UNIVERSIDAD DE COSTA RICA 

SISTEMA DE ESTUDIOS DE POSGRADO 

 

 

 

 

 

Restaurando el arrecife: estrategias para el crecimiento de 

corales en viveros in situ en Bahía Culebra, Pacífico Norte de 

Costa Rica 

 

 

 

 

 

Tesis sometida a la consideración de la Comisión del Programa de Estudios de Posgrado 

en Biología para optar al grado y título de Maestría Académica en Biología 

 

 

 

 

 

SÒNIA FABREGAT MALÉ 

 

 

 

 

 

Ciudad Universitaria Rodrigo Facio, Costa Rica 

2022 

  



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DEDICATORIA 

 

 

 

 

 

 

 

 

 

 

 

A la platja del Pas 

  



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AGRADECIMIENTOS 

En primer lugar, quiero agradecer a mi tutor Juan José Alvarado su asesoramiento y 

constante apoyo. Gracias por brindarme su experiencia y darme las oportunidades de 

crecer como científica. A mi comité de tesis, la Dra. Paola Rodríguez Troncoso y el Dr. 

Jeffrey Sibaja, por su retroalimentación y enriquecer mi trabajo con su conocimiento. 

Asimismo, agradezco al Dr. Carlos E. Jiménez y la Dra. Celeste Sánchez Noguera su 

revisión y aportes en el primer capítulo de mi tesis. Asimismo, doy las gracias a la Dra. 

Guaria Cardenes y al personal de la Escuela de Geología de la Universidad de Costa Rica 

por su colaboración en el corte de las colonias de coral en sus instalaciones. 

Este proyecto no habría sido posible sin la ayuda y el apoyo el equipo científico 

del Centro de Investigación en Ciencias del Mar y Limnología (CIMAR), tanto de 

estudiantes como investigadores que nos acompañaron en las salidas a campo mes a mes: 

Adriana Arce, Fabio Quesada, Ben Chomitz, Dra. Cindy Fernández, Juan Carlos 

Azofeifa, Dra. Celeste Sánchez Noguera, Sergio Madrigal, Andrea Bogantes, Camila 

Valverde, Maricruz Calvo y Randy Domínguez. Agradezco, sobre todo, a mis dos 

compañeros de gira más longevos, Sebastián Mena y Gabriela López, por todas las horas 

de trabajo compartidas y la complicidad que generamos bajo el agua. Gracias a Eddy 

Gómez y Juan Guillermo Sagot por los análisis de mis muestras, a Davis Morera, Eleazar 

Ruiz y el resto del equipo del CIMAR. Estoy tremendamente orgullosa de haber formado 

parte del centro y haber tenido el privilegio de trabajar con gente con tanto talento, pasión 

y dedicación. 

Agradezco el apoyo de Península Papagayo y de su increíble personal, desde 

Mónica Gutiérrez, Elsiana Luna, Miguel Sánchez a todo el staff del Dive Bar y Marina 

Papagayo. Especialmente, quiero darle las gracias a Carlos Marenco, por su constante 

atención y ayuda en campo. Gracias también a las compañías de buceo Almaco y BA 

Divers, junto con sus capitanes Michi, Alonso, Lenín y Harry, quienes nos llevaron tantas 

veces a los sitios de restauración durante las giras. Agradezco el apoyo de la Agencia 

Alemana para la Cooperación Internacional (GIZ) en el desarrollo del proyecto de 

restauración en Bahía Culebra. Gracias al equipo de Raising Coral Costa Rica por la 

ayuda técnica inicial y el establecimiento de los primeros tres viveros del proyecto, allá 

en 2019. Agradezco también a Lisa Combillet su ayuda en el procesamiento y medición 

inicial de imágenes que son ahora parte de los resultados de esta tesis. 



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A mis amigos de Catalunya, por su continua amistad pese a habernos podido ver 

tan poco cuando decidí quedarme del otro lado del Atlántico, y por haberme mantenido 

conectada con el lugar del que vengo. Gracias, sin ningún orden particular, Clara, 

Mariona, Alfred, Oriol, Arnau, Mel, Júlia y Ariadna. 

Quiero agradecer a Chepe el estar siempre presente, alentarme y ser una fuente de 

apoyo tanto en los buenos como en los malos momentos, prácticamente desde que llegué 

al país. Gracias por su amistad, oportunos consejos y hacerme creer en mí un poco más. 

Espero poder volver a compartir giras con usted en algún momento – sin despertarlo a las 

cinco de la mañana, lo prometo. 

Finalmente, quiero mostrar mi especial agradecimiento hacia las siguientes 

personas: a la familia Madrigal-Mora, por ser mi segunda familia en este país y hacerme 

sentir una más de ustedes. Gracias por cuidarme y ayudarme siempre que lo he necesitado. 

Siempre van a tener las puertas de mi casa abiertas, esté donde esté. Los quiero 

muchísimo. A mis dos mejores amigos durante mis cuatro años en Costa Rica, Sergio 

Madrigal y Ben Chomitz. Gracias por las giras llenas de risas, la complicidad, las noches 

y caminadas por San José, por escuchar y apoyarme en todo momento – sé que no siempre 

es fácil – y las conversaciones llenas de crisis existenciales. Muy probablemente no habría 

llegado hasta aquí sin su constante apoyo y amistad. Y a vos, Checho, por no desaparecer. 

Gracias a Xandri, por ser illa i recer, por haber aparecido en mi vida cuando menos 

lo esperaba, por estar siempre en este tramo final, entender, apoyarme y darme la energía 

necesaria para terminar finalmente esta carrera de fondo. A mi pequeña familia entera, 

pero especialmente a mis papás y a mi hermano, por su amor, consejos, oportunidades 

brindadas y apoyo incondicional en todo lo que hago, sobre todo desde que decidí irme a 

Costa Rica. Gràcies per fer-me obrir els ulls, cada estiu, a tot el que hi ha sota l’aigua i 

descobrir-me un nou món. Gràcies per fer-me qui sóc. Us estimo. 

