This common species is somewhat widely distributed. 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 highly 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 20% 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.

This species is found in the Red Sea, the Gulf of Aden, the northwest Indian Ocean, eastern India, northern Sri Lanka (Veron et al. 2016) and the Persian Gulf. It has also been confirmed from northern Madagascar (DeVantier and Turak 2017). Reports from the Lesser Sunda Islands and Savu Sea and Halmahera in Indonesia (DeVantier and Turak 2017) may require verification.

The depth range is 4-18 m (Ali et al. 2020).

This species is common and may form patch reefs (Veron et al. 2016, DeVantier and Turak 2017).

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 2036 for SSP2-4.5 assuming no level of adaptation and by the year 2058 for SSP5-8.5 and by 2058 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 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 4-18 m and is predominately found at depths less 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 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 highly susceptible to threats related to reef degradation. Therefore, past decline was inferred from the 80% percentile of estimated coral cover loss, resulting in a suspected global level decline of 20% since 1989, or over the past three generations.

This species is found in shallow water and is tolerant of high salinities, generally to depths of 15 m. Although this species lives generally in less than 5 m and tolerates salinities up to 48 ppt (Sheppard and Sheppard 1991). 

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 had a low susceptibility to bleaching in the Red Sea (Osman et al. 2020). There has been great degradation of this species along the coast of Pakistan, often uprooted for sale (Ali et al. 2021). The genus is not particularly susceptible to bleaching, but is more prone to disease than many other corals.

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.

Coral reefs are threatened by human and natural stressors at a range of scales. In general, the greatest large-scale threat to corals is from global climatic change, which is linked to lethal seawater temperature anomalies, along with increased frequency and severity of El Niño Southern Oscillation (ENSO) events and storms, and ocean acidification (Pandolfi et al. 2011, IPCC 2018), each a major threat to reefs in their own right. The most recent, and first, multi-year, global ‘bleaching’ event (spanning hundreds of kilometres or more) was from 2014 to 2017. Globally, 75% of reefs were affected by bleaching-level stress, with more than 50% of affected reef areas impacted at least twice over the period (Hartfield et al. 2018, Hughes et al. 2018, Eakin et al. 2019), and some localities experienced almost complete coral cover loss (Vargas-Ángel et al. 2019). The first global coral bleaching event was in 1997-98, however, this had also been preceded by multiple smaller regional and local scale bleaching events since at least 1982 (Goreau et al. 2000). While coral populations can be resilient to coral bleaching and bounce back (e.g. Diaz-Pulido et al. 2009, Pisapia et al. 2016), more frequent bleaching events in the future are expected to prevent full reef recovery and cause local extinctions of some species (Hooidonk et al. 2016, Sheppard et al. 2020). Heating episodes are also increasing in intensity with the 2014-2017 global bleaching event exposing more than three times as many reefs to bleaching-level heat stress than the 1998 event (Skirving et al. 2019). Almost all coral reefs are very likely to have degraded from their current state by 2100, even if global warming remains below 2^oC from pre-industrial levels (Frieler et al. 2012), meaning species composition will differ and diversity and extent will be reduced from present levels (IPCC 2018). There is limited scope for future latitudinal range extension of current reefs towards the poles (Muir et al. 2015), and severe bleaching episodes can also cause positive feedbacks, including impairment of larval recruitment via mortality of adult brood stock (Hughes et al. 2019).

