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Revista de Biología Tropical

On-line version ISSN 0034-7744

Rev. biol. trop vol.60  suppl.2 San José Apr. 2012

 

Reconstruction of Diadema mexicanum bioerosion impact on three Costa Rican Pacific coral reefs

Juan José Alvarado1*,2*, Jorge Cortés1,3 & Héctor Reyes-Bonilla4*

*Dirección para correspondencia


Abstract

The 1982-83 El Niño event produced a high coral mortality (50-90%) in several localities in the Eastern Tropical Pacific, which resulted in an outbreak of the sea urchin populations of Diadema mexicanum A. Agassiz, 1863 in some reefs, leading to an increase in coral framework bioerosion. In Costa Rica, El Niño impact varied among three of the most important coral reefs localities, being higher in Cocos Island, moderate in Caño Island, and lower in Culebra Bay; D. mexicanum densities followed the same pattern. To understand the historic role of this sea urchin on the balance between bioerosion and bioacretion, we made a reconstruction of bioerosion impact based on current patterns of carbonate ingestion by the sea urchins, growth rates and skeletal density of the main coral builders, and historical information of sea urchin population density and coral cover. The reconstruction model varied depending on locality. At Cocos Island, the effect on the reef carbonate budget ranged from negative to positive, improving coral recruitment and the recovery of the reef. At Caño Island, there was no apparent effect. In Culebra Bay, the effects ranged from a positive-neutral effect to a negative one, the latter possibly associated with an increase of eutrophic conditions that facilitated bioerosion of the coral framework. The importance of this sea urchin in reef dynamics varies with amount of reef protection, overfishing, and coastal management, and it has a large influence on the carbonate balances of the Pacific Costa Rican coral reefs.

Keywords: Density, coral cover, bioerosion, bioacretion, management, reef budget, Eastern Tropical Pacific.

Resumen

El  fenómeno de El Niño de 1982-83  produjo una alta  mortalidad  coralina  (50-90%)  en  varias  localidades del Pacífico Tropical Oriental, lo que en algunos arrecifes trajo como consecuencia una explosión en la poblaciones de erizos de mar, Diadema mexicanum, y por consiguiente un aumento en  la bioerosión del basamento coralino. En Costa Rica, el impacto fue diferencial en tres localidades arrecifales, siendo mayor en la Isla del Coco, intermedio en la Isla del Caño, y menor en Bahía Culebra, con similares patrones en la presencia del erizo D.  mexicanum. Con el fin de poder entender el  papel histórico que desempeña este erizo de  mar en el balance entre bioerosión y bioacreción, se reconstruyó el impacto  bioerosivo basándose en patrones actuales de ingestión de carbonatos por parte del erizo,  tasas de crecimiento y densidad del  esqueleto coralino, y datos históricos de  densidad poblacional del erizo y cobertura  coralina. Los resultados de las reconstrucciones variaron dependiendo de la  localidad.  En  la Isla del Coco, el efecto de los erizos de mar  varío de un efecto negativo sobre el balance arrecifal de carbonatos a un efecto positivo, favoreciendo el reclutamiento coralino y la recuperación del arrecife. En la Isla del Caño, Diadema presentó un efecto  neutro, al no tener una participación preponderante en el balance de carbonatos  de  esta isla. Mientras, que en Bahía Culebra, los efectos de los erizos de mar pasaron de tener un efecto positivo-neutro, a uno negativo, posiblemente asociado a un incremento en condiciones eutróficas de la bahía que están favoreciendo un incremento  en la bioerosión del basamento coralino. El valor de este erizo en la dinámica arrecifal y  su relación con la protección, sobrepesca, y manejo costero, posee una gran influencia en el balance de carbonatos en los arrecifes coralinos del Pacífico de Costa Rica.

Palabras  claves:  Densidad,  cobertura  coralina,  bioerosión,  bioacreción,  manejo,   balance  arrecifal,  Pacífico Tropical Oriental.