  





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ÍNDICE 

DEDICATORIA ........................................................................................................... ii 

AGRADECIMIENTOS ............................................................................................... iii 

ÍNDICE ....................................................................................................................... vi 

ÍNDICE DE FIGURAS .............................................................................................. viii 

ÍNDICE DE TABLAS ................................................................................................. xi 

RESUMEN ................................................................................................................. xii 

Chapter 1. A story of disturbance and loss: historical coral reef degradation in Bahía 

Culebra, North Pacific of Costa Rica ............................................................................. 1 

Abstract .................................................................................................................... 1 

1. Introduction ........................................................................................................... 1 

2. Methods ................................................................................................................ 4 

2.1. Study area....................................................................................................... 5 

3. Results and discussion ........................................................................................... 6 

3.1. The “pre-disturbed” period (1970-2000) ......................................................... 6 

3.2. The early degradation period (2000-2010) .................................................... 10 

3.3. The degraded period (2010–present) ............................................................. 14 

4. Is recovery possible for coral reefs in Bahía Culebra? .......................................... 20 

5. References........................................................................................................... 22 

Chapter 2. Testing the feasibility of coral nurseries in an upwelling area in the North 

Pacific of Costa Rica ................................................................................................... 39 

Abstract .................................................................................................................. 39 

1. Introduction ......................................................................................................... 40 

2. Materials and methods ......................................................................................... 42 

2.1. Study area..................................................................................................... 42 

2.2. Experimental design ..................................................................................... 43 

2.3. Data collection.............................................................................................. 47 

2.4. Data analysis ................................................................................................ 47 

3. Results ................................................................................................................ 48 

3.1. Coral fragment survival ................................................................................ 48 

3.2. Coral fragment growth .................................................................................. 51 

3.3. Nursery and fragment costs .......................................................................... 58 



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4. Discussion ........................................................................................................... 59 

5. References........................................................................................................... 66 

Chapter 3. Nursery-reared coral outplanting success in an upwelling-influenced area in 

Costa Rica................................................................................................................... 79 

Abstract .................................................................................................................. 79 

1. Introduction ......................................................................................................... 80 

2. Material & Methods ............................................................................................ 82 

2.1. Study area..................................................................................................... 82 

2.2. Experimental design ..................................................................................... 83 

2.3. Data analysis ................................................................................................ 85 

3. Results ................................................................................................................ 85 

3.1 Outplant survival ........................................................................................... 85 

3.2. Outplant growth............................................................................................ 87 

4. Discussion ........................................................................................................... 91 

5. References........................................................................................................... 95 

 

 

 

 

  



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ÍNDICE DE FIGURAS 

Fig. 1.1. Location of coral reefs in Bahía Culebra (North Pacific of Costa Rica) in 1995-

1996. ............................................................................................................................. 5 

Fig. 1.2. Proliferation and expansion of the macroalgae Caulerpa sertularioides in Bahía 

Culebra (2001-2014) based on patch size and abundance categories in Fernández 2007.

 ................................................................................................................................... 13 

Fig. 1.3. 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. ........................................................................................... 16 

Fig. 1.4. 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......................................................................................... 18 

Fig. 2.1. Location of nursery site (★) and donor colonies sites (●) in Bahía Culebra, North 

Pacific of Costa Rica. .................................................................................................. 43 

Fig. 2.4. Survivorship curves (%) of (A) Pocillopora spp. fragments according to nursery 

type (A-frames [n = 112 ●], rope line [n = 150 ●], and coral tree [n = 72 ●]); and (B) 

microfragments of massive Pavona clavus (n = 37 ●), Pavona gigantea (n = 148 ●), and 

Porites lobata (n = 66 ●) in the nursery site in Playa Jícaro, Bahía Culebra, North Pacific 

of Costa Rica, during nursery stage. ............................................................................ 49 

Fig. 2.5. Survivorship curves (%) of Pocillopora spp. fragments according to donor 

colony site in each type of nursery (A = coral tree, B = rope line, C = A-frame), during 

nursery stage in Playa Jícaro, Bahía Culebra, North Pacific of Costa Rica. Final 

survivorship at the end of nursery stage is shown next to each donor site. ................... 50 

Fig. 2.6. Presence (% of fragments affected) of bleaching or paleness (bars), and partial 

tissue loss (lines) in microfragments of massive species (Pavona clavus ■, Pavona 

gigantea ■, and Porites lobata ■) during nursery stage (September 2019-August 2020) in 

Playa Jícaro, Bahía Culebra, North Pacific of Costa Rica. No data is available for March 

2020 due to the COVID-19 pandemic national lockdown. ........................................... 51 



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Fig. 2.7. Mean growth rate (cm yr-1) of Pocillopora spp. fragments from shared donor 

sites (Jícaro and Matapalo) between nurseries. * = p < 0.05......................................... 53 

Fig. 2.8. Change of massive species microfragment area (cm2) from the start to the end 

of nursery stage (September 2019-August 2020), in the nursery site in Playa Jícaro, Bahía 

Culebra, North Pacific of Costa Rica. .......................................................................... 54 

Fig. 2.9. Mean (±SD) monthly seawater temperature (ºC) in Playa Jícaro restoration site 

(Bahía Culebra, North Pacific of Costa Rica) during nursery periods for coral fragments 

in coral trees and rope line nurseries (A) and A-frames (B). Shaded areas represent 

upwelling period (December to April). ........................................................................ 55 

Fig. 2.10. Monthly mean nutrient concentration (µM) in nursery site in Playa Jícaro, 

Bahía Culebra (Costa Rica), from December 2019 to August 2021. No data is available 

between March and May 2020 due to the COVID-19 pandemic national lockdown……56 

Fig. 2.11. Mean (±SD) monthly photosynthetic active radiation (PAR) (µM s-1 m-2) in 

Playa Jícaro restoration site (Bahía Culebra, North Pacific of Costa Rica) during nursery 

period.......................................................................................................................... 57 

Fig. 2.12. Mean linear growth rate (cm yr-1) of stained Pocillopora spp. fragments 

growing in the nursery and on the reef in Bahía Culebra, North Pacific of Costa Rica. 58 

Fig. 3.1. Location of Güiri-Güiri outplanting site (■), Jícaro nursery site (★), and donor 

colonies sites (●) in Bahía Culebra, North Pacific of Costa Rica. ................................ 83 

Fig. 3.2. Outplanting technique for nursery-reared Pocillopora spp. colonies (A) and for 

massive species (Porites lobata, Pavona gigantea and Pavona clavus) fragments (B) in 

Güiri-Güiri reef, Bahía Culebra, North Pacific of Costa Rica. ..................................... 84 

Fig. 3.3. Survivorship curves (%) for the massive coral fragments (Pavona clavus n = 8 

[⸺], Pavona gigantea n = 18 [⸺], and Porites lobata n = 5 [⸺]) in the outplanting site 

in Bahía Culebra, North Pacific of Costa Rica, through monitoring period (September 

