The taxonomic status of H. bermudensis is unclear and it was listed as Data Deficient (DD) in the previous assessment (Short et al., 2011). This species is endemic to Bermuda (Supplementary Figure S1), with the original description (Den Hartog, 1964) being the only published record of collected specimens of this morphological form within the genus (in 1913 at Gibbet Island and in 1922 at Walsingham Bay). It is possible that the genetic character of H. bermudensis could be determined from preserved type specimens (Ferrer-Gallego and Boisset, 2020). Inferred population reduction is likely ongoing, based on general decline of meadows across the Bermuda Platform (Fourqurean et al., 2019); therefore, it is listed as Critically Endangered (CR, Table 1).

The taxonomic status of H. ciliata is unclear and it was listed as Data Deficient (DD) in the previous assessment by Short et al. (2011). Plants with leaf tip morphology as described for H. ciliata have only been reported at Taboga Island, Panama (Den Hartog, 1970) (Supplementary Figure S1). No specimens of this seagrass have been seen for 20 years, despite concerted efforts to locate it and the only previously known location has been destroyed (F. Short and B. Sullivan personal communication). There have been no new reports of this species for the 2007–2020 time period. This species is listed as Critically Endangered (CR, Table 1). It may possibly be extinct because of its small extent of occurrence and area of occupancy (one location).

The taxonomic status of H. emarginata is unclear and it was listed as Data Deficient (DD) in the previous assessment (Short et al., 2011). Plants with leaf tips corresponding to H. emarginata have only been reported in Brazil, and stable populations have been found at 10 locations (Supplementary Figure S1). While new populations have been identified since the last assessment, these likely represent an increased effort and not a range expansion of the species. Presently, this species is found in a few fragmented populations and it is therefore classified as Near Threatened (NT, Table 1).

In its previous assessment, the clover grass H. baillonii was only known from approximately seven locations (Short et al., 2011). This is a small species that often grows in turbid and shallow waters with restricted visibility, and populations are often small and unstable, which is why it is potentially underreported. Its presence in shallow turbid estuaries, which are increasingly under human pressure, merit concern over its future distribution. Even though it has been reported at new sites since 2007 (Supplementary Figure S1), the area of occupancy (AOO) is still estimated to be less than 2,000 km² as reported in the previous assessment by Short et al. (2011), and in congruence with this assessment, it remains listed as Vulnerable (VU, Table 1).

Commonly known as star grass, H. engelmannii, a small opportunistic seagrass, usually has a patchy distribution in either shallow habitats (< 10m), or in deeper habitats (10 - >20 m) as sparse to dense meadows in the Gulf of Mexico, and Atlantic coastal waters of Cuba and US (Supplementary Figure S1). In shallow waters, various populations have declined or disappeared, and the deeper waters remain understudied. Because of reported ongoing population decline in shallow near-shore waters, this species remained Near Threatened (NT, Table 1). More attention on status and trends for this diminutive, patchily distributed species will be required in the coming years to better understand its extinction risk, including renewed efforts to map and understand the role of deeper, offshore populations.

This is the first assessment of this species, which was not included in Short et al. (2011), because at that time it was considered a synonym of R. maritima. However, its species status has since been recognized (Triest et al., 2018) and it was first reported for this bioregion in 1825. R. didyma was originally described from the Caribbean but is presently only known from a few locations in the Yucatan Peninsula, Mexico (Novelo, 1991) and Venezuela (Villarreal et al., 2014) (Supplementary Figure S1). It inhabits small waterbodies with clear, hypersaline water, generally near mangroves; and it has recently been found also in sinkholes with fresh to brackish water (Brigitta I. van Tussenbroek and Duarte Frade, unpublished data). R. didyma has a small area of occupancy, and occurs in 13–16 subpopulations that are likely to be severely fragmented given their disperse distribution and the large distance between them. The single known site in Puerto Rico, Guánica Lagoon, was drained in 1955 (Viqueira Rios et al., 2012), and there are no recent records from other Caribbean islands, where mangrove-dominated wetlands face continuing degradation, such as Saint Barthelemy and Guadeloupe (Ellison and Farnsworth, 1996; Jadot, 2016). In this first assessment the species is therefore categorized as Vulnerable (VU, Table 1).

