Estructura funcional de comunidades que habitan sobre los fondos duros de Isla de Pascua (Rapa Nui).
The occurrence of ecological regime shifts (also referred to as phase shifts, alternate stable states or tipping points), which evoke change in the internal governing dynamics and feedbacks of an ecosystem (Hughes, 1994; Scheffer et al., 2001; Levin & Lubchenco, 2008; Hughes et al., 2010; Barnosky et al., 2012), has become increasingly commonplace in all ecosystems of our planet. Since such rapid wholesale changes to community structure can dramatically alter the very ecosystem functions/services on which humans rely, there is good reason for the growing concern not only by scientists, but by the population at large. Causes are complex, usually involving the interaction of two or more major drivers of change, but often appear to be associated with weakening or outright loss of key functional groups of species, which through varied mechanisms can increase system resilience and/or vulnerability to human-related disturbance (Scheffer et al., 2001). The demonstrated relationship between species and functional group diversity on ecosystem functioning has reinforced the idea that a trait-based view of communities may be more meaningful/insightful than that based solely on species composition, especially when inferring ecological responses to environmental change (StuartSmith et al., 2013). Toward this endeavor, characterizing the temporal variability or persistence, as well as spatial organization of functional community structure, is indeed a critical step for early detection of warning signals of an imminent regime shift and/or propagation of localized regime shifts to larger scales (Hughes et al., 2010; Fung et al., 2013).
Coral reef ecosystems living near the limit of individual species thermal tolerances appear to be undergoing global degradation, as symbolized by the seemingly increasing occurrence of dramatic regime shifts from hard coral dominance to a state dominated by fleshy macroalgae (Done, 1992; Hughes, 1994; Nystrom et al., 2000; Bellwood et al., 2004). While studies have demonstrated the interaction of multiple causes (e.g., increased run off and sea surface temperature) induced by human damage, changes in grazing regimes associated with depleted herbivore abundances (e.g., from over-fishing) appear to be a common and critical denominator in most documented examples of extensive coral mortality that follows bleaching induced, for instance, by anomalously high temperature events (e.g., Hughes, 1994, Jackson et al., 2001; Wellington & Glynn, 2007; Baker et al., 2008; Edwards et al., 2014). Indeed, fast-growing macroalgae can rapidly colonize and inhibit coral recovery and thus reduced herbivory has been invoked as a critical symptomatic driver of eroding reef resilience and coralmacroalgal regime shifts (McCook et al., 2001; Jompa & McCook, 2002; Diaz-Pulido et al., 2009). Despite the demonstrated connection between variation in diversity and abundance of functional groups and community resilience (Cheal et al., 2010), few studies have documented changes to functional structure over relevant landscape and inter-annual timescales.
Rapa Nui, or Easter Island, is the most isolated island in the Pacific Ocean, so the existence there of relatively depauperate (species poor) communities with high levels of endemism is not surprising. The comparatively brief history of scientific studies of Rapa Nui's coral reef communities, along with compilations of taxonomic inventories, were previously reviewed for marine benthic flora by Santelices & Abbott (1987), for invertebrates by Castilla & Rozbaczylo (1987) and Glynn et al. (2003, 2007), and for nearshore fishes by Castilla & Rozbaczylo (1987); Ojeda & Aviles (1987) and Sepulveda (1987), with notable recent updates on zooxanthellate corals (Glynn et al., 2007) and fishes (Randall & Cea, 2011).
The ecology of shallow, subtidal, hard bottom communities around the island is still poorly studied. General benthic community surveys were first made by DiSalvo in the mid 1980's (DiSalvo et al., 1988), but were not quantitatively surveyed again until the late 1990's (Hubbard & Garcia, 2003; Glynn et al., 2007). Apart from our own irregular surveys since 1999 (see below), a recent expedition supported by the Chilean government and the National Geographic Society, and other foundations (Waitt Foundation, Oceana Chile), surveyed and compared benthic communities of Rapa Nui to nearby Salas y Gomez Island, documenting important between-island differences in community structure driven largely by the relatively lower fish abundances and scarce apex predators observed at Rapa Nui (Friedlander et al., 2013). Based on anecdotal reports, personal observations and differences among earlier quantitative surveys, Hubbard & Garcia (2003) suggested that shallow water communities around the island undertook an important shift in dominance from macroalgae (particularly Sargassum) to coral reefs sometime during the early-mid 1980s. This is particularly notable in light of the opposing trajectory of coral loss described for much of the world's coral reef communities over the past 30 years (Hughes et al., 2007). This further underscores the importance and urgency of establishing baselines against which future community changes and functional structure can be evaluated.
As a first step toward setting these baselines and their current spatial variability, we here describe geographic and among-habitat patterns in the functional structure of shallow hard bottom communities, including fish assemblages, across the northeast and west coasts of Rapa Nui. We also document temporal patterns of change in dominant benthic functional groups that have taken place over the past 15 years at selected sites on this Pacific island. The patterns reported here could also provide guidelines for the much-needed management and conservation plans for this isolated ecosystem.
Rapa Nui lies at the eastern edge of the Subtropical Counter Current and Polynesian Triangle, corresponding to the austral limit of coral distribution. Like many high-latitude subtropical reef-building coral communities (e.g., Veron & Done, 1979), significant limestone reef structures are rare. However, incipient reef formations are observed, with the largest (aprox. 7 m thick) found at depth on the northeast sector of the Poike Peninsula (Hubbard & Garcia, 2003). Yet, in comparison to other subtropical areas, there is high coral cover at most sites, dominated by two of the 13 recognized species for the island, Porites lobata and Pocillopora verrucosa (Glynn et al., 2003, 2007). Macroalgal cover is generally low in abundance around the island (Glynn et al., 2007), despite relatively high species richness (Santelices & Abbott, 1987).
The fish fauna at Rapa Nui is generally species poor, with only 164 shore and epipelagic species (those that live offshore and near the surface) described (Randall & Cea, 2011). Geographic isolation, lack of lagoonal ecosystems and few protected bays offer plausible explanations of low species richness. Moreover, smallsized individuals characterize fish assemblages and there is notable lack of apex predators (sharks and jacks), with planktivores and herbivores dominating trophic structure (Friedlander et al., 2013).