2020-September 2021). ............................................................................................... 86 

Fig. 3.4. Percentage (%) of outplanted fragments of massive species (Pavona clavus [■], 

Pavona gigantea [■], and Porites lobata [■]) exhibiting partial tissue loss through 

monitoring period (September 2020 to September 2021), in Güiri-Güiri reef, Bahía 

Culebra, North Pacific of Costa Rica. .......................................................................... 87 



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Fig. 3.5. Increase of Pocillopora spp. colony area (cm2) from September 2020 to 

September 2021 (A), and growth of the same Pocillopora spp. outplant at the start of 

monitoring period (left), after six months (middle), and after one year (right) (B), in the 

outplanting site in Güiri-Güiri, Bahía Culebra, North Pacific of Costa Rica. ................ 88 

Fig. 3.6 Changes in area (cm2) of outplanted fragments of massive fragments (Pavona 

clavus [⸺], Pavona gigantea [⸺], and Porites lobata [⸺]) over monitoring period 

(September 2020-September 2021) in Güiri-Güiri reef, Bahía Culebra, Costa Rica. .... 89 

Fig. 3.7. Differences of mean monthly growth rate (cm2 mo-1) of Pocillopora spp. 

outplants between different sites of origin of donor colonies at Güiri-Güiri restoration 

site, Bahía Culebra, Costa Rica. .................................................................................. 91 

Fig. 3.S1. Seawater temperature (ºC) recorded in outplanting site in Güiri-Güiri reef, 

Bahía Culebra, North Pacific of Costa Rica, according to season of upwelling (■) or non-

upwelling (■). ...……………………………………………………………………….107 

Fig. 3.S2. Surface seawater salinity recorded in outplanting site in Güiri-Güiri reef, Bahía 

Culebra, North Pacific of Costa Rica, according to season of non-upwelling (wet season, 

■) or upwelling (dry season, ■). ……………………………………………………….108 

Fig. 3.S3. Nutrient concentration (µM) of ammonium (NH4
+, ■), nitrates (NO3

-, ■), 

nitrites (NO2
-, ■), phosphates (PO4

3-, ■), and silicates (SiO4, ■) in outplanting site in Güiri-

Güiri, Bahía Culebra, Costa Rica. ……………………………………………………..109 

 

 

 

 

 

 

  



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ÍNDICE DE TABLAS 

Table 2.1. Number of donor colonies and total number of fragments from each donor site 

for each nursery and species cultivated in the nursery site in Playa Jícaro, Bahía Culebra, 

North Pacific of Costa Rica. ........................................................................................ 46 

Table 2.2. Initial and final sizes (mean length ± SD), and mean growth rate (± SD) of 

Pocillopora spp. fragments cultivated in the different in situ coral nurseries in Playa 

Jícaro (Bahía Culebra, North Pacific of Costa Rica). Growth rate was calculated from 

fragments that survived throughout the whole monitoring period. * p < 0.005. ............ 52 

Table 2.3. Initial and final sizes (mean area ± SD), and mean growth rate (± SD) of 

microfragments of massive species cultivated in in situ coral nurseries in Playa Jícaro 

(Bahía Culebra, North Pacific of Costa Rica). Growth rate was calculated from fragments 

that survived throughout the whole monitoring period. * = p < 0.05. ........................... 54 

Table 2.4. Nutrient concentration (µM) in nursery site in Playa Jícaro, Bahía Culebra 

(Costa Rica), from October 2019 to August 2021. ND = not detectable by method. ..... 57 

Table 3.1. Initial and final sizes (mean area and diameter ± SD), and mean growth rate 

(± SD) of outplanted species to Güiri-Güiri restoration site (Bahía Culebra, North Pacific 

of Costa Rica). ............................................................................................................ 89 

Table 3.2. Nutrient concentration (µM) in outplanting site Güiri-Güiri, Bahía Culebra 

(Costa Rica), from September 2020 to September 2021. ND = not detectable by method. 

* = p < 0.05. ................................................................................................................ 90 

 

  



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RESUMEN 

Los arrecifes de coral son ecosistemas de una elevadísima diversidad biológica, que 

brindan a millones de personas una gran variedad de servicios ecosistémicos. Sin 

embargo, en las últimas décadas, han sufrido una degradación alarmante debido al 

incremento de las temperaturas, la acidificación oceánica, la sobrepesca y el desarrollo 

costero no planificado, poniendo en riesgo su estructura y funcionamiento ecológicos. 

Esta pérdida sin precedentes ha favorecido el desarrollo de medidas de conservación 

activas, como la restauración ecológica, con el objetivo de revertir esta tendencia y 

recuperar las funciones y servicios de estos ecosistemas.   

 En los últimos años, el número de iniciativas de restauración coralina en el 

Pacífico Tropical Oriental (PTO) ha incrementado drásticamente. Sin embargo, la 

experiencia es todavía limitada y pocos proyectos usan la jardinería de coral. Este método, 

basado en la reproducción asexual por fragmentación de los corales, comprende una fase 

inicial de cultivo en vivero hasta que los fragmentos alcanzan cierto tamaño y son 

trasplantados a un arrecife a intervenir. El género Pocillopora es el más utilizado en 

proyectos en el PTO, debido a su rápido crecimiento y contribución a la complejidad 

estructural por su morfología ramificada, mientras que el uso especies masivas, de 

crecimiento lento, es aún escaso. Las particulares condiciones oceanográficas del PTO, 

con fenómenos periódicos de El Niño-Oscilación Sur (ENOS), áreas de afloramiento 

estacional, baja saturación de aragonita y amplias fluctuaciones de salinidad y nutrientes, 

crean la necesidad de adaptar las técnicas de restauración a las características de la región. 

En Costa Rica, los esfuerzos de restauración anteriores se han centrado en el 

Pacífico Sur. La menor precipitación y el afloramiento estacional en el Pacífico Norte 

marcan claras diferencias en las condiciones ambientales entre las regiones. Esto queda 

reflejado en una composición de especies de coral formadoras de arrecife y dinámicas 

arrecifales distintas. Además, las presiones antropogénicas que históricamente han 

afectado –y siguen afectando– los ecosistemas coralinos de ambas regiones son 

diferentes, la del norte mucho más influenciada por el desarrollo costero, sobrepesca e 

impactos derivados. Como consecuencia, estos factores deberán considerarse para 

adecuar los esfuerzos de restauración de cada una de las zonas. 

