This species is widespread and rare. 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 less than 25% over the next three generations, or by 2050. It is listed as Least Concern. 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 has a wide geographic range and it occurs throughout both the Indian Ocean and the Pacific Ocean (Glynn 1997, Glynn and Ault 2000, Stefani et al. 2008). In the Indo-West Pacific, this species is found in Mayotte, Gulf of Aden, Socotra (Benzoni et al. 2012), Seychelles, Indonesia, Taiwan (Stefani et al. 2008, Denis et al. 2015), the South China Sea (Huang et al. 2016), the Line Islands, the oceanic west Pacific (Kenyon 2010), the Hawaiian Islands and Johnston Atoll (Schopmeyer and Vargas-Angel 2010), and the far eastern Pacific, including Easter Island. It is also confirmed from Brunei, northern Palawan in the Philippines (Huang et al. 2015), the Sulu Sea, and Sunda Shelf (DeVantier and Turak 2017).

In the Eastern Tropical Pacific, it is recorded from Mexico, Costa Rica, Panama, Colombia, and Ecuador (Glynn et al. 2012); specific records include:
Mexico: Baja California Sur, Nayarit, Jalisco, Colima (including the Revillagigedo Islands), Michoacán and Guerrero (Glynn 1997, Reyes-Bonilla and López-Pérez 1998, Glynn and Ault 2000, Reyes-Bonilla 2002, Reyes-Bonilla 2003, Reyes-Bonilla et al. 2005, Pérez-Vivar et al. 2006, Ulate et al. 2016).

Costa Rica: Islas Murciélago Archipelago, Culebra Bay, La Penca, Brasilito Bay, Cabo Blanco, Curú Bay, Punta Mala, Manuel Antonio, Marino Ballena National Park, Osa Peninsula, Golfo Dulce, Caño Island, Cocos Island (Cortés 1990, Glynn 1997, Cortés and Guzmán 1998, Glynn and Ault 2000, Reyes-Bonilla 2002, Alvarado et al. 2005, Reyes-Bonilla et al. 2005, Bernadette et al. 2006), Clipperton Atoll (Glynn and Ault 2000).

Panama: Coiba Island, Uva Island, Unnamed Island in the Gulf of Chiriquí, and Iguana Island, Taboga Island and Saboga Island in the Gulf of Panama (Holst and Guzmán 1993, Glynn 1997, Reyes-Bonilla 2002, Glynn and Ault 2000, Maté 2003, Guzmán et al. 2004, Reyes-Bonilla et al. 2005).

Colombia: Gorgona Island, Ensenada de Utría and Tebada (Glynn 1997, Glynn and Ault 2000, Reyes-Bonilla 2002, Zapata and Vargas-Ángel 2003, Reyes-Bonilla et al. 2005).

Ecuador: La Libertad and Sucre Island at mainland Ecuador, and Devils Crown, Floreana; Wolf Island; Darwin Bay, Genovesa Island; Gardner Bay, Española; Marchena (Wells 1983, Glynn 1997, Glynn and Ault 2000, Reyes-Bonilla 2002, Glynn 2003, Hickman 2005, Reyes-Bonilla et al. 2005, Brown 2016).

The depth range is 0-51 m (Denis et al. 2015).

This species is rare in general (DeVantier and Turak 2017).

In the Eastern Tropical Pacific, the relative abundance has been categorized as common in many parts of its range including Mexico, Costa Rica, and Panama. It is considered rare in Clipperton Atoll and mainland Ecuador, and was common in the Galápagos before the 1982-83 El Niño event (Hickman 2005). In Galápagos, following the 1982-83 El Niño event, this species experienced extreme population reductions at Española Island (Glynn 1997); in Costa Rica, after the 1991-92 ENSO event, ~40% of all colonies were dead in Manuel Antonio (Jiménez and Cortés 2001), and after the 1997-98 El Niño event, it disappeared from one locality (Punta Cambial) in Costa Rica (Jiménez and Cortés 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 2037 for SSP5-8.5 and by 2040 for SSP2-4.5 assuming no level of adaptation and by the year 2063 for SSP5-8.5 and by 2072 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 0-51 m and is predominately found at depths greater than 10 m, generalized abundance is considered rare, 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 low susceptibility to crown-of-thorns starfish, is more resilient 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 mainly on shallow wave-washed rock, or at depths of 15-20 m on coarse sand bottoms (Hickman 2005). However, a relatively recent record of this species at a depth of ~50 m in Taiwan highlights that there might be additional overlooked populations of the species in mesophotic depths elsewhere.