MATERIALS AND METHODS
Shallow hard bottom communities at six sites spaced 1-10's km apart across the northeast and west coasts were surveyed during irregular visits to the island between 1999-2013 (Fig. 1). All surveys were conducted during austral autumn months (April-May). Benthic communities were surveyed annually between 1999-2001 and 2010-2013 as well as 2005 for a total of 8 years. Not all sites and depths were sampled every year and thus we present temporal patterns of benthic communities only for four locations with the most complete records. Fish assemblages were surveyed annually between 20112013. Due to common heavy swell and wave action, combined with difficult access, it was not possible to undertake quantitative samples along the exposed southeast coast. Sites were composed of rocky outcrops openly exposed to oceanic swells and were selected to be as similar as possible in terms of geomorphology, slope, and local orientation. The greatest heterogeneity among sites occurred between the northeast and west coasts, which differ in orientation and wave exposure.
At each site, local numerical density of mobile benthic invertebrates (e.g., sea urchins, snails, stars, crabs, limpets, etc.) and spatial cover of sessile species (e.g., corals, macroalgae, sponges, barnacles, etc.) at shallow (<8 m), mid (10-15 m), and deep (17-23 m) zones were quantified at 2-3 rocky reefs using the standard transect-quadrat method, as described in detail in Wieters et al. (2009). A minimum of ten 0.25 [m.sup.2] quadrats, each with 81 systematic point grids, regularly placed along 50 m transects stretched parallel to the shoreline (along depth contour) were sampled and photo graphed. Because of their large size and sparse or clumped distribution, densities of sea urchins were estimated by counting the number of individuals within a minimum of 15 randomly placed large 1 [m.sup.2] plots. Due to local geomorphology and topography, not all zones were present at all sites. Cover of species inhabiting primary space (attached to rock/reef surface) and secondary space (i.e., epibiotic such as algae on corals) was estimated separately, so total cover could exceed 100%. All organisms were resolved to the lowest possible taxonomic level in the field. Areas obviously influenced by sediment abrasion were avoided to standardize suitable substrate habitat and avoid confounding among-site variability with habitat differences. To minimize potential observer-induced biases, ~75% of quadrats across all years were recorded by a single observer (EAW).
At each study site, local composition and abundance of reef fish at shallow (5-15 m) and deep (15-25 m) zones were quantified using unbaited video (years, 2011 2013) and underwater visual census (2013). All locations and sites were characterized by similarly lowmedium-relief reefs (caves, ledges and overhang relatively scarce). Minimum underwater visibility during all surveys was 35 m. At each zone, an unbaited video camera (Go-Pro tm hero 2-3) was anchored with a 2 kg weight and deployed to record images parallel to the seafloor. A 10 m transect with weighted and vertically positioned, graduated PVC tubes of 30 cm height at every 1 m, was stretched directly in front of the camera to aid in quantification of individual fish size. Cameras were allowed to continuously record 25 to 35 min. Videos were reviewed on a large screen computer monitor. Species presence and the maximum number of individuals (MaxN, Willis et al., 2000) belonging to each species in the field of view at one time were recorded. This provides a conservative estimate of relative density and avoids repeated counts of the same individuals. For each video clip, we obtained subsamples and the calculated area as delimited by tape transects and 20 cm tall tubes placed 3 m apart. We divided by area to standardize estimates to [m.sup.2].
Underwater visual censuses (UVC) were conducted by a single SCUBA diver (APM) that swam along a randomly allocated 50 m strip transect placed perpendicular to the coastline. Observations were delayed 10-15 min after placement of transects to limit and standardize any potential disturbance to fish by the diver. Large (>20 cm TL) and small (<20 cm TL) fish were counted within 4 m or 2 m transect width, respectively, which was equally distributed to the left and right of the observer/diver. Underwater photography aided species identification (see Friedlander et al., 2013). Two replicate transects, separated by 5065 m, were conducted at each site x zone combination. Fishes were resolved to lowest possible taxonomic level possible and categorized as endemic or nonendemic species as per Randall & Cea (2011).
We categorized species into one of 8 (benthic) or 5 (fish) functional groups based on a) morphological and anatomical traits related to maximum possible rate of biomass production (mass-specific growth; ability to quickly assimilate nutrient availability) and resistance to herbivory [algae; largely following Steneck & Dethier (1994) with grouping modifications similar to Broitman et al. (2001) and Wieters et al. (2009)] or b) trophic level and foraging strategy [invertebrates and fishes; DiSalvo et al. (2007); Halpern & Floeter (2008); Randall & Cea (2011)]. Benthic: 1) ephemeral algae, 2) corticated algae, 3) leathery algae, 4) crustose algae, 5) corals, 6) filter feeders, 7) herbivores, and 8) predators. Fishes: 1) herbivores, 2) planktivores, 3) invertivores, 4) omnivores, and 5) piscivores.
To compare abundances of functional groups among habitats, 1-way ANOVAs, with depth zone considered as a fixed factor and abundance per site and year as replicates were used. A completely factorial design with depth, site, and year was not possible due to incomplete information for several sites and/or years. To evaluate spatial consistency of differences in functional group abundance across depth (zones), a 2way ANOVA was used, with zone considered a fixed factor and sites considered a random factor. Significant 2-way interaction required depth zone to be considered separately in further analyses of spatial patterns in functional group abundance (see Results).
To evaluate temporal consistency of among-site spatial patterns in mid depth (8-15 m) functional group abundance, a 2-way ANOVA with site and year (2012 or 2013) was used. Separate analyses were undertaken for individual functional groups. In all cases a log(n+1) transformation was applied to meet ANOVA assumptions. Multivariate characterization of variability in benthic functional structure across the island was performed with principal component analyses (PCA) and non-metric multidimensional scaling (NMDS) with individual functional groups as dependent variables.
For each fish sampling method (video vs visual surveys), fish species richness (total number of species recorded), diversity (Simpson's Diversity Index) and numerical abundances (relative density of each speciesindividuals [m.sup.-2]) were similar between depth zones, regardless of site and sampling year (P > 0.05 for main effect of Zone and Site x Year interactions in 3-way ANOVA) and thus zones were pooled in further analyses (see below).
