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. 2025 Jan 24;39(4):e14430. doi: 10.1111/cobi.14430

Effects of protection on large‐bodied reef fishes in the western Indian Ocean

Melita Samoilys 1,2,✉, Kennedy E Osuka 1,3, Ronan Roche 4, Heather Koldewey 5,6, Pascale Chabanet 7
PMCID: PMC12309643  PMID: 39853835

Abstract

Predatory and large‐bodied coral reef fishes have fundamental roles in the functioning and biodiversity of coral reef ecosystems, but their populations are declining, largely due to overexploitation in fisheries. These fishes include sharks, groupers, Humphead wrasse (Cheilinus undulatus), and Green Humphead parrotfish (Bolbometopon muricatum). In the western Indian Ocean, this situation is exacerbated by limited population data on these fishes, including from conventional visual census methods, which limit the surface area surveyed. We developed a rapid timed scuba swim survey approach for application over large areas for estimation of the abundance of large‐bodied reef fishes and assessment of the effectiveness of marine protected areas (MPAs) in maintaining these species’ populations. Using this method, we sampled 7 regions in the western central Indian Ocean and Gulf of Aden, including 2 remote reference locations where fishing is prohibited. Eight families were selected for the surveys from across 3 categories: pelagic, demersal, and large‐bodied single species. Sharks (Carcharhinidae) were absent in 5 of the 7 regions, observed only in Mozambique and the Chagos Archipelago. Tunas (Scombridae) and barracudas (Sphyraenidae) were rarely observed (none in Madagascar, Djibouti, and Iles Glorieuses). The Giant grouper (Epinephelus lanceolatus) was absent in all regions, Humphead wrasse was absent in Comoros and Iles Glorieuses, and Green Humphead parrotfish was observed at only one site in Tanzania. The MPAs were not effective in protecting these single large‐bodied species or the 4 pelagic families, except for sharks in the highly protected reference locations. However, MPAs with medium levels of protection were effective in maintaining the abundance of some demersal families, notably large‐bodied groupers. Our results support the hypothesis of local extirpation of these large‐bodied fishes on many coral reefs in the western Indian Ocean.

Keywords: coral reefs, endangered species, Indian Ocean, large‐bodied fishes, marine parks, predatory fishes, arrecifes de coral, especie en peligro, Océano Índico, parques marinos, peces depredadores, peces de talla grande

INTRODUCTION

Large‐bodied and predatory fishes that associate with coral reefs are declining (Hammerschlag et al., 2018; Osuka et al., 2022; Richards et al., 2012). Primary causes are overexploitation in fisheries due to poor management or lack of enforcement and destructive fishing gears (Graham et al., 2017; Sadovy et al., 2003; Samoilys et al., 2017). Biological vulnerability due to the life‐history strategies of some of these species, particularly sharks and groupers (Sadovy de Mitcheson et al., 2020; Sherman et al., 2023), exacerbates these issues. However, long‐term quantitative data on the population status of large‐bodied reef fish populations are scant, particularly in the western Indian Ocean (WIO). This region contains globally significant biodiversity hotspots (Myers et al., 2000) and faces challenges in monitoring and protecting marine ecosystems due to a lack of human, institutional, and technical capacity (Samoilys et al., 2022; Stefanoudis et al., 2022; Veech, 2003; Wafar et al., 2011). There is therefore an urgent need for species‐level population data in the WIO to improve understanding of how the coral reef ecosystems are functioning.

Ecological importance of large‐bodied reef fishes

Fish biomass is often used as a proxy for estimating energy flow, which can be loosely termed as ecosystem functioning or productivity of a reef (Mora et al., 2011; Morais et al., 2020). Large‐bodied fishes, such as sharks, barracuda, trevally, tunas, and groupers, play a critical role in reef ecosystem productivity (Kulbicki et al., 2015) and represent a significant energy store as piscivorous species at a high trophic level (Hammerschlag et al., 2019; Hussey et al., 2013). Top‐level predatory reef fishes associate with high coral cover and structural complexity (Sartori et al., 2021), and their absence releases lower trophic taxonomic groups, thereby disrupting fish community trophic structure (Heithaus et al., 2008; Hixon, 2015; Robinson et al., 2019). The importance of predatory reef fishes is also evidenced in their use as a metric in ecosystem extinction risk assessments for coral reef ecosystems (Obura et al., 2021).

IUCN Red List

The International Union for Conservation of Nature (IUCN) Red List is the most comprehensive standardized data source on the extinction risk of species and is widely used by governments, nongovernmental organizations (NGOs), researchers, and donors to inform and catalyze action for biodiversity conservation, policy change, and awareness raising (Betts et al., 2020). Three of the large‐bodied reef‐associated species we surveyed are threatened (IUCN, 2023): 2 Labridae—the endangered (EN) Humphead wrasse (Cheilinus undulatus) and the vulnerable (VU) Green Humphead parrotfish (Bolbometopon muricatum)—while the Giant grouper (Epinephelus lanceolatus), the largest bony fish in the world (Craig & Hastings, 2007), previously classified as vulnerable, is now classified as Data Deficient (DD). Despite this global recognition of the vulnerability of these 3 species, no published population surveys from the WIO were available to the IUCN when these assessments were conducted (Chan et al., 2012; Fennessy et al., 2018; Russell, 2004), largely because of limited funding and human capacity for in situ research surveys on coral reefs in this region (Samoilys et al., 2022; Stefanoudis et al., 2022; K.O., personal observation). Population data from the WIO are critical to raising awareness in the region of vulnerable species that are taken in fisheries and to assist member states in reporting progress toward international multilateral agreements on biodiversity conservation.

Fishing and economic value

Fishing is one of the key threats to coral reefs globally (Andrello et al., 2022; Hughes et al., 2017; Letessier et al., 2019). Large‐bodied reef fishes are often target fishery species because of their high market value (Sadovy de Mitcheson et al., 2020), though this is less so in the WIO, where artisanal fisheries dominate and are characterized by the use of gears that capture a wide range of species (McClanahan & Mangi, 2004; Samoilys et al., 2017). However, large‐bodied reef fishes in the WIO face increasing threats because the region is experiencing faster warming than other ocean basins (Hermes et al., 2019), has a rapidly growing and hence youthful human population (Davis & Balls, 2022; United Nations, 2011), and faces varying levels of technical and management capacity (Wafar et al., 2011).

Populations of reef sharks have been declining worldwide for decades (Jackson et al., 2001) and are now highly depleted, except in areas where gill net fishing is excluded, large marine sanctuaries are in place, or both (MacNeil et al., 2020), such as in New Caledonia and Palau (Ward‐Paige, 2017). In the WIO, Seychelles is the only country that actively protects sharks and where abundant reef sharks are reported (MacNeil et al., 2020). Understanding of natural population levels of shark species associated with coral reefs is poor. For example, even in the Chagos Archipelago, reef shark populations have declined by 75% from the 1970s to 2012 (Ferretti et al., 2018). However, in most countries in the WIO, there is no active management of sharks, and quantitative information on reef shark populations is lacking (Bennett et al., 2022). For example, many WIO fisheries reports to the Food and Agricultural Organization simply provide an aggregate value for “sharks and rays,” making targeted species management difficult.

