This species is widespread and uncommon. 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 8% 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 66% 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 A3ce.

This species occurs in the central Indo-Pacific, Australia, Japan, the East China Sea, Southeast Asia, Western Australia, the Great Barrier Reef, Coral Sea, the Solomon Islands, Fiji, Micronesia, the Marshall Islands and the oceanic west Pacific (Z. Richards pers. comm. 2008, Veron et al. 2016). It has also been confirmed from West Malaysia, southern Vietnam, Taiwan and western Luzon in the Philippines (Huang et al. 2015).

The depth range is 2-40 m (Muir et al. 2015), but the species is most common at 2-20 m (L. DeVantier pers. comm. 2024).

This species is generally uncommon (DeVantier and Turak 2017). It can be frequently found amongst branching Acropora stands at the Houtman Abrolhos Islands of central-west Australia and it is rare in the Cocos (Keeling) Islands (Z. Richards pers. comm. 2021). Dietzel et al. (2021) estimated that, between 1997 and 2006, the mean total abundance of this species in the northwest Pacific (Indonesia to French Polynesia) was approximately 20 million colonies, however, the current abundance in this region is not known. It is abundant on the Great Barrier Reef (Veron 1986, Díaz and Madin 2010).

There is no species-specific population information available across its entire range; however, there is evidence that overall cover of Acropora has declined in many regions, including Australia (Hughes et al. 2020, Dietzel et al. 2020), French Polynesia (Moritz et al. 2021), Mauritius (Elliott et al. 2018), Thailand (Yeemin et al. 2013), Chagos Archipelago (Head et al. 2019), Lakshadweep Archipelago (Yadav et al. 2018), Indonesia (Cleary et al. 2014), Japan (Hongo and Yamano 2013), Taiwan (Kuo et al. 2012), the Arabian/Persian Gulf (Riegl et al. 2018). These genus-level declines indicate population declines are highly likely for all Acropora species (with the possible exception of tabular Acropora which have been shown to recover quickly after disturbance; Johns et al. 2014, Mellin et al. 2019, Richards et al. 2021). 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 2033 for SSP5-8.5 and by 2036 for SSP2-4.5 assuming no level of adaptation and by the year 2061 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 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 66% 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 2-40 m and is predominately found at depths greater than 10 m, generalized abundance is considered uncommon, the overall population is not restricted or highly fragmented, does not occur off-reef, is moderately susceptible to disease, does not recover well from bleaching or disease, has a high susceptibility to crown-of-thorns starfish (COTS), is highly 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 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 8% since 1989, or over the past three generations.

This species is found on shallow reefs in lagoons or reef slopes protected from strong wave action. It is restricted to enclosed lagoons and protected reef slope habitats and thus may be missed on many reefs (Wallace 1999). 

The age at first maturity of most Acropora species 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). 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 is particularly susceptible to bleaching, disease, predation and other threats, and therefore, population declines are predicted across its entire range.

This species has a branching growth form and is susceptible to predation by the crown-of-thorns starfish (COTS) (Acanthaster planci), which are found throughout the Pacific and Indian Oceans, and the Red Sea. Crown-of-thorns are voracious predators of reef-building corals, with a preference for branching and tabular corals such as Acropora species. Populations of the crown-of-thorns starfish have greatly increased since the 1970s and have been known to remove large areas of coral reef habitat. Increased breakouts of COTS has become a major threat to some species, and has contributed to the overall decline and reef destruction in the Indo-Pacific. 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. Crown-of-thorn outbreaks are particularly concerning in coral communities that are recovering from disturbances such as coral bleaching as feeding on remnant survivors and juveniles can further inhibit community recovery (Haywood et al. 2019).

In general, the major threat to Acropora corals is global climate change, in particular, temperature extremes leading to bleaching induced mortality, and an increased susceptibility to disease (Hoegh-Guldberg et al. 2007, Hughes et al. 2017; 2018; 2019). Bleaching can lead to mortality and a reduction in both coral cover and effective population sizes. It also disrupts coral reproduction. Regional coral extinction events following thermally anomalous events are increasingly reported (Sheppard et al. 2020, Muir et al. 2021, Richards et al. 2021). In addition, climate change is predicted to cause an increased severity of ENSO (El Niño/La Niña Southern Oscillation) events and storm intensity, and longer-term changes in ocean chemistry impacting calcification, along with an increase in the severity of flood and fire events impacting coastal catchment areas.

Coral disease has emerged as a serious threat to coral reefs worldwide and a major cause of reef deterioration (Weil et al. 2006). The numbers of diseases and coral species affected, as well as the distribution of diseases have all increased dramatically within the last two decades (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 recently reported from the Great Barrier Reef (Willis et al. 2004, Haapkyla et al. 2013), Indonesia (Haapkyla et al. 2007, Subhan et al. 2020), Thailand (Lamb et al. 2014), Marshall Islands (Jacobson 2006), Micronesia (Myers and Raymundo 2009), American Samoa (Work and Rameyer 2005), and the northwestern Hawaiian Islands (Aeby et al. 2006), the Cocos (Keeling) Islands (Preston and Richards 2021), the Maldives (Montano et al. 2015), and the Persian Gulf (Aeby et al. 2020). Increased coral disease levels on the GBR were correlated with increased ocean temperatures (Boyett et al. 2007, Howells et al. 2020) supporting the prediction that disease levels will be increasing with higher sea surface temperatures. As environmental conditions continue to change, it is predicted that conditions on temperate reefs will become favourable for coral diseases and thermodependent bacteria (Bally and Garrabou 2007, Brodnicke et al. 2019, Vergés et al. 2019).

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 (Burke et al. 2011). The severity of these combined threats to the global population of this species is not known.

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, Indonesia and Malaysia currently (2020) have quotas for the number of wild Acropora species in general for export, which range from 3,000 to 377,500 pieces per annum depending on the country (UNEP-WCMC 2020).

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