To assess potential methodological biases in characterization of fish assemblages, overall sitespecific fish species richness was compared using unpaired t-test for the simultaneously recorded video and visual surveys conducted in 2013.
Multivariate characterization of variability in fish species composition across the island was performed with Jaccard or Bray-Curtis resemblance matrices with presence/absence (video and visual surveys) or abundance (visual surveys only) of individual species as dependent variables, respectively. To evaluate similarity in composition among sites, depth zones, and years, a three factor permutational multivariate analysis of variance was used for estimates based on presence/absence data. Among-site differences in composition based on abundance estimates were analyzed using a one factor permutational multivariate analysis of variance (PERMANOVA; Anderson, 2001).
We performed a redundancy analysis (RDA) of the abundant species and functional groups for fishes and sea urchins against major sessile benthic functional groups. We selected this constrained method due to our a priori hypothesis of the possible effects of coral reef structure on abundances of reef fishes and sea urchins (Diadema savignyi). To test for associations among fish abundances and benthic functional structure, we performed the PERMANOVA routine among sites. We used the Hellinger transformation of the fish and pointcount data; this transformation makes the data agreeable for the types of analyses performed here (Legendre & Legendre, 1998). We used RDA to graphically display the differences of the most abundant species among the study sites, and report the contribution of each constraint. All analyses were conducted using the R package 'vegan' (http://www.rproject.org/).
Benthic community functional structure Zonation
In general (all sites and years pooled), total abundances of mobile and sessile organisms were similar across depths (mobile: MS = 87.41, [F.sub.1,43] = 0.24, P = 0.63; sessile: MS = 191.30, [F.sub.1,43] = 2.65, P = 0.11), with sessile organisms covering on average 66-87% of available substrate space. However, the relative contribution of some key functional groups differed (Fig. 2). Dominant space-occupying corals presented lowest cover at shallow (<5 m) and highest cover at mid (9-13 m) depth, which largely reflected the near absence of Pocillopora near the surface and its increasing relative contribution with depth (Fig. 2 insert). Likewise, ephemeral algae (largely Lobophora sp.) exhibited increasing abundance with depth, though average covers never reached greater than 10%. In contrast, leathery (mainly Padina sp.) and overall scarce (<3%) articulated (predominently Halimeda sp.) algae were relatively more abundant at shallow depths. No clear pattern for benthic herbivores (predominantly the urchin Diadema savignyi) was observed across depth zones, whereas corallivore snails (Coralliophora violacea) were absent near the surface and most abundant at deeper locations.
Spatial structure across the island
Live coral consistently dominated substrates at all sites (60-80% cover), but showed significant among-site variation that persisted between years (Fig. 3, nonsignificant Site x Year interaction, see Table 1). Tukey tests revealed that differences were largely due to the significantly lower overall coral abundance at Taimana, which differed from Hanga Roa and northeast coast sites (Le Perouse, Anakena) that are characterized by large monospecific patches of Porites lobata among smaller patches of mixed assemblages. Relatively higher abundance of Pocillopora is found on the northwest coast (Tautara, Taimana, Omohi), where it commonly occurs in near pure stand patches and can reach similar abundance as Porites (e.g., Tautara; Fig. 3 insert). Algal abundance was generally low but varied significantly among sites and between years (Table 1). Regardless of year, dominant algal functional groups (primarily ephemerals but also articulated) presented peak abundances along the northwest tip of the island (Omohi, Taimana). Abundances of ephemerals and articulated macroalgae were higher during 2012 and 2013, respectively, at all sites (significant Time effect), but the spatial pattern peristed (non-significant Site x Time interaction). Crusts and leathery algae were absent from all but a few sites along the northwest coast, where they never reached greater than 2% cover (Fig. 3). No clear patterns were observed for benthic herbivores (mainly sea urchins), though densities varied greatly among sites.
The multivariate characterization of benthic functional community structure showed marked spatial differentiation across the island, with geographic neighbors not always showing obvious similarity (Fig. 4). Important separation was observed between three groups of sites: 1) Anakena, Le Perouse, and Hanga Roa, 2) Omohi and Taimana, and 3) Tautara alone. The first two principal components explained 54% and 33% of the variation, respectively, with groups primarily separated on component 1 and secondarily (but still important) on component 2. PC1 reflected a combination of most all functional groups, with all but Porites being relatively similarly weighted. PC2 was most strongly weighted by Porites and ephemeral algae, followed by Pocillopora.
Temporal variation at selected sites
Temporal trends in total coral cover differed among depths (Fig. 5). While coral increased over time at our two shallow sites (Anakena and Hanga Roa), cover abruptly decreased in 2005, and remained low, at the relatively deep (>15 m) location of Taimana. The patterns in total coral cover were largely driven by massive Porites, particularly after 2000 when a sharp decline in branching corals (Pocillopora) was observed at most sites (Fig. 6). No clear temporal trend was observed at mid-depth (10-15 m) sites, although a notable decline in coral cover was again observed in 2005 at Taimana (Fig. 5). A trend to increasing abundance of articulated/calcified algae was observed at all sites, though overall cover was consistently low (<5%). No consistent temporal patterns were observed for crustose algae, which showed relatively high abundances at Tautara and Taimana during 2001. At all sites, cover of ephemerals increased and reached peak abundances during 2010, after which cover quickly decreased to near-absent again at all sites but Taimana. At all sites, leathery algae presented elevated abundances of relatively short duration (few years) between 2000-2005, with marked peaks at mid and deep locations of Taimana during 2005.
Fish species composition
Our fish surveys documented a total of 45 species from 25 families, with 24% of sighted species being endemic to the island (Table 2). The total number of species observed was similar between sampling methods (P > 0.05), with 36 (80%) recorded by visual census and 39 (87%) by unbaited video. However, unbaited video better captured the presence of those species that typically tend to avoid divers (butterflyfishes, girellids), whereas visual census proved better in documenting the presence of relatively small cryptic species associated with complex microhabitats such as crevices (e.g., blennies). The distribution of species across functional groups was not even (Fig. 7). The numbers of species that comprised a given functional group (species richness) was strikingly lower for apex fish predators (piscivores), whereas invertebrate feeders were the most species-rich group.