Rapid underwater census methods

Insufficient data on large‐bodied reef‐associated fishes in the WIO are partly due to the difficulties in surveying these species underwater because many are shy and highly mobile or have naturally low population densities and are not often observed in commonly employed belt transect surveys with relatively small surface area (∼250 m2). Consequently, they are often not selected for regular coral reef monitoring (Samoilys & Carlos, 2000; Samoilys et al. 2019; Usseglio, 2015), such as in Global Coral Reef Monitoring Network (GCRMN) surveys (Obura et al., 2017; Wilkinson, 2008). Therefore, we developed a rapid, timed, long scuba swim approach to assess the population status of predatory and large‐bodied (attain >55 cm total length [TL]) reef fishes and to determine whether this method can be usefully added to the standard 250‐m2 belt transect surveys by GCRMN teams in the WIO. We deployed the timed swim method across 7 countries in the western and central Indian Ocean and Gulf of Aden. We then assessed the effects of protected areas on these selected taxonomic groups and hypothesized that the 2 remote regions (Chagos, Iles Glorieuses) support high populations of these species, possibly representing close to unfished natural populations, compared with reefs in the other 5 locations that are fished or less protected.

METHODS

Study area

Surveys of large‐bodied reef fishes were conducted in 27 locations that ranged from 11.8°N to 14.7°S to 42.5°–72.7°E and were spread across 7 countries (hereafter referred to as regions) (Figure 1; Appendix S1). Five regions were in the WIO: Comoros, Madagascar, Mozambique, Tanzania, and Iles Glorieuses (France). One was in the central Indian Ocean, Chagos Archipelago, and one was in Djibouti in the Gulf of Aden. Surveys were all conducted from 2009 to 2015, except in Comoros, where additional surveys were conducted in 2018 to supplement the limited survey in 2009.

FIGURE 1.

FIGURE 1

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Mean biomass for 10 taxonomic groups of large‐bodied reef fishes across 7 regions in the western and central Indian Ocean and Gulf of Aden (smallest open circles, zero individuals observed).

Marine protected area indices

Reference locations for unfished populations (Chabanet et al., 2016; Samoilys et al., 2018) were in 2 marine protected areas (MPAs) that are protected from fishing and have highly restricted access because they are military zones: the Chagos Archipelago and Iles Glorieuses (Figure 1). Sites were assigned one of 4 protection indices based on the management and enforcement in place (Appendix S1), following previous criteria for the region (Osuka et al., 2021): none, no management in place; low, MPA gazetted and established but enforcement considered poor; medium, MPA gazetted and established with some no‐take‐zone (NTZ) restrictions and good enforcement; and high, MPA with NTZs and strong enforcement. Survey sites were spread as much as possible across these protection levels in each location to maximize comparisons across the study area; however, high‐protection sites were present in only 2 reference locations. This index is relatively simple and does not incorporate the size or age of the MPA, which varied considerably across the study regions (Appendix S1).

Taxonomic groups

We estimated the density and size of a preselected group of reef‐associated fishes that attain a large body size (>55 cm TL) and are either high‐trophic‐level species (∼4.0; Hussey et al., 2013) or threatened by extinction (IUCN, 2023). These taxonomic groups spanned 8 families and one subfamily and fell into 3 types: pelagic, demersal, and single large‐bodied species (Table 1). Pelagic taxa included highly mobile or wide‐ranging reef‐associated predators: sharks (Carcharhinidae), tunas (Scombridae), trevally (Carangidae), and barracuda (Sphyraenidae). Demersal large‐bodied taxa included grouper (Epinephelidae, piscivores), snapper and emperor combined (Lutjanidae and Lethrinidae), largely omnivores, pooled for analyses), and parrotfishes (Scarinae [subfamily of Labridae]), herbivores). Single IUCN Red List threatened species included wrasse (Labridae): EN Humphead wrasse (C. undulatus), VU Green Humphead parrotfish (B. muricatum), and the DD Giant grouper (E. lanceolatus). The 4 demersal families are frequently surveyed in standard underwater visual census (UVC) belt transect surveys, but here we surveyed only those species that attain maximum length (L max) of greater than 55 cm Total Length (TL) (Paxton et al., 2020).

TABLE 1.

Species recorded using the timed long swim survey method in 7 regions spanning the western and central Indian Ocean and the Gulf of Aden (2009–2018).

Functional group Family or subfamily Species Common English name Maximum total length (cm) a Trophic level a IUCN RL status b
Reef associated predators (highly mobile and wide ranging) Carcharhinidae Carcharhinus amblyrhynchos Grey Reef Shark 255 4.1 EN 2020
Carcharhinus melanopterus Blacktip Reef Shark 200 3.9 VU 2020
Carcharhinus albimarginatus Silvertip Shark 300 4.2 VU 2015
Carcharhinus spp.
Triaenodon obesus Whitetip Reef Shark 213 4.2 VU 2020
Scombridae Euthynnus affinis Kawakawa/ little tuna 107 4.5 LC 2011
Grammatorcynus bilineatus Double‐lined Mackerel 105 4.2 LC 2011
Gymnosarda unicolor Dogtooth Tuna 272 4.5 LC 2011
Rastrelliger kanagurta Indian mackerel 42 3.2 DD 2011
Scomberomorus commerson Narrow barred King Mackerel 250 4.5 NT 2011
Sarda orientalis Striped bonito 108 4.2 LC 2009
Scombrid spp.

