Astreopora eliptica is now a synonym of this species (WoRMS online database accessed January 2022).

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 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 distributed from the Red Sea, Gulf of Aden, East Africa to South Africa (S.N. Porter pers. comm. 2023), southwest and northern Indian Ocean, Persian Gulf, central Indo-Pacific, Australia, Southeast Asia, Japan, the East China Sea, and oceanic west Pacific. It does not occur at Hawaii and Lord Howe Island. It has also been confirmed from eastern Thailand, northern Vietnam, southern China and Taiwan (Huang et al. 2015).

The depth range is 1-50 m, but the species primarily occurs from 3-30 m (L. DeVantier pers. comm. 2024).

This species is common (Veron et al. 2016, DeVantier and Turak 2017). In a survey of the relative abundance of reef building corals across the Indo-Pacific, this species was recorded at 51.61% of the 3,075 sites (DeVantier and Turak 2017) making it one of the most common corals in the world. At Kut Island, Thailand, the percentage cover of this species declined by 69% between 2007 (before bleaching) and 2012 (after bleaching) (Yeemin et al. 2013).

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. 2020). 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 2071 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 1°C of adaptation is lower given this species is primarily restricted to depths shallower than 30 m and is more resilient to bleaching. 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 1-50 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 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 moderately 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 less than 25% since 1989, or over the past three generations. Given that ASB is projected to occur beyond the three-generation length period, the final species decline was based on past coral cover loss.

This species is found in most shallow subtidal, tropical reef environments except very turbid water (Veron et al. 2016) and does seem to be able to exist in adverse conditions, although it does not tolerate sediment emersion such as that which occurs on reef flats (Lamberts 1982). It can be rare or relatively common (Lamberts 1982). This species is a hermaphroditic broadcast spawner.

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 whose age at first maturity is typically 4 years; however, it can vary between 3 and 8 years (Harrison and Wallace 1990, Iwao et al. 2010, Baria et al. 2012, Montoya-Maya et al. 2014, Ligson and Cabaitan 2021). 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 does not seem to be particularly susceptible to thermally induced bleaching but rather has a relatively low to moderate bleaching susceptibility (Davies et al. 1997, Marshall and Shuttenberg 2006, McClanahan et al. 2004; 2007, Sutthacheep et al. 2013, Raymundo et al. 2019). Astreopora stood out as the only genus of Acroporidae not severely affected by bleaching on the Great Barrier Reef during the 1998 bleaching event (Marshall and Baird 2000). McClanahan (2004) also found that Astreopora is likely to have high potential recovery, as although it bleached, it did survive. Nevertheless, during the 2016-2017 bleaching event, most reefs around the world exhibited significant levels of bleaching and over the past two decades the probability of bleaching has shown an increasing trend (Sully et al. 2019). In the western Indian Ocean, during the 2016 bleaching event, there was an approximate 20% decline in coral cover (Gudka et al. 2018). In the Gulf of Thailand, a severe bleaching event occurred in 1998 followed by another in 2010 which resulted in significant mortality of many corals, and caused a decline in the percentage cover of this species from 2007 to 2012 at the two sites it was found with little evidence of recovery (Yeemin et al. 2013). At Farallon de Medinilla, western Pacific, Carilli et al. (2020) found that there was severe bleaching in three of the most dominant corals, which included Astreopora, with more than 90% of colonies being affected. In Mauritius, Astreopora was also found to exhibit low levels of bleaching with an index of only 7 (McClanahan et al. 2005). On the reefs of Guam, this species was moderately susceptible to bleaching, but exhibited no mortality (Raymundo et al. 2019). During the 2016 global bleaching event, Astreopora spp. from the Maldives were one of the more bleaching resistant genera (Ibrahim et al. 2017). In the Maldives, this species has been predicted to have a low total susceptibility to mass bleaching and a low relative extinction risk (Muir et al. 2017). These findings suggest a high degree of intra-specific bleaching susceptibility in Astreopora that may be genetically controlled and or environmentally influenced (Bradbury 2013, Pisapia et al. 2014).