Species composition of local fish assemblages showed significant spatial structure among sites, regardless of whether estimates were based on presence/absence (Permanova, df = 6, pseudo F = 2.6, P < 0.001) or abundances (Permanova, df = 4, pseudo F = 2.8 , P < 0.001). In general, sites at the northwestern side of the island grouped together and differed from all other sites. Spatial differences were largely due to variable presence and abundances of Centropyge hotumatua, Acanthurus leucopareius and Chrysiptera rapanui.
Abundances of individual functional groups
In general, total abundances of fishes (species/ functional groups pooled) were low at all sites (0.4-1.5 ind [m.sup.-2]), but varied significantly across the island (Kruskal-Wallace test, df = 4, P < 0.05), with overall highest densities observed at Taimana (1.5 ind [m.sup.-2]) and Anakena (1.4 ind [m.sup.-2]) and lowest at Omohi (0.4 ind [m.sup.-2]). High among-site variability was common for individual functional groups (Fig. 8). Invertivores were most abundant at the northeast coast site of Anakena, where densities were at least twice that observed else<03_TB006> where. Herbivorous fishes were strikingly more abundant at sites along the northeast coast and northwest tip (Taimana, Omohi) of the island. Piscivores, omnivores, and invertivore/piscivores were scarce at all sites, with the exception of Taimana where invertivore/piscivore abundance was relatively higher and similar to that observed for other functional groups. No clear geographic pattern was observed for fishes with planktivorous diets, which were highly variable among sites, reaching greatest densities of any functional group at Anakena and Taimana.
Fish correspondence with benthic functional community structure
Redundancy analysis (RDA) and permutation tests revealed strong, significant associations between fish functional structure and benthic functional community structure (Fig. 9a; df = 4, pseudo F = 2.9; P-value = 0.005). In general, herbivorous fishes were associated with branching corals (Pocillopora), while invertivores were tightly associated with massive corals (Porites). Planktivores were associated with articulated and leathery algae. Benthic functional structure constrained more than 63% of the observed variation in total abundance of fish functional groups. No significant associations were observed for fish assemblage structure when based on individual species composition (Fig. 9b, df = 4, pseudo F = 1.2; P-value = 0.28).
Our results offer an important benchmark for understanding and recognizing future functional changes at the community level, placing them in light of existing spatial structure and temporal variability over the past 15 years. Our studies, like all available quantitative surveys for the island, were initiated/conducted after an apparent large shift in dominance from macroalgae to corals. Thus, we acknowledge our altered perception (i.e., "shifting baseline", sensu Pauly, 1995) and its limited potential use in determining some former "natural" or pristine condition (Dayton et al., 1998; Jackson et al., 2001; Myers & Worm, 2003). Instead, we view this work as an important starting point to provide context for localized "snapshot" studies. This can help to provide a framework to generate testable hypotheses about driving processes, assess subsequent change and, hopefully, trigger scientific support for locally-derived ecosystem-based conservation/management policy. Our results highlight the apparent resilience of the current system and the identification of geographic "compartments" of functional organization, providing a first-cut guideline as to where management subsytems might be prescribed so as to spatially match ecosystem characteristics.
We documented a consistent increase in coral abundance in shallow habitats over the past 15 years, despite naturally depauperate fish assemblages and intense fishing pressure that is likely most responsible for the low fish abundances, small body sizes, and skewed trophic structure (Friedlander et al., 2013). Although data are restricted to only two shallow sites located on opposing coasts, field observations suggest the pattern is general to at least the northeast and west coasts of the island (no comparable data are available for the southeast coast). In light of the mass bleaching event of 2000 that caused high mortality of Pocillopora and had notable effects on this and total coral cover at most sites (Wellington et al., 2001, Wieters, unpublished data, Fig. 6), the temporal trend reported here indicates high capacity for relatively rapid recovery of coral cover after disturbance. The rapid recovery from the large-scale disturbance was often associated with a change in relative coral composition, suggesting a compensatory response by massive Porites, which had suffered widespread bleaching but overall little mortality (Glynn et al., 2003). Similar changes in generic composition have been reported among other reefs showing capacity for rapid recovery from disturbance (e.g., Adjeroud et al., 2009). Increase in coral cover beyond the pre-bleaching event was additionally enhanced by new recruitment that became most apparent after 2005, although regrowth/ regeneration of remnants was also observed (Wieters, pers. obs.). In general, our results suggest that the current ecosystem at Rapa Nui is relatively "resilient", capable of absorbing stress and not losing its fundamental functional dynamics.
Small blooms of leathery macroalgae did follow coral mortality and disturbance, lasting at most a few years prior to about 2005. Given that all surveys were undertaken during the same months (Abril-May) each year, potential shorter-term seasonal effects are unlikely to confound the among-years pattern of algal abundances. The role of herbivory in recovery and community resilience must be interpreted with caution due to lack of temporal data for herbivorous fishes, as well as the difficulty in relating abundances of mobile fishes and benthic grazers (predominantly sea urchins Diadema) to actual grazing rates. However, the rather abrupt temporal changes with higher benthic grazer density following disturbance and during palatable macroalgal blooms suggest that sea urchin feeding activity may have been critical to control fast-growing macroalgae and enhance coral settlmenet and recovery. Indeed, on-going field experiments reveal a strong capacity for aggregative behavioral responses to macroalgal availability by urchins, adults of which can form slow-moving "fronts" (e.g., Wieters, unpublished data). Future experimental studies teasing apart the role of diversity and identity of herbivores and whether/how these are modified by climate-induced bleaching stress are necessary and would provide critical insight into mechanisms of recovery, as well as provide insight into manageable attributes that drive functional vulnerability under scenarios of climate change.