Hace apenas dos décadas, Bahía Culebra, en el Pacífico Norte de Costa Rica, 

albergaba algunos de los arrecifes de coral más extensos del Pacífico del país. Sin 

embargo, la situación sufrió un drástico cambio a inicios de los 2000s, debido a múltiples 



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perturbaciones que ocurrieron en un corto periodo de tiempo y tuvieron efectos en cascada 

sobre las comunidades asociadas. Como resultado, los arrecifes se degradaron 

severamente y muchos de ellos colapsaron y desaparecieron por completo, dejando atrás 

cementerios de esqueletos y arena. 

Los ecosistemas coralinos de Bahía Culebra siguen siendo de importancia 

socioeconómica para las comunidades costeras aledañas. Aunque, en la última década, 

varias de las perturbaciones que causaron la degradación cesaron, debido a su estado 

actual, la recuperación natural es poco probable. Por ende, su trayectoria a futuro va a 

depender de las acciones de manejo y conservación que permitan su preservación a largo 

plazo. En este contexto, en 2019 se inició un proyecto de restauración basado en la 

jardinería de coral. Entender qué perturbaciones han afectado históricamente el sitio a 

restaurar y cómo ha respondido el ecosistema es un paso previo clave para el desarrollo 

de un plan de restauración. Dicha información es especialmente relevante considerando 

los futuros cambios pronosticados en las condiciones ambientales. 

En el Capítulo 1, se analizó y sintetizó la evidencia sobre arrecifes y ecosistemas 

coralinos y su entorno en Bahía Culebra, con el objetivo de (i) identificar las causas y 

consecuencias de su deterioro y (ii) proponer medidas de gestión para promover la 

recuperación ecosistémica. Para ello, se realizó una búsqueda de publicaciones y 

literatura gris, además de incluir datos no publicados del Laboratorio de Ecología y 

Conservación de Ecosistemas Arrecifales del Centro de Investigación en Ciencias del 

Mar y Limnología de la Universidad de Costa Rica. 

Los Capítulos 2 y 3 se centran en la evaluación del primer proyecto de restauración 

en Bahía Culebra. El Capítulo 2 está enfocado en la fase de cultivo en viveros in situ de 

fragmentos de coral, con los objetivos de (i) monitorear la salud y supervivencia de los 

fragmentos de Pocillopora spp., Pavona gigantea (Verrill 1869), Pavona clavus (Dana 

1846) y Porites lobata (Dana 1846); (ii) cuantificar su tasa de crecimiento en los viveros 

y en el arrecife; y (iii) determinar el efecto del tipo de vivero y las variables ambientales 

en el crecimiento y la supervivencia. Para este propósito, se obtuvieron fragmentos de 

Pocillopora spp. (n = 334), P. gigantea (n = 148), P. clavus (n = 37) y P. lobata (n = 66) 

a partir de colonias donantes de distintos sitios en la bahía. Los fragmentos de Pocillopora 

spp. se colocaron en tres tipos de vivero diferentes: árbol de coral (n = 72), tendedero (n 

= 150) y marco en A (n = 112), mientras que los microfragmentos de las especies masivas, 

solo en el árbol. Durante un año, se tomaron fotografías de cada fragmento mensualmente, 



xiv 

 

que se analizaron con ImageJ para medir el largo, ancho (cm) y área (cm2) y establecer 

su supervivencia. Adicionalmente, se hizo una tinción con rojo de alizarina a fragmentos 

en los viveros y en el arrecife, se dejaron crecer durante 12 meses, y se colectaron y 

cortaron para medir el crecimiento desde la línea de tinción. 

Por otro lado, el Capítulo 3 está enfocado en la evaluación del programa piloto de 

trasplante de colonias cultivadas en vivero a un arrecife degradado, para (i) monitorear su 

salud y supervivencia, (ii) cuantificar sus tasas de crecimiento y (iii) determinar si el sitio 

donante de los fragmentos y la presencia de afloramiento influyen sobre el crecimiento. 

Para ello, se trasplantaron 30 colonias de Pocillopora spp. (de seis sitios donantes) y 31 

fragmentos de especies masivas (P. gigantea [n = 18], P. clavus [n = 8], and P. lobata [n 

= 5]) al arrecife de Güiri-Güiri. Mensualmente durante un año, se determinó su 

supervivencia y estado de salud. El crecimiento se midió mediante fotocuadrantes 

alrededor de las colonias, y se obtuvo el área (cm2), el porcentaje de cobertura del 

fotocuadrante (%) y el diámetro (cm) con PhotoQuad. Se cuantificó la densidad del erizo 

de mar Diadema mexicanum (A. Agassiz 1863) mediante censos visuales, por su papel 

como bioerosionador y herbívoro de competidores de coral (macroalgas y turf). 

Para caracterizar los sitios de cultivo y trasplante, se colocaron sensores in situ de 

temperatura y radiación fotosintéticamente activa. Además, se colectaron muestras de 

agua mensualmente para determinar la salinidad y concentración de nutrientes (nitratos, 

nitritos, amonios, fosfatos y silicatos). 

En el Capítulo 1, se dividió la historia reciente de los arrecifes de Bahía Culebra 

en tres periodos, según la intensidad de perturbaciones antropogénicas. En el 1r periodo 

(1970-2000), los arrecifes albergaban la riqueza de especies de coral más alta de la costa 

pacífica del país, formados principalmente por los géneros Pocillopora y Pavona. Pese a 

la poca incidencia de presiones antrópicas, destacaban la extracción de especies 

ornamentales y fenómenos ENOS, que generaron blanqueamiento y mortalidades 

masivas de coral. En el 2º periodo (2000-2010), el acelerado desarrollo costero y 

eutrofización de las aguas ocasionaron eventos recurrentes e intensos de floraciones 

algales nocivas que causaron una gran mortalidad coralina. A su vez, la invasión de la 

macroalga Caulerpa sertularioides (M. Howe 1905) y subsecuente explosión poblacional 

del erizo bioerosionador D. mexicanum marcan el inicio del cambio de fase hacia arrecifes 

dominados por macroalgas. En el último periodo (2010-presente), destaca la pérdida de 

cobertura de coral (<4%) y de complejidad estructural, así como una reducción en la 



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diversidad y abundancia de ictiofauna y cambios en la composición de criptofauna. Para 

impulsar la recuperación de los arrecifes de la bahía, se proponen el planeamiento espacial 

marino, la mitigación y monitoreo de factores estresantes y la restauración ecológica. 

