Recent genetic evidence suggests this species forms a species complex with Millepora complanata, M. alcicornis and M. complanata can only be differentiated morphologically, but not genetically (Ruiz-Ramos et al. 2014).

This widely distributed, coral species is 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 an inferred decline of <20% 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, this species is inferred to decline by at least 53% 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 A3c. The change in status from the previous assessment reflects updated declines calculated from improved data on modelled coral cover loss in the Caribbean and Brazil from 1978 to 2019 and projected date of annual severe bleaching, along with improved knowledge of species traits.

This species is distributed in the western Atlantic from Bermuda, the southwest Gulf of Mexico off Veracruz, Mexico (Tunnell 1988) and the northeastern Yucatan Peninsula to western Florida (Humann and DeLoach 2006) and Cuba southward throughout throughout the Caribbean to Trinidad and Tobago. It also occurs from northern Brazil (Hetzel et al. 1994) to São Paulo State, including Fernando de Noronha (F. Nunes and M. Kitahara pers. comm. 2021). In the eastern Atlantic, it occurs in the Cape Verde Islands (Laborel 1974), Canary Islands, Ascension Island (Hoeksema et al. 2014), Madeira and the Gulf of Guinea (F. Nunes pers. comm. 2021). A species in Bermuda is also referred to by this name but is very different from the rest of the Caribbean (Sterrer and Schoepfer-Sterrer 1986); however, molecular data shows that it is not genetically differentiated to individuals elsewhere in the Caribbean (de Souza et al. 2017). The depth range is 1–40 m.

This is the most common of the Millepora species in the Caribbean. Colonies may be locally abundant but usually comprise <10% of the overall surface cover. Widespread bleaching and mortality of millepores has been reported during mass bleaching events that have affected many coral reefs. This species is the most severely affected of the Millepora species and a recent study in Brazil, where it is one of the dominant reef-building corals, holds a functionally important role as the only branching species on reef crests in the Southwest Atlantic, recorded catastrophic declines in the local/regional populations (Duarte et al. 2020). That said, Millepora species are often the first to recover after bleaching events. Harmful effects of oil spills, chronic oil pollution and oil-spill detergents have been widely reported for millepores (Lewis 2006).

New records of this species are being reported (from the Canary Islands (28N) and even in the colder waters of Madeira Island (32N) for example) and genetic analysis suggested multiple independent long distance dispersal events had occurred from the Caribbean area (Wirtz and Zilberberg 2019). This may allow this species to have resiliency as high dispersal rates means it has the possibility of increasing its range as the climate changes. This species usually exhibits low mortality rates after bleaching (<8% in a Brazilian study for example) (Miranda et al. 2013, Teixeira et al. 2019). Recent evidence suggests a substantial increase in mortality with highs of ∼84 and 89% in Virada de Fora and Pedra do Silva, respectively, and ∼ 43% in Coroa Vermelha (Duarte 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 (Hughes et al. 2018). 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 corals 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 poorly understood and varies by species and locality (van Hooidonk et al. 2013, Logan et al. 2014). 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 2030 for SSP5-8.5 and by 2034 for SSP2-4.5 assuming no level of adaptation and by the year 2059 for SSP5-8.5 and by 2072 for SSP2-4.5 assuming 1°C of adaptation. We inferred that the uncertainty associated with the estimate of population decline based on no level of adaptation is lower given this species is not primarily restricted to depths shallower than 30 m and is highly susceptible to bleaching. Furthermore, since the primary depth range of this species is 10-40 m, population decline was estimated over 67% of its depth range (e.g. 10–30m). 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 understanding that a coral species with shallow depth preferences is more frequently exposed to extreme temperatures and is expected to decline at a faster global rate than species that also or primarily occur in deeper, cooler waters (Riegl and Piller 2003). 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 Atlantic, changes in aragonite saturation are expected to be most severe in the Mediterranean Sea (Flecha et al. 2015). Therefore, this species is inferred to experience a projected global level decline of at least 53% 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 was 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 five Caribbean and three Brazil subregions (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’ entire range in the Caribbean and Brazil. Sufficient estimates of coral cover loss are not available for the part of the range extending beyond the Caribbean and Brazil (i.e. eastern Atlantic); therefore, for the purposes of estimating global level decline per this Red List assessment, an assumption of 'no change' was applied for that area.

