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 is found in the Red Sea, Gulf of Aden, Madagascar, Gulf of Kachchh in India (Kumar et al. 2017, Marimuthu et al. 2018), southwest, northwest and central Indian Ocean, central Indo-Pacific, north and west Australia, Southeast Asia, Japan (Lien et al. 2012, Reimer et al. 2020), East and South China Sea, eastern Australia, the Marshall Islands (Richards and Beger 2013), oceanic West Pacific, and Southwest Pacific (Veron et al. 2016). It has also been confirmed from northern Vietnam and Taiwan (Huang et al. 2015).

The depth range is 2-40 m (Bouchon 1981, Sheppard and Sheppard 1991).

This species is common (Veron et al. 2016, DeVantier and Turak 2017). It can be locally abundant (Wallace and Lovell 1977). It was described as being common at Ishigaki Island, southwestern Japan (Fujioka 1998) and as being very common in Vanuatu (Veron 1990). It occurred at four of the seven sites investigated by Latypov (2011) in Vietnam and at all three sites investigated in Oman (Sheppard and Salm 1988). It was described as being rare at Reunion and Mauritius (Faure 1977), and was also rare at Guam (Raymundo et al. 2019). Poritidae 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).

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 2034 for SSP5-8.5 and by 2038 for SSP2-4.5 assuming no level of adaptation and by the year 2061 for SSP5-8.5 and by 2069 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 moderately 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 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 2-40 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 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 moderately 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 is found in intertidal and subtidal reef environments, especially lagoons, generally to depths of 30 m (Peach and Hoegh-Guldberg 1999, Veron et al. 2016). At Reunion and Mauritius it occurred on outer slopes at depths of 5-15 m (Faure 1977). It is also known to occur in tidal pools and on shallow reef slopes at Kikai Island, Japan, and on outer reef flats and upper reef slopes at 2-10 m at Ishigaki Island, Japan (Fujioka 1998, Sugihara et al. 1999). In the Arabian Sea, it is found in non-reef substrates in areas subjected to cold-water upwelling between 2-5 m deep (Sheppard and Sheppard 1991). 

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 was found to be moderately susceptible to bleaching in Kikai Island, Japan, though more so than several species of Porites from the same family (Sugihara et al. 1999). 

Goniopora is in the top-five genera in the aquarium trade (Wabnitz et al. 2003). Gonipora spp. were also one of the most popular imports of coral into the USA between 2003 and 2012 (Petrossian et al. 2020). Goniopora species were considered to be moderately vulnerable to overexploitation in the Queensland Coral Fishery (Roelofs 2018). 

Goniopora species are moderately susceptible to bleaching in the western Indian Ocean (McClanahan et al. 2005; 2007) and exhibit relatively low mortality (McClanahan et al. 2001, McClanahan 2004). They have a low bleaching susceptibility on the Great Barrier Reef as well (Marshall and Baird 2000, McClanahan et al. 2004). Goniopora spp. exhibited mild bleaching in Palau during the 1998 global bleaching event (Bruno et al. 2001). During the 2016 global bleaching event, Gonipora was one of the genera which displayed the highest resistance to bleaching in the Maldives (Ibrahim et al. 2017) and has been predicted to have a low total susceptibility to mass bleaching and a low relative extinction risk there (Muir et al. 2017). Goniopora also bleached extensively in 2010 at Palk Bay, India (Ravindran et al. 2012) and exhibited the highest levels of bleaching (80% of its cover) out of all genera surveyed at Thuwal, Saudi Arabia, Red Sea during the 2016 global bleaching event (Monroe et al. 2018).   

At least three types of disease have been recorded in Goniopora (Haapkyla et al. 2009, Montano et al. 2015, Aeby et al. 2020). Goniopora had the highest disease prevalence of all coral genera investigated in the Maldives, and was found to be particularly susceptible to black-band disease (Montano et al. 2012). Acropora, Goniopora, Hydnophora and Porites were the most susceptible coral genera to disease in Reunion (Séré et al. 2015). Crown-of-thorns starfish have been observed preying upon Goniopora lobata, but only after several other taxa had been consumed (Pratchett 2007). Campbell and Ormond (1970) and Randall (1972) also remarked that Goniopora was resistant to attack and not favoured. Goniopora did also not rank within the top-10 genera being fed on at Malapascua, Philippines, but it was fed on (Kensington 2019).

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, 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 consume 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 (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 recently reported from the Great Barrier Reef (Willis et al. 2004, Haapkyla et al. 2010). 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) 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 (Wells 2009), 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. 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.

The genus Goniopora was first recorded in the CITES trade database in 1989 with 120 wild harvested specimens exported globally (CITES 2020). Almost all (99.8%) of the wild harvested specimens were reported as Goniopora spp. in the CITES trade database, largely due to taxonomic species level uncertainty. Exports rose quickly and peaked in 2004 with a total of 1,054,426 specimens harvested from the wild. Since 2010, the median number of Goniopora spp. harvested from the wild was 215,897 per annum. First records specifically of Goniopora tenuidens in the CITES trade database appeared in 1997 with 40 specimens exported. Exports subsequently increased and peaked at 4,021 specimens in 2016.

All stony corals are listed on CITES Appendix II. All stony corals (Scleractinia) fall under Annex B of the European Union Wildlife Trade Regulations, 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. 

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, which now includes Aichi Biodiversity Target 11, calling for 10% of coastal and marine areas to be conserved by 2020. In 2016, the IUCN World Conservation Congress agreed upon a target of >30% global marine protection by 2030.

It is crucial that global warming is constrained well below 2°C (the goals of the Paris Agreement).