Eastern   Tropical   Pacific   (ETP)   reefs are thin CaCO3 accumulations in relation to their counterparts from the Indo-Pacific and Caribbean (Cortés 1997, Manzello et al. 2008). Most are relative small (1-2 ha), discontinuous, limited to shallow depths (10 m), built by few species, ephemeral in geological time (with low rates of carbonate production but fairly rapid framework accumulation), low cemetation, and have complex food webs (Glynn 1977, Glynn & Macintyre 1977, Macintyre et al. 1992, Cortés et al. 1994, Cortés 1997, Glynn 2004, 2008, Manzello et al. 2008). Two main types of reef structures can be observed in the region (Cortés 1997, 2003): reefs in Mexico, Panama, Colombia and some areas in Costa Rica and Ecuador are built by species of the genus Pocillopora. The other type of reef consists of Porites lobata or Pavona clavus massive corals, with the best developed being at Cocos, Clipperton and Galapagos islands.

The Eastern Tropical Pacific has been one of the most affected regions by sea warming as a result of El Niño events (Glynn & Colley 2001), with significant live coral cover losses at most localities (50-100%) due to bleaching (Glynn  1988,1990,  1996,  Guzman  &  Cortés 1992, 2001, 2007, Glynn et al. 2001, Jiménez &  Cortés  2001,  Jiménez  et  al.  2001,  Baker et al. 2008), which facilitated the growth of macroalgae and increased bioerosion by sea urchins (Glynn et al. 1979, 1988, 1997, Eakin 1992, 1996, 2001, Reaka-Kudla et al. 1996, Guzman  &  Cortés  1992,  2007).  During  the 1982-83 El Niño event in Costa Rica, a high coral mortality (90%) was observed at Cocos Island (Guzman & Cortés 1992), 50% mortality at Caño Island (Guzman et al. 1987), and death  of  coral  colonies  of  Pocillopora  spp. was  observed  in  Culebra  Bay  (Cortés  et  al. 1984). During the 1997-98 El Niño event, the coral reefs of these three localities were less impacted (~5% coral mortality) than other ETP reefs (Guzman & Cortés 2001, Jiménez et al. 2001), despite the fact that this El Niño was considered to be the most intense one of the last century (Enfield 2001).

After the 1982-83 El Niño event in Cocos Island, Guzman and Cortés (1992) predicted that coral recovery was going to take centuries due  to  the  high  densities  of  bioeroders  and the low coral reproduction rates. Guzman and Cortés (1992, 2007) indicated that part of the deterioration of the reef structures on Cocos island was due to the bioerosive action of Diadema mexicanum A. Agassiz, 1863. In 2002, there was a five-fold increased of coral cover and a notable reduction in sea urchins. Guzman & Cortés (2007), determined that D. mexicanum was not playing a relevant role in the reef bioerosion. However, the urchin still could fulfill its key role in assisting the recruitment of corals, as has been observed in other reefs in the Caribbean (Sammarco et al. 1974, Sammarco 1980, 1982a, b, Mumby et al. 2006).

We reconstruct the behavior of the bioerosion and bioacretion rates on the reef at Cocos and Caño islands and from Culebra Bay from 1980  until  2009,  using  published  and  field information of coral cover, growth and skeletal density, as well as population densities and daily carbonates  ingestions of D. mexicanum. Our goal was to understand the role that D. mexicanum plays in the balance of reef growth and destruction through time

Materials and Methods

To reconstruct a reef budget it was necessary to estimated the current bioacretion and bioerosion rates. For this, we visited three reef sites per study area (Fig. 1), where transects to  measure  coral  cover  and  sea  urchin  density were deployed. Three field studies were done in 2009: 1) 1-5 March, Cocos Island (5°31’45’’N-87°03’37’’W); 2) 18-20 July, Caño Island (8°41’36’’N-83°52’05’’W); and 3) 29-31 July, Culebra Bay (10°35’19’’N-85°40’27’’W).



Three 10 m long and 1 m wide transects were deployed parallel to the coast, between 4 and 8 m deep. A 1 m quadrat, divided into 0.01 m2 cells was placed every meter, on the right side of the transect line (Weinberg 1981). The amount of live, dead, and bleached coral cover, and also macroalgae, crustose coralline algae, rocks and sand cover were determined for each quadrat. In addition, sea urchin (D. mexicanum)  abundances  were  counted  along the 20m2  belt transect.