Tropical coral reef biomes are also at particular risk from localised human pressures, with 58% of coral reefs < 30 minutes from the nearest human settlements (Maire et al. 2016). Localised threats to corals include over-intensive fisheries, coastal development (industry, settlement, tourism, and transportation), changes in native species dynamics (competitors, predators, pathogens and parasites), introduction of invasive species (competitors, predators, pathogens and parasites), destructive fishing (e.g. using dynamite), chemical fishing, pollution from agriculture and industry, domestic pollution, and recreation and tourism activities  and global trade (Burke et al. 2012). Some of these threats impact corals directly, such as being physically disturbed and smothered with sediment during a construction project (Erftemeijer et al. 2012), while others operate indirectly via ecosystem processes and linkages between corals and other reef organisms. Macroalgae is a major competitor with corals that reduces growth, causes disease, prevents new coral recruitment and can tip the entire ecosystem into a less diverse and less productive ‘algal-dominated’ reef (Hughes 1994, Bellwood et al. 2004). Macroalgal levels are controlled by both bottom-up provision of nutrients (Fabricius 2005), and top-down herbivory by parrotfish (Mumby et al. 2007), hence, while the immediate threat to the coral is the algae, the ultimate threat may be sewage, fertiliser from agriculture or overfishing of herbivorous fish. The complex nature of the coral reef ecosystem means that while the immediate threat may be obvious (e.g. macroalgae, disease, crown-of-thorns outbreak), the ultimate threat is often less clear (Nyström et al. 2008, Anthony et al. 2015).

Fishing activities can affect corals directly and indirectly. Direct threats include destructive fishing techniques using dynamite or poison that can kill corals and trawling and net entanglement that can break colonies and disturb sediment. Indirectly, fishing affects corals by disrupting the food-web and removing key ecological roles. Overfishing of parrotfish and other herbivorous fish removes top-down control of macroalgae, allowing this seaweed to overgrow and out-compete corals (Bellwood et al. 2004, Mumby et al. 2007). Removal of predators allows urchin populations to grow, which then predate on coral larvae as they graze the reef (McClanahan 2000).

Increased sediment and nutrient-rich run-off into the sea through catchment-level land use change can significantly affect coastal coral reef communities at local to regional scales (Halpern et al. 2008). Run-off can cause increased turbidity, smothering, inhibited recruitment, and reduced growth of corals, as well as introducing toxic pollutants, pathogens, and competition with algae (Koop et al. 2001, Fabricius 2005).

Crown-of-thorns starfish (COTS) (Acanthaster planci), found throughout the Indo-Pacific, can undergo massive outbreaks that rapidly devastate reefs on a local and regional level, triggered through human impacts such as enhanced nutrient loads (Pratchett et al. 2014). Populations of COTS have greatly increased since the 1970s and have been known to kill large areas of coral reef habitat, and have contributed to the overall decline and destruction of reefs in the Indo-Pacific region (Pratchett et al. 2017). 

Cyclones can affect reefs at latitudes of 7^o to 25^o north and south of the equator (Scoffin 1993), but very rarely occur in equatorial regions (Puotinen et al. 2020). They mainly affect shallow water corals down to ~20 m depth through large waves that can break or remove colonies, with branching species being most susceptible to storm damage (Scoffin 1993, Madin and Connolly 2006). Storm effects can be widespread, with 48% of the coral cover losses in the Great Barrier Reef (GBR) from 1985 to 2012 reportedly from storm damage (De’ath et al. 2012). Ecologically, the impact of cyclones is similar to that of coral bleaching, where an acute impact rapidly reduces the coral population (Harmelin-Vivien 1994, Mumby and Steneck 2008). As with coral bleaching, the frequency of cyclones is expected to increase with climate change (Emanuel 2013, Puotinen et al. 2020), meaning the time for reefs to recovery between these acute impacts is reduced and may prevent recovery. Storm strength is also increasing (Emanuel 2005), and repeated storm damage without sufficient recovery can lead to a reduction in coral species diversity, coral cover, reef complexity and can result in a phase-shift to a macroalgae or rubble-dominated state (Hughes 1994, Vercelloni et al. 2020). The resultant physical damage to reefs leads to an increase in mobile rubble, which typically inhibits recruitment and regrowth of corals for many years (Scoffin 1993, Viehman et al. 2018). Counterintuitively, storms could also mitigate against the effects of bleaching in some instances by large-scale mixing and cooling of heated waters (Carrigan and Puotinen 2014).

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).