En el Capítulo 2, el crecimiento de Pocillopora spp. en los viveros fue un 107% 

mayor que en el arrecife, lo que podría indicar que los viveros favorecen el crecimiento. 

La supervivencia final fue de 49.4% para Pocillopora spp. y de 54.1% en P. clavus, 

31.8% en P. gigantea y 12.1% en P. lobata. La supervivencia sufrió una drástica 

disminución posteriormente al cese temporal del mantenimiento debido al confinamiento 

nacional por la pandemia de COVID-19. El crecimiento de Pocillopora spp. fue mayor 

en los viveros flotantes que en los marcos en A. Los microfragmentos de las especies 

masivas crecieron más lentamente y se vieron afectados por dos episodios de pérdida de 

pigmentación, causando una mortalidad y pérdida de tejido masivas. Estos episodios no 

afectaron a Pocillopora spp., lo que podría indicar una mayor tolerancia a anomalías 

térmicas, contrariamente a lo que se creía previamente. 

En el Capítulo 3, la supervivencia final de los trasplantes fue del 100% para 

Pocillopora spp., pero menor para las especies masivas (71.4% en P. clavus, 47.5% en 

P. gigantea y 20% en P. lobata). En estas últimas, se observó pérdida de tejido y marcas 

de depredación. Los trasplantes de Pocillopora spp. incrementaron su tamaño inicial un 

333.1%, a 9.98 ± 1.69 cm año-1, influenciado por el sitio donante del fragmento original, 

pero no por la presencia de afloramiento. Los fragmentos trasplantados de especies 

masivas crecieron más lentamente, entre 1.35 y 0.61 cm año-1. 

Los resultados indican un proceso de aclimatación a las condiciones locales de 

afloramiento. Asimismo, demuestran que la jardinería de coral es una técnica viable para 

el crecimiento de Pocillopora spp., pero requiere modificaciones para las especies de 

crecimiento masivo. El sitio de trasplante también permitió un buen desempeño de las 

colonias de Pocillopora spp., pero demostró no ser apto para los microfragmentos, 

demasiado pequeños y vulnerables ante la interacción con otras especies arrecifales. Por 

ende, se recomienda trasplantar fragmentos de mayor tamaño (>25 cm2), lo que 

disminuiría su susceptibilidad a perturbaciones. Considerando las diferentes respuestas 

entre especies frente a perturbaciones, es crucial no centrar los esfuerzos únicamente en 

Pocillopora, sino desarrollar un enfoque multi-especie, para así maximizar el éxito de la 

restauración a largo plazo. Esta información es útil para guiar futuros esfuerzos de 

restauración en zonas del PTO y de afloramiento estacional.   



1 

 

Chapter 1. A story of disturbance and loss: historical coral reef degradation in 

Bahía Culebra, North Pacific of Costa Rica 

Abstract 

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. In this chapter, 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. 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). 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 

unprecedented 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. In order 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. 

Keywords: coral cover, Eastern Tropical Pacific, ecosystem recovery, historical ecology, 

resilience, phase shift. 

1. Introduction 

Worldwide, live coral cover is declining and coral reef ecosystems are collapsing at 

unprecedented rates due to several stressors acting in tandem (Hughes et al. 2017, 

Knowlton et al. 2021, Dixon et al. 2022). These disturbances (e.g., ocean warming, 

acidification, overfishing, unplanned coastal development) can jeopardize the structure 



2 

 

and ecological functioning of coral reefs (Graham et al. 2011), by compromising coral 

growth and survival, reef composition, and functional diversity (Hughes et al. 2017, 

Dietzel et al. 2020, McWilliam et al. 2020). In some cases, several acute and/or 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 

ecosystem recovery or force it to remain below a certain coral-cover threshold (Hughes 

et al. 2010, Zaneveld et al. 2016, de Bakker et al. 2016, Bozec et al. 2021). A new stable 

benthic 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 have the ability to 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 (Hughes 

et al. 2017, Pratchett et al. 2020, Dixon et al. 2022). 

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 (Norström et al. 2009, Arias-Godínez et al. 2019, 2021). Algal-dominated states 

generally may reduce habitat heterogeneity and structural complexity (Pratchett et al. 

2014), which will also affect reef-associated species (Stella et al. 2011, Chong-Seng et al. 

2012, Ainsworth & Mumby 2015, Salas-Moya et al. 2021), 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 services (Norström et 

al. 2009, Cheal et al. 2013, 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 

(Lamy et al. 2016, Hughes et al. 2018, Romero-Torres et al. 2020). Nonetheless, 

predicting such trajectories is highly complex due to the numerous factors (e.g., 

disturbance intensity and frequency, ecosystem connectivity and ecological 

characteristics) that can influence reef resilience and determine trajectories of recovery 

(Hughes & Tanner 2000, Done et al. 2010, Graham et al. 2011, McClanahan et al. 2014). 

In the Eastern Tropical Pacific (ETP), coral reefs have shown particular resilience 

in response to historical disturbances, and coral loss events have not translated into 



3 

 

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), 

considered their primary threat and main driver of reduction in coral cover (Manzello et 

al. 2008, Zapata et al. 2010, Romero-Torres et al. 2020). 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 recovery 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 differences could be 

attributed to the degree and isolation from anthropogenic impacts, and protection status 

(Cortés et al. 2010, Alvarado et al. 2019). Whilst the first two areas are protected and 

relatively isolated from anthropogenic disturbances, Bahía Culebra has no level of 

protection and is located in an extensive coastal development 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 (Jiménez 2001a, Arias-Godínez et 

al. 2019, 2021). However, due to ENSO events (e.g., 1997-1998), harmful algal blooms 

(HABs), macroalgal invasion mainly by Caulerpa sertularioides (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 (Morales-Ramírez et al. 2001, Fernández 2007, Jiménez 2007a, Vargas-



4 

 

Montero et al. 2008, Fernández-García et al. 2012, Sánchez-Noguera et al. 2018b, Arias-

Godínez et al. 2019, 2021). Even though in the last decade (2010s) some stressors have 

ceased (J.J. Alvarado pers. comm.), 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 previous environmental disturbances is 

especially relevant considering the predicted future changes in environmental conditions. 

Identifying and understanding coral reef decline drivers and their ability to recover from 

past disturbance through historical data is essential for conservation and management 

measures, such as coral restoration (Knowlton & Jackson 2008, Cheal et al. 2010, 

Kittinger et al. 2011, Bruno et al. 2014, Godfray & May 2014). 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. 