To determine the bioerosion rate per locality (removal of CaCO3  per time unit), 30 sea urchins  were  collected  (between  08:00  and 10:00 hours) and placed in plastic buckets (5 l), leaving them to evacuated the gut content for a period of 24 hours (Glynn 1988, Reyes-Bonilla & Calderón-Aguilera 1999). The evacuated material was collected and later dried in an oven at 60ºC for 24 h. The samples were weighted on an analytic balance (0.0001 g) and transferred to a furnace at 500ºC for 6 h. After this period, the samples (ashes) were measured again. The difference between the weights determined the amount of organic matter that is present in the sea urchin’s gut (Carreiro-Silva & McClanahan 2001). The remaining portion represents the  inorganic  fraction,  composed  by  CaCO3 and non-soluble residues (silica fragments like quartz grains, sponge spicules, diatoms, radiolarians and lime). Inorganic fractions were digested with 10% HCl. The remaining material, after dissolution of the carbonates, was filtered with a pre-weighed fiberglass filter. The weight of the residue material retained on the filter corresponds with the non-carbonate fraction (NCF). The difference between the ashes and the weight of the residue equals the CaCO3 (Carreiro-Silva & McClanahan 2001). This gives us the amount of carbonates removed by each sea urchin (CaCO3 g ind-1  d-1).

The bioerosion rates per urchin (kg m-2 y-1) at the different study areas were calculated  by  multiplying  the  daily  carbonate  content (CaCO3 g ind-1  d-1) by density of the sea urchins (ind m-2) over 365 days (modified from Carreiro-Silva & McClanahan 2001).

To  obtain  the  carbonate  deposition  rate of the coral, we used the method proposed by Chave et al. (1972), which combines the coral growth rate (cm y-1) and the coral skeletal density (g cm-3) to calculate the gross carbonate production. We  multiplied  this  result  by  the amount of live coral cover (LCC) to estimate the net deposition of carbonates on the reef (CaCO3 kg m-2 y-1).

To  obtain  the  skeletal  and  coral  growth rates,  we  sampled  10  fragments  of  colonies of Porites lobata from Cocos Island, five of P. lobata from Caño Island, five of Pocillopora elegans from Caño Island, and five of P. elegans from Culebra Bay. The samples were sent to the laboratory of  Ecología Marina, Centro de Investigación Científica y Educación Superior de Ensenada (CICESE), Baja California, México, where the Carricart-Gavinet and Barnes (2007) protocol was applied to obtain density values. The values for coral growth were taken from the literature (Guzman & Cortés 1989, Jiménez & Cortés 2003). For Cocos Island, 10 colonies were stained with Alizarin Red (Lamberts 1978) in August 2007, removed in September 2008, then sliced, measured and the growth rate calculated.

To reconstruct the carbonate budgets, we assumed that coral growth and skeletal density were the same for all the analyzed years, as well as the carbonate content present in the sea urchin’s guts. Coral growth can varied between years or seasons, so our reconstruction assumed an “ideal scenario” to asses only the impact  of  D.  mexicanum  on  reef  accretion. To  determine  the  past  bioerosion  rates,  we used literature information about historical D. mexicanum densities (Guzman 1986, Guzman & Cortés 1992, 2007, Lessios et al. 1996, Jiménez 1998, Alvarado & Chiriboga 2008, J.J. Alvarado unpubl. data) and present densities. In addition, to determine past bioacretion rates, we  used  historical  live  coral  cover  information (Guzman et al. 1987, Guzman & Cortés 1992, 2007, Jiménez 2001, J.J. Alvarado & J. Cortés unpubl. data), and present covers. The carbonate balance is a quantitative measured of the functional state of the reef (Perry et al. 2008). As used here, balance refers to the net carbonate budget change in each reef system by comparing the production and the erosion. It is positive if the reef is growing and negative if the reef is eroding (Eakin 2001).