Carangidae

 Carangoides bajad Orangespotted Trevally 55 4.0 LC 2016
Carangoides ferdau Blue Trevally 70 4.3 LC 2016
Carangoides fulvoguttatus Yellowspotted Trevally 133 4.4 LC 2016
Carangoides gymnostethus Bludger 90 4.1 LC2016
Caranx ignobilis Giant Trevally 170 4.2 LC 2016
Caranx heberi Blacktip Trevally 88 3.7 LC 2016
Caranx melampygus Bluefin Trevally 127 4.5 LC 2016
Caranx sexfasciatus Bigeye Trevally 120 4.5 LC 2019
Decapterus sp. (cf russelli) c looks like Indian scad 45 3.7 N/A
Gnathanodon speciosus Golden Trevally 120 3.8 LC 2016
Elagatis bipinnulata Rainbow Runner 180 4.3 LC 2016
Scomberoides commersonnianus Talang queenfish 120 4.4 LC 2016
Scomberoides lysan Doublespotted Queenfish 110 4.0 LC 2016
Scomberoides tol Needlescaled queenfish 60 4.1 LC 2016
Selar crumenophthalmus Bigeye Scad 70 3.8 LC 2015
Trachinotus blochii Snubnose pompano 110 3.7 LC 2016
Carangids
Sphyraenidae Sphyraena acutipinnis Sharpfin Barracuda 80 4.5 Not evaluated
Sphyraena barracuda Great Barracuda 200 4.5 LC 2015
Large‐bodied rare and threatened Labridae Labridae d (Scarinae) Cheilinus undulatus Humphead Wrasse 270 4.0 EN 2004
Bolbometopon muricatum Green Humphead Parrotfish 130 2.7 VU 2012
Demersal piscivores Epinephelidae Aethaloperca rogaa Redmouth Grouper 60 4.2 LC 2018
Anyperodon leucogrammicus Slender Grouper 65 3.9 LC 2018
Cephalopholis argus Peacock Grouper 60 4.5 LC 2018
Dermatolepis striolata Smooth Grouper 85 4.0 LC 2018
Epinephelus caeruleopunctatus Whitespotted Grouper 76 3.7 LC 2018
Epinephelus coioides Orange‐spotted Grouper 120 4.0 LC 2018
Epinephelus flavocaeruleus Blue and Yellow Grouper 90 4.2 LC 2018
Epinephelus fuscoguttatus Brown‐marbled Grouper 120 4.1 VU 2018
Epinephelus malabaricus Malabar grouper 234 4.2 LC 2018
Epinephelus polyphekadion Camouflage Grouper 75 4.0 VU 2018
Epinephelus tukula Potato Grouper 200 4.2 LC 2018
Epinephelus multinotatus Whiteblotched Grouper 100 3.9 LC 2018
Plectropomus areolatus Squaretail Coral Grouper 80 4.5 VU 2018
Plectropomus laevis Blacksaddled Coral Grouper 125 4.1 LC 2018
Plectropomus punctatus Marbled Coral Grouper 96 4.5 LC 2018
Variola albimarginata White‐edged Lyre Tail 65 4.5 LC 2018
Variola louti Yellow‐edged Lyretail 83 4.3 LC 2018
Demersal omnivores/some piscivores Lutjanidae Aphareus furca Small‐toothed Jobfish 70 4.1 LC 2016
Aprion virescens Green Jobfish 112 4.3 LC 2016
Lutjanus argentimaculatus Mangrove Red Snapper 150 3.6 LC 2016
Lutjanus bohar Two‐spot Red Snapper 90 4.3 LC 2016
Lutjanus monostigma One‐spot Snapper 60 4.3 LC 2016
Lutjanus rivulatus Blubberlip Snapper 80 4.1 LC 2016
Macolor macularis Midnight Snapper 60 4.0 LC 2016
Macolor niger Black and White Snapper 75 4.0 LC 2016
Lethrinidae Gymnocranius grandoculis Blue‐lined Large‐eye Bream 80 3.8 LC 2016
Lethrinus erythracanthus Orange‐spotted Emperor 70 3.4 LC 2016
Lethrinus mahsena Sky emperor 65 3.4 EN 2019
Lethrinus microdon Smalltooth Emperor 80 3.8 LC 2016
Lethrinus nebulosus Spangled emperor 87 3.8 LC 2016
Lethrinus obsoletus Orange‐striped Emperor 60 3.9 LC 2016
Lethrinus olivaceus Longface emperor 100 4.0 LC 2016
Lethrinus xanthochilus Yellowlip Emperor 70 3.8 LC 2016
Monotaxis grandoculis Humpnose Big‐eye Bream 60 3.4 LC 2016
Demersal herbivores (excavators and scrapers) Scarinae b Cetoscarus ocellatus Bicolor Parrotfish 80 2.0 LC 2016
Chlorurus strongylocephalus Indian Ocean Steephead Parrotfish 70 2.0 LC 2012
Hipposcarus harid Candelamoa Parrotfish 75 2.0 LC 2012
Scarus falcipinnis Sicklefin Parrotfish 60 2.0 LC 2012
Scarus ghobban Blue‐barred Parrotfish 75 2.0 LC 2012
Scarus prasiognathos Singapore parrotfish 70 2.0 LC 2012
Scarus rubroviolaceus Ember Parrotfish 70 2.0 LC 2012
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a

Maximum total length and trophic levels from FishBase.

b

International Union for Conservation of Nature Red List (IUCN RL) source (www.iucnredlist.org).

c

Species of Decapterus not identified. Decapterus russelli (Indian Scad) maximum total length and trophic group values used as proxy.

d

In the Labridae family, 3 taxonomic groups were separated for analyses: Cheilinus undulatus; Bolbometopon muricatum (because of their huge size and rarity); and the remaining 7 large‐bodied Scarinae subfamily species.

UVC method

A timed long swim (LS) of 10 min on scuba (Choat & Pears, 2003; Robbins et al., 2006) was used to estimate densities and fish sizes of the 10 preselected taxonomic groups (Table 1). All species (TLmax >55 cm, where TLmax is the maximum size the species attains) in each of the pelagic and demersal families were searched for; thus, missing species from sites represented zero observations. The UVC survey consisted of a timed swim parallel to the reef crest ranging over the depth of the reef but ≤30 m while counting fish within a ∼20‐m‐wide swathe (Appendix S2). Mean visibility was 14.7 m; no swims were conducted in <7‐m visibility. The swim was designed to cover as much distance as possible underwater parallel to the reef edge. At each reef site, 2 LS replicate transects were completed in 2 dives within ∼500–1000 m of each other. To estimate the average length of an LS, GPS readings were taken to measure a subset of 16 LS surveys at Glorieuses Island. The mean length of an LS was 147.4 m (SD 54.6). Thus, fish densities and biomass per LS were calculated based on a sampled reef area of ∼3000 m2.

Data analyses

Biomass (kilograms per hectare) was estimated from fish size census estimates based on published length–weight relationships (Froese & Pauly, 2023; Kulbicki et al., 2005), and medians and means were calculated (Appendix S3). Biomass was used as the standard metric because it is generally more sensitive to fish population change (Graham et al., 2017). Density values (number of individuals per hectare) were also calculated for the zero‐inflated negative binomial (ZINB) modeling.

Biomass of the 10 taxonomic groups was compared across the 7 regions and 4 protection levels with a Kruskal–Wallis test because data did not conform to parametric assumptions. Regions that showed no variance in biomass (i.e., zeros) were excluded from the test. This was followed by a post hoc Mann–Whitney test to identify significant differences.

A ZINB model was fitted to assess the influence of protection on the density of each of the taxonomic groups. We selected a ZINB based on the likelihood ratio test that showed better model performance than a zero‐inflated Poisson model (Lewsey & Thomson, 2004). The intercept was omitted from the ZINB model to allow assessment of the influence of each level of protection on fish density. Moran's I test was used to check the effect of spatial autocorrelation in each taxonomic group, and where an effect was detected, the glm model was fitted using spatial correlation structures, with Akaike information criterion (AIC) selection methods applied to select the best model. Spatial autocorrelation was detected in the Epinephelidae and Scarinae; therefore, the model was corrected using rational quadratic and Gaussian spatial correlation structures, respectively. To reduce the influence of uneven sampling effort, we conducted all our analyses at a country level. We also ran the ZINB model with the offset function within the model formula for selected taxa and found no major differences in the results.

RESULTS

Population biomass between geographic locations

Carcharhinidae were absent in 5 of the 7 regions, occurring only in Mozambique and the Chagos Archipelago (Figure 1; Table 2). Regional absences also occurred in 2 other pelagic taxa, Scombridae and Sphyraenidae, which were missing in Madagascar, Djibouti, and Iles Glorieuses. Giant grouper was absent in all regions, Humphead wrasse was absent in Comoros and Glorieuses, and Green Humphead parrotfish was only observed in Tanzania (Figure 1).

