This species is widespread and 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 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 less than 25% over the next three generations, or by 2050. It is listed as Least Concern.

This species occurs throughout most of the western Indian Ocean, the Red Sea, the Gulf of Aden, the Persian Gulf, central and northern Indian Ocean, the central Indo-Pacific, west, north and east Australia, Southeast Asia, Japan, the South China Sea, the oceanic West Pacific, the central and eastern Pacific, including Johnston Atoll, Pitcairn Island, Clipperton Island and the west coast of Mexico (Veron et al. 2016).

The depth range is 0-70 m, but the species primarily occurs from 0-40 m (Muir and Pichon 2019, L. DeVantier pers. comm. 2024).

This species is common (Veron et al. 2016, DeVantier and Turak 2017). In the Red Sea, this species is abundant in sheltered areas where huge colonies can cover over 75% of the substrate over thousands of square metres. 

On the crest and slope areas of the Great Barrier Reef have decreased by 45 and 15% respectively since 1995/1996 (Dietzel et al. 2020). At Kut Island, Thailand, the percentage cover of this species declined by 22% between 2007 (before bleaching) and 2012 (after bleaching) (Yeemin et al. 2013). In the Chagos Archipelago, percentage cover of this species declined by 42% between 1979 to 1999 (Sheppard 1999).

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. 2021). 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 2035 for SSP5-8.5 and by 2039 for SSP2-4.5 assuming no level of adaptation and by the year 2062 for SSP5-8.5 and by 2070 for SSP2-4.5 assuming 1°C of adaptation. For species where the onset of 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 1°C of adaptation is lower given this species is not primarily restricted to depths shallower than 30 m and is more resilient 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 less than 25% by the year 2050, 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 0-70 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 or does not recover well from bleaching or disease, has a low susceptibility to crown-of-thorns starfish, is more resilient 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 with Porites lobata and P. australiensis on back reef margins, lagoons and fringing reefs, generally to depths of 30 m. This species is commonly found from 1-15 m, with massive colonies at 3-5 m, in the South China Sea and Gulf of Siam (Titlyanov and Titlyanova 2002). In American Samoa, one colony is known to be 7 m tall and 41 m circumference (Brown et al. 2009). 

It thrives in oligotrophic, in turbid backreef and in coastal environments subjected to terrestrial runoff of sediments and nutrients (Scoffin et al. 1992, Veron 2000). Colonies may live for more than 1,000 years (Soong et al. 1999). This species grows in extreme mangrove lagoons on the Great Barrier Reef (Camp et al. 2019).

The age at first maturity of most reef-building corals is typically three to eight years (Wallace 1999). Based on this, we infer that the average age of mature individuals of this species is greater than eight years. 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 has a low susceptibility to bleaching (Burt et al. 2008, Grimsditch et al. 2010) and a moderate recovery potential (Buerger et al. 2015). It had a moderate bleaching level and moderate estimated mortality (10-40%) during a bleaching event in Palau (Bruno et al. 2001).

Porites species are generally considered to be relatively resistant to bleaching and suffer significantly less bleaching-induced mortality compared to other species (Gleason 1993, Hoegh-Guldberg and Salvat 1995, Baird and Marshall 2002, McClanahan and Maina 2003, Baker et al. 2008). Porites on the Great Barrier Reef were considered to be bleaching tolerant during the 2016 bleaching event as even at shallow depths of 5 m, less than 40% of the colonies bleached and mortality was  ̴ 5% (Frade et al. 2018). In the central Red Sea, the 2015-2016 third global bleaching event resulted in 30% of the cover of Porites spp. bleaching (Monroe et al. 2018). In the Gulf, colonies of this species were able to calcify even when exposed to summertime temperatures of 32.5°C, suggesting that they may have adapted to living in such extreme conditions there (Behbehani et al. 2019). Furthermore, when high temperatures were coupled with decreased pH, calcification was still observed, suggesting some potential resilience to climate change (Behbehani et al. 2019). This species suffered high mortality (̴ 40%) at 3 m due to bleaching in 1998 in southern Japan (van Woesik et al. 2004). Between 25 and 60% of massive Porites colonies exhibited moderate and severe signs of bleaching in Indonesia in 2010 (Guest et al. 2012). Rising temperatures have also been linked to reduced skeletal growth in this species from Thailand (Tanzil et al. 2009). Nevertheless, in the Maldives, massive species of Porites have been predicted to have a low total susceptibility to mass bleaching and a low relative extinction risk (Muir et al. 2017). Furthermore, this species is extremely widespread in tropical habitats and across a range of depths, providing a possible degree of resilience to threats relating to global warming.

This species in the South Yemen is strongly attacked and destroyed by a sponge belonging to the genus Clathria (Microciona) (Benzoni et al. 2008). This species appears to be a good bioindicator in monitoring studies of metal pollution (Jafarabadi et al. 2018). This species is very susceptible to pollution damage from chlorinated paraffins, which lower symbiodinium density (Jafarabadi et al. 2021). Land-based disturbance is a great threat to this species off the coast of Hainan, China (Roder et al. 2013, J. Crabbe pers. comm. 2022). 