Finally, we analyzed the relationship between  D.  mexicanum  density  and  the  results of reef balance with a logarithmic regression. The regression equation was used to calculate possible scenarios of reef balance in Culebra Bay at different urchin densities (3, 4, 5 and 11 ind m-2).

Results

Table 1 displays: 1) carbonate amounts present in the D. mexicanum‘s gut, and 2) coral growth and skeletal density for the main reefbuilder species. The highest amount of carbonate present in the urchin guts was at Cocos Island, and lowest at Caño Island (Table 1). Pocillopora elegans possesses a higher growth rate than Porites lobata (Table 1). Coral skeletal density was similar in the three study areas, being higher at the species level for P. lobata than for P. elegans (Table 1).

At Cocos Island, we observed a reduction of D. mexicanum density from 1987 to 2009, and a gradual increase in live coral cover (Table 2, Fig. 2A). At this island, the highest sea urchin densities were reported (11.4 ind m-2) in 1987, whereas they were ≤ 1 ind m-2  (Table 2) from 2002 to 2009. At Caño Island, the densities were  below  0.50  ind  m-2,  while  in  Culebra Bay they increased 900% from 2006 to 2009 (Table 2). In the three study areas, live coral cover was high (~20-50%) during times when urchin densities were low (< 1 ind m-2, except for Caño in 1984), and low (~3%) when the densities of D. mexicanum were close to 3 ind m-2 (Table 2). From the three study areas, Caño Island  maintained  a  relatively  constant  coral cover  and  urchin  density,  except  that  coral cover decreased in 1984, with a slight increase in D. mexicanum density (Table 2).



In Cocos Island, as the density of Diadema decreased, its bioerosion impact decreased (Fig. 2A). In 1987, intense bioerosion (9.3 kg m-2  y-1) and very low bioacretion (0.4 kg m-2 y-1) were observed, resulting in a negative balance in the carbonate budget (Fig. 2A). At this locality  there  was  a  positive  balance  in 2002, which was maintained until 2009, with net accretion of 3.1 kg m-2  y-1  (Fig. 2A).

In Caño Island, D. mexicanum had little impact on bioerosion (<1 kg m-2  y-1) in comparison to Cocos Island (Fig. 2A-B). Also, bioacretion has been high (~11 kg m-2 y-1), producing a positive carbonate balance (Fig. 2B).

Finally, in Culebra Bay the carbonate balance was highly positive, with bioacretion rates of 18 kg m-2  y-1 and low bioerosion rates (0.07 kg m-2 y-1) until 2006. After that  date, there  was  a  drastic  change  in  the  carbonate budget,  becoming  almost  negative    by  2009 (0.43 kg m-2  y-1; Fig. 2C). The bioerosion rate increased from 0.07 to 0.75  kg  m-2 y-1, as a result of the increase in sea urchin density (Fig. 2C). Loss of coral cover (Table 2) resulted from the  proliferations  of  harmful  phytoplankton that caused high mortality in the area (Jiménez 2007, Jiménez et al. in prep).

The relationship between D. mexicanum density and reef carbonate balance is negative (Fig. 3). When there are few sea urchins, carbonate balance is highly positive (Culebra Bay: 1996 and 2006). When the density of sea urchins exceeds ~1.5 ind m-2, reef balance start to be negative, becoming highly erosive when urchin density exceeds 4 ind m-2. Figure 3 help to explain how quickly can be a negative process, such as in Culebra Bay, where in less than 3 years there was a fast loss on reef balance. Also, this analysis help to visualized the slow process of recovery took Cocos Island after a strong disturbance until now days. In the case of Caño Island, it seems that D. mexicanum does not possess a key role in reef balance. For Culebra Bay, a continuing increase in sea urchin density will produce a negative balance. The 2009 densities were 2.19 ind m-2, pushing the reefs near negative balance. An increase to  3  ind  m-2 would produce a considerable reduction on reef erosion, and an increase to 11 ind m-2  would produce a similar impact to reef balance as the one seen in  Cocos Island after the 1982-1983 El Niño event.