TABLE 2.

Median and mean biomass (kilograms per hectare) for 10 taxonomic groups across countries in the western and central Indian Ocean and Gulf of Aden.

Taxa Tanzania (n = 28) Mozambique (n = 49) Comoros (n = 29) Madagascar (n = 16) Djibouti (n = 26) Iles Glorieuses (n = 16) Chagos (n = 22)
Carcharhinidae              
Median (interquartile range [IQR]) 0.0 (0.0–0.0) 0.0 (0.0–0.0) 0.0 (0.0–0.0) 0.0 (0.0–0.0) 0.0 (0.0–0.0) 0.0 (0.0–0.0) 178.1 (0.0–641.4)
Mean (SD) 0.0 (0.0) 116.1 (813.0) 0.0 (0.0) 0.0 (0.0) 0.0 (0.0) 0.0 (0.0) 503.0 (849.3)
Scombridae              
Median (IQR) 0.0 (0.0–0.0) 0.0 (0.0–0.0) 0.0 (0.0–0.0) 0.0 (0.0–0.0) 0.0 (0.0–0.0) 0.0 (0.0–0.0) 0.0 (0.0–0.0)
Mean (SD) 2.9 (11.4) 0.0 (0.0) 0.7 (2.3) 0.0 (0.0) 1.5 (7.5) 0.0 (0.0) 13.8 (39.0)
Carangidae              
Median (IQR) 0.0 (0.0–4.8) 0.0 (0.0–0.0) 0.0 (0.0–0.4) 0.0 (0.0–0.0) 8.6 (2.9–17.9) 0.0 (0.0–10.1) 13.3 (0.0–34.8)
Mean (SD) 21.3 (64.3) 41.6 (248.0) 6.2 (17.2) 5.0 (15.1) 15.3 (20.8) 25.8 (69.8) 35.8 (65.7)
Sphyraenidae              
Median (IQR) 0.0 (0.0–0.0) 0.0 (0.0–0.0) 0.0 (0.0–0.0) 0.0 (0.0–0.0) 0.0 (0.0–0.0) 0.0 (0.0–0.0) 0.0 (0.0–0.0)
Mean (SD) 0.6 (3.3) 0.2 (1.5) 1.9 (6.3) 0.0 (0.0) 0.0 (0.0) 0.0 (0.0) 7.6 (35.9)
Epinephelidae              
Median (IQR) 0.0 (0.0–13.7) 0.0 (0.0–10.1) 3.3 (1.9–6.3) 0.0 (0.0–13.0) 0.0 (0.0–0.0) 35.5 (4.1–57.5) 42.0 (25.0–64.5)
Mean (SD) 12.3 (20.0) 12.9 (30.4) 6.9 (10.6) 8.6 (15.3) 1.0 (3.8) 47.1 (65.4) 53.0 (49.1)
Lutjanidae/Lethrinidae              
Median (IQR) 4.5 (0.0–23.3) 0.0 (0.0–4.4) 14.3 (6.4–27.4) 34 (0.0–18.7) 0.0 (0.0–0.0) 46.3 (27.4–87.0) 57.8 (21.8–264.0)
Mean (SD) 38.6 (117.5) 38.0 (215.9) 28.2 (40.1) 20.2 (35.1) 2.8 (6.4) 77.9 (79.4) 173.8 (259.5)
Scarinae              
Median (IQR) 0.9 (0.0–22.9) 0.0 (0.0–5.9) 4.5 (0.0–12.4) 6.6 (0.0–24.1) 0.0 (0.0–10.3) 6.8 (0.3–16.4) 6.9 (0.0–25.7)
Mean (SD) 14.5 (26.2) 6.6 (13.1) 17.4 (41.0) 16.6 (23.4) 5.9 (11.6) 12.4 (15.4) 17.3 (27.2)
Epinephelus lanceolatus              
Median (IQR) 0.0 (0.0–0.0) 0.0 (0.0–0.0) 0.0 (0.0–0.0) 0.0 (0.0–0.0) 0.0 (0.0–0.0) 0.0 (0.0–0.0) 0.0 (0.0–0.0)
Mean (SD) 0.0 (0.0) 0.0 (0.0) 0.0 (0.0) 0.0 (0.0) 0.0 (0.0) 0.0 (0.0) 0.0 (0.0)
Cheilinus undulatus              
Median (IQR) 0.0 (0.0–0.0) 0.0 (0.0–0.0) 0.0 (0.0–0.0) 00 (0.0–0.1) 0.0 (0.0–9.6) 0.0 (0.0–0.0) 0.0 (0.0–0.0)
Mean (SD) 6.8 (20.8) 7.6 (27.1) 0.0 (0.0) 4.0 (10.4) 20.0 (53.6) 0.0 (0.0) 5.0 (13.2)
Bolbometopon muricatum              
Median (IQR) 0.0 (0.0–0.0) 0.0 (0.0–0.0) 0.0 (0.0–0.0) 0.0 (0.0–0.0) 0.0 (0.0–0.0) 0.0 (0.0–0.0) 0.0 (0.0–0.0)
Mean (SD) 2.2 (9.4) 0.0 (0.0) 0.0 (0.0) 0.0 (0.0) 0.0 (0.0) 0.0 (0.0) 0.0 (0.0)
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Significant differences between geographic locations in pelagic taxa were found only in Carcharhinidae and Carangidae (Figure 2a); median biomass of Carcharhinidae was significantly higher in Chagos than in Mozambique (H = 19.64, p < 0.001) (Figure 2a). Carangidae had significantly higher median biomass in Djibouti (2.6 kg/ha) and Chagos (4.0 kg/ha) compared with other locations (H = 21.75, p < 0.001), except Iles Glorieuses (Figure 2a; Table 2). For Scombridae, zero observations were recorded in Mozambique, Madagascar, and Iles Glorieuses; biomass was low elsewhere and highest in Chagos, where Euthynnus affinis (Kawakawa tuna), Gymnosarda unicolor (Dogtooth tuna), and other unidentified species were present. Other large‐bodied scombrids included Grammatorcynus bilineatus (double‐lined mackerel), which was observed rarely in Tanzania. In Comoros, only the small scombrid Rastrelliger kanagurta (Indian mackerel) was observed.

FIGURE 2.

FIGURE 2

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Median (interquartile range) and mean (red point) biomass for (a) pelagic taxa, (b) demersal taxa, and (c) large‐bodied single species across the 7 regions in the western and central Indian Ocean and Gulf of Aden based on Kruskal–Wallace tests (biomass values in Table 2) (letters, Mann–Whitney pairwise comparisons; NA, not available).

Outliers illustrate the high spatial variation in these data largely driven by pelagic species. For example, in Mozambique, trevally were substantially more abundant in Nacala and sharks were only seen at one reef location, Vamizi (Figure 2). In the Chagos Archipelago, outliers represented more abundant scombrids and barracuda on the Chagos Bank. These outliers did not prevent the detection of significant differences between geographic locations or protection effects but did show variability, which should be considered in local‐ or national‐level population management.