Similar to that observed in shallow habitats, strong bleaching-induced declines in branching Pocillopora in 2000 with rapid, compensatory recovery of total coral cover was also observed at mid- to deep-locations at Taimana. However, recovery trajectories at depth were abruptly disrupted by 2005, when notable coral decline and peak abundances in leathery macroalgae were again observed. This time, the coral decline was largely due to decreased abundance in massive Porites. These results suggest important depth-dependent disturbance regimes with potentially greater types and frequency of disturbance at depth. Unfortunately, we have little idea of drivers and can only surmise that high-frequency upwelling or shoaling cold events, such as those we have recently begun to document at Rapa Nui (Wieters, unpubl. data), might have been at play. Severe shortduration cold water events can indeed cause large-scale coral mortality (e.g., Lirman et al., 2011) and have been observed with increasing depth elsewhere (Sheppard, 2009). Recovery also appeared depth-dependent, as corals at mid-depth had recovered to pre-2005 disturbance conditions again by 2010, whereas Porites (and total coral) abundances at the deeper location (aprox. 20 m) have consistently remained lower (albeit similar to that prior to the 2000 bleaching event). Such hypotheses are amenable to experimentation.
Long-term monitoring is critical to evaluate fluctuations in community functioning and the impacts of current and future disruptive stresses. Our limited dataset covering 15-year time span is clearly short and requires longer view for interpretation. For example, the trend to increasing coral abundance in shallow habitats is even more dramatic when considering the covers reported by DiSalvo et al. (1988) from their surveys in mid 1980's. According to long-term residents of the island, the rise to coral dominance was due to a combination of released competition and escape from predation following abrupt, large-scale mortality of once-thriving brown macroalgae (Sargassum and Lobophora) that had dominated space and provided habitat to corallivore fishes. The single corallivore fish species, the parrotfish Leptoscarus vaigiensis, was reportedly abundant in the late 1960s but was last sighted at the island in 1994 (Randall & Cea, 2011). As opposed to the now relatively common reports and demonstrations of coral-to-macroalgal phase shifts on coral reefs worldwide, persistent shifts of macroalgaeto-coral on coral reefs are, in general, extremely rare (but see Done, 1992; Idjadi et al., 2006). Unfortunately, no quantitative data exist prior to the shift in dominance at Rapa Nui and we have little idea of potential driving mechanisms. Studies testing for alternate ecological feedback mechanisms that sustain coral or algal dominance at Rapa Nui are currently underway (Wieters, unpublished data).
Spatial structure: vertical zonation
Rapa Nui is directly exposed to open ocean conditions and large storm swells. Limestone reef structures do not reach the surface to buffer wave forces, and the coastline is characteristically steep. Shores are frequently wave-beaten and localized shelter habitat is scarce. Indeed, it is common to observe toppled and broken colonies (pers. obs.), which led Hubbard & Garcia (2003) to suggest waves are the principal factor limiting reef-building. Thus, extreme wave energy in shallow waters is likely responsible for the observed reduced coral cover and absence of branching growth forms, which are known to be structurally more susceptible to wave-induced damage (e.g., Madin & Connolly, 2006). Moreover, colonies of massive Porites in shallow waters are small relative to those found deeper, suggesting routine mortality and/or stunted growth. Alternatively, or additionally, Porites and Pocillopora may host specific symbiotic zooxanthellae (Symbiodinium spp.) that are differentially adapted to the depth-mediated light gradient (e.g., Iglesias-Prieto et al., 2004), and thus inhibitory physiological performance may play an important role in influencing the upper limit of depth distribution of Pocillopora. Indeed, photo damage and bleaching are more readily observed in Pocillopora than Porites under the same depth and environmental conditions in the field (Wieters, unpublished data). Unfortunately, we currently have no information regarding the relative abundance of different Symbiodinium clades or how they vary across depths among species or functional groups. Reciprocal transplantation among depth zones could provide insight into the relative importance of holosymbiont characteristcs and local environmental/physical factors influencing coral distribution.
In contrast to what is commonly described for coral reefs in similar depth ranges to those explored here (Huston, 1985), we did not observe important reductions in total coral abundance (cover) at deeper sites (>20 m). The persistently clear oceanic waters of Rapa Nui offer high light penetration and visibility (often >50 m) and thus our surveys were unlikely deep enough for irradiance to limit photosynthesis. Indeed, corals are frequently observed at >60 m depth (H. & M. Garcia, pers. comm).
Neither wave action nor light availability is likely to explain the reduced cover of ephemeral macroalgae in shallow waters, since these early colonizers are usually favored by repeated disturbances that prevent succesional development. Instead, herbivory and/or increased competition with the relatively more resistent and wave-tolerant articulated/calcified algae -that reach higher abundances in the shallows- seems more plausible. Alternatively, deeper sites could receive higher rates of nutrient replenishment, such as that supplied by upwelling/bores (e.g., Leichter & Genovese, 2006), and promote ephemeral macroalgal growth (Smith et al., 2004).
Spatial structure: among-site variability
Spatial patterns of functional community structure and similarity reflect shifts in the relative abundance of the same set of species, rather than compositional changes. Site grouping sructure did not strictly follow geographic characteristics, as neighboring sites were not always grouped together. For example, benthic communities of Hanga Roa grouped tightly with sites on the northeast coast (Anakena, Le Perouse) and Tautara was an outlier by itself despite its close proximity to Taimana, which grouped more tightly with Omohi. The surprisingly similar spatial grouping of fish functional assemblage structure further reinforces geographic patterns and suggests tight benthic-fish associations. Identification of distinct benthic regions or groups of sites can offer important insight into regularities and causes of change, as well as indicate the limits to which locally-derived ecological models apply (Navarrete et al., 2005). Over large scales, the variability of the physical environment might "set the stage" for ecological interactions and functioning. Unfortunately, very little is known of the nearshore oceanographic conditions and how they vary around Rapa Nui. However, Andrade et al. (2014) recently described variation in satellite-derived temperature and chlorophyll-a, documenting important differences between the northwest and northeast coasts. The similar spatial change in benthic and fish functional community structure documented here suggests there may be strong coupling with coastal oceanographic conditions that varies on scales of 1s-10s of km. Such phenomena are of growing recognition and increasing interest around the world and warrant further study with emphasis on local in situ conditions.