This is the first assessment of R. mexicana as it was split from R. maritima based on chromosome counts and morphological evidence (Den Hartog et al., 2016), and is supported by recent molecular microsatellite and chloroplast haplotype data (Triest et al., 2018). Although at present it is currently known only from the Yucatan Peninsula and adjoining parts of Mexico (Supplementary Figure S1), it likely occurs more widely. R. mexicana is assessed here for the first time and occupies a wide range of habitats, apparently being persistent in eutrophic locations. The species is therefore assessed as Least Concern (LC, Table 1).

Manatee grass S. filiforme is a widespread and abundant species throughout its range (Supplementary Figure S1), often found mixed with other seagrass species. But it also forms dense and extensive monospecific meadows. Its overall spatial cover appears stable, and it is increasing in abundance in some places (Van Tussenbroek et al., 2014). Such gains may balance losses elsewhere. The species has a great capacity for recovery once conditions improve. In congruence with the previous assessment for this species (Short et al., 2011) it remains of Least Concern (LC, Table 1).

Turtle grass T. testudinum, is the most robust species in the Tropical Atlantic Bioregion, where it forms extensive, dense and stable seagrass beds, and is considered to be the most important habitat-forming seagrass species throughout a large part of the bioregion (Supplementary Figure S1). The previous Red List Status LC (Least Concern, Short et al., 2011) of this species has been changed to NT (Near Threatened) due to localized threats throughout the region causing declines in condition and loss of many populations since the last assessment in 2007 (Table 1). Localized threats in the region included those driven by coastal urbanization and pressure from nutrient enrichment (Van Tussenbroek et al., 2014; Hall et al., 2016), algal blooms (e.g., holopelagic Sargassum spp., Van Tussenbroek et al., 2017), overgrazing by green sea turtles (Fourqurean et al., 2019) and competitive exclusion by the non-native seagrass H. stipulacea (Winters et al., 2020). These local pressures are all expected to increase, and in synergy they may result in unexpected collapse of stable T. testudinum dominated meadows (Van Tussenbroek et al., 2014). For these reasons the species has now changed category from Least Concern to Near Threatened (NT) under criterion A (population reduction, Table 1).

C. nodosa is found throughout the Mediterranean Sea, extending northwards into the Atlantic Ocean to mid-Portugal and southwards to Madeira, the Canary Islands, as well as the African continental coasts of Mauritania, Senegal, The Gambia and the archipelago of Cabo Verde (but see discussion in Creed et al., 2016). The Tropical Atlantic Bioregion contains the southernmost populations of the species (Supplementary Figure S1), which are very genetically distinct from the Mediterranean populations (Alberto et al., 2008). The overall population in this bioregion is thought to be presently stable, but ecological niche models predicted that the entire distribution area in the Tropical Atlantic Bioregion may be lost by 2050 under the RCP 2.6 and RCP 8.5 climate change scenarios and considering sea level rise (Chefaoui et al., 2018; 2021). The species is therefore listed as Vulnerable (VU) for this bioregion under criterion E (quantitative extinction analysis; Table 1).

The taxonomic status of H. beaudettei is unclear and it was listed as Data Deficient (DD) in the previous assessment (Short et al., 2011). Since then, the known range of the species has been expanded based on recent and previously published records (Supplementary Figure S1), mostly correcting past misidentifications. This is not a true species range expansion, yet illustrative of current taxonomic uncertainty. As this species has a wider distribution than reported in the previous assessment period (Short et al., 2011), and extending into Bioregion 1 (Temperate North Atlantic), it is now listed as Least Concern (LC, Table 1).

Commonly referred to as shoal grass, H. wrightii is a widespread and locally abundant species in the Caribbean Sea, Gulf of Mexico and Eastern Tropical Pacific (Supplementary Figure S1), where it is often found together with other seagrass species such as Thalassia testudinum, Syringodium filiforme, or Halophila spp., but it can also form dense and extensive monospecific meadows in habitats considered challenging for other major meadow forming species such as T. testudinum or S. filiforme. It is the dominant species in Brazil and the tropical coast of western Africa from Mauritania to Sierra Leone as well as Cabo Verde, São Tomé e Príncipe, and Angola. It extends into neighboring seagrass bioregions, being, Bioregion 1 (Temperate North Atlantic), Bioregion 4 (Temperate North Pacific) and Bioregion 6 (Temperate Southern Oceans) (Supplementary Figure S1). The overall population trend for this species is stable, and possibly increasing in some parts of its range, such as in Brazil, due to variations in water temperature and salinity (Copertino et al., 2016). Given its wide distribution and abundance, as well as its resilience to multiple stressors, in congruence with the previous assessment for this species (Short et al., 2011), in this regional assessment it remains of Least Concern (LC, Table 1).

