This species is widespread and common and is highly susceptible to bleaching. 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 <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 A3c. The change in status from the previous assessment reflects updated declines calculated from improved data on modelled coral cover loss and projected date of annual severe bleaching, along with improved knowledge of species traits.

This species is distributed from the Red Sea, Gulf of Aden, East Africa, southwestern and northern Indian Ocean, central Indo-Pacific, north, west and eastern Australia, Japan, East China Sea, and oceanic West Pacific (Veron 2000). It has also been confirmed from eastern Thailand, Taiwan (Huang et al. 2015), the Lakshadweep Islands and Yap (DeVantier and Turak 2017).

The depth range is 1-90 m, but the species primarily occurs from 2-30 m (Muir and Pichon 2019, L. DeVantier pers. comm. 2024).

This species is common (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 (Hughes et al. 2018). 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. Annual severe bleaching 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). Annual severe bleaching 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 Intergovernmental Panel on Climate Change (IPCC) CMIP6 global climate models to estimate the projected onset of ASB for the years 2015–2100 on a 27 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 poorly understood and varies by species and locality (van Hooidonk et al. 2013, Logan et al. 2014). 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 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 understanding that a coral species with shallow depth preferences is more frequently exposed to extreme temperatures and is expected to decline at a faster global rate than species that also or primarily occur in deeper, cooler waters (Riegl and Piller 2003). 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 future 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’ entire 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 1–90 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 more resilient to disease, does recover well from bleaching or disease, has a low susceptibility to crown-of-thorns starfish, is highly susceptible to bleaching, has an unknown 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 inferred to be more resilient to threats related to reef degradation. Therefore, past decline was suspected from the 20% percentile of estimated coral cover loss, resulting in a global level decline of <25% since 1989, or over the past three generations.

This 'robust' species is found on reef slopes and on reef flats or lagoons as a free-living single polyp (Veron 2000). The maximum size is 30 cm in diameter. Corals are circular, and have a thick flat or strongly arched form. It prefers clear waters protected from wave exposure. This species is common in onshore and offshore reefs in Indonesia along reef flat and slopes (Hoeksema 2012). Mushroom corals serve as habitats to shrimps, flatworms, fishes, ctenophores, etc. (Hoeksema et al. 2012, 2013). This species is a gonochoric spawner.

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.

This species is highly susceptible to bleaching and has a low susceptibility to disease. In one study, 75% of the individuals of this species were recorded as bleached (Hoeksema 1991). In a later study (2005) in the same area, this appears to have had no mortality effect on population density (B. Hoeksema pers. comm. 2008).

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 (Blunden 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 1997–1998, 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°C 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 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).

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. Over-fishing 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). 

Coral disease is emerging as a serious threat to coral reefs worldwide, causing large-scale reef deterioration (Weil et al. 2006), and may be as likely to cause mortality as bleaching in the coming decades (Maynard et al. 2015). Increased coral disease prevalence and mortality has been linked to increased thermal stress, reduced water quality and clarity, dredging associated sedimentation, and plastic pollution (Sokolow 2009, Ghoora et al. 2018, Lamb et al. 2018).

Tropical cyclones can affect reefs at latitudes most commonly between 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). Coral is harvested and traded to supply building materials domestically, and curios, jewellery, and ornamental organisms for the aquarium industry globally (Green and Shirley 1999, Bruckner 2001) Indonesia and Fiji dominate the export market while the United States of America are the biggest importers (Pavitt et al. 2021). Although not considered a major threat to most species, population declines associated with trade has led to the listing and monitoring of more than 2000 species of coral by the Convention on International Trade in Endangered Species of Wild Fauna and Flora (CITES) (Green and Shirley 1999). Almost all the CITES monitored exported coral between 2007 and 2016 was recorded as either wild-sourced live (transported in water, recorded by number of pieces only) or raw coral (dead coral, base rock, and live rock, reported by kg) totalling 19.8 million pieces, and 24 million kg respectively (Pavitt et al. 2021). Most species in trade are Scleractinia species with raw corals nearly all traded as Scleractinia spp., while reported trade in live corals is much more taxonomically diverse with 445 taxa recorded. Accurate identification of species in trade is however highly specialised and taxa are often identified at higher levels than species in many trade records (Green and Shirley 1999). 

The severity of these combined threats to the global population of each individual species is not known. However, more than 60% of the world’s reefs are immediately threatened by local pressures (Burke et al. 2012), and only 0.3% of global coastal coral ecosystems are at very low or no risk from these combined anthropogenic pressures, or considered to be ‘wilderness’ areas (Jones et al. 2018). The presence of these multiple local stressors is shown to directly negatively impact reefs as well as magnify the impacts from prolonged marine heatwaves (Donovan et al. 2021).

All stony corals are listed on CITES Appendix II. All stony corals (Scleractinia) fall under Annex B of the European Union Wildlife Trade Regulations (EU 2019), and have done so since 1997. Furthermore, several countries (India, Israel, Somalia, Djibouti, Solomon Islands and the Philippines) have at various stages banned either the trade or export of CITES II listed species (UNEP 2020).

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. Parts of the species’ range overlaps with Marine Protected Areas.

Recommended measures for conserving this species include research in taxonomy, population, abundance and trends, ecology and habitat status and reproduction, 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.

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).