This common species has a relatively small range. 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, distribution and bleaching vulnerability, this species is suspected to decline by at least 71% 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 is distributed in the Red Sea, Socotra, Mayotte and Chagos (Veron et al. 2016).

The depth range is 0-30 m, but the species primarily occurs from 8-20 m (L. DeVantier pers. comm. 2024).

This species is common (DeVantier and Turak 2017).

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 2031 for SSP5-8.5 and by 2033 for SSP2-4.5 assuming no level of adaptation and by the year 2056 for SSP5-8.5 and by 2063 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 71% 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 0-20 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 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 moderately susceptible to threats related to reef degradation. Therefore, past decline was inferred from the 50% 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 in shallow, tropical reef environments. It occurs in most reef environments (Veron et al. 2016).

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.

Members of this genus have a low resistance and low tolerance to bleaching and disease, and are slow to recover. Montipora species have a high susceptibility to thermally induced bleaching and have high subsequent mortality (Marshall and Baird 2000, McClanahan 2004, McClanahan et al. 2007, Hughes et al. 2018, Dietzel et al. 2020). 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 central Red Sea, the 2015-2016 global bleaching event resulted in 35% of the cover of Montipora spp. bleaching (Monroe et al. 2018). In the Maldives, the majority of Montipora species have been predicted to have a very high total susceptibility to mass bleaching and a very high relative extinction risk (Muir et al. 2017). Similarly, on the Great Barrier Reef, Montipora was considered to be bleaching sensitive as there was major bleaching (̴ 60%) and even mortality (̴ 10%) to colonies at 40 m depth (Frade et al. 2018).

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 range of this species, such as coral bleaching from thermal stress (Furby et al. 2013, Monroe et al. 2018), disease (Hadaidi et al. 2018), predation by crown of thorns (Riegl et al. 2012), pollution (Ziegler et al. 2016; 2019), sedimentation (Mohammed and Mohamed 2005) as well as dredging and land filling (Hilmi et al. 2012). 

The northern Red Sea from Rabigh to the Sinai Peninsula escaped most of the bleaching and the mortality of the last couple of decades (Osman et al. 2018). However, although northern Red Sea corals exhibit remarkably high thermal resistance, the rapidly rising incidence of marine heat waves of high intensity indicates this region may not remain a thermal refuge much longer (Genevier et al. 2019). The central Red Sea has experienced several bleaching incidences and the frequency and severity of bleaching events since 1998 on nearshore reefs is unprecedented over the past century (DeCarlo 2020). The southern Red Sea has not escaped bleaching events and the Gulf of Aqaba and the Hurghada regions are affected by numerous direct impacts from coastal development and industry (Barakat et al. 2015). The Red Sea is extremely vulnerable to environmental threats posed by large coastal populations, overfishing, pollution, and coastal development and nearly 60% of the reefs in the Red Sea are at risk from landfilling and dredging, port activities, sewage and other pollution, and tourism (Burke et al. 2011).

Montipora spp. may be particularly susceptible to disease (Haapkyla et al. 2013). At least five different diseases have been recorded in Montipora spp. (Bourne 2005, Vargas-Ángel 2009, Sato et al. 2010, Burns and Takabayashi 2011, Ushijima et al. 2012, Johan et al. 2014, Aeby et al. 2016, Chen et al. 2017). Disease in Montipora spp. from the Red Sea is known to be prevalent (Winkler et al. 2004, Barneah et al. 2007) and there have been recent outbreaks of blackband disease which have affected Montipora spp. there (Hadaidi et al. 2018). However, in terms of microbial-based diseases, Red Sea corals display many typical disorders, including white syndromes, skeletal eroding band, black band disease, and growth anomalies, but these are considered to be rare within Red Sea waters (Neave et al. 2019). 

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). 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).  

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 (Green and Bruckner 2000, Porter et al. 2001, Sutherland et al. 2004, Weil 2004). 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 (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, Chen et al. 2017, Shore-Maggio et al. 2018) which have placed coral reefs in the Indo-Pacific at high risk of collapse.

Crown-of-thorns starfish (COTS) (Acanthaster planci) are found throughout the Pacific and Indian Oceans, and the Red Sea (Campbell and Ormond 1970). 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 destry 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.

Other more localized threats include disturbance by 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, storms (Babcock and Davies 1991, Erftemeijer et al. 2012, Cunning et al. 2019), and human recreation and tourism activities (Lamb et al. 2014).

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).

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).