The paddle grass H. decipiens has a tropical and subtropical, circumglobal distribution and can be locally abundant. In deeper waters, H. decipiens has few major threats; however, coastal development, reduced water quality (Short et al., 2011), invasive species, overgrazing, storms and hurricanes can affect seagrass beds in shallower and some deeper (>20m) areas. The population trend in the Tropical Atlantic Bioregion is stable, with range expansion such as in Brazil (Gorman et al., 2016) and eastern Florida, USA (Virnstein and Hall, 2009) (Supplementary Figure S1). In congruence with the previous assessment for this species (Short et al., 2011), H. decipiens remains of Least Concern (LC, Table 1) in this regional assessment.

R. maritima (sensu lato, see Discussion) is a colonizing seagrass that tolerates a wide range of environmental conditions, is capable of fast growth and reproduction (Strazisar et al., 2015), and has a wide distribution, extending into bioregions north and south of the tropical Atlantic (Short et al., 2007). The general population trend in the bioregion is stable or increasing (e.g. Martínez-Garrido et al., 2017), with R. maritima showing resilience to large disturbance events (hurricanes, oil spills), and occasionally replacing other seagrasses in the northern Gulf of Mexico following disturbance (Cho et al., 2009, 2017). Although parts of the region (West Africa and the Caribbean) remain poorly studied, the species is tolerant of disturbance, and is still widely distributed and locally common in the bioregion (Supplementary Figure S1). It is therefore listed as Least Concern in the Tropical Atlantic Bioregion (LC, Table 1).

Worldwide, Z. noltei occurs in the eastern Atlantic as well as the Baltic, Mediterranean, Black, Caspian and Aral Seas. The Tropical Atlantic Bioregion contains populations of this species at their southernmost limits of distribution, with large areas in Mauritania and smaller sites (~ 40 km diameter) in Senegal Saloum Delta (Sidi Cheikh et al., 2023) (Supplementary Figure S1). There, it is predicted to be negatively affected in the future from ongoing climate change, especially extreme climate events and sea-level rise as the entire distribution in the Tropical Atlantic Bioregion may be lost by 2050 under the RCP 2.6 and RCP 8.5 climate change scenarios and considering sea-level rise (Chefaoui et al., 2021). Large-scale mortality events have already been reported in Mauritania (De Fouw et al., 2016). The species is therefore listed as Vulnerable (VU, Table 1) for this bioregion, under criterion E (quantitative extinction analysis).

H. ovalis is wide-spread in the Indo-Pacific and Red Sea, and it is an introduced species in Florida and the Caribbean (Short et al., 2011) (Supplementary Figure S1). It is a rapidly growing opportunistic species with high shoot turnover rate. The Florida subpopulation was previously described as Halophila johnsonii (Eiseman and McMillan, 1980), an endemic species known only from portions of the Indian River Lagoon (Virnstein and Hall, 2009) south to Virginia Key (Gelber et al., 2000; Christian and Hall, 2012), but is now recognized as a synonym of H. ovalis subsp. ovalis (Waycott et al., 2021). Although H. ovalis populations are sparse and fragmented; the geographical range is gradually expanding in the Caribbean (Waycott et al., 2021). As a non-native species it is listed as Not Applicable (NA, Table 1) in this regional assessment.

H. stipulacea is native in the Indo-Pacific Ocean and Red Sea and is an introduced species in the Caribbean, where it is rapidly spreading, colonizing the Caribbean islands and mainland Venezuela (Vera et al., 2014; Willette et al., 2014) (Supplementary Figure S1). This small species has a high shoot turnover rate, and is therefore a rapidly growing species. At places it poses a threat to native species, but in others it does not (Winters et al., 2020). More recently in 2024, this species was also reported on the mainland of the United States (Key Biscayne, Florida) for the first time (Campbell et al., 2025). As a non-native species H. stipulacea in this regional assessment is listed as Not Applicable (NA, Table 1).