All 3 groups of demersal large‐bodied taxa showed significant regional differences. Higher median biomass of Epinephelidae (H = 52.50, p < 0.001) was recorded in Chagos (12.6 kg/ha) and Iles Glorieuses (10.7 kg/ha) compared with other countries, and these biomass values were significantly higher than in Tanzania and Djibouti (Figure 2b; Table 2). Higher median biomass of Lutjanidae and Lethrinidae combined was recorded in Chagos (17.3 kg/ha) and Iles Glorieuses (13.9 kg/ha), and these were significantly higher than in all other locations, except Comoros (H = 68.96, p < 0.001). For Scarinae, no significant differences in median biomass were detected (H = 9.77, p > 0.05). Medians ranged from 0 (Mozambique and Djibouti) to 2.1 kg/ha (Iles Glorieuses).

No statistical analyses were possible for Giant grouper due to a lack of observations. Green Humphead parrotfish was only sighted in Tanzania, and there were no significant differences (H = 0.36, p > 0.05) in the median biomass across all sites (Figure 2c). For Humphead wrasse, significant differences were detected across sites; median biomass in Tanzania, Mozambique, Madagascar, Djibouti, and Chagos was significantly higher than in Comoros (H = 4.38, p < 0.05) (Figure 2c). Relatively high biomass was recorded in Djibouti, but there were zero observations in Iles Glorieuses and Comoros (Table 2). However, significant differences were not detected between Iles Glorieuses and Chagos or between Tanzania and Mozambique, suggesting these median values were not significantly higher than zero (Figure 2a; Table 2).

Effects of protection

Protection effects were detected in the median fish biomass of 2 pelagic families (Carcharhinidae and Carangidae) (Figure 3a) and 3 demersal taxonomic groups (Epinephelidae, Lutjanidae and Lethrinidae combined, and Scarinae) (Figure 3b). No significant protection effects were detected for Sphyraenidae and Scombridae (Figure 3a) or for Humphead wrasse and Green Humphead parrotfish (Figure 3c), largely due to many zero observations of these taxa. For the 5 taxonomic groups that showed protection effects, differences were not always aligned with significantly higher median biomass with greater protection except for Carcharhinidae (Figure 3). For Carangidae, biomass was greater in high, medium, and none protection areas compared with areas with low levels of protection (Mann–Whitney p < 0.05; Figure 3a; Appendix S3). The biomass of Epinephelidae was higher in high‐protection sites than in other levels of protection, and medium protection showed significantly higher biomass levels than in sites with low and no protection (Figure 3b). For the Lutjanidae and Lethrinidae combined, significant differences in median biomass with protection level were detected, but results were not straightforward. High‐protection sites had significantly higher biomass than medium, low, and none sites, whereas medium‐protection sites were only greater than no‐protection sites (Figure 3b). For the Scarinae, sites with high, medium, and low protection had statistically similar median biomass, although they were all significantly higher than the no‐protection sites (Figure 3b).

FIGURE 3.

FIGURE 3

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Median biomass for (a) pelagic taxa, (b) demersal taxa, and (c) large‐bodied single species in the western and central Indian Ocean and Gulf of Aden across 4 protection levels based on Kruskal–Wallace tests. Abbreviations as for Figure 2.

Several significant predictions of protection effects across the taxonomic groups were found through the ZINB modeling (Table 3), although few were consistent with an increase in density as protection level increased. For example, higher densities were predicted in medium‐protection sites (11.99 individuals/ha) relative to high‐protection sites (5.63 individuals/ha) for Carcharhinidae. However, the odds of observing excess zero counts of sharks were significantly higher in medium levels of protection. For Scombridae, all protection levels significantly influenced predicted densities (Table 3): high‐protection sites (14 fish/ha) had the most effect, medium‐protection sites (6 fish/ha) had the least effect, and low‐protection sites (10–11 fish/ha) had a moderate effect. The odds of observing excess zero counts were significant at all protection levels (highest for medium‐protection sites). For Carangidae, there was no clear effect of increasing protection level on population densities (Table 3). All protection levels significantly influenced predicted densities. Sites with no protection had the most influence (52 individuals/ha). Estimates for low, medium, and high protection, although significantly >0, were similar. Carangidae were not affected by protection, or the medium‐ or high‐protection sites were not suitable for Carangidae. High densities (180 fish/ha) of Sphyraenidae were predicted in high‐protection sites relative to medium‐ and no‐protection sites, which had 20–60 times lower predicted densities. The highest odds of observing excess zero counts were for sites with no protection; together, these were evidence for a protection effect in Sphyraenidae.

TABLE 3.

Results of a zero‐inflated negative binomial (ZINB) model predicting influence of protection on the density of 4 pelagic fish taxa (Carcharhinidae, Scombridae, Carangidae, and Sphyraenidae), 3 demersal taxa (Epinephelidae, Lutjanidae/Lethrinidae, and Scarinae), and one large‐bodied species (Cheilinus undulatus).

Model a Protection (predictor) Estimate CI p Estimate CI p Estimate CI p
Carcharhinidae Scombridae Carangidae
Count None – – – 10.84 4.87–24.09 <0.001 52.67 22.43–123.71 <0.001
Low – – – 10.33 3.86–27.61 <0.001 10.13 1.91–53.80 0.007
Medium 11.99 6.93–20.74 <0.001 5.69 1.25–26.04 0.025 12.32 3.99–37.98 <0.001
High 5.63 3.98–7.96 <0.001 14.88 6.86–32.30 <0.001 12.28 4.43–34.06 <0.001
Zero‐inflated None – – – 18.04 5.63–57.80 <0.001 0.48 0.09–2.58 0.396
Low – – – 6.87 1.55–30.37 0.011 0.80 0.09–6.84 0.839
Medium 11.65 3.58–37.89 <0.001 35.07 4.76–258.23 <0.001 0.70 0.12–4.03 0.686
High 2.22 1.08–4.54 0.029 10.91 3.34–35.61 <0.001 0.23 0.01–8.10 0.417
Sphyraenidae Epinephelidae Lutjanidae/Lethrinidae
Count None 9.00 4.68–17.30 <0.001 5.41 4.07–7.18 <0.001 9.18 5.85–14.40 <0.001
Low – – – 6.60 3.45–12.60 <0.001 13.83 6.40–29.90 <0.001
Medium 2.82 1.17–6.82 0.021 9.23 7.33–11.62 <0.001 56.41 35.10–90.67 <0.001
High 180.00 155.53–208.31 <0.001 8.50 6.78–10.64 <0.001 33.78 22.11–51.62 <0.001
Zero‐inflated None 56.99 7.89–411.58 <0.001 1.43 0.83–2.46 0.194 0.56 0.26–1.18 0.127
Low – – – 2.91 0.93–9.11 0.067 0.39 0.10–1.61 0.194
Medium 16.87 4.00–71.10 <0.001 0.34 0.16–0.72 0.005 0.38 0.17–0.86 0.020
High 35.00 4.80–255.47 <0.001 0.19 0.07–0.47 <0.001 0.01 0.00–777.96 0.428
Scarinae Cheilinus undulatus
Count None 9.15 6.13–13.66 <0.001 4.09 2.08–8.06 <0.001
Low 20.14 11.42–35.52 <0.001 6.03 3.62–10.05 <0.001
Medium 12.23 8.63–17.33 <0.001 5.05 2.49–10.24 <0.001
High 8.60 5.75–12.86 <0.001 4.63 1.83–11.69 0.001
Zero‐inflated None 1.46 0.83–2.56 0.192 7.79 3.26–18.61 <0.001
Low 0.56 0.19–1.62 0.281 0.69 0.24–2.03 0.504
Medium 0.29 0.12–0.69 0.005 6.11 2.34–15.96 <0.001
High 0.52 0.23–1.14 0.101 10.11 3.04–33.62 <0.001
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Note: The estimate represents the expected change in the outcome variable (density) associated with a change in the predictor variable (protection). Unavailable estimates (–) indicate zeros were observed and therefore the modeled estimates are not reliable.