To inform the choice of the best (i.e., lowest level of uncertainty) out of the three percentile declines, we considered nine species-specific traits related to vulnerability to coral cover loss. Given this species’ depth range is 1–40 m and is predominately found at depths greater than 10 m, generalized abundance is considered common, overall population is not highly fragmented,  does not occur off-reef, is moderately susceptible to disease, does recover well from bleaching or disease and is highly susceptible to bleaching, it is overall inferred 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 global level decline of <20% since 1989, or over the past three generations.

This species is the only Caribbean fire coral that can commonly be found deeper than 10 m, and is relatively uncommon in shallow surge zones (Humann and DeLoach 2006). Millepora species are generally found in inshore areas characterized by turbidity, and exhibit a tolerance for siltation. They often occur in clear offshore sites (Lovell pers. comm. 2008), commonly encrusting and overgrowing other species (W. Precht pers. comm. 2008).

Fire corals are gonochoric broadcast spawners that reproduce sexually by producing medusoids and planula larvae. Despite their ecological importance to the ecosystem functioning of coral reefs, Millepora hydrocorals have been relatively understudied and information regarding their reproduction and dispersal patterns remain scarce. Millepora colonies become mature during the spring or summer. Sexual reproduction period is usually correlated with the increase of sea water temperature however some studies suggest reproduction may occur throughout the year (de Weerdt 1984). Millepora larvae have their symbiotic algae present in the planula larvae i.e. are zooxanthellate and therefore have the potential to live for several weeks before settlement. This feature is certainly a character to keep in mind to explain the large distribution in all oceans. Reproductive output (ampulla density) is variable from colony to colony. Amaral et al. (2008) found an average of 10 ampullae/cm² for Millepora species occurring on Brazilian reefs.

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 is moderately susceptible to bleaching. Species of Millepora are generally susceptible to bleaching and are some of the first hard corals to bleach but are resilient, being some of the first to recruit after the bleaching. This species has been reduced from historical baselines, but probably not much more than 10% except locally on some reefs. This species has been affected in past bleaching events but appears to recover more rapidly than most scleractinian species (W. Precht pers. comm. 2008).

In Brazil, this species has been collected for the souvenir and aquarium trade, with some locations suffering high pressure of exploitation (Leão 1994, Gasparini et al. 2005), which could, if unregulated, lead to population decline. Artisanal fishing has also had an impact on this species in some areas of Brazil, where fishing nets become entangled and break colonies (C. Ferreira pers. comm. 2020).

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. In addition to global climate change, corals are also threatened by disease, and a number of localized threats. 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 coral species is not known.

Coral disease has emerged as a serious threat to coral reefs worldwide and is 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 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), Marshall Islands (Jacobson 2006) and the northwestern Hawaiian Islands (Aeby et al. 2006). Increased coral disease levels on the Great Barrier Reef were correlated with increased ocean temperatures (Boyett et al. 2007) supporting the prediction that disease levels will be increasing with higher sea surface temperatures.

This genus is generally not found in aquarium trade, but is sometimes collected for curio and jewelry trade. In 2005, 117 pieces of live Millepora alcicornis were exported for the aquarium and curio trade (E. Wood pers. comm. 2008). There are no records in the CITES database of exports of non-scleractinians by weight. Atlantic reef corals are not part of the international trade for the aquarium industry (B. Hoeksema pers. comm. 2022).

This non-scleractinian coral is listed under Appendix I and II of CITES. Parts of this species' distribution overlaps with several marine protected areas (MPAs).

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.

Having timely access to national-level trade data for CITES analysis reports would be valuable for monitoring trends this species. The species is targeted by collectors for the aquarium trade and fisheries management is required for the species, e.g., MPAs, quotas, size limits, etc. Consideration of the suitability of species for aquaria should also be included as part of fisheries management, and population surveys should be carried out to monitor the effects of harvesting.