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 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, Gulf of Aden, the southwest and northern Indian Ocean, central Indo-Pacific, Australia, Southeast Asia, Japan, Vietnam (Latypov 2011), East China Sea, western Taiwan (Veron et al. 2016), Marshall Islands (Richards and Beger 2013), oceanic west Pacific, and the central Pacific. It has also been confirmed from Hong Kong (DeVantier and Turak 2017).

The depth range is 0-45 m, but the species primarily occurs from 1-30 m (Muir and Pichon 2019, L. DeVantier pers. comm. 2024).

This species is common and may be a dominant species on subtropical 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 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 1°C of adaptation is lower given this species is primarily restricted to depths shallower than 30 m and is moderately 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 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 0-45 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 occur off-reef, is highly susceptible to disease, does or does not recover well from bleaching or disease, has a low susceptibility to crown-of-thorns starfish, is moderately susceptible 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 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 tropical, shallow and deeper clear-water reef environments on subtidal rock and rocky reefs, on the back and foreslopes, and in lagoons. It may be found in the outer reef channel. Colony growth forms are submassive, solid plates, or are encrusting (Veron et al. 2016). Species in this genus are within the 'Robust' clade grouping, and have a preference for clear water and protected exposure conditions (Madin et al. 2016).

Reproduction has been observed in colonies with an arithmetic mean radius of 1.5-2.0 cm and aged 4+ years (Kojis and Quinn 1981).

While there is some information regarding the age in which corals reach sexual maturation, it is largely based on measurements of size as a proxy for age (Harrison and Wallace 1990), 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). 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 likely to be more than ten years. Therefore, any population decline rates for the Red List assessment are measured over at least 30 years.

Bleached colonies of this species were observed after the 2016 marine cold spell in Western Australia (Tuckett and Wernberg 2018). This species appeared to survive the hyposaline disturbance in shallow waters that was a consequence of the 1991 floods in the Keppel Islands, Australia (van Woesik 1991).

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 locations 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 (van 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).

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. 

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).

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 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 be 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).

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. 

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, preferably below 1.5°C, compared to pre-industrial levels (meeting the goals of the Paris Agreement).

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.