Revista de Biología Tropical, ISSN: 2215-2075, Vol. 73(S1): e63624, enero-diciembre 2025 (Publicado Mar. 03, 2025)

A story of disturbance and loss: historical coral reef degradation

in Bahía Culebra, North Pacific of Costa Rica

Sònia Fabregat-Malé1; https://orcid.org/0000-0001-6764-0502

Juan José Alvarado1,2,3*; https://orcid.org/0000-0002-2620-9115

1. Centro de Investigación en Ciencias del Mar y Limnología (CIMAR), Universidad de Costa Rica, San Pedro, San José

11501-2060, Costa Rica; soniafabregat.m@gmail.com, juan.alvarado@ucr.ac.cr (*correspondence)

2. Centro de Investigación en Biodiversidad y Ecología Tropical (CIBET), Escuela de Biología, Universidad de Costa Rica,

San Pedro, San José 11501-2060, Costa Rica.

3. Escuela de Biología, Universidad de Costa Rica, San Pedro, San José 11501-2060, Costa Rica.

Received 05-III-2024. Corrected 20-VIII-2024. Accepted 24-I-2025.

ABSTRACT

Introduction: Climate change and other multiple stressors have globally caused the collapse of many coral reefs.

Understanding how coral reefs have responded to previous disturbances is key to identify possible trajectories in

the face of future more frequent and intense disturbances.

Objective: We review the ecological history of coral reefs in Bahía Culebra, a historically important area for coral

reef development in the North Pacific of Costa Rica, which has suffered extreme deterioration in the last decades.

Methods: We assessed historical traits of coral reefs using both historical and recent data, divided as follows: (i)

the “pre-disturbed” period (1970–2000), (ii) the early degradation period (2000–2010), and the degraded period

(2010–present day).

Results: Forty years ago, Bahía Culebra harbored the highest coral species richness in the Costa Rican Pacific,

with high live coral cover (> 40 %). Signs of early degradation were observed after El Niño events and unprec-

edented coastal development that caused anthropic eutrophication, which led to coral death and a shift to

macroalgae-dominated reefs. In the last decade, a steep decline in live coral cover (1–4 %), the loss of many reefs,

and a decrease in reef fish diversity and abundance were recorded.

Conclusions: To promote the recovery of coral reefs in the bay, we propose management actions such as marine

spatial planning, mitigation and monitoring of stressors, and ecological restoration. The latter could help turn

the tide by increasing live coral cover, eventually leading to ecosystem functionality recovery, with spill-over

effects on reef-associated communities, including local coastal communities. Nonetheless, such actions need

governmental and local support; thus, raising awareness through environmental education and citizen science

programs is key for the long-needed conservation of coral reefs in Bahía Culebra.

Key words: coral cover; Eastern Tropical Pacific; ecosystem recovery; historical ecology; resilience; phase shift.

RESUMEN

Una historia de perturbación y pérdida: degradación histórica de los arrecifes de coral

en Bahía Culebra, Pacífico Norte de Costa Rica

Introducción: El cambio climático y otros múltiples factores estresantes han provocado a nivel mundial el

colapso de muchos arrecifes de coral. Comprender cómo han respondido los arrecifes de coral a perturbaciones

anteriores es clave para identificar posibles trayectorias ante perturbaciones futuras más frecuentes e intensas.

https://doi.org/10.15517/rev.biol.trop..v73iS1.63624

SUPPLEMENT

2Revista de Biología Tropical, ISSN: 2215-2075 Vol. 73(S1): e63624, enero-diciembre 2025 (Publicado Mar. 03, 2025)

INTRODUCTION

Worldwide, live coral cover is declining and

coral reef ecosystems are collapsing at unprec-

edented rates due to several stressors acting in

tandem (Dixon et al., 2022; Hughes et al., 2017;

Knowlton et al., 2021). These disturbances

(e.g., ocean warming, acidification, overfishing,

unplanned coastal development) can jeopardize

the structure and ecological functioning of

coral reefs (Graham et al., 2011), by compro-

mising coral growth and survival, reef com-

position, and functional diversity (Dietzel et

al., 2020; Hughes et al., 2017; McWilliam et al.,

2020). In some cases, several acute and chronic

stressors occurring at once or in short periods

can act as reinforcing feedback mechanisms

which can impede the ability of corals to cope

with these disturbances, and thus prevent eco-

system recovery or force it to remain below

a certain coral-cover threshold (Bozec et al.,

2021; de Bakker et al., 2016; Hughes et al.,

2010; Zaneveld et al., 2016). A new stable ben-

thic assemblage, such as those dominated by

turf, cyanobacteria, and fleshy macroalgae, may

establish and cause a phase shift, modifying

ecosystem functionality (Bruno et al., 2009;

Dudgeon et al., 2010; Fung et al., 2011; Lesser,

2021). As highly dynamic ecosystems, reefs

could bounce back and recover (Rohr et al.,

2018; Romero-Torres et al., 2020). However, as

these disturbance events become more frequent

and intense, they can compromise reef recovery

time and thus, their resilience (Dixon et al.,

2022; Hughes et al., 2017; Pratchett et al., 2020).

In the long term, coral reef degradation

and ecological phase shifts to algal-dominated

states modify the whole coral reef ecosystem,

from benthic groups to top predators (Arias-

Godínez et al., 2019; Arias-Godínez et al., 2021;

Norström et al., 2009). Algal-dominated states

generally may reduce habitat heterogeneity and

structural complexity (Pratchett et al., 2014),

which will also affect reef-associated species

(Ainsworth & Mumby, 2015; Chong-Seng et al.,

2012; Salas-Moya et al., 2021; Stella et al., 2011),

which depend on corals for feeding, nursery,

and shelter (Pratchett et al., 2014). Hence, these

alternative stable states can lead to the local

extinction of some species, weaken fisheries

productivity (Ainsworth & Mumby, 2015), and

undermine critical ecosystem functions and

Objetivo: revisamos la historia ecológica de los arrecifes de coral en Bahía Culebra, un área históricamente

importante para el desarrollo de arrecifes de coral en el Pacífico Norte de Costa Rica, que ha sufrido un deterioro

extremo en las últimas décadas.

Métodos: Evaluamos los rasgos históricos de los arrecifes de coral utilizando datos históricos y recientes, dividi-

dos de la siguiente manera: (i) el período “pre-disturbio” (1970–2000), (ii) el período de degradación temprana

(2000–2010) y el período degradado (2010–actualidad).

Resultados: Hace cuarenta años, Bahía Culebra albergaba la mayor riqueza de especies de coral en el Pacífico cos-

tarricense, con una alta cobertura de coral vivo (> 40 %). Se observaron signos de degradación temprana después

de los fenómenos de El Niño y un desarrollo costero sin precedentes que provocó una eutrofización antrópica,

que provocó la muerte de los corales y un cambio hacia arrecifes dominados por macroalgas. En la última década,

se registró una fuerte disminución de la cobertura de coral vivo (1–4 %), la pérdida de muchos arrecifes y una

disminución en la diversidad y abundancia de peces de arrecife.