Regardless of the underlying causes, identification of areas of distinct functional community structure has important implications for conservation and management, particularly when charged with the goal of upholding ecosystem functioning. Since spatial differences may reflect changes in regulating ecological processes, such information can help identify where to focus efforts or where different strategies might be most effective. Currently, the Chilean government is embarking on the implementation of a large no-take marine protected area east of Rapa Nui, Motu Motiro Hiva Marine Park, which includes the closest island of Salas y Gomez (see Friedlander et al., 2013). Morevoer, the native Rapa Nui population, with governmental and nongovernmental agency support, is debating and developing cooperative conservation-driven management plans for the nearshore coastal habitats of the island. We hope this study can offer a rough, first-cut contribution to building ecological literacy for these decisions.
Received: 10 March 2014; Accepted: 18 August 2014
We thank friends that helped us in the field through the years, including S. Navarrete, K. McDaniel, R. Finke,
B. Bulzar and others. We particularly appreciate the committed help, logistical support, and thoughtful perspective provided by Michel Garcia, without whom this work would not be possible. Enthusiastic conversations and underwater natural history explorations with H. Garcia, C. Rapu and A. Cea were always enlightening. E.A. Wieters thanks P.W. Glynn and G.M. Wellington. Financial support was provided by FONDECYT 1100920 and 1130167 to E.A. Wieters. We acknowledge additional support by the Center for Marine Conservation Nucleo Milenio Initiative P10-033F. A.P.M. acknowledges support by FONDECYT 11110351. E.A. Wieters appreciates support from the U.S. National Geographic Society grant to P.W. Glynn and G.M. Wellington that provided funding for initial surveys.
Adjeroud, M., F. Michonneau, P.J. Edmunds, Y. Chancerelle, T.L. Loma, L. Penin, L.Thibaut, J. VidalDupiol, B. Salvat & R. Galzin. 2009. Recurrent disturbances, recovery trajectories, and resilience of coral assemblages on a South Central Pacific reef. Coral Reefs, 28: 775-780.
Anderson, M.J. 2001. A new method for non-parametric multivariate analysis of variance. Austral Ecol., 26: 32-46.
Andrade, I., S. Hormazabal & M. Correa-Ramirez. 2014. Time-space variability of satellite chlorophyll-n in the Easter Island Province, Southeastern Pacific Ocean. Lat. Am. J. Aquat. Res., 42(4): 871-887.
Baker, A.C., P.W. Glynn & B. Riegl. 2008. Climate change and coral reef bleaching: an ecological assessment of long-term impacts, recovery trends and future outlook. Estuar. Coast. Shelf Sci., 80: 435-471.
Barnosky, A.D., E.A. Hadly, J. Bascompte, E.L. Berlow J.H. Brown, M. Fortelius,W.M. Getz, J. Harte, A. Hastings, P.A. Marquet, N.D. Martinez, A. Mooers, P. Roopnarine, G. Vermeij, J.W. Williams, R. Gillespie, J. Kitzes, C. Marshall, N. Matzke, D.P. Mindell, E. Revilla & A. B. Smith. 2012. Approaching a state shift in Earth's biosphere. Nature, 486: 52-58.
Bellwood, D.R., T.P. Hughes, C. Folke & M. Nystrom. 2004. Confronting the coral reef crisis. Nature, 429: 827-833.
Broitman, B.R., S.A. Navarrete, F. Smith & S.D. Gaines. 2001. Geographic variation of southeastern Pacific intertidal communities. Mar. Ecol. Prog. Ser., 224: 2134.
Castilla, J.C. & N. Rozbaczylo. 1987. Invertebrados marinos de Isla de Pascua y Salas y Gomez. In: J.C. Castilla (ed.). Islas oceanicas chilenas: conocimiento
cientifico y necesidades de investigaciones. Ediciones Universidad Catolica de Chile, Santiago, pp. 191-215.
Cheal, A.J., M.A. MacNeil, E. Cripps, M.J. Emslie, M. Jonker, B. Schaffelke & H. Weatman. 2010. Coralmacroalgal phase shifts or reef resilience: links with diversity and functional roles of herbivorous fishes on the Great Barrier Reef. Coral Reefs, 29: 1005-1015.
Dayton, P.K., M.J. Tegner, P.B. Edwards & K.L. Riser. 1998. Sliding baselines, ghosts, and reduced expectations in kelp forest communities. Ecological applications. Bull. Ecol. Soc. Am., 8: 309-322.
Diaz-Pulido, G., L.J. McCook, S. Dove, R. Berkelmans, G. Roff, D.I. Kline, S. Weeks, R.D. Evans, D.H. Williamson & O. Hoegh-Guldberg. 2009. Doom and boom on a resilient reef: climate change, algal overgrowth and coral recovery. PLoS ONE, 4(4): e5239.
DiSalvo, L.H., J. Randall & A. Cea. 1988. Ecological reconnaissance of the Easter Island sublittoral marine environment. Nat. Geogr. Res., 4: 451-473.
DiSalvo, L.H., J. Randall & A. Cea. 2007. Stomach contents and feeding observations of some Easter Island fishes. Atoll Res. Bull., 548: 1-22.
Done, T.J. 1992. Phase shifts in coral reef communities and their ecological significance. Hydrobiologia, 247: 121-132.
Edwards, C.B., A.M. Friedlander, A.G. Green, M.J. Hardt, E. Sala, H.P. Sweatman, I.D. Williams, B. Zgliczynski, S.A. Sandin & J.E. Smith. 2014. Global assessment of the status of coral reef herbivorous fishes: evidence for fishing effects. Proc. Roy. Soc., 281 pp.
Friedlander, A.M., E. Ballesteros, J. Beets, E. Berkenpas, C.F. Gaymer, M. Gorny & E. Sala. 2013. Effects of isolation and fishing on the marine ecosystems of Easter Island and Salas y Gomez Island, Chile. Aquat. Conserv., 23: 515-531.
Fung, T., R.M. Seymour & C.R. Johnson. 2013. Warning signals of regime shifts as intrinsic properties of endogenous dynamics. Am. Nat., 182: 208-222.
Glynn, P.W., G.M. Wellington, E.A. Wieters & S.A. Navarrete. 2003. Reef-building coral communities of Easter Island (Rapa Nui), Chile. In: J. Cortes (ed.). Latin American Coral Reefs. Elsevier Science, Amsterdam. pp. 473-494.