2. Methods 

To determine drivers and historical coral reef development in Bahía Culebra, we 

identified published and grey literature data focusing on coral reef ecology in the area. 

Additionally, we included unpublished data from the Laboratorio de Ecología y 

Conservación de Ecosistemas Arrecifales from Centro de Investigación en Ciencias del 

Mar y Limnología, 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 degradation in Bahía Culebra, and divided their recent history into three 

periods, according to the level of intensity of anthropogenic disturbance 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. 



5 

 

2.1. 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. 1.1). For the purpose of 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). 

 

Fig. 1.1. Location of coral reefs in Bahía Culebra (North Pacific of Costa Rica) in 1995-

1996.  



6 

 

3. Results and discussion 

3.1. The “pre-disturbed” period (1970-2000) 

3.1.1. Coral & benthic composition 

Historical data show 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 upwelling 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 papyracea (Dana 1846), Cycloseris curvata (Hoeksema 1989), and 

Pocillopora meandrina (Dana 1846) (Cortés & Murillo 1985, Cortés 2012). However, 

intense non-controlled extraction of adult coral colonies in the early 1980s, especially 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. Jiménez pers. comm.) 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, Jiménez pers. comm.). Other coral species, 

such as Porites lobata (Dana 1846), Pavona clavus (Dana 1846) and L. papyracea 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 (Jiménez 1998, 2001, 2007a,b, Cortés & Jiménez 2003). The results showed a 

total of nine coral reefs in the bay (Fig. 1.1) (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 (Glynn et al. 2014, Arias-Godínez et al. 2019). Even 



7 

 

though 20 coral species were reported during that time (Jiménez 1997, 2001), most sites 

were considered as monospecific, as they were dominated by pocilloporid corals (mainly 

Pocillopora elegans Dana 1846), forming flat carpets that extended in between rocky 

reefs and sandy bottoms, and contributed to 42% of all live coral cover (Jiménez 1998, 

2001). In the deeper parts of the reef (>13 m depth), P. clavus was the most abundant 

species (13.0 ± 3.6% cover), followed by Pavona gigantea (Verrill 1869), Psammocora 

spp. and L. papyracea (Jiménez 2001a, Cortés & Jiménez 2003). 

Back then, Bahía Culebra was also home to unique coral formations not 

commonly found in other areas of the ETP. The largest P. clavus 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. 1.1), there used to be the only reef in 

the ETP built by L. papyracea, which extended over an area of 2,500 m2 (Jiménez 1997, 

Cortés & Guzmán 1998), and the site also held the only live population of C. curvata 

known in Costa Rica (Jiménez 1998). Additionally, Bahía Culebra harbored a relatively 

large (0.3 ha) coral patch built mainly by Psammocora profundacella (Gardiner 1898) 

and Psammocora stellata (Verrill 1866) in Playa La Penca (Fig. 1.1), with a live coral 

cover of 42.8 ± 20.8% (Jiménez 1998). Psammocora-dominated environments are rare 

throughout the ETP, and only four used to exist in Costa Rica (Jiménez 1998, Bezy et al. 

2006). Despite the fact that 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, crustaceans, 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 (Guzmán & Cortés 2001, Jiménez et al. 2001, Cortés & 

Jiménez 2003, Alvarado et al. 2012), 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 



8 

 

most likely because of the absence of chronic anthropogenic stressors, the rapid growth 

rates of corals in the bay, which are higher than in other localities of the ETP (Jiménez & 

Cortés 2003), and low densities (0.2 ± 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 (Jiménez 1998, Alvarado et al. 2012). 

Despite densities of D. mexicanum being low during this period (Jiménez 2001a), 

telltale 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 (Jiménez unpublished data). 

3.1.2. Reef fish communities 

During this period, a total of 78 reef fish species from 32 different families were found in 

Bahía Culebra (Dominici-Arosemena et al. 2005, Arias-Godínez et al. 2019), which is 

similar to the species richness estimated for other coral reefs in the ETP region (Cortés et 

al. 2017). Regarding trophic composition, invertebrate feeders and planktivores were the 

most abundant groups, followed by mesopredators, some of which were commercially 

important (Dominici-Arosemena et al. 2005, Arias-Godínez et al. 2021). Several 

predatory species were recorded during this period, such as the scalloped hammerhead 

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 cover (Dominici 1999), and higher diversity 

occurred in shallow Pocillopora reefs (Dominici-Arosemena et al. 2005). These habitats 

contributed to high structural complexity and presence of microhabitats, which provide 

multiple food sources, shelter and protection from predators for many reef species 

(Depczynski & Bellwood 2004, Komyakova et al. 2013, Russ et al. 2020). 

3.1.3. Anthropic disturbances 

Bahía Culebra’s terrestrial and marine ecosystems have been exposed to both natural – 

such as ENSO events – and anthropogenic stressors over time (Sánchez-Noguera 2012a). 

The strong interest in the bay’s marine resources at the end of the 20th century led to 

conflict between the different users with contrasting interests to represent (Jiménez 1997, 

1998, 2001, Dominici 1999, Sánchez-Noguera 2012b, Sánchez-Noguera et al. 2018b). 



9 

 

The lack of a management plan to protect marine ecosystems from anthropic activities 

brought localized degradation of some coral reefs and some of their associated ecological 

communities, such as species used in ornamental trade (Jiménez 1998, 2001) and artisanal 

fishing (Gutiérrez 1994, Jiménez 1997, Dominici-Arosemena et al. 2005, Sánchez-

Noguera et al. 2018b). 

The coral species most affected by ornamental trade were Pocillopora grandis 

(Dana 1846), P. meandrina, and T. coccinea, as well as certain anemone species, 

mollusks, algae, and colorful invertebrates (Jiménez 1997, 2001, Cortés & Jiménez 2003). 

Around 35-45 ornamental reef fish were also extracted for aquarium trade, but extraction 

was particularly intense for Thalassoma lucasanum (Gill 1862), Pomachantus zonipectus 

(Gill 1862), Holacanthus 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 terminal 

phases of these species. Moreover, patches of coral reefs up to 25 m2 were destroyed in 

order 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 (Jiménez 1998, Cortés & Jiménez 2003). 