Conclusiones: Para promover la recuperación de los arrecifes de coral en la bahía, proponemos acciones de

manejo como la planificación espacial marina, la mitigación y monitoreo de factores estresantes y la restauración

ecológica. Esto último podría ayudar a cambiar la tendencia al aumentar la cobertura de coral vivo, lo que even-

tualmente conduciría a la recuperación de la funcionalidad del ecosistema, con efectos indirectos en las comuni-

dades asociadas a los arrecifes, incluidas las comunidades costeras locales. No obstante, tales acciones necesitan

apoyo gubernamental y local; por lo tanto, crear conciencia a través de programas de educación ambiental y

ciencia ciudadana es clave para la tan necesaria conservación de los arrecifes de coral en Bahía Culebra.

Palabras clave: cobertura coralina; Pacífico Tropical Oriental; recuperación de ecosistemas; ecología histórica;

resiliencia; cambio de fase.

Revista de Biología Tropical, ISSN: 2215-2075, Vol. 73(S1): e63624, enero-diciembre 2025 (Publicado Mar. 03, 2025)

services (Cheal et al., 2013; Norström et al.,

2009; Pratchett et al., 2014).

Multiple ecosystem trajectories can lead to

coral reef recovery if disturbances are managed

or cease entirely, although returning to their

original states might not be possible (Hughes

et al., 2018; Lamy et al., 2016; Romero-Torres et

al., 2020). Nonetheless, predicting such trajec-

tories is highly complex due to the numerous

factors (e.g., disturbance intensity and frequen-

cy, ecosystem connectivity and ecological char-

acteristics) that can influence reef resilience

and determine trajectories of recovery (Done et

al., 2010; Graham et al., 2011; Hughes & Tan-

ner, 2000; McClanahan et al., 2014).

In the Eastern Tropical Pacific (ETP), coral

reefs have shown resilience in response to

historical disturbances, and coral loss events

have not translated into lasting, region-wide

decline, but to long-term cycles of loss and

recovery (Romero-Torres et al., 2020). In this

region, coral reefs are small, discontinuous, and

dominated by few coral species (Glynn et al.,

2017). They are influenced by a low aragonite

saturation, fluctuations in nutrient and salinity

levels (Rixen et al., 2012; Sánchez-Noguera et

al., 2018a), and are periodically impacted by

El Niño-Southern Oscillation (ENSO), con-

sidered their primary threat and main driver

of reduction in coral cover (Manzello et al.,

2008; Romero-Torres et al., 2020; Zapata et

al., 2010). However, some regions have in fact

shown coral reef recovery after ENSO events

in Colombia (Zapata, 2017), Panama (Glynn

et al., 2014), Galápagos Islands (Glynn et al.,

2015), Mexico (Martínez-Castillo et al., 2022)

and Costa Rica (Guzmán & Cortés, 2007),

which recovered to pre-disturbance coral cover

levels. The observed resilience of reefs in the

ETP is promoted by (i) the fast-growth strategy

of the dominant coral species, (ii) presence of

thermotolerant symbionts, (iii) heterogeneous

incidence of non-optimal conditions such as

high irradiance and high-temperature stresses

across the ETP, and (iv) the possible existence

of ecological memory, by which the trajectory

of frequently stressed coral reef ecosystems is

shaped by previous conditions (Romero-Torres

et al., 2020).

Within the Pacific of Costa Rica, the recov-

ery of coral reefs has been disparate. While

coral reefs in areas such as Isla del Coco (an

oceanic island) or Isla del Caño (a continental

island in the South Pacific of the country) have

shown recovery after disturbance events, reefs

in mainland areas, like Bahía Culebra (North

Pacific coast), have not (Alvarado et al., 2012;

Sánchez-Noguera et al., 2018b). These differ-

ences could be attributed to the degree and

isolation from anthropogenic impacts, and pro-

tection status (Alvarado et al., 2019; Cortés et

al., 2010). Whilst the first two areas are protect-

ed and relatively isolated from anthropogenic

disturbances, Bahía Culebra has no level of

protection and is in an extensive coastal devel-

opment area (Sánchez-Noguera et al., 2018b).

Merely two decades ago, coral reefs in Bahía

Culebra were among the most extensive in the

North Pacific of Costa Rica, had high live coral

cover (44 % on average) and diversity of reef

fish species and ecological roles (Arias-Godínez

et al., 2019; Arias-Godínez et al., 2021; Jiménez,

2001a). However, due to ENSO events (e.g.,

1997–1998), harmful algal blooms (HABs),

macroalgal invasion mainly by Caulerpa sertu-

larioides (S.G.Gmelin) M.Howe 1905, and other

disturbances occurring in a short period of time

in the early 2000s, coral reefs in the bay severely

deteriorated, and many eventually collapsed,

with effects on reef-associated communities

(Arias-Godínez et al., 2019; Arias-Godínez et

al., 2021; Fernández, 2007; Fernández-García

et al., 2012; Jiménez, 2007a; Morales-Ramírez

et al., 2001; Sánchez-Noguera et al., 2018b;

Vargas-Montero et al., 2008). Even though in

the last decade (2010s) some stressors have

ceased (J. J. Alvarado, personal observation,

November 2023), the future trajectory of these

ecosystems is still uncertain and will depend

on management actions that support natural

recovery (Alvarado et al., 2018).

The response of coral ecosystems to previ-

ous environmental disturbances is especially

relevant considering the predicted future chang-

es in environmental conditions. Identifying and

4Revista de Biología Tropical, ISSN: 2215-2075 Vol. 73(S1): e63624, enero-diciembre 2025 (Publicado Mar. 03, 2025)

understanding coral reef decline drivers and

their ability to recover from past disturbance

through historical data is essential for con-

servation and management measures, such as

coral restoration (Bruno et al., 2014; Cheal et

al., 2010; Godfray & May, 2014; Kittinger et

al., 2011; Knowlton & Jackson, 2008). This is

especially pertinent in areas rapidly affected

by coastal development (Cortés, 2012), like

Bahía Culebra.

Here, we analyze the available evidence

on coral reefs, coral ecosystems, and their

environment in Bahía Culebra, and synthesize

current knowledge as a tool to (1) identify

the causes of coral reef decline over time and

the recorded ecological consequences of such

demise and (2) propose management actions

for ecosystem recovery.

MATERIALS AND METHODS

To determine drivers and historical coral

reef development in Bahía Culebra, we identi-

fied published and grey literature data focusing

on coral reef ecology in the area. Additionally,

we included unpublished data from the Labo-

ratorio de Ecología y Conservación de Eco-

sistemas Arrecifales Neotropicales (LECAN)

from Centro de Investigación en Ciencias del

Mar y Limnología (CIMAR), Universidad de

Costa Rica. Literature search was conducted via

Web of Science, Google Scholar and SCOPUS,

using the following keywords: “coral”, “reefs”,

and “Bahía Culebra”, “Golfo de Papagayo”, “fish”,

“bioerosion” and “algae”. We compiled and

chronologically arranged information on the

state of coral reefs and their drivers of deg-

radation in Bahía Culebra, and divided their

recent history into three periods, according to

the level of intensity of anthropogenic distur-

bance and degradation state of the reef: (i) the

“pre-disturbed” period (1970–2000s, when low

anthropogenic pressure was recorded); (ii) the

early degradation period (2000–2010, when

anthropogenic stressors intensified); and (iii)

the degraded and most recent period (2010s,

when reef framework collapsed). For each

period, we collected information on benthic

cover, natural and anthropic disturbances on

coral reefs, and their impacts on reef-associated

communities.