Glynn, P.W., G.M. Wellington, B. Riegl, D.B. Olson, E. Borneman & E.A. Wieters. 2007. Diversity and biogeography of the scleractinian coral fauna of Easter Island (Rapa Nui). Pac. Sci., 61: 67-90.
Halpern, B.S. & S.R. Floeter. 2008. Functional diversity responses to changing species richness in reef fish communities. Mar. Ecol. Prog. Ser., 364:147-156.
Hubbard, D.K. & M. Garcia. 2003. The corals and coral reefs of Easter Island -a preliminary look. In: J. Loret & J.T. Tanacredi (eds.). Easter Island: Scientific exploration into the world's environmental problems in microcosm. Kluwer Academic/Plenum Publishers. New York, pp. 3-77.
Hughes, T.P. 1994. Catastrophes, phase-shifts, and largescale degradation of a Caribbean Coral-Reef. Science, 265: 1547-1551.
Hughes, T.P., N.A.J. Graham, J.B.C. Jackson, P.J. Mumby & R.S. Steneck. 2010. Rising to the challenge of sustaining coral reef resilience. Trends Ecol. Evol., 25: 633-642.
Hughes, T.P., M.J. Rodrigues, D.R. Bellwood, D. Ceccarelli, O. Hoegh-Guldberg, L. McCook, N. Moltschaniwskyj, M.S. Pratchett, R.S. Steneck & B. Willis. 2007. Phase shifts, herbivory, and the resilience of coral reefs to climate change. Curr. Biol. CB, 17: 360-365.
Huston, M.A. 1985. Patterns of species diversity on coral reefs. Annu. Rev. Ecol. Syst., 16: 149-177.
Idjadi, J.A., S.C. Lee, J.F. Bruno, W.F. Precht, L. AllenRequa & P.J. Edmunds. 2006. Rapid phase-shift reversal on a Jamaican coral reef. Coral Reefs, 25: 209-211.
Iglesias-Prieto, R., V.H. Beltran, T.C. LaJeunesse, H. Reyes-Bonilla & P.E. Thome. 2004. Different algal symbionts explain the vertical distribution of dominant reef corals in the eastern Pacific. Proc. Roy. Soc., 271: 1757-1763.
Jackson, J.B.C., M.X. Kirby, W.H. Berger, L. Botsford, B.J. Bourque, R.H. Bradbury, R. Cooke, J. Erlandson, J.A. Estes, T.P. Hughes, S. Kidwell, C.B. Lange, H.S. Lenihan, J.M. Pandolfi, C. M. Peterson, R.S. Steneck, M.J. Tegner & R.R. Warner. 2001. Historical overfishing and the recent collapse of coastal ecosystems. Science, 293: 629-637.
Jompa, J. & L.J. McCook. 2002. Effects of competition and herbivory on interactions between a hard coral and a brown alga. J. Exp. Mar. Biol. Ecol., 271: 25-39.
Legendre, P. & L. Legendre. 1998. Developments in environmental modelling. Numerical ecology. Elsevier Sci., 20: 1-853. doi. 10.1016/S0167-8892(12)70002-7.
Leichter, J.L. & S.J. Genovese. 2006. Intermittent upwelling and subsidized growth of the scleractinian coral Madracis mirabilis on the deep fore-reef slope of Discovery Bay, Jamaica. Mar. Ecol. Prog. Ser., 316: 95-103.
Levin, S.A. & J. Lubchenco. 2008. Resilience, robustness, and marine ecosystem based management. Bioscience, 58: 27.
Lirman, D., S. Schopmeyer, D. Manzello, L.J. Gramer, W.F. Precht, F. Muller-Karger, K. Banks, B. Barnes, E. Bartels, A. Bourque, J. Byrne, S. Donahue, J.
Duquesnel, L. Fisher, D. Gilliam, J. Hendee, M. Johnson, K. Maxwell, E. McDevitt, J. Monty, D. Rueda, R. Ruzicka & S. Thanner. 2011. Severe 2010 cold-water event caused unprecedented mortality to corals of the Florida reef tract and reversed previous survivorship patterns. PLoS ONE, 6: e23047.
Madin, J.S. & S.R. Connolly. 2006. Ecological consequences of major hydrodynamic disturbances on coral reefs. Nature, 444: 477-480.
McCook, L.J., J. Jompa & G. Diaz-Pulido. 2001. Competition between corals and algae on coral reefs: a review of available evidence and mechanisms. Coral Reefs, 19: 400-417.
Myers, R.A. & B. Worm. 2003. Rapid worldwide depletion of predatory fish communities. Nature, 423: 280-283.
Navarrete, S.A., E.A. Wieters, B.R. Broitman & J.C. Castilla. 2005. Scales of benthic-pelagic coupling and the intensity of species interactions: from recruitment limitation to top-down control. P. Natl. Acad. Sci. USA, 102: 18046-18051.
Nystrom, M., C. Folke & F. Moberg. 2000. Coral reef disturbance and resilience in a human-dominated environment. Trends Ecol. Evol., 15: 413-417.
Ojeda, F. & S. Aviles. 1987. Peces oceanicos chilenos. In: J.C. Castilla (ed.). Islas oceanicas chilenas: conocimiento cientifico y necesidades de investigaciones. Ediciones Universidad Catolica de Chile, Santiago, pp.
Pauly, D. 1995. Anecdotes and the shifting baseline syndrome of fisheries. Trends Ecol. Evol., 10: 430.
Randall, J.E. & A. Cea. 2011. Shore fishes of Easter Island. Ref. 89467. University of Hawaii Press, Honolulu, 164 pp.
Santelices, B. & I.A. Abbott. 1987. Geographic and marine isolation: an assessment of the marine algae of Easter Island. Pac. Sci., 41: 1-20.
Scheffer, M., S. Carpenter, J.A. Foley, C. Folke & B. Walker. 2001. Catastrophic shifts in ecosystems. Nature, 413: 591-596.
Sepulveda, J. 1987. Peces de las islas oceanicas chilenas. In: J.C. Castilla (ed.). Islas oceanicas Chilenas: conocimiento cientifico y necesidades de investigaciones. Ediciones Universidad Catolica de Chile, Santiago, pp. 225-270.