Discussion

Guzman   and   Cortés   (2001)   observed that after severe disturbance, the recovery of ETP reefs has been slow, due to the lack of recruitment and the continuous predation by corallivorous  organisms.  In  the  case  of  the 1982-83 El Niño, the recovery was limited by 1)  the  extreme  oceanographic  conditions  of the region, 2) the high coral mortality suffered during the El Niño, 3) the intense herbivory resulting from high sea urchin concentrations produced high rates of bieorosion, 4) low abundance of recruits due to reduced sexual activity of the main reef builder species, and 5) low potential for successful recruitment due to the low abundance of crustose coralline algae (Guzman & Cortés 2001). Baker et al. (2008) indicated that, for a reef affected by bleaching to recover, a change in the balance between reef accumulation and erosion, the ability to maintain healthy levels of herbivory, and macroalgae cover and coral recruitment are required.

After the 1982-83 El Niño, there was a high incidence of coral mortality and bleaching at Cocos Island. During this period there were high sea urchin densities (Guzman & Cortés 1992), producing a negative carbonate budget, and reef erosion in 1987. It took approximately 15 years for those reefs to switch back to a positive balance. Probably, as Guzman and Cortés (2007) indicated, the impact of the sea urchin as bioeroder declined, and its action as a herbivore facilitated coral recruitment. Thus, the role of the sea urchin can change in accordance with its density.

Glynn et al. (2009) indicated a similar recovery pattern for the coral reefs at Darwin Island, northern Galapagos Archipelago, which underwent  lower  bioerosion  in  comparison with  other  reefs  in  the  Archipelago  (Glynn 1988). This resulted in an intact reef framework and, together with successful sexual and asexual  coral  recruitment  (Glynn  et  al.  2009), there was a similar recovery as observed at Cocos Island.

The recovery in reef balance observed at Cocos Island, like at Darwin Island, could be due in part to the positive role that Diadema can have through creating spaces for settlement (Carleton & Sammarco 1987) and increasing the cover of crustose coralline algae, which enhances successful settlement (Morse et al. 1988). Vance (1979) found that the sea urchin Centrostephanus coronatus, a member of the Diadematidae family, avoids consuming crustose coralline algae when eating other algae, favoring development of coralline algae. Nevertheless, the sea urchin in search of its food can erode the substrate simply by scraping areas where algae grow (Toro-Farmer 1998). Therefore, in the case of Cocos Island, the herbivorous activity of D. mexicanum possibly resulted in a larger abundance of calcareous algae, and a greater availability of substrate for recruitment of new coral recruits, favoring its recovery.

Eakin (2001) indicated that the reef budget in Uva Island, Panama, between 1985 and 2000,  was  negative,  without  showing  signs of recovery. In some places the densities of Diadema diminished due to the reduction in the reef complexity due to the urchin’s erosion, diminishing their refuge. In Cocos Island, reef complexity has remained high (Alvarado et al. in prep.), but density of sea urchins has diminished (Fig. 2A, Table 2), which is probably related to predation. This island has been a no-take Marine Protected Area for the last 30 years, and protection has been strictly enforced for the last 10 years. This has favored the recovery of trophic webs. This contrasts with Panama, where the reefs were under strong fishing pressure for many years, diminishing the presence of predators on D. mexicanum. It has been observed that inside marine reserves, where the fishing pressure has been reduced, the trophic interactions are re-established between the sea urchins and their predators (McClanahan et al. 1999), as opposed to places where intense fishing is occurring and densities of Diadema are high (Harborne et al. 2009).

At Caño Island, from the historic densities of D. mexicanum (Fonseca 1999, Guzman & Cortés 2001, Fonseca et al. 2006), it seems that this sea urchin has not been an important bioeroder and that its impact has been minimum (Fig. 2B). Scott et al. (1988) indicated that after the 1982-83 El Niño, this island experienced intense bioerosion (9 kg m-2  y-1). Later, in 1996, Fonseca (1999) determined a lower bioerosion rate (0.05 kg m-2 y-1) for Platanillo Reef, as well as high bioacretion (7.1 kg m-2 y-1). Guzman and Cortés (2001) indicated that these reefs in the 1997-98 El Niño had less effect than prior events because of high crustose coralline cover (80-95%) and successful coral recruitment.