In the western Indian Ocean, a white patch syndrome associated with infection by fungal hyphae has been reported on Astreopora (McClanahan et al. 2004, Sere et al. 2015). Astreopora was also reported to be one of the most susceptible genera to disease by Sere et al. (2015) in the western Indian Ocean and by Aeby et al. (2016) on reefs in New Caledonia. Growth anomalies affecting tabulate acroporids and Astreopora sp. were the most common disease condition in the Kerama Islands of Japan (Weil et al. 2012). Coral disease has emerged as a serious threat to coral reefs worldwide and a major cause of reef deterioration (Weil et al. 2006, Ruiz-Moreno et al. 2012). The numbers of diseases and coral species affected, as well as the distribution of diseases have all increased dramatically (Porter et al. 2001, Green and Bruckner 2000, Sutherland et al. 2004, Weil 2004). Coral disease epizootics have resulted in significant losses of coral cover and were implicated in the dramatic decline of acroporids in the Florida Keys (Aronson and Precht 2001, Porter et al. 2001, Patterson et al. 2002). In the Indo-Pacific, disease is also on the rise with disease outbreaks reported from the Great Barrier Reef (Willis et al. 2004, Haapkyla et al. 2010), Marshall Islands (Jacobson 2006) and the northwestern Hawaiian Islands (Aeby et al. 2006). White syndrome has been reported from numerous locations throughout the Indo-Pacific and constitutes a growing threat to coral reef ecosystems (Sussman et al. 2008, Bourne et al. 2015). Increased coral disease levels on the GBR were correlated with increased ocean temperatures (Boyett et al. 2007, Miller and Richardson 2015, Maynard et al. 2015, Aeby et al. 2020) supporting the prediction that disease levels will be increasing with higher sea surface temperatures. In most instances, disease is a symptom of  escalating anthropogenic stresses such as thermal stress, increased turbidity, nutrient enrichment and even SCUBA diving and tourist activities (Sutherland et al. 2004, Ruiz-Moreno et al. 2012, Lamb et al. 2014, Pollock et al. 2014, Vega Thurber et al. 2014) which have placed coral reefs in the Indo-Pacific at high risk of collapse. 

Astreopora species are elected over many other species by the crown of thorns starfish (Tokeshi and Daud 2011), and under certain circumstances, can be considered common prey such as was observed on a high-latitude reef in South Africa (Celliers and Schleyer 2006). 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 (Pratchett 2010, Kayal et al. 2012, Baird et al. 2013). 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 has become a major threat to some species, and have contributed to the overall decline and reef destruction in the Indo-Pacific region (Sweatman et al. 2011, Baird et al. 2013, Montano et al. 2014, Pratchett et al. 2014). 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.

In general, the major threat to corals is global climate change, in particular, temperature extremes and marine heatwaves leading to bleaching and increased susceptibility to disease, increased severity of marine ENSO events and storms, and ocean acidification. Global warming is significantly altering coral reef ecosystems through an increasing frequency and magnitude of coral bleaching events (Graham et al. 2007; 2015; Hughes et al. 2017). 

Localized threats to corals include fisheries, human development (industry, settlement, tourism, and transportation) (Nguyen et al. 2013), changes in native species dynamics (competitors, predators, pathogens and parasites), invasive species (competitors, predators, pathogens and parasites) (Hume et al. 2014), dynamite fishing (Wells 2009), chemical fishing (Madeira et al. 2020), pollution from agriculture and industry (Bruno et al. 2003), domestic pollution, sedimentation (Babcock and Davies 1991, Cunning et al. 2019), and human recreation and tourism activities (Lamb et al. 2014). The severity of these combined threats to the global population of each individual species is not known. However, due to the wide range of this species, these threats are known to occur within its distribution. Furthermore, these threats are known to negatively affect the species to varying degrees and may have pronounced synergistic effects.

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) at various stages have banned either the trade or export of CITES II listed species, which includes all stony corals, since 1999 (UNEP 2020). Fiji and Malaysia currently (2020) have quotas for the number of wild harvested Astreopora species in general for export, which range from 2,100 to 1,000 pieces per annum, respectively (UNEP-WCMC 2020).

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, 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 (CBD 2010), 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 (IUCN 2016).

It is crucial that global warming is constrained well below 2°C (the goals of the Paris Agreement) (Hoegh-Guldberg et al. 2018).