For additional taxonomic information, please refer to Reyes-Bonilla et al. (2017). According to Veron et al. (2016), this species is probably a synonym of Pocillopora damicornis (Combosch et al. 2008). It is accepted in the WoRMS online database (accessed February 2023). This species is difficult to identify morphologically (Paz-Garcia et al. 2015).

There is high taxonomic uncertainty associated with this rare species that may be restricted to the eastern tropical Pacific. Little is known on its distribution, population, ecology and potential threats; therefore, it is listed as Data Deficient.

If valid, this species is distributed in the eastern tropical Pacific from the continental and oceanic islands of Mexico, Costa Rica, Panama and Ecuador (Glynn et al. 2001, Reyes-Bonilla et al. 2017).

Mexico: Gulf of California and nearby areas, Guerrero and Oaxaca (Reyes-Bonilla 2003, Reyes-Bonilla et al. 2005, Lopez-Perez et al. 2012).
Costa Rica: Santa Elena (Cortés and Jiménez 2003); Bahia Culebra (Jiménez unpublished data, Cortés and Jiménez 2003); and San Pedrito Island, Islas Muricelago (Glynn 1999, Cortes 2017).
Panama: Saboga Island, and Contadora Island in the Gulf of Panama (Mate 2003, Glynn 1999), Pedro Gonzalez Island (Las Perlas) and Gulf of Chiriqui (Guzmán et al. 2004).
Ecuador: throughout the Galapagos Archipelago (Glynn 1999).

According to Obura and Stone (2002), this species is tentatively identified in the central tropical Pacific in the Republic of Kiribati. 

The depth range is 1-10 m.

The relative abundance of this species has been categorized as:

Rare: in the Gulf of California and nearby areas, from Nayarit to Oaxaca, Mexico (Glynn and Ault 2000, Reyes Bonilla 2003), and Costa Rica (Glynn and Ault 2000).

Uncommon: in Panama, and in the Galápagos Archipelago, Ecuador (Glynn and Ault 2000).

According to Guzmán et al. (2004), this is a rare species at Coiba Archipelago, Panama; found in less than 25% of the studied sites. At seven sites in Las Perlas and nine sites in the Gulf of Chiriqui (Guzmán et al. in prep. 2008).

Glynn (1999) indicates that the abundance of pocilloporid species shows marked interannual fluctuation. At three sites in the Galápagos Islands (Cormorant Bay, Floreana; Northeast anchorage, Santa Fe; and Caleta Robinson, Santa Cruz) in surveys undertaken from 1993 to 1997, it was present during some surveys, but absent after 1 to 3 yr in subsequent surveys (Glynn 1999).

Although this species is widely distributed in the Galapagos Islands, it has a low relative abundance (1.9 to 16.7% of all pocilloporid corals), as well as very low densities, ranging from 0.2 to 2.5 colonies per ha (Glynn 1999). According to Glynn (1999), this low abundance probably reflects the severe coral mortality (97%) during the 1982-83 ENSO event. Since them coral recovery has been negligible to slow (Glynn 1994); however, it is likely that all live colonies of this species have recruited since 1983 (Glynn 1999).

This species occurs in coral reef and coral communities on shallow rocky substrata (H. Guzmán pers. comm. 2008) shallower than 10 m depth (Glynn 1999, Guzmán et al. in prep. 2008). It is usually found intermixed with other pocilloporid corals in Costa Rica and Panama (Jiménez and Cortés 2003, Guzmán et al. in prep. 2008).

According to Glynn (1999), this species may have evolved within the eastern Pacific and is presently distributed widely over much of the region. The presence of a large dead colony in the Galápagos Urvina Bay uplift, where a coral community was elevated during a volcano eruption in 1954, indicates it is not a recent ENSO-associated immigrant from the Indo-West Pacific and that this species has been present in the Galápagos Islands for more than four decades (Glynn 1999).

Amongst the reef building corals in the Eastern Tropical Pacific, pocilloporid species have the fastest growth rates (Guzmán and Cortes 1993). The growth rates of this species ranges between 2.0 and 4.4 cm per year, with a mean growth rate of 3.15 cm per year at Bahia Culebra Costa Rica (Jiménez and Cortés 2003, Cortés and Jiménez 2003b). 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). Pocillopora species are preyed on by at least nine groups of consumers, including pufferfishes (Arothron), parrotfishes (Scaridae), filefishes (Monacanthidae) (Glynn 2002), hermit crabs (Trizopagurus, Aniculus, and Calcinus) (Glynn 2002), gastropods (Jenneria pustulata and Quoyula sp.; Glynn 2002), butterflyfishes, angelfishes, damselfish (Stegastes acapulcoensis), and Acanthaster planci (Glynn 2002), and 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.

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 (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 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 3 m depth (Guzmán et al. 1990).

According to Glynn (2001), pocilloporid coral harvesting is an important threat in the Eastern Tropical Pacific region, especially 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. All stony corals (Scleractinia) fall under Annex B of the European Union Wildlife Trade Regulations, and have done so since 1997. Furthermore, several countries (India, Israel, Somalia, Djibouti, Solomon Islands and the Philippines) at various stages have banned either the trade or export of CITES II listed species, which includes all stony corals, since 1999. 

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. 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).

Taxonomic research is needed (Reyes-Bonilla et al. 2017).