Pocillopora danae is now a synonym of this species (WoRMS accessed January 2022).

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 at least 62% 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.

This species is distributed in the Red Sea, the Gulf of Aden, the southwest, northwest and central Indian Ocean, the Andaman and Nicobar Islands (Mondal et al. 2019), South Africa (Kruger and Schleyer 1998, Ridgway et al. 2001), the central Indo-Pacific, tropical Australia, southern Japan, the South China Sea (Yang et al. 2021), the oceanic West Pacific, the Marshall Islands (Richards and Beger 2013), the central Pacific, Palmyra Atoll (Williams et al. 2008), Hawaiian Islands, Johnston Atoll, and the Far Eastern Pacific. It is also confirmed from northern Vietnam and Taiwan (Huang et al. 2015).

In the Eastern Tropical Pacific, this species is found in Mexico: Baja California Sur, Nayarit, Jalisco, Colima, Michoacán, Guerrero and Oaxaca (Glynn and Ault 2000, Reyes-Bonilla 2003, Reyes-Bonilla et al. 2005); Costa Rica: mainland Costa Rica including Cocos Island (Glynn and Ault 2000); Panama (Glynn and Ault 2000); Colombia: mainland Colombia including Malpelo Island (Glynn and Ault 2000); Ecuador: mainland Ecuador (Glynn and Ault 2000, Glynn 2003) and the Galapagos (Veron et al. 2016).

Glynn and Ault (2000) reported this species throughout the Eastern Tropical Pacific, while Glynn (2003) later recorded this species as present on the mainland coast of Ecuador. It is not reported by Cortes and Jimenez (2003), Mate (2003), or Zapata and Vargas-Angel (2003) in Costa Rica, Panama, or Colombia, respectively. Additionally, Reyes-Bonilla (2003) considered this species as a synonym of P. elegans. Reyes-Bonilla (2002) discussed the uncertain taxonomic status of P. verrucosa and P. elegans, and suggested that more detailed morphological and genetic studies are needed to clarify the issue of whether they should be synonymised. However, Reyes-Bonilla et al. (2005) reported Pocillopora verrucosa and Pocillopora elegans as valid species. 

The depth range is 1-54 m, but the species primarily occurs from 1-30 m (Reyes-Bonilla et al. 2005, Roberts et al. 2019, L. DeVantier pers. comm. 2024). 

This species is common (Veron et al. 2016, DeVantier and Turak 2017). The absence of population subdivision driven by environmental factors and over large geographic distances suggests efficient larval dispersal and successful settlement of recruits from a wide range of reef sites. It also advocates broadcast spawning as the main reproductive strategy of this species in the Red Sea as reflected by the absence of clones in sampled colonies. These factors might explain the success of Pocillopora species throughout the Indo-Pacific and Arabian Seas (Robitzch et al. 2015).

According to Glynn and Ault (2000), the relative abundance of Pocillopora verrucosa in the Eastern Tropical Pacific region has been categorized as follows:

Abundant: Gulf of California (Mexico); Revillagigedo Islands (Mexico), and Panama.

Common: mainland Mexico, and mainland Colombia and Malpelo Island.

Uncommon: mainland Costa Rica, and mainland Ecuador.

Rare: Cocos Island

Reyes-Bonilla (2003) recorded this species as abundant throughout Mexico. By contrast, according to H. Guzmán (pers. comm. 2008), this is an uncommon species throughout its range within the region.

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 2038 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 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 62% 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 1-54 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 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 less than 25% since 1989, or over the past three generations.

This species occurs in shallow reef environments from exposed reef fronts to protected fringing reefs and coral communities on rocky substrata (H. Guzmán pers. comm. 2008). It is commonly found from 1-15 m, rarely 18-20 m, in the South China Sea and the Gulf of Siam (Titlyanov and Titlyanova 2002). The maximum size is 30 cm across.

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 2001).

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 (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). This species is susceptible to bleaching (Hoegh-Guldberg and Salvat 1995, Richier et al. 2008, Yeemin et al. 2012). However, some studies found it to be bleaching resistant (Rodríguez-Troncoso et al. 2014; 2016). This species may be vulnerable to sudden changes in underwater light fields resulting from processes such as increased turbidity caused by coastal development along the Saudi Arabian Red Sea coast (Ziegler et al. 2014). The bacterial community of this species may be rather inflexible and thereby unable to respond or acclimatize to rapid changes in the environment, contrary to what was previously observed in other corals (Pogoreutz et al. 2018). In Kenya, this species bleached less and experienced only partial mortality as compared to other species (Montano et al. 2010).

The collection of this species for the aquarium trade may lead to overharvest and localised reductions in abundance, especially for populations of naturally rare species (Bruckner and Borneman 2006). However, the wild collection of corals is highly selective and considered low impact in the long-term relative to other activities such as coral mining and dynamite fishing (Green and Shirley 1999, Pratchett et al. 2020). 

In the Eastern Tropical Pacific, El Niño and presumably climate change are threats. 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). According to Glynn et al.(1988), 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). In the Eastern Tropical Pacific, 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 3m 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 (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.

This species is moderately traded for aquaria, with 1,000-10,000 pieces being collected from the wild and exported annually between 2010-2020 (CITES 2021).

All stony corals (Order: Scleractinia) are listed on CITES Appendix II, and under Annex B of the European Union Wildlife Trade Regulations. Moreover, several countries (e.g., India, Israel, Jordan, Djibouti, Fiji and the Philippines) at various stages have banned either the trade or the export of CITES II listed species, which includes all stony corals. Other countries such as Indonesia, trade maricultured corals, with quotas, production limit and regulations in place to ensure the trade is sustainable. Having timely access to national-level trade data from CITES is valuable for monitoring trends for this species. Consideration of the suitability of species for aquaria should also be included as part of fisheries management, and population surveys should be carried out to monitor the effects of harvesting alongside other population trends.