This species is widely distributed 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 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 with a recommendation to monitor the eastern Pacific population, which tends to occur in shallow waters and is at higher risk.

This species is found in the southwest and central Indian Ocean, the central Indo-Pacific, tropical Australia, southern Japan, the South China Sea, the oceanic West Pacific, the Hawaiian Islands, Johnston Atoll, and the eastern Pacific. It has also been confirmed from eastern Thailand (Huang et al. 2015).

It is recorded throughout the Eastern Tropical Pacific region: México: San Lorenzo Channel, Baja California Sur, and in the locations of Santa Cruz Bay and La Entrega Bay at Oaxaca (Layte-Morales et al. 2001, Reyes-Bonilla 2003, Reyes-Bonilla et al. 2005); Costa Rica: Bahía Culebra and Cocos Island (Jiménez 1997, Jiménez et al. 2001, Cortés and Guzmán 1998, Cortés and Jiménez 2003); Panamá: Gulf of Panamá* (Glynn and Maté 1997, Maté 2003); Colombia: Gorgona Island, Colombia* (Glynn and Ault 2000, Zapata and Vargas-Ángel 2003); Ecuador: La Plata Island* (Glynn 2003).

(*only dead L. papyracea found) 

The depth range is 1-176 m, but the species primarily occurs from 15-150 m (L. DeVantier pers. comm. 2024). In the western Pacific, the species prefers deeper depths than in the eastern Pacific (Dinesen 1980, Pochon et al. 2015, Padilla‐Gamiño et al. 2019, Tavakoli-Kolour et al. 2023). In the Eastern Tropical Pacific, it is usually found at mean depths of 13 m (Leyte-Morales et al. 2001). 

This species is uncommon (DeVantier and Turak 2017).

There is no specific population information available for the Indo-West Pacific. According to Glynn and Ault (2000), populations seem to be declining in Eastern Tropical Pacific region. Since this species was previously recorded from several sites throughout the Eastern Tropical Pacific.

As of 2008, this species was only found in two countries in the Eastern Tropical Pacific region in Costa Rica (Jiménez 1997, Cortés and Guzmán 1998, Jiménez 2001, Jiménez et al. 2001, Cortés and Jiménez 2003) and México (Leyte-Morales et al. 2001, Reyes-Bonilla 2003, Reyes-Bonilla et al. 2005).

In Mexico, this species was found in the states of Baja California Sur (San Lorenzo Channel) and in Oaxaca (Santa Cruz and La Entrega). Populations from Oaxaca have disappeared, only dead fragments are visible (Leyte-Morales et al. 2001, Reyes-Bonilla pers. comm. 2008). In addition, populations at La Entrega and in San Lorenzo Channel are very small with most colonies dead (Leyte-Morales et al. 2001). According to Leyte-Morales et al. (2001), this is evidence that populations in these sites are failed, transient or marginal, which hampers the possibility of sexual reproduction.

Before the 1997-98 ENSO event, at the Esmeralda reef, Bahía Culebra (Costa Rica), the deepest (i.e., mean depth of 13 m) sections of the reef were completely dominated and constructed by this species (Jiménez 1997). This living stand was approximately 2,500 m² (Jiménez 2001). Nevertheless, this small patch reef was reduced from 68% to 3.6% live coral cover when the 1997-98 ENSO event began (Jiménez et al. 2001, Cortés and Jiménez 2003); after the event, the total live coral cover was reduced by more than 90% (Jiménez et al. 2001).

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 2071 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 1-176 m and is predominately found at depths greater than 10 m, generalized abundance is considered uncommon, overall population is not restricted or highly fragmented, does occur off-reef, is more resilient to disease, does not 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 usually occurs on lower reef slopes, on horizontal sandy bottoms and inter-reef soft substrate areas. It can occasionally found in shallow sandy flats. This species is typically found on hard surfaces amongst soft substrata with little sediment movement (Jiménez 1997; 2001). It usually forms large fields. Fragments can easily break off as a mechanism of asexual reproduction. The ecophysiology of this species at mesophotic depths in Hawaii is influenced by symbiont-host associations, photoacclimatization, trophic plasticity, and adaptation (Padilla-Gamiño et al. 2019). 

According to Leyte-Morales (2001), it occurs in shallow depths in the Eastern Tropical Pacific because of two main factors: reduced light and lower temperatures at shallow depths, both of which are associated with upwelling and imitate deeper areas in the central and western Pacific where the species more typically occurs. At both Bahía Culebra, Oaxaca and Baja California Sur, this coral is found at a mean depth of approximately 13 m (Jiménez 1997, Jiménez 2001, Leyte-Morales et al. 2001). Considerable sections of the Esmeralda Reef are constructed of this species; although frame construction by this species is rare (Jiménez 1997, Jiménez 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.

This species exhibited high bleaching and mortality in the 1998 bleaching event in Palau (Bruno et al. 2001). Local extinctions of this species have occurred in some areas of the Eastern Tropical Pacific (e.g., Colombia and Panama; see Glynn and Ault 2000) and they have not been linked to known causes (Glynn 1997). Nevertheless, the ENSO event presumably had disastrous consequences for small populations of eastern Pacific reef corals (Glynn 1988).

Jiménez (2001) reported a frequent and intensive use of Esmeralda Reef, Bahía Culebra (Costa Rica), by commercial divers for the aquarium trade. The impact of this activity on the coral community and coral reef is considerable, having disturbed entire areas (>25 m²) of the Esmeralda reef (Jiménez 1997). Another major threat at Esmeralda reef is coastal development, since a resort marina may be built over this reef (Jimenez 2001). 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 may degrade coral reefs in Costa Rica, Colombia and 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.

The most recent, and first, multi-year, global bleaching event (spanning hundreds of kilometres or more) was from 2014 to 2017. Nearly 30% of reefs suffered mortality level-stress, more than 50% of affected reef areas were impacted at least twice, and some locations saw almost complete coral cover loss (Eakin et al. 2019). The average interval between bleaching events is now more than 50% less than before, preventing full reef recovery (Hughes et al. 2018). Bleaching events, leading to coral mortality, are predicted to become more frequent and severe. 

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.

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. And 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).