Study area: Bahía Culebra (10°37’N,

85°39’W) is a semi-enclosed bay in the Gulf of

Papagayo, North Pacific of Costa Rica (Fig. 1).

For this review, we consider Bahía Culebra as

the section of the coast from Islas Palmitas in

the north to Punta Cacique (Playa La Penca) in

the south, including the inner part of the bay,

based on previously published studies from the

same region. The bay extends for more than 20

km2 and reaches its maximum depth at about

42 m (Rodríguez-Sáenz & Rodríguez-Fonseca,

2004). The area is influenced by one of the

three seasonal upwelling systems occurring in

the ETP (Tehuantepec, Papagayo and Panama),

which affects the region from December to

April, lowering water temperature up to 10 ºC

from the annual mean (27 ± 0.1 ºC) (Jiménez,

2001b; Jiménez et al., 2010).

RESULTS AND DISCUSSION

The “pre-disturbed” period (1970–2000):

Coral & benthic composition: Historical

data that reef development in Bahía Culebra

occurred 350 years ≈ ago (Glynn et al., 1983).

Nonetheless, it is hypothesized that mass coral

mortality was caused by the Little Ice Age

(1675–1800 A.D.), which led to a decrease

in seawater temperature, and intensified and

extended the duration of the seasonal upwell-

ing in the Gulf of Papagayo (Glynn et al., 1983).

A few anecdotic reports exist on coral reefs in

Bahía Culebra between the 1930s and the 1940s

(Beebe, 1942; Fraser, 1943). In the late 1970s,

the coral framework was covered by algae, with

small and highly dispersed colonies (Glynn et

al., 1983). It is not until the 1980s when national

scientists from the Universidad de Costa Rica

reported the area as important for coral reef

development, with the highest richness on

coral species in the Costa Rican Pacific and the

presence of rare species, such as Leptoseris pap-

yracea (Dana, 1846), Cycloseris curvata (Hoek-

sema, 1989), and Pocillopora meandrina Dana,

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1846 (Cortés, 2012; Cortés & Murillo, 1985).

However, intense non-controlled extraction of

adult coral colonies in the early 1980s, especial-

ly for aquarism, led what once were abundant

species in the bay to become rare (Cortés &

Murillo, 1985). The extraction continued well

into the 1990s, with more groups such as black

corals and octocorals being targeted (C.E. Jimé-

nez, personal communication, June 2021) Later,

the area was affected by the particularly intense

El Niño event in 1982–83, which caused mass

coral mortality across the ETP (Glynn, 1984).

Barren and dead coral platforms were observed,

as well as dead coral colonies of several species

of Pocillopora –which were the most affected–

covered by algae (Cortés et al., 1984; C. E.

Jiménez, personal communication, June 2021).

Other coral species, such as Porites lobata Dana,

1846, Pavona clavus (Dana, 1846) and L. papy-

racea were also affected (Jiménez, 1998). Even

though coral mortality in Bahía Culebra after

the 1982–83 El Niño was enormous, it went

largely understudied and only qualitative data

exists (Jiménez, 2002).

It was not until the 1990s when coral

reefs in Bahía Culebra were mapped for the

first time (Cortés & Jiménez, 2003; Jiménez,

1998; Jiménez, 2001a; Jiménez, 2007a; Jimé-

nez, 2007b). The results showed a total of nine

coral reefs in the bay (Fig. 1) (Jiménez, 2001a),

with a mean live coral cover of 44.0 ± 3.3 %

(reaching up to 90 %), and an area that ranged

from 0.8 to 2.3 hectares (Cortés & Jiménez,

2003; Jiménez et al., 2010), which is similar to

other pre-disturbed fringing Pocillopora reefs

in the ETP (Arias-Godínez et al., 2019; Glynn

et al., 2014). Even though 20 coral species

were reported during that time (Jiménez, 1997;

Jiménez, 2001b), most sites were considered

as monospecific, as they were dominated by

pocilloporid corals (mainly Pocillopora elegans

Dana, 1846), forming flat carpets that extended

Fig. 1. Location of coral reefs in Bahía Culebra (North Pacific of Costa Rica) in 1995-1996.

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in between rocky reefs and sandy bottoms,

and contributed to 42 % of all live coral cover

(Jiménez, 1998; Jiménez, 2001a). In the deeper

parts of the reef (> 13 m depth), P. c l a v u s was

the most abundant species (13.0 ± 3.6 % cover),

followed by Pavona gigantea (Verrill 1869),

Psammocora spp. and L. papyracea (Cortés &

Jiménez, 2003; Jiménez, 2001a).

Back then, Bahía Culebra was also home to

unique coral formations not commonly found

in other areas of the ETP. The largest P. c l a v u s

reef in the ETP was found in Güiri-Güiri, with

colonies up to 10 m in diameter (Cortés &

Jiménez, 2003). On the other side of the bay,

in Punta Esmeralda (Fig. 1), there used to be

the only reef in the ETP built by L. papyra-

cea, which extended over an area of 2 500 m2

(Cortés & Guzmán, 1998; Jiménez, 1997), and

the site also held the only live population of C.

curvata known in Costa Rica (Jiménez, 1998).

Additionally, Bahía Culebra harbored a rela-

tively large (0.3 ha) coral patch built mainly by

Psammocora profundacella Gardiner, 1898 and

Psammocora stellata (Verrill, 1866) in Playa La

Penca (Fig. 1), with a live coral cover of 42.8 ±

20.8 % (Jiménez, 1998). Psammocora-dominat-

ed environments are rare throughout the ETP,

and only four used to exist in Costa Rica (Bezy

et al., 2006; Jiménez, 1998). Even though they

do not provide high habitat heterogeneity and

structural complexity (Bezy et al., 2006), they

shelter high biodiversity of invertebrates such

as polychaetes, amphipods, decapods, crus-

taceans, and mollusks, which attracts a broad

range of predator fauna (Glynn, 1974).

