This widely distributed species is 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 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 21% 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 65% 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. The change in status from the previous assessment reflects updated declines calculated from improved data on modeled coral cover loss and projected date of annual severe bleaching, along with improved knowledge of species traits.

This species occurs in Australia, including the Great Barrier Reef, Lord Howe Island, Northern Territory and Western Australia from Shark Bay, Dampier and Ningaloo Reef (Atlas of Living Australia online database accessed October 2020). It also occurs in New Caledonia, the Solomon Islands and the Marshall Islands (Z. Richards pers. comm. 2008). Its presence in Palau requires verification (Veron et al. 2016). There are unconfirmed reports at Myanmar by Ko Ko (2019) and at Aldabra by Sheppard (1987). It has also been confirmed from southern Vietnam and the Spratly Islands (Huang et al. 2015).

The depth range is 0-10 m, but the species is most common at 1-7 m (L. DeVantier pers. comm. 2024).

This species is common (DeVantier and Turak 2017). It is considered to be rare on the Great Barrier Reef and further south in Moreton Bay, Australia (DeVantier 2010, DeVantier and Turak 2017). It was recorded at only one of seven localities in the Kimberly region of Western Australia but listed as cf. indicating taxonomic uncertainty (Richards et al. 2014) and at two out of 12 localities in Hervey Bay, eastern Australia (Butler et al. 2013). In a survey of the relative abundance of reef building corals across the Indo-Pacific, DeVantier and Turak (2017) categorised the species as exhibiting diffusive rarity (as opposed to geo-suffusive and habitat-suffusive) as it is distributed sparsely across a broad geographic range and only occurred at 0.1% of the 3,075 sites surveyed. This species was considered to be regionally vulnerable in the Marshall Islands (Richards and Beger 2013). It was also considered by Roelofs (2018) to be vulnerable to overexploitation in the Queensland coral fishery largely due to its rarity. It is rare on the Great Barrier Reef (Veron 1986, Díaz and Madin 2010).

There is limited species-specific population information available; however, there has been a substantial regional-scale decline of both plate and branching Acropora corals on the Great Barrier Reef over the past century (Clark et al. 2017, Dietzel et al. 2020). Non-tabular Acropora on the crest and slope areas of the Great Barrier Reef have decreased by 63 and 46% respectively since 1995/1996 (Dietzel et al. 2020). 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 2060 for SSP5-8.5 and by 2068 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 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 65% 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 0-10 m and is predominately found at depths less 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 (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 21% 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 occurs on shallow reef flats and in lagoons (Veron et al. 2016). For example, Loh et al. (1998) collected this species at depths of 0-5 m at One Tree Island on the Great Barrier Reef. This species occurs predominantly along continental or island coastlines, rather than on patch reefs. In Eastern Australia, it occurs in sub-tropical latitudes, in coastal shoals, rocky reefs and on some cay-bearing reefs offshore, but is rarely seen on the middle and outer reefs of the Great Barrier Reef (Wallace 1999). This species inhabits islands/cays rather than middle and outer reefs, and thus its habitat range seems quite narrow (C.I. Nuñez Lendo pers. comm. 2023).

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 rare and only found in very shallow depths (0-5 m), thus making this species highly susceptible to bleaching (C.I. Nuñez Lendo pers. comm. 2023). Members of this genus have a low resistance and low tolerance to bleaching and disease, and are slow to recover. However, at Heron Island, it was the only reef flat species to survive the 2002 bleaching event (M. Pratchett pers. comm. 2020), and therefore, may be resistant to bleaching. This species disappeared from the only two sites out of six that it was recorded from after a flood event at Hervey Bay, Australia (Butler et al. 2013).

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 (El Niño/La Niña Southern Oscillation) 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, Graham et al. 2015, Hughes et al. 2017, Dietzel et al. 2020). Marine heatwaves have resulted in widespread coral bleaching and mortality (Hughes et al. 2017). Acropora species are particularly susceptible to thermally induced bleaching and have high subsequent mortality (Marshall and Baird 2000, McClanahan et al. 2007, Hughes et al. 2018a). 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).

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, 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 (Baird et al. 2013). 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. 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 (Keesing et al. 2019). 

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 within the last decade (Green and Bruckner 2000, Porter et al. 2001, 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. 2011). 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). Several diseases have been found to have extended geographic distribution to Japan in the northern Pacific (Weil et al. 2012). 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 et al. 2014), which have placed coral reefs in the Indo-Pacific at high risk of collapse.

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 (Albert et al. 2012), 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, many of the general threats listed above are known to occur within the distribution of this species, such as coral bleaching from thermal stress (Beger et al. 2008, Hughes et al. 2018b), disease (Beger et al. 2008), predation by crown of thorns (Beger et al. 2008), pollution (Kroon et al. 2016), sedimentation (Bainbridge et al. 2010), harvesting for the aquarium trade (Brewer et al. 2012) as well as dynamite and cyanide fishing (Brewer et al. 2013).

This species is not specifically targeted or well known in the aquarium trade due to its rarity and most are likely to be mistaken for Acropora digitifera, which it superficially resembles (Roelofs 2018). In 2006, there were a total of 55 specimens exported globally, and in 2009, a total of 3 specimens imported globally (CITES 2020).

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, 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).