a

For the count model, the estimate is density; for the zero‐inflated model, the estimate is the rate ratio (log odds).

All protection levels significantly influenced predicted densities of Epinephelidae (Table 3). Predicted densities were slightly lower for sites with no or low protection (5.4–6.6 individuals/ha) than for high‐ and medium‐protection sites (8.5–9.2 individuals/ha). The odds of observing excess zero counts of Epinephelidae decreased significantly only in high‐protection sites (by 0.19). For the Lutjanidae and Lethrinidae combined, all protection levels significantly influenced predicted densities (Table 3). The highest densities were predicted in sites with medium (56 individuals/ha) and high (34 individuals/ha) protection. The odds of observing excess zero counts of Lutjanidae and Lethrinidae were not significant. For large‐bodied Scarinae, all protection levels significantly influenced predicted densities, but the effect was inconsistent among no‐protection (9 individuals/ha), low‐protection (20 individuals/ha), medium‐protection (12 individuals/ha), and high‐protection (8 individuals/ha) sites. Protection level did not affect the odds of observing excess zeros.

For Humphead wrasse, all protection levels significantly influenced predicted densities, but the highest predicted density was in low‐protection sites. Significant odds of observing excess zero counts were seen in all protection levels except low protection, and the highest odds of observing excess zero counts were predicted in high‐protection sites.

DISCUSSION

Zero observations

Local extirpation of large‐bodied reef‐associated fishes, including sharks, Giant grouper, Green Humphead parrotfish, and Humphead wrasse, on many WIO coral reefs is suggested by our results. We found reef sharks at only one reef in northern Mozambique. In contrast, reef sharks were regularly observed in the Chagos Archipelago. However, with a mean biomass of only 503 kg/ha (SD 849), compared with 3147–6663 kg/ha at 2 other Iles Eparses islands (Chabanet et al., 2016), even Chagos shark populations appeared depleted. The 3 large‐bodied species vulnerable to extinction (IUCN, 2023)—Green Humphead parrotfish (VU), Humphead wrasse (EN), and Giant grouper (DD)—were rarely observed or not observed at all throughout the 7 countries. Ours is the first study to demonstrate that the population abundance of large‐bodied reef fishes from 8 functionally important families appears to be extremely low across the WIO region.

The pelagic tunas and barracudas also appeared to be depleted; there were multiple zero observations at sites in Madagascar, Iles Glorieuses, Mozambique (tuna), and Djibouti (barracuda) and low biomass elsewhere except within Chagos. The scombrid tuna and tuna‐like species contribute to some of the most valuable commercial fisheries in the Indo‐Pacific (Dahlet et al., 2019; Gillett & Cartwright, 2010), but they are threatened by intense fishing pressure and climate change (Bell et al., 2016; Kumar et al., 2012). Although many of these species spend a large portion of their time foraging away from coral reefs, the low numbers we recorded support concerns that scombrids are depleted in the WIO (Tidd et al., 2023).

Barracuda are circumglobal, large piscivorous fishes that move from nearshore mangroves and seagrass beds to coral reefs for their adult phase (de Sylva, 1963). Two barracuda species were recorded in the present study: Sphyraena barracuda (Great Barracuda) in WIO locations, whereas Sphyraena acutipinnis (Sharpfin) was the only barracuda sighted in the Chagos Archipelago. Great Barracuda biomass values ranged from 0.2 to 1.9 kg/ha in eastern Africa, suggesting their populations may be depleted. However, knowledge of the ecology of Giant Barracuda is sparse, limited to Florida and the Bahamas. Although adult Giant Barracuda commonly occur in <10 m near coral reefs (O'Toole et al., 2011), they also move distances of hundreds of kilometers over the continental shelf and open water to >175‐m depth (Hansen, 2015). The zero observations in the WIO may therefore represent false zeros (Blasco‐Moreno et al., 2019) because our rapid timed surveys did not adequately sample these species.

Trevally (jacks) are naturally abundant predators, often occurring in schools on or near coral reefs (Anderson, 2005), and associate positively with high and low coral complexity and strong currents (Sartori et al., 2021). Exceptionally high biomass values (up to 626 kg/ha) of trevally were found in 250‐m2 transect area surveys at Iles Glorieuses (Chabanet et al., 2016), whereas biomass was 26 kg/ha in Iles Glorieuses and 36 kg/ha in Chagos in our ∼3000‐m2 surveys (both data sets from 2014). In Madagascar, Comoros, and Djibouti, trevally biomass was exceptionally low. Population abundance values are limited for trevally in the Indo‐Pacific, making our results difficult to interpret. The large discrepancy in biomass values at Iles Glorieuses could have been due to the different UVC methods or aggregation behavior in Carangidae (Daly et al., 2018). Nevertheless, the low biomass in eastern Africa of these predatory species is of concern for coral reef ecosystem functioning and local fisheries.

Threatened species

Awareness of endangered species and the IUCN Red List is still relatively poor in many WIO countries, and this is linked to limited data. For example, previous IUCN assessments of the Green Humphead parrotfish, Humphead wrasse, and Giant grouper contained little or no population data from the WIO (Chan et al., 2012; Fennessy et al., 2018; Russell, 2004). We found Giant grouper was not sighted throughout the study sites, despite reports of their occurrence across a range of locations (Fricke et al., 2022), including in Chagos in 1999 (Winterbottom & Anderson, 1999). However, the species is poorly studied and its early life history is not well understood. Other ad hoc underwater observations in the region by one of us (M.S.) are from Zanzibar (3 adults) on inner reefs in northern Kenya (one juvenile) and at Iles Glorieuses (one adult). Giant grouper is also reported from Juan de Nova in the Iles Eparses (Chabanet et al., 2016). The lack of observations of Giant grouper suggests severe depletion, but with no baselines and limited ecological understanding, the status of this species in the WIO remains unknown.