While the moderate 1991–92 El Niño event

had no notable effect on corals in Bahía Culebra

(Jiménez & Cortés, 2001), the intense 1997–98

event did cause thermal anomalies of +0.2 to

3.9 ºC, which led to mass coral bleaching that

mostly affected Pavona varians (Verrill, 1864),

P. lobata, and Tubastraea coccinea Lesson, 1830

(Jiménez et al., 2001). Coral mortality was

recorded in eight different species, such as

Pocillopora spp. colonies (> 60 % of all beached

and dead colonies), and a loss of > 90 % of L.

papyracea. Despite the magnitude of the event,

overall mortality was lower (7.2 %) than in

other areas of the ETP (Alvarado et al., 2012;

Cortés & Jiménez, 2003; Guzmán & Cortés,

2001; Jiménez et al., 2001), most likely due to

the combined effects of upwelling and high

cloud cover, which off-set the effects of high

temperatures (Jiménez, 2002; Palmer et al.,

2022). The natural recovery of Bahía Culebra’s

coral reefs was possible most likely because of

the absence of chronic anthropogenic stress-

ors, the rapid growth rates of corals in the bay,

which are higher than in other localities of the

ETP (Jiménez & Cortés, 2003), and low densi-

ties (0.20 ± 0.02 ind. m-2) of the bioeroder sea

urchin Diadema mexicanum A. Agassiz, 1863,

which translated to high bioaccretion rates (18

kg CaCO3 m-2) and low bioerosion (Alvarado et

al., 2012; Jiménez, 1998).

Despite densities of D. mexicanum being

low during this period (Jiménez, 2001a), tell-

tale signs of past population explosions were

observed. During visits to Bahía Culebra in

the late 1980s and early 1990s, remains of

highly bioeroded coral colonies were reported

in several sites, with undoubted signs of sea

urchin effects. Such bioerosion mostly affected

colonies of massive species, with little to no

effect on the Pocillopora reefs (C. E. Jiménez,

unpublished data).

Reef fish communities: During this period,

a total of 78 reef fish species from 32 different

families were found in Bahía Culebra (Arias-

Godínez et al., 2019; Dominici-Arosemena et

al., 2005), which is like the species richness

estimated for other coral reefs in the ETP

region (Cortés et al., 2017). Regarding trophic

composition, invertebrate feeders and plank-

tivores were the most abundant groups, fol-

lowed by mesopredators, some of which were

commercially important (Arias-Godínez et al.,

2021; Dominici-Arosemena et al., 2005). Sev-

eral predatory species were recorded during

this period, such as the scalloped hammer-

head shark (Sphyrna lewini (Griffith & Smith,

1834)) and the whitetip reef shark (Triaenodon

obesus (Rüppell, 1837)) (Arias-Godínez et al.,

2019). Most reef fish species in Bahía Culebra

showed a positive correlation with live coral

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cover (Dominici, 1999), and higher diversity

occurred in shallow Pocillopora reefs (Domi-

nici-Arosemena et al., 2005). These habitats

contributed to high structural complexity and

presence of microhabitats, which provide mul-

tiple food sources, shelter and protection from

predators for many reef species (Depczyn-

ski & Bellwood, 2004; Komyakova et al., 2013;

Russ et al., 2020).

Anthropic disturbances: Bahía Culebra’s

terrestrial and marine ecosystems have been

exposed to both natural – such as ENSO events

– and anthropogenic stressors over time (Sán-

chez-Noguera, 2012a). The strong interest in

the bay’s marine resources at the end of the

20th century led to conflict between the dif-

ferent users with contrasting interests to repre-

sent (Dominici, 1999; Jiménez, 1997; Jiménez,

1998; Jiménez, 2001a; Sánchez-Noguera, 2012b;

Sánchez-Noguera et al., 2018b). The lack of a

management plan to protect marine ecosystems

from anthropic activities brought localized deg-

radation of some coral reefs and some of their

associated ecological communities, such as spe-

cies used in ornamental trade (Jiménez, 1998;

Jiménez, 2001a) and artisanal fishing (Domi-

nici-Arosemena et al., 2005; Gutiérrez, 1994;

Jiménez, 1997; Sánchez-Noguera et al., 2018b).

The coral species most affected by orna-

mental trade were Pocillopora grandis (Dana,

1846), P. meandrina, and T. coccinea, as well as

certain anemone species, mollusks, algae, and

colorful invertebrates (Cortés & Jiménez, 2003;

Jiménez, 1997; Jiménez, 2001a). Around 35–45

ornamental reef fish were also extracted for

aquarium trade, but extraction was particularly

intense for Thalassoma lucasanum (Gill, 1862),

Pomacanthus zonipectus (Gill, 1862), Holacan-

thus passer Valenciennes, 1846, and Diodon

holocanthus Linnaeus, 1758 (Dominici, 1999;

Fournier, 2004). Dominici (1999) reported

declines in catch between 1999 and 2000, and

selective extraction of juveniles and male termi-

nal phases of these species. Moreover, patches

of coral reefs up to 25 m2 were destroyed to

extract reef fish and shrimps for aquarium

trade (Jiménez, 1997). Bahía Culebra’s coral

reefs also suffered mechanical impacts, as it was

already a snorkeling and diving hotspot in the

1990s (Jiménez, 1997), stirring up sediments,

breaking coral colonies or turning them over

(Cortés & Jiménez, 2003; Jiménez, 1998).

Finally, being the only large bay in the

area, protected from wave action and relatively

deep, Bahía Culebra had been a focal point for

coastal development in the North Pacific of

Costa Rica for many decades (Jiménez, 1998;

Sánchez-Noguera, 2012b). The intensive devel-

opment began in the 1990s, when the largest

tourist complex in Central America was built

(Jiménez, 1998), with no consideration of the

potential negative effects on coral areas. For

instance, urbanization caused the burial of a

100 m2 patch of a P. g i g a n t e a reef in Güiri-Güiri,

causing 84 % coral mortality (Jiménez, 2001a).

The following year, a P. c l a v u s reef was also

affected by sediment input caused by the ero-

sion of an unpaved road, resulting in high sedi-

ment suspension and lower coral growth rates

(Jiménez, 1998; Jiménez, 2001a). Mortality of L.

papyracea was observed after land movements

in Playa Panamá (Jiménez, 1998). In addition,

sea currents carried sediments, sewage, and fuel

leftover from the construction, and discharges

of the new tourist development to coral reef

patches, even relatively distant ones (Jiménez,

1998). This negative effect can be intensified

due to the strong trade winds during upwell-

ing season (Jiménez, 2001a; Jiménez, 2001b)

and the semi-enclosed morphology of the bay,

which promotes particle dispersion (García-

Céspedes et al., 2004).

The early degradation period (2000–2010):

Anthropic disturbances: The picture for

coral reefs in Bahía Culebra undoubtedly

started changing in the early 2000s, when

coral degradation was not localized but was

observed throughout the area. The unprec-

edented coastal development and an increase in

the number of visitors (Cortés & Reyes-Bonil-

la, 2017; Jiménez, 2001a; Sánchez-Noguera,

2012b) caused anthropic eutrophication and

modified water quality (Alvarado et al., 2018;

Beita-Jiménez et al., 2019; Fernández, 2007;

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Fernández-García et al., 2012). In addition, the

synergistic effect of El Niño events (1997, 2003,

2007, 2009) triggered a series of reinforcing

feedback mechanisms that acted in synergy and

prevented coral reef ecosystems from recover-

ing. The short time between disturbance events

(proliferation of HABs, increase in algae cover,

the propagation of the invasive macroalgae C.

sertularioides, and a subsequent increase in the

sea urchin D. mexicanum density), promoted a

general decline in live coral cover, undermining

the health of the whole ecosystem (Alvarado

et al., 2012; Alvarado et al., 2016; Fernández-

García et al., 2012; Jiménez, 2007a; Sánchez-

Noguera, 2012a). Even though during the early

2000s (2003–2006) some sites (e.g., Cacique,

Islas Palmitas, Pelonas, Playa La Penca) experi-

enced a most impressive pulse of reef expansion

and growth (C. E. Jiménez, unpublished data),

they were nonetheless eventually affected by

this generalized degradation.