In Culebra Bay, bioerosion increased after 2009, when a massive explosion of D. mexicanum occurred. The reefs from this area underwent intense mortality due to recurrent proliferations of phytoplankton between 2006 and 2009; these produced intense bleaching resulting from the anoxic conditions of the water  and  the  lack  of  sun  light  penetrating to the bottom (Jiménez et al. en prep.). Dead corals were replaced by turf algae. This availability of algae could have caused an increase in the recruitment of sea urchins, increasing their densities, and their bioerosional effect. If this bioerosion continues, the coral framework will be weakened and may collapse (Colgan & Glynn 1990). Figure 4 illustrates the difference in conditions in 2005 compared to now.

Bioerosion intensity depends on several environmental factors, like depth, light and nutrient supply (Chazottes et al. 1995). Baker et al. (2008) indicated that a variety of disturbances can cause a significant reduction in live coral cover. Eutrophication, sedimentation and bleaching can quickly initiate to an erosive phase, resulting in a loss of structural integrity and topographical relief. Moreover, Edinger et al. (2000) mentioned that coral reefs growing in eutrophic coastal environments are exposed to higher bioerosion than those in clear oceanic waters, resulting in negative carbonate budgets. In Dominican Republic, Tewfik et al. (2005) found that a nutrient enrichment together with overfishing accelerated the environment deterioration, and facilitated an increase in sea urchin densities (Lytechinus variegatus). As consequence of this population increase there was a considerable reduction on the diversity on the affected areas.

In Culebra Bay, the environmental conditions have deteriorated in recent years due to anthropogenic eutrophication (Fernández 2007, Fernández et al. in prep.). This has resulted in invasions by Caulerpa sertularoides (Fernández & Cortés 2005, Fernández 2007) and phytoplankton  proliferations  (Jiménez  2007, Jiménez et al. in prep.). As a consequence, live coral cover in the Bay has declined (Jiménez 2007, Bezy et al. 2006, Cortés et al. 2010, Fernández et al. in prep.). When nutrients become  abundant,  reef  carbonate  producers  tend to be overgrown by fast-growing fleshy and filamentous algae. Meanwhile the bioeroders tend to increase due to the food availability, magnifying the negative effects (Chazottes et al. 1995). Additionally, successful recruitment by juvenile sea urchins have been attributed to algal proliferation because of the increase in nutrient availability, favoring a greater density of sea urchins in organically enriched areas (Eklöf et al. 2008).

Another negative factor that could have promoted an increase in D. mexicanum high densities  at  Culebra  Bay,  is  that  this  area suffers  from  a  strong  fishing  pressure,  with few controls and complete lack of protection of the resources. With few or no sea urchin predators and with more food available, Diadema reached densities never before reported for the area (Table 2), causing an increase in bioerosion of the coral framework. On Marías Island in Mexico, and in the Canary, Galapagos and Hawai’i archipelagos, it has been observed that reductions in the populations of predator fishes by overfishing resulted in an increase in sea urchin abundance (Torrejón-Arellano et al. 2008, Clemente et al. 2009, Sonnenholzner et al. 2009, Vermeij et al. 2010).