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 development began in the 1990s, when the largest tourist complex in Central 

America was built (Jiménez 1998), with no consideration on the potential negative effects 

on coral areas. For instance, urbanization caused the burial of a 100 m2 patch of a 

P. gigantea reef in Güiri-Güiri, causing 84% coral mortality (Jiménez 2001a). The 

following year, a P. clavus reef was also affected by sediment input caused by the erosion 

of an unpaved road, resulting in high sediment suspension and lower coral growth rates 

(Jiménez 1998, 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 upwelling season (Jiménez 2001a,b) and 



10 

 

the semi-enclosed morphology of the bay, which promotes particle dispersion (García-

Céspedes et al. 2004). 

3.2. The early degradation period (2000-2010) 

3.2.1. 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 unprecedented coastal development and an increase in the number of visitors 

(Jiménez 2001a, Sánchez-Noguera 2012b, Cortés & Reyes-Bonilla 2017) caused 

anthropic eutrophication and modified water quality (Fernández 2007, Fernández-García 

et al. 2012, Alvarado et al. 2018, Beita-Jiménez et al. 2019). 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 

recovering. 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 (Jiménez 2007, 

Alvarado et al. 2012, 2016, Fernández-García et al. 2012, Sánchez-Noguera 2012a). Even 

though during the early 2000s (2003-2006) some sites (e.g., Cacique, Islas Palmitas, 

Pelonas, Playa La Penca) experienced a most impressive pulse of reef expansion and 

growth (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-Céspedes et al. 

2004, García et al. 2006, Vargas et al. 2015, Vargas-Zamora et al. 2018). Moreover, 

petroleum hydrocarbons were absent in the bay’s waters (Acuña-González et al. 2004), 

and sedimentation and suspended matter (4.12 mg/L) 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. 



11 

 

3.2.2. Harmful algal blooms 

When coastal development first escalated, many tourist facilities lacked waste water 

treatment plants (Jiménez 1998, Cortés & Jiménez 2003, Fernández-García et al. 2012). 

With changes in land use, nutrient wash-off during rainy season, and seasonal upwelling, 

nutrients in the bay’s waters and nearby areas increased (Fernández 2007, Sánchez-

Noguera 2012a, Stuhldreier et al. 2015a, Alvarado et al. 2018). In consequence, HABs 

increased in frequency, magnitude and duration along the North Pacific coast of Costa 

Rica, especially between 2006 and 2009 (Jiménez 2001a, 2007a,b, Morales-Ramírez et 

al. 2001, Vargas-Montero et al. 2008, Cortés & Reyes-Bonilla 2017, Sánchez-Noguera et 

al. 2018b). These episodes can induce coral stress, bleaching, and mortality, as well as in 

fish and other marine taxa (Jiménez 2007a,b, Sánchez-Noguera et al. 2018b), through 

different mechanisms, such as the reduction of light penetration, direct toxicity, and the 

decrease in oxygen availability (Guzmán et al. 1990, Bauman et al. 2010). During the 

continuous episodes of HABs in 2007 in Bahía Culebra, which affected over 100 km of 

coastline, coral habitats in the bay experienced a most devastating decline in coral cover 

(Jiménez 2007b). 

3.2.3. 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 

(Fabricius 2005, Lapointe et al. 2005, De’ath & Fabricius 2010, Fernández-García et al. 

2012, Adam et al. 2021). This resulted in their highest cover in the early 2000s (Alvarado 

et al. 2012, Fernández-García et al. 2012), at the expense of reef-building corals, 

outcompeted then by algae, which limit –and, in some cases, inhibit– coral larval 

settlement (McCook et al. 2001, Kuffner et al. 2006, 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 density of herbivores in Bahía Culebra (Dominici-Arosemena et al. 2005, 

Villalobos-Rojas et al. 2014), therefore reducing top-down control on macroalgae and 

turf. Thus, previously coral-dominated reefs became dominated by turf and macroalgae, 

and coral cover decreased from 38.5% in 2006 to 2.5% in 2009 (Sánchez-Noguera 2012a, 

Alvarado et al. 2012). 



12 

 

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 (Meinesz 1999, 

Lapointe et al. 2005). Even though this macroalgae has historically 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. 1.2), 

particularly where the substrate consisted of Pocillopora spp. fragments, which provide 

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 (Jiménez 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), coinciding with the increase in nutrients 

that occurs in upwelling events (Rixen et al. 2012, Stuhldreier et al. 2015b,c). 

Additionally, its high asexual reproduction rate through fragmentation and low herbivory 

rates, partly due to the presence of toxins in its fronds (Doty & Aguilar-Santos 1966, 

Davis et al. 2005), favored the dispersal and settlement of this macroalgae on the coral 

areas of the bay (Fernández & Cortés 2005, Fernández 2007). 



13 

 

 

Fig. 1.2. Proliferation and expansion of the macroalgae Caulerpa sertularioides in Bahía 

Culebra (2001-2014) based on patch size and abundance categories in Fernández 2007. 

This accelerated spreading caused a 40.5% reduction in the growth rate of 

P. elegans (Fernández 2007), as C. sertularioides can affect coral growth through 

different mechanisms: (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 highly costly 

physiological processes such as growth (River & Edmunds 2001), (ii) overgrowing of 

live coral tissue, and (iii) sediment trapping, which decreases available light (Sato 1985, 

Fabricius 2005). 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 corals and thick beds of 

C. sertularioides (Bezy et al. 2006). In 2006, the Psammocora reef had completely 



14 

 

disappeared (J.J. Alvarado pers. comm.). Thus, where once were coral-dominated 

environments, a new benthic assemblage established, forming 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). 

3.2.4. 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, Villalobos-Rojas et al. 2014, Sánchez-Noguera et al. 2018b). 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, 2001, Alvarado et al. 2012, Glynn & 

Manzello 2015). 

The role of sea urchins in coral reef ecology 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 & Manzello 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 framework takes place. In consequence, calcareous structures 

weakened, and reef structural complexity eventually diminished (Alvarado et al. 2012, 

2016). 

3.3. The degraded period (2010–present) 

After almost a decade of being affected by several types of disturbances, ecological 

consequences 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. 



15 

 

3.3.1. 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. papyracea, C. curvata, P. meandrina), while dominant species in the 

1990s were the ones that remained present in the bay’s coral reefs (Sánchez-Noguera 

2012a, Sánchez-Noguera et al. 2018b). Additionally, drastic reductions in live coral cover 

in many coral reefs were reported 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. 1.3). 

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 pers. comm.) (Fig. 1.3). 



16 

 

 

Fig. 1.3. 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. 