At the beginning of this period (2000–

2002), Bahía Culebra still had clear waters and

low presence of heavy metals and total fecal

coliforms, which were 3 μg/g in lead and < 4

MPN/100 ml, respectively (Acuña-González et

al., 2004; García et al., 2006; García-Céspedes

et al., 2004; Vargas et al., 2015; Vargas-Zamora

et al., 2018). Moreover, petroleum hydrocar-

bons were absent in the bay’s waters (Acuña-

González et al., 2004), and sedimentation and

suspended matter (4.12 mg l-1) was also low

(Vargas-Zamora et al., 2018). The minimal

runoff discharges and wastes into the bay could

explain these low concentrations (Vargas-

Zamora et al., 2018). Despite this, intense

HABs (also known as red tides) occurred and

major disturbances occurred, but no specific

data were recorded.

Harmful algal blooms; When coastal devel-

opment first escalated, many tourist facilities

lacked wastewater treatment plants (Cortés &

Jiménez, 2003; Fernández-García et al., 2012;

Jiménez, 1998). With changes in land use, nutri-

ent wash-off during rainy season, and seasonal

upwelling, nutrients in the bay’s waters and

nearby areas increased (Alvarado et al., 2018;

Fernández, 2007; Sánchez-Noguera, 2012a;

Stuhldreier et al., 2015a). In consequence,

HABs increased in frequency, magnitude and

duration along the North Pacific coast of Costa

Rica, especially between 2006 and 2009 (Cortés

& Reyes-Bonilla, 2017; Jiménez 2001a; Jiménez,

2007a; Jiménez, 2007b; Morales-Ramírez et

al., 2001; Sánchez-Noguera et al., 2018b; Var-

gas-Montero et al., 2008). These episodes can

induce coral stress, bleaching, and mortality, as

well as in fish and other marine taxa (Jiménez,

2007a; Jiménez, 2007b; Sánchez-Noguera et al.,

2018b), through different mechanisms, such

as the reduction of light penetration, direct

toxicity, and the decrease in oxygen availabil-

ity (Bauman et al., 2010; Guzmán et al., 1990).

During the continuous episodes of HABs in

2007 in Bahía Culebra, which affected over 100

km of coastline, coral habitats in the bay experi-

enced a most devastating decline in coral cover

(Jiménez, 2007b).

Coral & benthic composition: Once corals

died, their skeletons became available space

for the recruitment of fleshy and filamentous

macroalgae and turf. Both groups thrive in

high-nutrient environments, particularly with

phosphates and nitrates, which promote and

accelerate their growth (Adam et al., 2021;

De’ath & Fabricius, 2010; Fabricius, 2005;

Fernández-García et al., 2012; Lapointe et al.,

2005). This resulted in their highest cover in the

early 2000s (Alvarado et al., 2012; Fernández-

García et al., 2012), at the expense of reef-build-

ing corals, outcompeted then by algae, which

limit –and, in some cases, inhibit– coral larval

settlement (Kuffner et al., 2006; McCook et al.,

2001; Roth et al., 2018). Hence, rates of coral

recruitment are inversely correlated to algal

abundance and cover (Kuffner & Paul, 2004),

which, in the long term, can impede coral reef

recovery after disturbance events (Adam et al.,

2021). Furthermore, the synergistic effect of

fishing pressure caused a decline in the den-

sity of herbivores in Bahía Culebra (Dominici-

Arosemena et al., 2005; Villalobos-Rojas et

al., 2014), therefore reducing top-down con-

trol on macroalgae and turf. Thus, previously

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coral-dominated reefs became dominated by

turf and macroalgae, and coral cover decreased

from 38.5 % in 2006 to 2.5 % in 2009 (Alvarado

et al., 2012; Sánchez-Noguera, 2012a).

Another stressor which promoted the

decline on live coral cover in the area was the

spread of the green invasive macroalgae C.

sertularioides (Fernández-García et al., 2012), a

well-known species for its invasive-like growth

in coastal waters (Lapointe et al., 2005; Meinesz,

1999). Even though this macroalgae has histori-

cally been found in Bahía Culebra, its densities

had always remained low until 2001, when its

proliferation began (Fernández, 2007). The first

rise was recorded in areas where many boats

anchored (i.e., Playas del Coco and Ocotal)

and close to coastal development like hotels

and human communities (Fernández, 2007).

From there, it began to spread along the bay

(Fig. 2), particularly where the substrate con-

sisted of Pocillopora spp. fragments, which pro-

vide a suitable surface for the attachment of C.

sertularioides stolons and rhizoids (Fernández-

García et al., 2012). In 2001, several hectares

of very dense mats were reported, smothering

coral colonies > 50 cm in height (C. E. Jimé-

nez, unpublished data). The maximum peak of

C. sertularioides cover was recorded in 2005,

when it reached 40 % during wet season and

70.7 % during the dry upwelling season, with

a growth rate of 1.1 cm day-1 (Fernández,

2007; Fernández-García et al., 2012), coincid-

ing with the increase in nutrients that occurs

in upwelling events (Rixen et al., 2012; Stuhl-

dreier et al., 2015b; Stuhldreier et al., 2015c).

Additionally, its high asexual reproduction rate

Fig. 2. Proliferation and expansion of the macroalgae Caulerpa sertularioides in Bahía Culebra (2001–2014) based on patch

size and abundance categories in Fernández (2007).

10 Revista de Biología Tropical, ISSN: 2215-2075 Vol. 73(S1): e63624, enero-diciembre 2025 (Publicado Mar. 03, 2025)

through fragmentation and low herbivory rates,

partly due to the presence of toxins in its fronds

(Davis et al., 2005; Doty & Aguilar-Santos,

1966), favored the dispersal and settlement of

this macroalgae on the coral areas of the bay

(Fernández, 2007; Fernández & Cortés, 2005).

This accelerated spreading caused a

40.5 % reduction in the growth rate of P. el e -

gans (Fernández, 2007), as C. sertularioides can

affect coral growth through different mecha-

nisms: (i) constant abrasion of fronds against

coral tissue, which damages it and makes coral

polyps retract, thus allocating more energy for

tissue repair, and limiting access to energy and

metabolites, crucial for non-essential and high-

ly costly physiological processes such as growth

(River & Edmunds, 2001), (ii) overgrowing of

live coral tissue, and (iii) sediment trapping,

which decreases available light (Fabricius, 2005;

Sato, 1985). Under these stressful conditions,

coral colonies may be partially damaged and

eventually die, resulting in a reduction of live

coral cover. For instance, the Psammocora-

dominated coral reef in Playa Penca lost a 95

% of its live coral cover and was reduced to

small patches (1–3 m2) found among dead cor-

als and thick beds of C. sertularioides (Bezy et

al., 2006). In 2006, the Psammocora reef had

completely disappeared (J. J. Alvarado, personal

observation, November 2023). Thus, where

once were coral-dominated environments, a

new benthic assemblage was established, form-

ing Caulerpa prairies (Alvarado et al., 2018).