Although Porites spp. are generally not favored by crown-of-thorns, Porites spp. ranked second in terms of the top-10 most favored genera for crown of thorns at Malapascua, Philippines (Kensington 2019). Similarly, although generic preference for Acropora and Pocillopora species was found in Moorea, to a lesser extent massive Porites spp. were also consumed (Kayal et al. 2011; 2012). There are also records dating back to the 1960s on the Great Barrier Reef of predation by adult crown-of-thorn on several species of massive Porites, suggesting that this species is not an uncommon prey species (DeVantier and Done 2007). For example, in the 1980s, it was estimated that a quarter of the massive Porites spp. corals surveyed at five reefs on the Great Barrier Reef had been killed by crown-of-thorns, with recovery estimates in excess of 50 years (Done 1988).  

Numerous studies have reported that crown-of-thorns have certain food preferences, mainly Acropora and Montipora, while rarely feeding on other taxa such as Porites, and massive colonies of Porites spp. such as this species are no exception (Brauer et al. 1970, Pearson 1973, Moran 1986, Birkeland and Lucas 1990, Pratchett et al. 2014). Juvenile crown of thorns also showed no preference for this species based on food-choice experiments (Johansson et al. 2016). Crown-of-thorns starfish (COTS) (Acanthaster planci) are found throughout the Pacific and Indian Oceans, and the Red Sea (Campbell and Ormond 1970). 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. Increased breakouts of COTS have 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.

At least five different types of disease have been recorded in this species including white syndrome, ulcerative white spot, focal bleaching, non-focal bleaching and pink spot (Priess et al. 2000, Ravindran et al. 2001, Ravindran and Raghukumar 2006, Putchim et al. 2012, Lawrence et al. 2015, Séré et al. 2016). In the Andaman Sea, diseased individuals of this species occurred at 16 out of the 17 sites assessed, prior to a bleaching episode in 2010 (Putchim et al. 2012). After the bleaching event, the frequency of ulcerative white spot was higher, but other diseases were still categorized as being rare. A crustose coralline alga belonging to the genus Hydrolithon has also been known to overgrow and kill this species. At Saint-Leu, Reunion Island, as many as 66% of the colonies of this species were infested with it (Séré et al 2012).  

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 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). 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). In the western Indian Ocean, during the 2016 bleaching event, there was an approximate 20% decline in coral cover (Gudka et al. 2018).

Coral disease has emerged as a serious threat to coral reefs worldwide with increases in numbers of diseases, coral species affected, and geographic extent (Ward et al. 2004, Sutherland et al. 2004, Sokolow et al. 2009). Outbreaks of coral diseases have damaged coral reefs worldwide with the most widespread, virulent, and longest running coral disease outbreak currently occurring on the Florida Reef Tract and throughout the Caribbean. The disease, stony coral tissue loss disease, has been ongoing since 2014 (Precht et al. 2016) and has devastated affected reefs along Florida (Walton et al. 2018, Williams et al. 2021) and throughout the Caribbean (Alvarez-Filip et al. 2019, Kramer et al. 2019). Numerous disease outbreaks have also occurred in the Indo-Pacific (Willis et al. 2004, Aeby et al. 2011; 2016), Indian Ocean (Raj et al. 2016) and Persian Gulf (Howells et al. 2020). Escalating anthropogenic stressors combined with the threats associated with global climate change of increases in coral disease, frequency and duration of coral bleaching and ocean acidification place 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 (Wells 2009), chemical fishing (Hughes et al. 2013, Madeira et al. 2020), pollution from agriculture and industry (Bruno et al. 2003, Zamani et al. 2020), domestic pollution (Nguyen et al. 2013), sedimentation (Babcock and Davies 1991, Erftemeijer et al. 2012, Hughes et al. 2013, 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 but all threats covered above occur within the range of this species.

The first record of Porites lutea in the CITES Trade database was in 1992, when 20 wild harvested specimens were imported into Great Britain (CITES 2020). Subsequent trade has been erratic with a peak in 2011 of  3,218 specimens harvested from the wild globally, according to export data. Since 2010, the annual median number of wild harvested specimens exported globally was 221. However, in many instances (56%), the wild harvested specimens traded are simply referred to as Porites spp. and therefore it is probable that specimens of P. lutea may comprise some of these. In terms of trade in Porites species in general, there has been a steady increase since 1989 with the number of wild harvested specimens peaking at 136,146 in 2013, according to export data. Since 2010, the median global number of annual exports has been 81,702 wild harvested specimens. Porites was also the most illegally imported genus of coral into the USA between 2003-2012 according to LEMIS data (Petrossian et al. 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). Parts of this species distribution overlaps with several Marine Protected Areas.

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).