Various seagrass species share similar morphological vegetative characteristics, making it difficult to distinguish between them based on vegetative morphology alone, which is further hampered by reduced reproductive parts in adaptation to underwater fertilization (Kuo and Den Hartog, 2001). In addition, seagrasses exhibit significant phenotypic plasticity (the ability of an organism to change its characteristics in response to the environment; McDonald et al., 2016; Pazzaglia et al., 2021), and dimensions and morphologies of plant parts of related species can overlap, leading to confusion regarding species boundaries. Seagrasses are found in a wide range of environments, and species´ morphologies and dimensions may vary geographically (McDonald et al., 2016), which can lead to erroneous recognition of different (sub)species. In contrast, in some cases, what appears to be a single species may actually consist of two or more cryptic species that are genetically distinct but morphologically indistinguishable (Martínez-Garrido et al., 2016; Triest et al., 2018). Interbreeding between some congeneric seagrass species has also been reported (e.g. Ito and Tanaka, 2011; Sinclair et al., 2019; Liu et al., 2020), making it difficult to determine clear taxonomic distinctions. While genetic analysis is a powerful tool for distinguishing species, the genetic data available for seagrasses is still relatively limited (e.g., Dilipan et al., 2018; Tuya et al., 2024). In the past, clear seagrass identification guides or keys were not widely available, and during this assessment we were able to reveal various cases of species misidentification, requiring adjustments in geographical distributions.

Although all Halophila species in this bioregion have distinct morphologies, misidentifications have occurred in the past; possibly due to a combination of limited visibility and historical lack of seagrass identification guides. H. baillonii, for example, has been misidentified as H. decipiens, H. johnsonii, or H. engelmannii (Short et al., 2006; Samper-Villarreal et al., 2018; Creed and Samper-Villarreal, 2019). This species is often found in deep or turbid waters with low visibility, where it is covered with fine sediments, which hampers field identification. H. baillonii is currently not considered to be present in Florida and surrounding area, where it was previously reported and most likely wrongly identified. Similarly, H. engelmannii has been confused with H. baillonii, even though there is no taxonomic uncertainty regarding the identity of either one of these species (Kuo, 2020). Reports of H. engelmannii from the southern and southwestern Caribbean (Trinidad, Puerto Rico, Venezuela and Costa Rica) are also misidentifications of either H. baillonii or H. decipiens. In this assessment, we adjusted to the EOO (minimum convex polygon around all present native occurrences of a species) and AOO (Area of Occurrence) of both species.

Recent molecular advances have clarified several taxonomic uncertainties regarding Halophila species in the region, such as confirming that H. johnsonii is now considered a synonym of H. ovalis subs. ovalis. This species was likely first observed in Florida, USA in the 1950s, but at the time was not distinguished from H. decipiens and also misidentified as H. baillonii (Phillips, 1960). In 1972, specimens from the Indian River Lagoon, Florida, were considered to be H. johnsonii (Eiseman, 1973), and Johnson’s seagrass was formally described soon after (Eiseman and McMillan, 1980). Since then, the taxonomic status of H. johnsonii has been repeatedly called into question (Short et al., 2010; Fonseca and Olsen, 2016). Early molecular studies showed H. johnsonii was genetically similar to and closely allied with a number of its congeners (Short et al., 2010; Waycott et al., 2015) forming what was proposed as the “Halophila ovalis species complex” (sensu Den Hartog, 1970). Some argued that H. johnsonii was H. ovalis introduced from the Indo-Pacific (Short et al., 2010; Waycott et al., 2015, but see Kuo, 2020), an assertion supported by its observed asexual (York et al., 2008) and monoclonal (Waycott et al., 2015) population. Populations of this species were originally restricted to Florida (NOAA-NMFS, National Oceanic and Atmospheric Administration, National Marine Fisheries Service, 2007), but recently appeared in the broader Caribbean (Short et al., 2010; Waycott et al., 2015). DNA sequences, microsatellites, and ddRAD SNPs from H. johnsonii and H. ovalis throughout their respective ranges, revealed that the Florida and Antiguan populations of H. johnsonii were (1) identical, (2) entirely monoclonal, and (3) well resolved with east African subpopulations of H. ovalis (Waycott et al., 2021). In response, the United States’ National Marine Fisheries Service (NMFS) removed H. johnsonii from the List of Threatened and Endangered Species and eliminated its critical habitat designation in 2022 (Federal Register, 2022). It is now generally regarded as a synonym of H. ovalis subsp. ovalis (e.g., https://wfoplantlist.org, https://powo.science.kew.org, https://florida.plantatlas.usf.edu).