Such a rapid spreading of a macroalgae, causing

abrupt declines in coral cover, was at the time

unheard of in the ETP (Bezy et al., 2006).

Diadema mexicanum bioerosion: The

increase of turf and macroalgae cover resulted

in a population explosion of the sea urchin D.

mexicanum, due to the additional availability of

food sources and strong fishing pressure in the

area, which caused a reduction of sea urchin

predators (Alvarado et al., 2012; Sánchez-

Noguera et al., 2018a; Villalobos-Rojas et al.,

2014). The densities of this herbivore bioeroder

began to increase in 2009 (over 900 % increase

since 2006) (Alvarado et al., 2012), not only due

to anthropogenic factors but to the 2009–2010

ENSO event, as reported for other areas of the

ETP (Eakin, 1996; Eakin, 2001; Alvarado et al.,

2012; Glynn & Manzello, 2015).

The role of sea urchins in coral reef ecol-

ogy and their carbonate budgets can change in

accordance with their density. When sea urchins

graze on algae growing on dead coral substrate,

they reduce macroalgae and turf cover and

facilitate the growth of crustose coralline algae,

which promote coral larval recruitment, hence

favoring ecosystem recovery (Alvarado et al.,

2012; Glynn & Manzello, 2015). However, when

found in moderate and high densities, they can

cause significant erosion of reef framework

(from 0.07 to 0.75 kg CaCO3 m-2 yr-1 in Bahía

Culebra) (Alvarado et al., 2012; Glynn & Man-

zello, 2015). When their density surpasses 1.5

ind m-2, such as in the case of Bahía Culebra

during this period, carbonate balance becomes

negative and active bioerosion of reef frame-

work takes place. In consequence, calcareous

structures weakened, and reef structural com-

plexity eventually diminished (Alvarado et al.,

2012; Alvarado et al., 2016).

The degraded period (2010–present):

After almost a decade of being affected by

several types of disturbances, ecological con-

sequences of coral reef deterioration in Bahía

Culebra started being notorious. Effects were

seen not only in terms of changes in benthic

cover and drastic reductions of live coral cover,

but also in the loss of coral reef framework and

structural complexity, which in turn affected

the different coral-associated communities.

Coral & benthic composition: The most

direct effects of coral reef deterioration in Bahía

Culebra were a reduction in live coral cover and

the number of coral species. Several rare coral

species suffered local extinction (e.g., L. papy-

racea, C. curvata, P. meandrina), while domi-

nant species in the 1990s were the ones that

remained present in the bay’s coral reefs (Sán-

chez-Noguera, 2012a; Sánchez-Noguera et al.,

2018b). Additionally, drastic reductions in live

coral cover in many coral reefs were reported

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after the disturbance period: in 2010–2011,

live coral cover was between 1–4 % (Sánchez-

Noguera, 2012a), and during 2014–2016, it

remained 1.3 ± 2.4 % (Arias-Godínez et al.,

2019). Most recent reevaluations reveal that live

coral cover is currently 1.1 ± 0.7 % (Alvarado et

al., unpublished data) (Fig. 3).

While in 2010–2011 the dominant substrate

cover was dead coral, which exceeded 85 % of

total cover in some cases (Sánchez-Noguera,

2012a), it was later (2014–2016) replaced by

macroalgae, turf and sand (Arias-Godínez et

al., 2019). During this period, turf reached 46.1

± 15.6 % of substrate cover in the bay, similar

to values reported for reefs in the ETP (Fong

et al., 2017), and macroalgae cover (mainly C.

sertularioides) was 14.0 ± 7.2 % (Alvarado et al.,

unpublished data). Even though there was an

increase in herbivore abundance (Alvarado et

al., 2012; Arias-Godínez et al., 2021), the high

growth and dispersal rates of C. sertularioides

ensured that it remained abundant in many

coral reefs in Bahía Culebra and prevailed over

coral cover (Arias-Godínez et al., 2019). Thus,

the previously reported shifts from coral to

macroalgae-dominated ecosystems intensified

and took place in most coral reefs in the bay.

For instance, the existing Pocillopora spp. reef

in Playa Blanca was replaced by extensive mats

of C. sertularioides (< 60 %) (Arias-Godínez,

2017). However, C. sertularioides cover has

been declining during the last years (2021: 0.8

± 0.3 %), especially after hurricane Otto (2016)

and tropical storm Nate (2017) hit the Pacific

coast of Costa Rica (J. J. Alvarado, personal

observation, November 2023) (Fig. 3).

Moreover, while cyanobacteria were absent

in the 1990s, they appeared as a new category of

substrate cover during reef evaluations in 2010–

2011, although their cover was < 1 % (Sánchez-

Noguera, 2012a). While cyanobacteria are a

natural component of coral reefs (Charpy et al.,

2012), they can form dense and extensive mats

in deteriorated reefs. Their presence inhibits

coral larval settlement and recruitment, and

they can act as pathogens for scleractinian cor-

als, and produce secondary metabolites to deter

grazing (Brocke et al., 2015; Charpy et al., 2012;

Kuffner & Paul, 2004; Kuffner et al., 2006).

Hence, when found in high densities, they are

considered indicators of nutrient enrichment

and declining reef health (Albert et al., 2005;

Charpy et al., 2012; Paerl & Paul, 2012).

Coral mortality in the 2000s also caused a

shift on composition and diversity of Pocillo-

pora-associated fauna (i.e., cryptofauna), from

Fig. 3. Mean (±SE) substrate cover (%) in Bahía Culebra from 2013 to 2021 (North Pacific, Costa Rica) in the four reefs

surveyed in Jiménez (1998) and Sánchez-Noguera et al. (2018b) (Güiri-Güiri, Punta Esmeralda, Islas Palmitas and Playa

Blanca). CCA: Crustose calcareous algae.

12 Revista de Biología Tropical, ISSN: 2215-2075 Vol. 73(S1): e63624, enero-diciembre 2025 (Publicado Mar. 03, 2025)

obligate symbiotic species to boring opportu-

nistic and facultative species (Salas-Moya et al.,

2021). An increase in the abundance of bivalves

of the genus Lithophaga (Leiosolenus) can con-

tribute to a rise in internal bioerosion of coral

colonies and carbonate substrate, particularly

during upwelling season (Salas-Moya et al.,

2021; Wizemann et al., 2018), thus weakening

the little remaining reef framework. Shifts in

cryptofauna communities can be promoted by

changes in food resources and chemical cues

once coral colonies die (Lecchini et al., 2014;

Wee et al., 2019). Furthermore, obligate symbi-

otic species found in live coral colonies in the

early 2000s, such as Trapezia sp. and Alpheus

sp., are known for their territorial behavior,

which could have prevented the settlement and

recruitment of other coral-associated organ-

isms (Tóth & Duffy, 2005; Wee et al., 2019), and

thus explain the increase of cryptofauna species

observed after coral death and reef degradation

in Bahía Culebra (Salas-Moya et al., 2021). In

the long term, however, species richness will

abruptly drop after the high erosion of car-

bonate structures (Enochs & Manzello, 2012a;

Enochs & Manzello, 2012b). Additionally, such

a shift in cryptofauna composition could com-

promise reef recovery, since they are a key

component for host colony health, and the

maintenance and recovery of reef framework

(Salas-Moya et al., 2021; Stella et al., 2011).

