Frailes, Sucre Island and La Plata Island, and throughout the Galapagos Archipelago (except for Fernandina and the west side of Isabela) (Glynn et al. 2001, Glynn 2003).

The depth range is 0-40 m (Veron et al. 2000), with a preferred depth range of 0-8 m (Glynn et al. 1982). 

This species is uncommon (DeVantier and Turak 2017).

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

Common: Gulf of California, and from Nayarit to Oaxaca, Mexico (Reyes-Bonilla 2003).

Uncommon: Costa Rica (Cortés and Guzmán 1998); Panama (Glynn and Ault 2000); found at 45 sites in the Gulf of Chiriqui and 17 sites in Las Perlas Archipelago (Guzmán et al. in prep. 2008). Colombia (Glynn and Ault 2000), and Ecuador including the Galápagos Archipelago (Glynn and Ault 2000, Glynn 2003).

Rare: Del Amor beach, El Salvador (Reyes-Bonilla and Barraza 2003) and Revillagigedo Islands, Mexico (Reyes-Bonilla et al. 2005).

H. Guzmán and Chiriboga (pers. comm. 2008) consider that P. capitata is common in some localities of Panama (Gulfs of Chiriqui and Panama) and the Galápagos Islands. Moreover, according to H. Guzmán (pers. comm. 2008), P. capitata populations seem to be recovering at Panamá.

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 this high coral mortality, numerous pocilloporid recruits have been observed at some sites in Costa Rica, Panama and the Galápagos Islands (P. Glynn unpublished data in Glynn et al. 1991).

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 recovery or recruitment in these areas since (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 2029 for SSP5-8.5 and by 2028 for SSP2-4.5 assuming no level of adaptation and by the year 2055 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 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 70% 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 0-40 m and is predominately found at depths less than 10 m, generalized abundance is considered uncommon, 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 19% since 1989, or over the past three generations.

This species occurs in shallow, tropical reef environments in shallow rocky foreshores and is found on coral reefs, and in lagoons, platforms and coral communities on rock at sites with high energy; from shallow depths to around 20 m (Cortes and Guzmán 1998, H. Guzmán pers. comm. 2008). The maximum size is 50 cm.

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 (1-15 m). 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).

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)), buterflyfishes, 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 A. planci (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.

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, Gardineroseris) 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). 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) 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 Galápagos 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 3 m 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).

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.

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.