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 corals, and produce secondary metabolites to deter grazing 

(Kuffner & Paul 2004, Kuffner et al. 2006, Charpy et al. 2012, Brocke et al. 2015). 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 

Pocillopora-associated fauna (i.e., cryptofauna), from obligate symbiotic species to 

boring opportunistic and facultative species (Salas-Moya et al. 2021). An increase in the 

abundance of bivalves of the genus Lithophaga (Leiosolenus) can contribute to a rise in 

internal bioerosion of coral colonies and carbonate substrate, particularly during 

upwelling season (Wizemann et al. 2018, Salas-Moya et al. 2021), 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 symbiotic species found in live coral colonies in 



17 

 

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 organisms (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 carbonate structures (Enochs & Manzello 2012a,b). Additionally, 

such a shift in cryptofauna composition could compromise reef recovery, since they are 

a key component for host colony health, and the maintenance and recovery of reef 

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

3.3.2. Diadema mexicanum bioerosion 

After the population explosion in 2009, D. mexicanum densities kept increasing for some 

years (Fig. 1.4), 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. unpublished 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 (Eakin 2001, Sánchez-Noguera 2012a, Arias-

Godínez et al. 2019). 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 entirely destroyed, and no reef 

framework exists nowadays (pers. obs.). 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 pers. comm.). Instead, substrate is now covered in coral rubble and sandy 

bottoms (Sánchez-Noguera 2012a, Arias-Godínez et al. 2019). 

When no more carbonate substrate was available to erode, D. mexicanum 

populations started to decline (Fig. 1.4a), and with them their bioerosive impact, 

especially in sites where densities were highest, like Playa Blanca (Sánchez-Noguera et 

al. 2018b, Alvarado et al. unpublished data), 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, 2001). Nevertheless, densities of 

D. mexicanum 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 (Chapter 3), which represents over a 60,000% increase (Fig. 1.4b). 



18 

 

 

Fig. 1.4. 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. 

3.3.3. Reef fish communities 

The loss of structural complexity caused by intense sea urchin bioerosion and loss of live 

coral cover had an effect on reef fish communities in Bahía Culebra (Arias-Godínez et al. 

2019, 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 families), and 49% of species sighted in the 1990s were not present in 



19 

 

the latter degraded period (Arias-Godínez et al. 2021). Habitat composition and structural 

complexity are important drivers of reef fish abundance and diversity (Ferrari et al. 2017, 

Eisele et al. 2021). 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 ecological interactions (Gratwicke & Speight 2005, 

Coker et al. 2012). 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), particularly after the invasion of a sole 

macroalgae species, which could explain the observed reduction in both species diversity 

and abundance (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, Pratchett et al. 2015). Hence, 

their reduction can be considered an indicator of early coral reef degradation (Pratchett et 

al. 2006, Flower et al. 2017). 

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 mesopredator species, which were 

more abundant on pre-disturbed coral reefs (Arias-Godínez et al. 2021). In fact, over half 

of the mesopredatory 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 (Villalobos-Rojas et al. 2014, Beita-Jiménez 

et al. 2019), and potentially declining reef health (Aburto‐Oropeza et al. 2015). Declines 

in mesopredators and top predators 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). 

 

 



20 

 

4. Is recovery possible for coral reefs in Bahía Culebra? 

Historical insights of ecosystem changes can shed light on factors influencing ecosystem 

resilience and causing phase shifts among ecological states (Fong et al. 2006). How an 

ecosystem responded to previous environmental disturbances is particularly relevant 

considering future global change, in the face of which coral reefs are extremely vulnerable 

(Knowlton & Jackson 2008, Kittinger et al. 2011, Hughes et al. 2017, Hoegh-Guldberg 

et al. 2018). Thus, understanding how changing environmental conditions and anthropic 

disturbances shaped current coral reefs is key in order to implement informed 

management actions (Bruno et al. 2014, zu Ermgassen et al. 2015). 

Bahía Culebra is one of the most intensively studied regions in the Pacific coast 

of Costa Rica (Cortés 2012). Its marine environments, particularly coral reefs, have 

always received considerable attention for their ecologic and economic resources 

(Fernández 2007, Sánchez-Noguera 2012a). However, even though the first studies 

focusing on coral ecosystems 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 baselines 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 (Sánchez-Noguera et al. 2018b, Arias-Godínez et 

al. 2019), 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 column, HABs, invasion by the 

macroalgae C. sertularioides, and increased bioerosion triggered a series of reinforcing 

feedbacks that led to the collapse of coral reefs in less than 15 years (Jiménez 2001a, 

2007a, Fernández & Cortés 2005, Fernández 2007, Alvarado et al. 2012, 2016, Sánchez-

Noguera 2012a, Graham et al. 2013). This makes it difficult to identify the main drivers 

of degradation, 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 (Mumby 2009, Graham et al. 2015, 

Johns et al. 2018), which made recovery of coral cover to pre-disturbed states impossible. 



21 

 

But is the recovery of coral reefs in Bahía Culebra possible or should we lose all 

hope? Existing evidence points that, after major disturbances, coral recovery can occur in 

relatively short periods when stressors cease or are managed (Guzmán & Cortés 2001, 

Emslie et al. 2008, Graham et al. 2011, Gilmour et al. 2013). However, ecosystem 

recovery is much less likely 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 limited 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 capacity after disturbances (Guzmán & Cortés 2001, Bezy 2009). 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 in order 

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 protected areas, except for Bahía Culebra (Cortés & Jiménez 2003). This 

is particularly relevant considering the many reef resources used by different stakeholders 

in the bay (Jiménez 1997, 1998, 2001, Dominici 1999, 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 

in order 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 concentration, benthic cover of C. sertularioides and turf, and 

sea urchin populations, requires special attention, since they can serve as early warnings 



22 

 

of deterioration (Cooper et al. 2009, Gil et al. 2016, Flower et al. 2017). Nonetheless, 

while monitoring local stressors in the bay is relevant, upstream and larger-scale stressors 

that may have caused local coral deterioration require special attention. Integrated 

management actions like improving the region’s wastewater management and coastal 

land use are key to decrease HABs’ recurrence and intensity (Palmer et al. 2022). 

However, the current situation in the bay calls for the implementation of ecological 

restoration efforts so as 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 fisheries, coastal protection, tourism, and 

enhanced recreation opportunities (De Groot et al. 2013). Coupled with strong academic 

research and public enforcement, environmental education 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 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. 

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