This species is considered a synonym of Acropora multiacuta by Veron and Hodgson (1989) and Wallace (1999) and is not included in the revision of Acropora species by Wallace et al. (2012). It is valid according to the WoRMS online database (accessed January 2023) and Veron et al. (2016).

There is high taxonomic uncertainty associated with this species. Little is known on its distribution, population, ecology and potential threats; therefore, it is listed as Data Deficient.

If valid, this species occurs in Indonesia, the Philippines and Timor-Leste (Veron et al. 2016). A report of this species from the Andaman Islands (Sarkar and Ghosh 2013, Mondal et al. 2014) requires confirmation.

The depth range is 2-20 m.

This species is rare (Veron et al. 2016, DeVantier and Turak 2017). Of the 700 samples collected from 13 localities surveyed during the Snellius Expedition (1929-30) in eastern Indonesia and the Philippines, this species comprised 3 samples from only 2 localities (van der Meij and Visser 2011). It was recorded at 11 out of 47 sites in the Togean and Banggai Islands of Indonesia (Allen and McKenna 2001). In a survey of the relative abundance of reef building corals across the Indo-Pacific, it only occurred at 0.26% of the 3,075 sites surveyed (DeVantier and Turak 2017). A single colony was observed at Lizard Island, Great Barrier Reef in 2011, but it has not been observed since (Z. Richards pers. comm. 2021). 

There is no species-specific population information available for this species. 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) and there is evidence that overall coral reef habitat has declined globally.

This species occurs in shallow tropical reef environments (Veron 2000) on sheltered fringing reefs and offshore barrier reefs, platform reefs, pinnacles and coral cays and tends to avoid shallow atoll reefs and exposed reefs on the windward side of large islands and peninsulas (Allen and McKenna 2001). 

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.

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. 2018). 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 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. 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; Dietzel et al. 2020). Marine heatwaves have resulted in widespread coral bleaching and mortality (Hughes et al. 2017). 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.

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 (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). 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, Pollock et al. 2011, Ruiz-Moreno et al. 2012, Lamb et al. 2014, Vega Thurber 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 (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 to corals listed above are known to occur within the distribution range of this species, such as coral bleaching from thermal stress (Lalamentic 2013, Faliciano et al. 2018), disease (Kaczmarsky 2006), predation by crown of thorns (Verdadero et al. 2017), pollution (Verdadero et al. 2017), sedimentation (Verdadero et al. 2017), harvesting for the aquarium trade (Ferse et al. 2012) as well as destructive fishing practices (Verdadero et al. 2017). Reefs, for example, in northern Palawan, Philippines are considered to be in poor health (Verdadero 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 (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.

Taxonomic research is needed.