In general, some Psammocora species are slow-growing corals (Guzmán and Cortés 1989). Sexual reproduction is important, but asexual reproduction and fragmentation might be more effective strategies for colonizing free areas within the reef in some localities (Cortés and Guzmán 1998). Psammocora species were considered to be amongst the most opportunistic species because of their capacity to rapidly recolonize open areas after disturbances (Guzmán and Cortés 2001). Although Kolinski and Cox (2003) noted a tentative classification of this species as a brooder, a further, more in-depth examination strongly suggests that this species is broadcast-spawner and gonochoric (Glynn et al. 2012)

While there is some information regarding the age at which corals reach sexual maturation, it is largely based on measurements of size as a proxy for age (Harrison and Wallace 1990, Rapuano et al. 2023), which can be problematic in modular animals because of processes such as partial mortality and fission (Hughes and Jackson 1980). Nonetheless, it appears that many brooding coral species tend to reach puberty at about 1-2 years of age, which is much earlier than many broadcast-spawners that become reproductive at 3-8 years or more (Harrison and Wallace 1990, Wallace 1999, Rapuano et al. 2023). Therefore, we assume that the average age of mature individuals on a given reef is greater than eight years. Furthermore, based on average sizes and growth rates, we assume that the average generation length is 10 years unless otherwise stated. Total longevity is not known for any coral but is likely to be more than ten years. Therefore, any population decline rates for the Red List assessment are measured over at least 30 years.

This species is more resilient to some of the threats faced by corals, such as bleaching, and has a high recovery potential (Feingold 2001, Bezy et al. 2006, Glynn et al. 2012). The sea star Acanthaster planci and the fish Arothron meleagris feed on Psammocora species (Cortés and Guzmán 1998, Reyes-Bonilla et al. 1999). According to Cortés and Guzmán (1998), the puffer fish Arothron meleagris is capable of reducing populations of Psammocora species if other preferred coral species such as Porites lobata are absent. Despite the fact that this species is affected by increased seawater temperatures, this species is more resistant to bleaching than some other corals (Feingold 1996). Additionally, this species has also shown resistance to anomalous conditions during ENSO events (Feingold 1995, Feingold 1996, Jiménez et al. 2001). At Devils Crown, Floreana (Galápagos), this species bleached during the 1982-83 El Niño event but did not die (Robinson 1985).

Algae overgrowth has also been reported to cause mortality on Psammocora (Glynn 1997, Jiménez and Cortés 2001, Bernadette et al. 2006). At Uva Island, Panama, large aggregations of this species at 8-10 m depth that survived the 1982-83 El Niño event were overgrown by thick mats of Caulerpa species (Glynn 1997); additionally, in Manuel Antonio, Costa Rica, Psammocora species were completely overgrown by a brown algae (Jimenez and Cortés 2001). At La Penca (Costa Rica), Psammocora-dominated reef that once covered 0.3 ha is now reduced to several small (1-3 m²) patches of coral among a dense bed of Caulerpa (Bernadette et al. 2006).

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.

Generally, the biggest threat to the persistence of corals is climate change (Hoegh-Guldberg et al. 2017, Hughes et al. 2017, Sully et al. 2019), and more specifically - ocean warming and marine heatwaves that are leading to an increase in the frequency and intensity of events of anomalously high water temperatures (Hoegh-Guldberg et al. 2019, Laufkötter et al. 2020). Under anomalously high temperatures, the symbiotic relationship between corals and their photosynthetic symbionts is disrupted, and many corals begin to bleach (Glynn 1996, Hoegh-Guldberg et al. 1999, Warner et al. 1999, Loya et al. 2001, van Woesik et al. 2022). Mass bleaching events resulting from thermal stress have become increasingly common in the last two decades and may lead to widespread coral mortality and changes in the overall reef community over large areas (Loya et al. 2001, Graham et al. 2015, Hughes et al. 2018, Safaie et al. 2018, Stuart-Smith et al. 2018, McClanahan et al. 2019, Sully et al. 2019).

Superimposed on thermal stress and bleaching are additional stressors that can either directly threaten corals or exacerbate coral mortality after thermal stress (Kennedy et al. 2013, MacNeil et al. 2019, Abelson et al. 2020, Donovan et al. 2021, Knowlton et al. 2021). For example, an increasing number of storms per season, overfishing, high levels of nutrients, and other kinds of pollution are steadily increasing in magnitude and threatening coral reefs (Wiedenmann et al. 2013, Zaneveld et al. 2016, MacNeil et al. 2019, Donovan et al. 2020). Moreover, in some localities, increased amounts of outbreaks of the corallivorous sea star, crown of thorns, can cause substantial damage to the reef, contributing to the overall decline and reef destruction (Saponari et al. 2015, Pratchett et al. 2017).

The prevalence of coral disease is also rising (Aronson and Precht 2001, Rosenberg and Loya 2004, Sutherland et al. 2004, Weil et al. 2012, Maynard et al. 2015), especially in the Caribbean (Aronson and Precht 2001, Precht et al. 2016, Aeby et al. 2019, Alvarez-Filip et al. 2019, Muller et al. 2020). The increasing spatial spread and extent of diseases are associated with ocean warming (Muller et al. 2008, Ruiz-Moreno et al. 2012, Randall and van Woesik 2015) and additional anthropogenic stressors (Vega Thurber et al. 2014, Maynard et al. 2015). The escalating impacts of global warming alongside the ongoing increases in local anthropogenic stressors and diseases are causing fundamental changes to coral reefs and placing entire reef systems at a high risk of collapse.

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 have 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.

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.

All stony corals are listed on CITES Appendix II. Parts of this species distribution overlaps with several 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.