Diadema mexicanum bioerosion: After

the population explosion in 2009, D. mexica-

num densities kept increasing for some years

(Fig. 4), which produced a subsequent increase

in bioerosion rates, from 0.75 CaCO3 m-2 yr-1

in 2009 to 6.95 kg CaCO3 m-2 yr-1 in 2013

(Alvarado et al., 2012; Alvarado et al., unpub-

lished data). Such bioerosion rates exceeded

the bioaccretion capacity of coral reefs in Bahía

Culebra (< 0.01 kg CaCO3 m-2 yr-1), which

resulted in a debilitated structure and collapse

of the reef (Arias-Godínez et al., 2019; Eakin,

2001; Sánchez-Noguera, 2012a). For instance,

in Playa Blanca, where the highest sea urchin

densities (4.12 ± 0.83 ind. m-2) occurred during

2010-2011 surveys (Sánchez-Noguera, 2012a;

Sánchez-Noguera et al., 2018b), carbonate

structure was destroyed, and no reef framework

exists nowadays (J.J. Alvarado, personal obser-

vation, November 2023). The same occurred

in other sites, where nearly vertical 1.8 m

high coral scarps were grazed away, and reef

framework crumbled (C.E. Jiménez, personal

communication, June 2021). Instead, the sub-

strate is now covered in coral rubble and sandy

bottoms (Arias-Godínez et al., 2019; Sánchez-

Noguera, 2012a).

When no more carbonate substrate was

available to erode, D. mexicanum populations

started to decline (Fig. 4a), and with them their

bioerosive impact, especially in sites where den-

sities were highest, like Playa Blanca (Alvarado

et al., unpublished data; Sánchez-Noguera et al.,

2018b), potentially allowing for coral recovery

(Eakin, 2001; Guzmán & Cortés, 2007). This

follows what has been established in other

studies for coral reefs in the ETP (Eakin, 1996;

Eakin, 2001). Nevertheless, densities of D. mex-

icanum are still high in the bay, particularly

in reefs where carbonate framework remains,

such as Güiri-Güiri, where during 2010–2011,

its densities were 0.03 ± 0.01 ind. m-2 (Sánchez-

Noguera et al., 2018b), and have increased to

18.1 ± 1.16 ind. m-2 ten years later (Fabregat-

Malé et al., 2023) (Fig. 4b).

Reef fish communities: The loss of structur-

al complexity caused by intense sea urchin bio-

erosion and loss of live coral cover influenced

reef fish communities in Bahía Culebra (Arias-

Godínez et al., 2019; Arias-Godínez et al.,

2021). During 2014–2016 surveys, 56 reef fish

species from 24 families were detected, which

represents a significant reduction from species

richness in the 1990s (78 species from 32 fami-

lies), and 49 % of species sighted in the 1990s

were not present in the latter degraded period

(Arias-Godínez et al., 2021). Habitat composi-

tion and structural complexity are important

drivers of reef fish abundance and diversity

(Eisele et al., 2021; Ferrari et al., 2017). Habitats

with high live coral cover tend to hold a larger

diversity and abundance of reef fish, since

they provide shelter and different resources,

and thus promote species coexistence and key

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ecological interactions (Coker et al., 2012; Grat-

wicke & Speight, 2005). The severe degradation

and phase shift to macroalgal-dominated states

suffered by coral reefs in Bahía Culebra implied

a reduction of microhabitats and food sources

for reef fish (Arias-Godínez et al., 2019), par-

ticularly after the invasion of a sole macroal-

gae species, which could explain the observed

reduction in both species’ diversity and abun-

dance (Arias-Godínez et al., 2019). Coral-

dependent species were particularly affected,

such as the butterflyfish Chaetodon humeralis

(Günther, 1860) (Arias-Godínez et al., 2019).

These are one of the first groups affected by

loss of live coral cover, as they depend on it for

feeding and settlement (Graham et al., 2009;

Fig. 4. Populations of the sea urchin Diadema mexicanum in Bahía Culebra coral reefs over time. (A) Density of D.

mexicanum (ind. m-2) from April 2013 to August 2015; (B) Mean (±SE) density of D. mexicanum (ind. m-2) in four surveyed

coral reefs in Bahía Culebra, from 2014 to 2019.

14 Revista de Biología Tropical, ISSN: 2215-2075 Vol. 73(S1): e63624, enero-diciembre 2025 (Publicado Mar. 03, 2025)

Pratchett et al., 2015). Hence, their reduc-

tion can be considered an indicator of early

coral reef degradation (Flower et al., 2017;

Pratchett et al., 2006).

Trophic structure of reef fish assemblages

from this degraded period suffered a shift

from those back in the 1990s (1995–1996).

Even though planktivore fish remained as the

dominant group, other trophic groups were

greatly affected by coral reef degradation, such

as coral-dependent omnivore and mesopreda-

tor species, which were more abundant on

pre-disturbed coral reefs (Arias-Godínez et

al., 2021). In fact, over half of the mesopreda-

tory species (57 %) detected in the 1990s were

not sighted in the 2014–2016 surveys (Arias-

Godínez et al., 2021), including whitetip reef

shark T. obesus and scalloped hammerhead

shark S. lewini (Arias-Godínez et al., 2019).

Their decrease in abundance could be seen

as an indicator of strong fishing pressure in

the area, since this group is primarily targeted

by local fisheries (Beita-Jiménez et al., 2019;

Villalobos-Rojas et al., 2014), and potentially

declining reef health (Aburto-Oropeza et al.,

2015). Declines in mesopredators and top pred-

ators could have had cascading effects on the

whole ecosystem (Heithaus et al., 2008; Roff et

al., 2016; Sandin et al., 2022), by prey-release

mechanisms that led to predominance of lower

trophic level consumers (macroalgae-feeders,

herbivores-detritivores and invertivores) in the

now algae-dominated environments in Bahía

Culebra (Arias-Godínez et al., 2021).

Is recovery possible for coral reefs

in Bahía Culebra?:

Historical insights of ecosystem changes

can shed light on factors influencing ecosys-

tem resilience and causing phase shifts among

ecological states (Fong et al., 2006). How an

ecosystem responded to previous environ-

mental disturbances is particularly relevant

considering future global change, in the face

of which coral reefs are extremely vulnerable

(Hoegh-Guldberg et al., 2018; Hughes et al.,

2017; Kittinger et al., 2011; Knowlton & Jack-

son, 2008). Thus, understanding how changing

environmental conditions and anthropic dis-

turbances shaped current coral reefs is key

to implement informed management actions

(Bruno et al., 2014; Zu Ermgassen et al., 2015).