The Green Humphead parrotfish was observed only in Tanzania at one site. This species is not known from oceanic locations (Chabanet et al., 2016; Hamilton et al., 2008; Winterbottom & Anderson, 1997, 1999) and therefore might not be expected in Comoros or Iles Glorieuses. However, a recent observation of 15 individuals in Grande Comore (M.S., November 2023) refutes this assumption. Early records show the Humphead parrotfish was widely distributed in the Red Sea and Gulf of Aden (Ormond & Edwards, 1987). No observations in Djibouti support previous reports that this species is highly depleted in the Gulf of Aden (Hassan et al., 2002). In summary, zero observations of Green Humphead parrotfish in all but one site suggest severe depletion in its population in the WIO.

The Humphead wrasse occurs across Indo‐Pacific coral reefs (Donaldson & Sadovy, 2001; Sadovy et al., 2003) and in the WIO as far east as Reunion and Mauritius (Fricke et al., 2009; P.C. personal observation). The species prefers deep reef drop‐offs and is reported close to seagrass beds (Daly et al., 2020); therefore, it would be expected on many of the sites we surveyed. The species was not observed in Iles Glorieuses, a fully protected uninhabited French military zone. This contrasts with the relatively high biomass of Humphead wrasse in other islands in the Mozambique Channel (Chabanet et al., 2016). Elsewhere, the biomass of Humphead wrasse was low and variable, with the highest abundance observed in Djibouti. It is known to have high site fidelity, which makes it vulnerable to local fishing pressure but also amenable to protection through MPAs (Daly et al., 2020). Except for Daly et al.’s (2020) study in Seychelles, spatial and ecological knowledge is sparse, with only 2 studies from New Caledonia and Palmyra Atoll (Chateau & Wantiez, 2007; Weng et al., 2015). Research is needed urgently on this endangered species (Fricke et al., 2009) globally and in the WIO to improve existing conservation approaches.

Our data can improve IUCN Red List global assessments, raise awareness of the threats vulnerable reef fish species face in the WIO region, and improve funding (Betts et al., 2020). The status of the WIO populations of the 3 large‐bodied threatened species will be provided to regional and national red lists under the IUCN National Red List project, where a different threat status may be decided. This would help nations measure and report their progress toward international multilateral agreements and raise awareness locally. However, in the WIO region, only South Africa has completed a national red list, likely because of a lack of capacity and resources in marine biodiversity research (Wafar et al., 2011).

Effectiveness of protected areas for large‐bodied reef fishes

Although protection level significantly influenced population biomass and predicted density of these large‐bodied fishes, the predictions were not consistent with higher population abundance associated with increasing levels of protection. The Chagos Archipelago and Iles Glorieuses can provide high protection from fishing primarily because there are no resident fishing communities and fishers coming from outside these territories are excluded (Hays et al., 2020). In addition, with great size and age (Appendix S1), Chagos Marine Park would be expected to contain the greatest population densities and biomass of large‐bodied fishes (Ziegler et al., 2024). Although some illegal fishing occurs (Collins et al., 2021), these remote MPAs did contain the highest population biomass of sharks (Carcharhinidae) and large‐bodied groupers (Epinephelidae) compared with all other protection levels. However, there was little evidence of positive protection effects for tuna, barracuda, trevally, and parrotfishes, though barracuda and tuna were predicted to have the highest population densities in high‐protection sites. Little conservation impact on Indian Ocean tuna populations from the Chagos MPA has been reported previously (Curnick et al., 2020).

Protection effects were not detected throughout the study region for the large‐bodied Humphead wrasse; predicted densities were similar for all protection levels. No conclusions could be made on the effectiveness of protection for Giant grouper and Green Humphead parrotfish because either population densities were zero throughout or individuals were observed only in medium protection sites. Because high‐protection sites in this study were isolated oceanic reefs where Green Humphead parrotfish are not expected (Hamilton et al., 2008), these areas cannot be considered refuges for this species. In contrast, the lack of sightings of Giant grouper in the Chagos Archipelago suggests that a previous fishery on groupers that closed in ∼2010 may have had long‐term impacts on this species (Samoilys et al., 2018). The lack of knowledge of the population status of this species is illustrated by its DD listing on the IUCN Red List (Fennessy et al., 2018) and highlights a research gap.

Two taxonomic groups important in WIO artisanal fisheries revealed a particularly poor response to protection—the trevally and the 7 large‐bodied parrotfish species. The highest population densities of trevally were predicted in sites with no protection status, and biomass varied 20 times in remote highly protected sites in the Mozambique Channel (this study; Chabanet et al., 2016). Further ecological research is needed to better understand the dynamics of this large family (Skinner et al., 2020). Little evidence of positive protection effects on large‐bodied parrotfishes is surprising; the highest population densities were predicted in low‐protection sites. Possibly, this reflects that other processes besides protection drive parrotfish populations, such as climate‐change‐induced changes in habitat and productivity and species‐specific responses to coral mortality (McClanahan et al., 2019; Robinson et al., 2019; Samoilys et al., 2018; Williams et al. 2015). In contrast, medium‐ and low‐protection sites, typical of MPAs in the WIO (Osuka et al., 2021), were effective for 2 of the 3 demersal large‐bodied taxonomic groups—the grouper, and snapper and emperor combined. This may be related to the long period that some of the medium‐protection MPAs have been established (Ziegler et al., 2024), such as those in Tanzania. This has positive implications for coastal livelihoods because these taxonomic groups are important in WIO artisanal fisheries (Rehren et al., 2022).

Usefulness of rapid timed swim UVC method

The characteristics of sites with the highest levels of protection in this study, in the Chagos Archipelago and Iles Glorieuses, were not replicated on the eastern African coast in terms of being remote and uninhabited but also in terms of their geomorphology and oceanic characteristics, which can influence taxa presence and abundance. However, inclusion of these locations represents the best available method to examine differences in the abundance of large‐bodied and high‐trophic‐level fish across large gradients of fishing pressure within the WIO. Generally, we based our interpretations on the assumption that zero observations of a taxonomic group are true zeros and not false zeros caused by sampling errors. This assumption is based on Blasco‐Moreno et al.’s (2019) definitions of structural zeros (related to ecological or evolutionary restrictions of the hypothesis being examined) and sampling zeros (If sampling is inadequate, the zero is false.). Where a species’ range is conclusively known to exclude certain locations, such as oceanic reefs, then zero observations at these sites would represent structural zeros, such as Green Humphead parrotfish (B. muricatum) not occurring in the Chagos Archipelago. In contrast, the several zero observations of Giant barracuda may be false zeros due to inadequate sampling based on their movements and behavior. The zero observations throughout all sites of the widely distributed Giant grouper likely represent true zeros, though movement and behavior of this species are poorly known.

Remote camera methods, particularly baited remote underwater videos, are increasingly deployed to estimate populations of large roving reef predators and fish assemblages (Harvey et al., 2021; Osuka et al., 2022; Simpfendorfer et al., 2023; Skinner et al., 2020; Stefanoudis et al., 2023). However, they are not readily accessible in many WIO countries, and the capacity to analyze such video data is still low. The timed swim we used was designed to be a rapid add‐on to regular GCRMN fish transect surveys that many nations deploy regularly in the WIO. It built on historic UVC attempts to estimate the abundance of large mobile and diver‐shy fish species (Kimmel, 1985; Thresher & Gunn, 1986). Our results showed that the rapid timed scuba swim could detect significant differences between geographic locations and protection levels in large‐bodied reef‐associated fishes, thereby providing valuable conservation information for these lesser‐known species.