In the previous assessments, H. beaudettei, H. bermudensis, H. ciliata, and H. emaginata were listed as data deficient (DD, Short et al., 2011) due to taxonomic uncertainties, which remain until today. Halodule wrightii is the only well-established species of this genus in the Tropical Atlantic Bioregion. The taxonomy of Halodule spp. is based primarily on leaf tip morphology, because observations of its sexual reproductive structures have been lacking or infrequent, and they are highly reduced (Kuo and Den Hartog, 2001). Furthermore, Halodule spp. are phenotypically plastic, and the use of leaf tip morphology as a diagnostic character has been repeatedly questioned (Phillips, 1967; Ito and Tanaka, 2011; Wagey and Calumpong, 2013). In some cases, the leaf tip shape, used to separate the species, has been shown to vary among shoots from the same rhizome (Phillips, 1967; Wheeler et al., 2020). Genetic studies have indicated that most Atlantic Halodule spp. populations are closely related, from the Caribbean and Brazil to all of Africa, with higher differentiation inside the Gulf of Mexico, and future synonymizing of several species with H. wrightii is possible (e.g. Tavares et al., 2023; Sousa et al., 2025).

The taxonomy of Ruppia remains problematic and is undergoing revision (Den Hartog and Triest, 2020). R. maritima has traditionally been considered cosmopolitan (Short et al., 2011), but recent molecular studies demonstrated that it is a complex of multiple species, including polyploids and recent and ancient hybrid lineages (Ito et al., 2010; 2013; 2015; Triest et al., 2018). R. maritima sensu stricto (i.e. the diploid species in this complex that retains the original name) was described from northern Europe (Den Hartog and Triest, 2020), and has a wide but patchy distribution throughout Europe and the Mediterranean Basin (Triest and Sierens, 2014). It also occurs in the Tropical Atlantic Bioregion as it is present in Cabo Verde (Martínez-Garrido et al., 2017). In other parts of the world, several populations previously treated as R. maritima have been described as distinct species (Novelo, 1991; Yu and Den Hartog, 2014; Den Hartog et al., 2016; Kurniawan et al., 2024). Further cryptic species may have been identified but await formal description (Ito et al., 2010; 2015; Martínez-Garrido et al., 2016; Triest et al., 2018). In the Tropical Atlantic Bioregion, considerable morphologically diversity has been reported (Fernald and Wiegand, 1914; Koch and Seeliger, 1988; Oliveira Filho et al., 1983), but the taxonomy and distribution of these potentially cryptic species remain insufficiently studied. Most records lack sufficient information to be assigned to the distinct species, and therefore they are collectively considered as R. maritima in the present assessment, pending further studies. Morphological, cytological, and molecular studies indicated that R. mexicana and R. didyma are considered distinct species (Novelo, 1991; Den Hartog et al., 2016; Triest et al., 2018), and their conservation status is determined for the first time in these assessments.

In the previous IUCN global assessment, both C. nodosa and Z. noltei were reported for Cabo Verde islands (IUCN, 2010). But recent large-scale surveys revealed only the presence of H. wrightii and R. maritima in these islands (Creed et al., 2016; Martínez-Garrido et al., 2017). Although not registered by latter authors, C. nodosa was recently documented in Santiago Island, confirming its presence in Cabo Verde (Associação Caboverdiana de Ecoturismo ECOCV, pers. com.). But, at present, there is insufficient evidence of occurrence of Z. noltei in Cabo Verde islands. On one hand, it is possible that H. wrightii and R. maritima were wrongly identified as Z. noltei. But, on the other hand, records for Cabo Verde islands may also stem from confusion with the similarly named Cape Vert, in Senegal which is an upwelling area and within the range of this species. The two places have previously been confused in the literature (e.g., records of Ecklonia sp. in Price et al., 1978). Thus, adjustments were made to Z. noltei EOO and AOO in the Tropical Atlantic Bioregion.

