This species was first illustrated by Veron (2000) and later described in Veron (2002) from the Calamian Islands, Palawan, Philippines (Wallace et al. 2012). Since then, live colonies of the species are only known from this type locality.

This rare species has a restricted range in the Calamian Islands, Palawan, Philippines. 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 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 80% over the next three generations, or by 2050. The estimated extent of occurrence (EOO) is less than 20,000 km² but greater than 5,000 km² and the area of occupancy (AOO) is less than 500 km² but greater than 10 km². The number of locations is one based on the threat from bleaching events. Though species-specific information is limited, declines are inferred based on observed reductions in coral cover in the Philippines as well as severe bleaching events that have been recorded in the Calamians. Though bleaching is considered the primary threat that will increase in frequency as climate change progresses, reefs in the Philippines are also impacted by overfishing, destructive fishing methods, and siltation. This species is listed as Endangered B2ab(iii) with a recommendation to conduct species-specific studies to better understand its population, ecology and distribution. Though the species may qualify for a higher category of Critically Endangered under A3ce, the level of uncertainty associated with a categorisation under criterion B is lower based on the availability of additional species-specific information in that justification.

This species has a restricted distribution in the Calamian Islands of the Philippines. Recent studies of the fossil record of East Borneo (Indonesia) have recorded this species from one locality dated from the late Miocene (~10 Ma) (Santodomingo et al. 2015); this fossil occurrence only suggests that the species has a long geological persistence in the region, but it cannot be taken as evidence for its modern occurrence in the entire region. The extended distribution of the species according to Veron et al. (2016) to the entire marine ecoregion of Palawan extending to the northeast coast of Borneo (Malaysia and Indonesia), should be taken with caution as it is only due to the format chosen by Corals of the World database to display their data using the standard polygons of Marine Ecoregions of the World (MEOWs, Spalding et al. 2007). Predicted records in the Northeast Sulawesi Marine Ecoregion based on habitat requirements (Veron et al. 2016) should be confirmed in future assessments of the species.

The Calamian Islands are located in the province of Palawan, Philippines and have an estimated area of 10,000 km². Coral reefs in these islands cover 222.7 km² (ADB 2014). This is a rare species that occurs in shallow reefs with calm waters protected from wave action. At this time, it is challenging to estimate the proportion of the reefs that comply with the habitat requirements of the species as detailed mapping of the reefs are not yet available. For this reason, the estimated extent of occurrence (EOO) for this species is evaluated under the broad estimation of the extent of the Calamian Islands (10,000 km²). The estimated area of occurrence (AOO) is less than 500 km² and greater than 10 km² based on the available coral reef area in the Calamians and the species' natural rarity. The number of locations is one based on the major threat from severe coral bleaching events which are increasing in frequency as climate change progresses and acts on the entirety of this species' range.

The depth range is 8-40 m, but the species is most common at 8-30 m (L. DeVantier pers. comm. 2024).

This species is rare (Veron et al. 2016). This species has only been found in the Calamian Islands (Palawan Province, Philippines). In general, the Philippine coral reefs have been on a decline since the first national assessment in the 1970s (Gomez and Alcala 1979), when only 30% of the reefs had more than 50% coral cover. Multiple stressors, including overfishing, destructive fishing methods, and siltation have contributed to the further decline of reefs in this region, where currently only 10% of the reefs had more than 50% coral cover (Licuanan et al. 2017; 2020). With increasing demand for fish, destructive fishing methods still in place, and further effects of climate change, the decline of coral reefs in the Philippines is expected to continue over the next few years (Kimura et al. 2018, Licuanan et al. 2017; 2019; Panga et al. 2021). The average loss of coral cover in the past two decades has been estimated to be around 11% and it is predicted to continue in decline over the next 20 years (Souter et al. 2021).

There is no species specific population information available for this species. However, there is evidence that overall coral reef habitat has declined, and this is used as a proxy for population decline for this species. Acropora species are particularly susceptible to bleaching, disease, and other threats (Hoogenboom et al. 2017). 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 2034 for SSP5-8.5 and by 2037 for SSP2-4.5 assuming no level of adaptation and by the year 2061 for SSP5-8.5 and by 2079 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 80% 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 8-40 m and is predominately found at depths greater than 10 m, generalized abundance is considered rare, overall population is 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 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 less than 25% since 1989, or over the past three generations.

This species occurs with Acropora derawanensis in shallow reef environments protected from wave action (Veron et al. 2016). 

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. Given that bleaching events in the Philippines have been reported for the years 1998 (Arceo et al. 2001), 2010 and 2016-2017 (Panga et al. 2021), and 2020 (Philippine Coral Bleaching Watch 2022), it is inferred that the species have not had enough time for recovery between bleaching events. The last bleaching event in the Calamian Islands was high to very high (4 to 5 in a scale 1 to 5) (Philippine Coral Bleaching Watch 2022). Thus, it is also estimated that populations of Acropora species, including this species, experienced high levels of decline in this region, and it is projected that populations may continue to decrease if bleaching events occur more frequently and more severely over the next two decades (Hughes et al. 2018).

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, Richards et al. 2021, Muir et al. 2021). In addition, climate change is predicted to lead to 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 catchments. Acropora species have, in general, high bleaching susceptibility as it has been observed in multiple assessments in the Great Barrier Reef (Muir et al. 2021, Richards et al. 2021).

This species has a branching growth form and may be susceptible to predation by crown-of-thorns seastar. Crown-of-thorns (COTS) (Acanthaster spp.) 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 consume 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. 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). 

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). Three diseases (white syndrome, black band disease and an unidentified syndrome) have been recorded in Acropora populations on the Great Barrier Reef (Willis et al. 2004). 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) and the geographical range of tropical coral diseases will extend (Vergés et al. 2019).

Ocean acidification represents a key threat to coral reefs by reducing the calcification rate of framework builders (Langdon et al. 2000). Acidification may act in synergy with global warming to increase bleaching response and decrease productivity in corals (Anthony et al. 2008). Ocean acidification has also been shown to negatively affect reproduction and multiple early developmental stages of corals that are critical to reef persistence and resilience (Albright 2011, Webster et al. 2013).

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. The severity of these combined threats to the global population of each individual 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). At this time, this species appears in only two trade import records in Japan (2007) and the United States (2014) as live specimens with commercial purpose (UNEP-WCMC 2021). 

Within the distribution of this species in the Calamian Islands, there are the following Protected Areas: 1. the “Calauit Island Game Preserve and Wildlife Sanctuary” with 37.17 km² but mostly covering terrestrial areas, 2. Sagrada Bogtong MPA, 3. Siete Pecados Marine Park; 4. Balisungan MPA; 5. Cuaming MPA; 6. Bintuang Sangat Marine Park; and 7. San Miguel MPA. Despite some efforts of conservation of coral reefs ecosystems and the species that live in, most of these MPAs in the Philippines contain only small portions of the reefs under their jurisdiction.

As very little information is available on its distribution, abundance, habitat preferences, and susceptibility to threats, it is highly recommended to perform further surveys of modern reefs to confirm the extension and status of this species in the region.

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.