Keshavmurthy et al. (2013) reports that there are multiple lineages and clades associated with this species. Meziere et al. (2022), has reported that different lineages and clades have different bleaching responses.

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 at least 28% 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 Near Threatened A3ce.

This species is found in the Red Sea, the Gulf of Aqaba (Kteifan et al. 2017), the Gulf of Aden, the southwest, northwest and central Indian Ocean, the Persian Gulf, the central Indo-Pacific, tropical Australia, Guam (Maynard et al. 2017), southern Japan, the South China Sea, the oceanic West Pacific, and the central Pacific. It is also confirmed from northern Vietnam and Taiwan (Huang et al. 2015).

The depth range is 0-70 m, but the species primarily occurs from 1-40 m (Muir and Pichon 2019, Turak and DeVantier 2019, L. DeVantier pers. comm. 2024). This is a common deep specialist (Montgomery et al. 2019, Roberts et al. 2019).

This species is common and may be a dominant species on exposed reef fronts (Veron et al. 2016, DeVantier and Turak 2017). It has been extirpated from northern Kenya (McClanahan et al. 2008, 2020) and the Maldives (McClanahan and Muthiga 2014).

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 no level of adaptation is lower given this species is not primarily restricted to depths shallower than 30 m and is highly susceptible to bleaching. Furthermore, since the depth range of this species is 0-65 m, population decline was estimated over 46% of its depth range. 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 at least 28% 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’ 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 0-70 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 not occur off-reef, is highly susceptible to disease, does not recover well from bleaching or disease, has a high susceptibility to crown-of-thorns starfish, is highly susceptible 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 is primarily found in shallow water reef environments exposed to strong wave action. Sparse colonies are found from 3-11 m in the South China Sea and Gulf of Siam (Titlyanov and Titlyanova 2002). This species is considered to be a main reef-framework builder and is found from 0-4 m off Eilat (Sheppard 1982). It is also found on mesophotic reefs where the planulae are significantly smaller and produce less fluorescent proteins (Scucchia et al. 2020). The maximum size is approximately 30 cm across.

Endozoicomonas bacteria are abundant in the endodermal tissues of this species and appear to have an intimate relationship with the coral, possibly involving quorum sensing (Bayer et al. 2013). In biomineralization of this species, minerals are first precipitated as amorphous calcium carbonate and small aragonite crystallites in the pre-settled larva, which then evolve into the more mature aragonitic fibers characteristic of the stony coral skeleton (Akiva et al. 2018). The process is accompanied by modulation of proteins and ions within these minerals (Akiva et al. 2018, Bernardet et al. 2019). Heat stress on this species leads to starvation of the coral host due to the collapse of symbiotic nutrient cycling well before algal symbionts are actually lost. This means the loss of algal symbionts is not the direct cause of coral mortality during bleaching; rather, mortality is caused by the starvation of the coral host that follows a decrease in phototrophic carbon input by the algae (Rädecker et al. 2021). Survival of juvenile Stylophora pistillata and S. kuehlmanni corals translocated to the sea was significantly lower when not at parental depths, indicating that local adaptations and parental effects alongside larval selectivity and phenotype-environment mismatches combine to create invisible semipermeable barriers to coral dispersal and connectivity, leading to habitat-dependent population segregation (Shlesinger and Loya 2021).

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.

In general, species from this genus are considered to have higher bleaching vulnerability (Khen et al. 2023) and low recovery (Meziere et al. 2022), it may recover in the Red Sea more rapidly than in other areas of its range. In Lord Howe, there was little evidence of recovery of this species (Moriarty et al. 2023). It is a model species for stress experiments.

The collection of this species for the aquarium trade may lead to overharvest and localised reductions in abundance, especially for populations of naturally rare species (Bruckner and Borneman 2006). However, the wild collection of corals is highly selective and considered low impact in the long-term relative to other activities such as coral mining and dynamite fishing (Green and Shirley 1999, Pratchett et al. 2020). 

The susceptibility of this species to bleaching has been related to decreased capacity of its zooxanthellae species to cope with photo damage (Yakovleva and Hidaka 2004). The species is also sensitive to UV sunscreen filter chemicals, which cause bleaching (Downs et al. 2014, Fel et al. 2019), and to copper pollution, which reduces photosynthesis (Banc-Prandi et al. 2021). Thermal stress reduces the coral resilience of this species to ocean acidification by impairing control over calcifying fluid chemistry (Guillermic et al. 2021). This species is very sensitive to eutrophication (Hall et al. 2018) and undergoes severe oxidative stress and reduced aerobic scope when exposed to NO3- enrichment combined with thermal stress. Such conditions resulted in increased bleaching intensity compared to a low-nitrogen condition. However, NH4+ enrichment may amend the deleterious effects of thermal stress by favoring the oxidative status and energy metabolism of the coral holobiont (Fernandes de Barros Marangoni et al. 2020).

Like many branching corals, this species is highly susceptible to bleaching and with severely reduced populations and even local extinction of the species following bleaching events (Loya et al. 2001, van Woesik et al. 2011, McClanahan and Muthiga 2014). In Watamu Marine National Park, Kenya, this species was present prior to severe bleaching in 1998 (Lemmens 1993), but has not been found since, suggesting it is locally extinct (Cowburn et al. 2018). Ocean acidification reduces feeding rates in this species (Houlbrèque et al. 2015).

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.

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 (see e.g., Crabbe 2019). The severity of these combined threats to the global population of each individual species is not known.

This species is moderately traded for aquaria, with 1,000-10,000 pieces being collected from the wild and exported annually between 2010-2020 (CITES 2021).

Single and mixed cultivation methods of transplanted individuals of this species have been successful in Serangan planting areas, Bali, Indonesia (Sriwijayanti and Probosunu 2019). In the Gulf of Aqaba, this species shows temperature resistance of corals as a result of evolutionary history and might provide a genetic reservoir in the future, capable of restocking decimated coral reefs in other parts of the Red Sea (Krueger et al. 2017).

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. 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).