Bahía Culebra is one of the most inten-

sively studied regions in the Pacific coast of

Costa Rica (Cortés, 2012). Its marine environ-

ments, particularly coral reefs, have always

received considerable attention for their eco-

logic and economic resources (Fernández,

2007; Sánchez-Noguera, 2012a). However, even

though the first studies focusing on coral eco-

systems occurred previously to intensive coastal

development (Cortés & Murillo, 1985), human

settlement in the bay, and thus disturbances to

marine ecosystems, happened centuries before

(Sánchez-Noguera, 2012b). Hence, even if such

studies cannot be considered as historical base-

lines of pristine coral environments, they were

nevertheless carried out before the major and

intensive marine degradation that took place

in the 2000s.

The extreme and rapid degradation

observed in Bahía Culebra, with shifts from

coral to macroalgal-dominated states (Arias-

Godínez et al., 2019; Sánchez-Noguera et al.,

2018b), was most likely caused by numerous

disturbances acting synergistically and thus

diminishing reef resilience (Jiménez, 2007a;

Sánchez-Noguera et al., 2018b). The combined

action of increased nutrients in the water col-

umn, HABs, invasion by the macroalgae C.

sertularioides, and increased bioerosion trig-

gered a series of reinforcing feedbacks that led

to the collapse of coral reefs in less than 15 years

(Alvarado et al., 2012; Alvarado et al., 2016;

Fernández, 2007; Fernández & Cortés, 2005;

Graham et al., 2013; Jiménez, 2001a; Jiménez,

2007a; Sánchez-Noguera, 2012a). This makes it

difficult to identify the main drivers of degra-

dation, and which factors have maintained the

observed phase shifts and prevented natural

ecological recovery. The change in dominance

to fleshy macroalgal-states has led to stable

regime shifts in other regions (Graham et al.,

2015; Johns et al., 2018; Mumby, 2009), which

made recovery of coral cover to pre-disturbed

states impossible.

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But is the recovery of coral reefs in Bahía

Culebra possible or should we lose all hope?

Existing evidence points that, after major dis-

turbances, coral recovery can occur in relatively

short periods when stressors cease or are man-

aged (Emslie et al., 2008; Gilmour et al., 2013;

Graham et al., 2011; Guzmán & Cortés, 2001).

However, ecosystem recovery is much less like-

ly if live coral cover is < 5 % and reef framework

has collapsed (Graham et al., 2011). This is the

exact case of Bahía Culebra’s coral reefs, where

coral recruitment, and hence recovery, is lim-

ited by different factors. First, macroalgae and

turf cover is still high in the bay, inhibiting coral

larval settlement (Kuffner et al., 2006; Roth

et al., 2018). The loss of carbonate framework

due to intense bioerosion further contributes

to the limitation of available substrate for coral

larvae to recruit (Alvarado et al., 2012; Graham

& Nash, 2012). Moreover, sexual reproduction

rates of the main reef-building coral species

of the ETP are lower than in other regions,

which directly impacts natural recovery capac-

ity after disturbances (Bezy, 2009; Guzmán

& Cortés, 2001). Successful coral recruitment

is essential for ecosystem recovery, and its

inhibition or limitation by such factors could

prevent reversion from phase shifts (Hughes

& Tanner, 2000; Kuffner & Paul, 2004). It has

also been observed that reef recovery tends to

be slowest in the ETP, due to its geographic

isolation from other oceanographic regions

(through the 5 000–8 000 km Central Pacific

Barrier) and low functional diversity of corals

and reef fish (Graham et al., 2011). Compared

to other regions, ETP’s coral reefs are formed by

few coral species (Cortés, 1997; Veron, 2014),

which translates to low functional redundancy

(de Bakker et al., 2016).

Where we can go from here partly depends

on management actions taken to enhance coral

populations in Bahía Culebra and increase

their chances of natural recovery. In Costa

Rica, most coral reefs and important areas of

coral development can be found within pro-

tected areas, except for Bahía Culebra (Cortés

& Jiménez, 2003). This is particularly relevant

considering the many reef resources used by

different stakeholders in the bay (Dominici,

1999; Jiménez, 1997; Jiménez, 1998; Jiménez,

2001a; Sánchez-Noguera et al., 2012a), which

can lead to conflict between users (Sánchez-

Noguera et al., 2018b). The bay has no current

management plan of its resources, and marine

spatial planning is urgently needed to regulate

marine activities that could potentially clash

with coral reef conservation (Naranjo-Arriola,

2021). Continuing with existing monitoring

programs for stressors such as nutrient con-

centration, benthic cover of C. sertularioides

and turf, and sea urchin populations, requires

special attention, since they can serve as early

warnings of deterioration (Cooper et al., 2009;

Flower et al., 2017; Gil et al., 2016). Nonethe-

less, while monitoring local stressors in the bay

is relevant, upstream, and larger-scale stressors

that may have caused local coral deterioration

to require special attention. Integrated manage-

ment actions like improving the region’s waste-

water management and coastal land use are

key to decrease HABs’ recurrence and inten-

sity (Palmer et al., 2022). However, the current

situation in the bay calls for the implementation

of ecological restoration efforts to maintain

remaining diversity and increase live coral

cover back to a threshold where ecosystem

functionality is enhanced, and effects can be

seen on reef-associated communities. Coral reef

restoration also allows us to maintain existing

genetic diversity and coral populations while

climate change and other local and regional

anthropic stressors are managed or mitigated

(Baums et al., 2019). The recovery of coral reefs,

and thus restoration, can potentially provide

a wide range of economic benefits due to its

impact on ecosystem services, such as fisher-

ies, coastal protection, tourism, and enhanced

recreation opportunities (De Groot et al., 2013).

Coupled with strong academic research and

public enforcement, environmental educa-

tion and citizen science programs could also

increase public awareness and support of coral

conservation efforts, long needed for coral reefs

in Bahía Culebra (Dickinson et al., 2012; Hesley

et al., 2017). This review sets a guide for coral

restoration activities by providing a baseline

16 Revista de Biología Tropical, ISSN: 2215-2075 Vol. 73(S1): e63624, enero-diciembre 2025 (Publicado Mar. 03, 2025)

and describing how coral reefs in Bahía Culebra

have responded to past disturbances, and how

these have shaped the structure and function of

present-day coral reefs.

Ethical statement: the authors declare that

they all agree with this publication and made

significant contributions; that there is no con-

flict of interest of any kind; and that we fol-

lowed all pertinent ethical and legal procedures

and requirements. All financial sources are fully

and clearly stated in the acknowledgments sec-

tion. A signed document has been filed in the

journal archives.

ACKNOWLEDGMENTS

We are very grateful to all the people

who contributed information and data to this

work. We thank Carlos Jiménez, Jorge Cortés

and Celeste Sanchez for their collaboration.

Likewise, we thank Jeffrey Sibaja-Cordero and

Alma Paola Rodriguez Troncoso for comments

on a first version of this publication.

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