Short et al. (2007) established six seagrass bioregions based on species assemblages, their distribution ranges, and influences from tropical and temperate climates, along with physical barriers set by oceans and seas. These authors considered the distinction between temperate and tropical zones crucial, as well as the restriction of certain species assemblages to specific sea or ocean basins. The Tropical Atlantic Bioregion was characterized as having clear water with a high diversity of seagrasses on reefs and shallow banks, dominated by T. testudinum. The Eastern Tropical Pacific and the western tropical African coastline, included by these authors, do not entirely fit this characterization, but they share common species with the tropical western Atlantic, and their inclusion in this bioregion is therefore considered as justified.

However, the temperate-tropical divide is not clearly defined, and tropical species, such as H. wrightii, may spread into temperate waters up to the coasts of North Carolina in the western Atlantic (Stevenson et al., 2025). Typical temperate species, such as C. nodosa, Zostera marina and Z. noltei have marginal populations co-occurring with tropical species such as H. wrightii. Considering the diffuse divide between temperate and tropical species extreme distributions, we propose that the boundaries between adjacent bioregions could be overlapping. The overlapping transition boundaries between adjacent bioregions will maintain the original principal for the creation of the seagrass bioregions, based on temperate-tropical species assemblages. At the same time, they provide more flexibility in the assessments, reducing the overlap of regional assessments for species with marginal populations in a bioregion with main distribution in its neighboring bioregion(s). For the Tropical Atlantic Bioregion, for instance, such an overlapping transitional boundary includes the marginal populations at extreme distribution, such as Z. noltei and C. nodosa. These species have their main distribution in the Mediterranean Bioregion (C. nodosa) or Mediterranean and Temperate North Atlantic Bioregion (Z. noltei; Short et al., 2011). The establishment of an overlapping transition boundary, excludes the need for their regional assessments in the Tropical Atlantic Bioregion. H. beaudettei and H. wrightii have their main distributions in the Tropical Atlantic Bioregion, but they also extend into neighboring bioregions (Figures 1, S1), and their global assessments depend at present on regional evaluations of the marginal populations in the neighboring bioregions. In establishing overlapping transition boundaries, the assessment of these species would be global and final, with EOO of 10,827,677 km² for H. beaudettei, and 50,655,025 km² and H. wrightii.

We suggest to use of the global distribution of H. wrightii (rounded to half a degree) to define the transitional overlapping boundaries of the Tropical Atlantic Bioregion, and we propose the following overlapping transition zones between adjacent seagrass bioregions (Figure 1):

Considering ongoing tropicalization (Hyndes et al., 2017; 2022), the overlapping transition boundaries of the neighboring bioregions may be of special interest, as tropical species may ingress into more temperate neighboring regions. Also, on the African coast, temperate species such as C. nodosa and Z. noltei (Hyndes et al., 2017; 2022), and (annual) populations of Z. marina may disappear or transgress towards more temperate waters due to tropicalization (Bartenfelder et al., 2022). Overlapping transition boundaries between tropical and temperate bioregions may thus be useful in future assessments of the conservation status of seagrass species, which may also apply to corals, mangroves or saltmarsh species. The proposed overlapping transition boundaries may need adjustments in the future, which may by itself also be an indicator of species shifts around the globe.

The original seagrass bioregions (Short et al., 2007, Figure 1) did not include large sections of the western African and eastern Pacific coasts. Along the west African coast, the southern limit of the Tropical Atlantic Bioregion extends beyond the southern extreme known distribution of H. wrightii in Angola (Figure 1); further studies on seagrass distribution may allow better definition of this boundary. Further south, the coast of Namibia is more temperate, characterized by low water temperatures due to the Benguela Current and strong coastal upwelling, harboring coastal lagoons with Ruppia sp. meadows (Clarke and Klaasen, 2001). With exception of Zostera capensis (Adams, 2016), seagrass meadows near the southern tip of Africa remain poorly studied. Along the eastern Pacific coast, the southern limit of the Tropical Atlantic Bioregion was established at the southern border of Panama (UNEP-WCMC (United Nations Environment Programme World Conservation Monitoring Centre) and Short, 2005). However, this region has been poorly studied (Samper-Villarreal, 2024). Nearby Panama, Colombia and Ecuador have potentially suitable sheltered estuarine habitats for seagrasses, meriting more studies. Further south, the coastline lacks sheltered embayments and is influenced by the cold Humboldt current. Although seagrasses have not been registered along these coastlines to date, the presence of sparse seagrass populations can´t be ruled out. We therefore recommend that unassigned coastal sections falling between current seagrass bioregions be absorbed during the pending updated global assessment of the seagrass conservation status.