According to Muthiga and McClanahan (2007) it is of interest to record the historical level of the Diadema populations, as it may indicate the degree to which the sea urchin are required for the maintenance of the coral reef ecology or a negative factor that has been released  from  predation  by  overfishing.  Through the reconstructions presented here (Fig. 2), it is possible to visualize how Diadema mexicanum can have either a positive or a negative effect on coral reef development in the Eastern Tropical Pacific. Moreover, it was shown that for the sea urchin to have a positive or negative role, a series of synergies are required to switch it in one direction or another. Bad coastal management (anthropogenic eutropfication) and overfishing (lack of sea urchin predators), can result in this sea urchin eroding the coral framework, as seen in Culebra Bay. However, where the trophic webs are “healthy”, Diadema mexicanum facilitates the growth of crustose coralline algae, which at the same time helps the recovery of coral cover. As Veron et al. (2009) pointed out, “a reef far from additional human stress can realize a rapid recovery, returning to its own diversity in less than a decade”. Cocos Island, is an example of this statement, while Culebra Bay, being subjected to additional stress, likely will take longer to recover. MPAs provide the means for effectively limiting the number of urchins on coral reefs and subsequently returning the calcium carbonate cycle to a more balanced state (Brown-Saracino et al. 2007). However, if the anthropogenic eutrophied conditions diminish, there is fishery management, and there is stronger protection of the reef environments, it is possible that Diadema can help in the recovery of those reefs.  To  assess  the  future  development  of reef structures and their relationship with sea urchins it is important to continue monitoring the populations on all of these reefs. However, it is currently extremely important to do this in Culebra Bay, so that management measures can be initiated if D. mexicanum attain densities of > 3 ind m-2. Also, it is extremely important to begin studying recruitment and reproduction patterns, as well as to initiate efficient coastal and fishery management practices. Glynn and Fong (2006) stated that if corals survive total mortality and if environmental conditions return  to  pre-disturbance  levels,  there  exits the potential for rapid coral recovery in a few years. For Culebra Bay, this statement is extremely important, because its highlight the need for a better coastal management that would improve coastal ecosystems and their services.

Acknowlegments

We  acknowledge  the  following  persons and institutions that collaborate during the development of this work: C. Fernández, O. Breedy, C. Sánchez, S. Martínez, E. Gómez,  A. Planas, V. Flores, O. Norzagaray, L.E. Calderón-Aguilera, Centro de Investigación en Ciencias del Mar y Limnología (CIMAR, Universidad de Costa Rica), Universidad Autónoma de Baja California Sur (UABCS), Reserva Biológica Isla del Caño and Parque Nacional Isla del Coco park-rangers, MY Adventure crew, Instituto Costarricense de Turismo, Hotel Punta Marenco Lodge, Águila de Osa Inn, Centro de Investigación Científica y Educación Superior de Ensenada (CICESE). A special acknowledgement for their economic support of this research: Vicerrectoría de Investigación of the Universidad de Costa Rica, Ministerio de Ciencia y Tecnología de Costa Rica (MICIT), Consejo Nacional para Investigaciones Científicas y Tecnológicas de Costa Rica (CONICIT), Consejo Nacional de Ciencia y Tecnología de México (CONACYT), Fonds Français pour l’Environnement Mondial (FFEM) and Ecodesarrollo Papagayo. We appreciated the comments made by four anonymous reviewers that enrich this paper.

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*Correspondencia a:
Juan José Alvarado. Centro de Investigación en Ciencias del Mar y Limnología (CIMAR), Universidad de Costa Rica, San Pedro, 11501-2060 San José, Costa Rica; juanalva76@yahoo.com Posgrado en Ciencias Marinas y Costeras, Universidad Autónoma de Baja California Sur, La Paz, México.
Jorge Cortés.
Centro de Investigación en Ciencias del Mar y Limnología (CIMAR), Universidad de Costa Rica, San Pedro, 11501-2060 San José, Costa Rica; jorge.cortes@ucr.ac.cr. Escuela de Biología, Universidad de Costa Rica.
Héctor Reyes-Bonilla.
Departamento de Biología Marina, Universidad Autónoma de Baja California Sur, La Paz, México; hreyes@uabcs.mx
1. Centro de Investigación en Ciencias del Mar y Limnología (CIMAR), Universidad de Costa Rica, San Pedro, 11501-2060 San José, Costa Rica; juanalva76@yahoo.com; jorge.cortes@ucr.ac.cr
2. Posgrado en Ciencias Marinas y Costeras, Universidad Autónoma de Baja California Sur, La Paz, México.
3. Escuela de Biología, Universidad de Costa Rica.
4. Departamento de Biología Marina, Universidad Autónoma de Baja California Sur, La Paz, México; hreyes@uabcs.mx

Received 23-II-2011.    Corrected 05-X-2011.Accepted 29-II-2012.

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