Implications for ecosystem functioning and fisheries management

Many large‐bodied fishes on reefs perform a critical role in the functioning of coral reef ecosystems. Depletions in populations of large‐bodied and high‐trophic‐level demersal reef fishes affect reef ecosystems directly and indirectly (Frisch et al., 2014; Heithaus et al., 2008), whereas high grouper biomass has been linked to more complex coral reef systems (Karkarey et al., 2014; Sartori et al., 2021). Our data suggest that many large‐bodied predatory fishes are at or close to functional extinction (MacNeil et al., 2020) in the WIO and that populations of several taxa are depleted even in remote, highly protected locations of the Chagos Archipelago and Iles Glorieuses. When fishing removes these large‐bodied species, the fish community changes with cascading effects. For example, depletion in populations of Black tip (Carcharhinus melanopterus) and Gray reef sharks (Carcharhinus amblyrhynchos) has been clearly linked to changes in prey communities (Hammerschlag et al., 2018). However, these relationships are complex; for example, scale dependency (Heenan et al., 2020), inverted trophic pyramids (Sandin & Zgliczynski, 2015), and low‐trophic‐level fishes can enhance fish biomass under heavy fishing (Graham et al., 2017). Productivity and energy flow on coral reefs are also influenced by large‐bodied fishes that feed in pelagic or deep waters beyond the reef, such as the highly mobile scombrids, barracuda, trevally, and sharks. Such allochthonous pelagic subsidies (Morais & Bellwood, 2019) enhance fish biomass production on a coral reef, with obvious deleterious consequences if these fishes are removed.

The effectiveness of a protected area improves substantially if it is surrounded by strong fisheries management (Goetze et al., 2024; Roos et al., 2020). The gillnet is widely used in the WIO but is one of the most damaging fishing gears because it is unselective, capturing a wide range of taxa including large‐bodied species of snapper, emperor, grouper, parrotfish, and Humphead wrasse (Samoilys et al., 2017; Samoilys, Osuka, et al., 2019). Small‐mesh gillnets capture high proportions of juveniles, whereas large mesh sizes capture adults, including sharks and rays (Osuka et al., 2021). Prohibiting gill nets in buffer zones around protected areas or at least within locally managed protected areas (Kawaka et al., 2017) is recommended to enhance ecosystem function and conservation benefits (MacNeil et al., 2020, Goetze et al., 2024). Such prohibition of gillnets does, however, pose substantial socioeconomic challenges. A successful ban would require minimizing trade‐offs between ecological goals and the economic well‐being of fishers, ideally through working with fishers in gear modification trials and enforcement (Osuka et al., 2021) while facilitating livelihood diversification (Roscher et al., 2022).

We found that the best‐protected MPAs in the WIO, those with medium levels of protection, were not effective for several large‐bodied coral reef fishes, including sharks and 3 globally threatened species. Only the highly protected reefs of Iles Glorieuses and Chagos in the central Indian Ocean were effective in protecting these species. Medium‐level protection included nationally gazetted or privately established MPAs in Tanzania (some over 20 years old) and Mozambique and some reefs in Chagos (Diego Garcia) and Iles Glorieuses (Jeyser), where recreational fishing occurs. The medium‐protection MPAs in eastern Africa were, however, effective for the demersal large‐bodied groupers, snappers, and emperors. Eastern Africa is one of 5 regions with the lowest numbers of reef sharks in the world (MacNeil et al., 2020). Our results support these findings and suggest that sharks will disappear from reefs in the WIO if their management continues to be neglected. Our results also highlight the refuge that the Chagos Archipelago is providing to reef sharks in the WIO, despite the poaching (Collins et al., 2021). Based on the zero observations in this study of key reef fish species, including globally threatened large‐bodied species (Green Humphead parrotfish, Humphead wrasse, and Giant grouper), we recommend urgent action to initiate protection for these 3 species in the WIO and to expand standard coral reef monitoring to include the timed LS for large‐bodied reef fishes. We recommend additional conservation actions include targeted fisheries restrictions, for example, on gill nets, and species‐specific protection of critical habitats to improve the effectiveness of protected areas.

Supporting information

Appendix S1: Number of rapid long swim transects for large‐bodied reef fishes per location in each region in the western and central Indian Ocean and Gulf of Aden with Protection index, size of MPA and year established.

COBI-39-e14430-s001.docx (2.5MB, docx)

Appendix S2. Schematic diagram of a 10 minute long swim (LS) survey along a coral reef slope.

COBI-39-e14430-s003.jpg (44.9KB, jpg)

Appendix S3: Median (interquartile range—iqr) and mean (standard deviation ‐sd) biomass for 10 taxa groups across protection levels in the Western and Central Indian Ocean and Gulf of Aden.

COBI-39-e14430-s002.docx (16.4KB, docx)

ACKNOWLEDGMENTS

Field surveys were funded by a combination of grants to CORDIO East Africa from WIOMSA's Marine Science for Management Programme, Conservation International Madagascar, and the European Union's Biodiversity Management Programmes in Mozambique through ZSL and in the Horn of Africa through IGAD and IUCN. Surveys in Chagos were supported by the BIOT Section Administration of the Foreign Office, UK Government. We thank the captain, officers, and crew of the BIOT patrol vessel Pacific Marlin for invaluable operational support during the expedition. Surveys in Iles Glorieuses were supported by France's Iles Eparses program, through the Institut National des Sciences de l'Univers, Institut de Recherche pour le Developpement, Agence des Aires Marines Protegees, and logistic support of Terres Australes et Antarctiques Françaises. M.S. thanks B. Randriamanantsoa, S. Yayha, I. da Silva, and H. Mahudi for invaluable diving support and sharing of local knowledge.

Samoilys, M. , Osuka, K. E. , Roche, R. , Koldewey, H. , & Chabanet, P. (2025). Effects of protection on large‐bodied reef fishes in the western Indian Ocean. Conservation Biology, 39, e14430. 10.1111/cobi.14430

Article impact statement: Marine protected areas in the western Indian Ocean are not effective for large‐bodied fishes (e.g., shark, trevally, Green Humphead parrotfish).

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Associated Data

This section collects any data citations, data availability statements, or supplementary materials included in this article.

Supplementary Materials

Appendix S1: Number of rapid long swim transects for large‐bodied reef fishes per location in each region in the western and central Indian Ocean and Gulf of Aden with Protection index, size of MPA and year established.

COBI-39-e14430-s001.docx (2.5MB, docx)

Appendix S2. Schematic diagram of a 10 minute long swim (LS) survey along a coral reef slope.

COBI-39-e14430-s003.jpg (44.9KB, jpg)

Appendix S3: Median (interquartile range—iqr) and mean (standard deviation ‐sd) biomass for 10 taxa groups across protection levels in the Western and Central Indian Ocean and Gulf of Aden.

COBI-39-e14430-s002.docx (16.4KB, docx)

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