Sheppard, C. 2009. Large temperature plunges recorded by data loggers at different depths on an Indian Ocean atoll: comparison with satellite data and relevance to coral refuges. Coral Reefs, 28: 399-403.
Smith, J.E., C.M. Smith, P.S. Vroom, K.L. Beach & S. Miller. 2004. Nutrient and growth dynamics of Halimeda tuna on Conch Reef, Florida Keys: possible
influence of internal tides on nutrient status and physiology. Limnol. Oceanogr., 49: 1923-1936.
Steneck, R.S. & M.N. Dethier. 1994. A functional group approach to the structure of algal-dominated communities. Oikos, 69: 476-498.
Stuart-Smith, R.D., A.E. Bates, J.S. Lefcheck, J.E. Duffy, S.C. Baker, R.J. Thomson, J.F. Stuart-Smith, N.A. Hill, S.J. Kininmonth, L. Airoldi, M.A. Becerro, S.J. Campbell, T.P. Dawson, S.A. Navarrete, G.A. Soler, E.M.A. Strain, T.J. Willis & G.J. Edgar. 2013. Integrating abundance and functional traits reveals new global hotspots of fish diversity. Nature, 501: 539-542.
Veron, J.E.N. & T.J. Done. 1979. Corals and coral communities of Lord Howe Island. Mar. Fresh. Res., 30: 203-236.
Wellington, G.M., P.W. Glynn, A.E. Strong, S.A. Navarrete, E.A. Wieters & D. Hubbard. 2001. Crisis on coral reefs linked to climate change. Trans. Am. Geophys. Union, 82: 1-5.
Wellington, G.M. & P.W. Glynn. 2007. Responses of coral reefs to El Nino-Southern Oscillation seawarming events. In: R.B. Aronson (ed.). Geological approaches to coral reef ecology. Springer, New York, pp. 342-385.
Wieters, E.A., B.R. Broitman & G.M. Branch. 2009. Benthic community structure and spatiotemporal thermal regimes in two upwelling ecosystems: comparisons between South Africa and Chile. Limnol. Oceanogr., 54: 1060-1072.
Willis, T., R. Miller & R. Babcock. 2000. Detection of spatial variability in relative density of fishes: comparison of visual census, angling, and baited underwater video. Mar. Ecol. Prog. Ser., 198: 249-260.
Evie A. Wieters (1), Alba Medrano1 & Alejandro Perez-Matus (2)
(1) Centro de Conservacion Marina & Departamento de Ecologia, Estacion Costera de Investigaciones Marinas Facultad de Ciencias Biologicas, Pontificia Universidad Catolica de Chile P.O. Box 114-D, Santiago, Chile (2) Subtidal Ecology Laboratory & Center for Marine Conservation, Estacion Costera de Investigaciones Marinas,
Pontificia Universidad Catolica de Chile P.O. Box 114-D, Santiago, Chile
Corresponding author: Evie A. Wieters (email@example.com)
Table 1. Two-way ANOVA tests of functional group abundances in relation to site and year. Bold values are significant at [alpha] = 0.05. df SS F P Ephemerals Site 4 1226.64 3.94 0.0049 Year 1 1127.31 14.49 0.0002# Site x Year 4 415.81 1.34 0.2606 Error 116 11764.76 Articulated Site 4 85.46 5.49 0.0005# Year 1 78.75 20.11 <0.0001# Site x Year 4 21.77 1.40 0.24 Error 107 416.65 Crusts Site 4 5.05 1.87 0.1198 Year 1 3.19 4.73 0.0317# Site x Year 5 5.05 1.87 0.1198 Error 116 78.24 Leathery Site 4 2.68 2.91 0.0244# Year 1 0.07 0.31 0.5774 Site x Year 4 3.11 3.38 0.0118# Error 116 26.67 Corals Site 4 9974.70 6.59 <0.0001# Year 1 33.52 0.09 0.7665 Site x Year 4 740.54 0.49 0.7435 Error 116 43876.63 Herbivores Site 4 0.10 0.52 0.7216 Year 1 0.02 0.45 0.5054 Site x Year 4 0.10 0.52 0.7216 Error 116 5.76 Table 2. List of fish species identified by contrasting survey methods and their assigned trophic functional groups. Family Species Muraenidae Gymnothorax eurostus Gymnothorax nasuta Congridae Conger cinereus Synodontidae Synodus isolatus Holocentridae Sargocentron wilhelmi Fistulariidae Fistularia commersonii Aulostomidae A ulos to mus chinensis Scorpaenidae Scorpaena orgila Serranidae Acanthistius fuscus Trachypoma macracanthus Priacanthidae Heteropriacanthus cruentatus Carangidae Pseudocaranx cheilio Mullidae Mulloidichthys vanicolensis Kyphosidae Girella nebulosa Kyphosus sandwicensis Chaetodontidae Chaetodon flavirostris Chaetodon litus Chaetodon pelewensis Forcipiger flavissim us Hem itaurichthys multispinosus Pomacanthidae Centropyge flavissimus Centropyge hotumatua Cirrhitidae Itycirrhitus wilhelmi Pomacentridae Chromis randalli Chrysiptera rapanui Stegastes fasciolatus Labridae Anampses caeruleopunctatus A nampses femininus Bodianus unimaculatus Cheilio inermis Coris dehueni Pseudolabrus fuentesi Thalassoma lutescens Thalassoma purpureum Gobiidae Pascua caudUinea Priolepis squamogena Acanthuridae Acanthurus leucopareius Soleidae Aseraggodes bahamondei Monacanthidae Aluterus scriptus Canther hines dumerilii Cantherhines rapanui Balistidae Xanthichthys mento Ostraciidae Lactoria diaphana Tetraodontidae Arothron meleagris Diodontidae Diodon holocanthus Total (%) Family Rapanui or common name Trophic category Muraenidae Koreha puhi, Morena Piscivore+invertivore Koreha, Morena Piscivore+invertivore Congridae Ko'iro Inverti vore+pi scivore Synodontidae Papa hakatara Piscivore Holocentridae Marau hiva Carnivore Fistulariidae Toto amo hiku kio'e, Piscivore+invertivore Pez Corneta Aulostomidae Toto amo Inverti vore+pi scivore Scorpaenidae Nohu Piscivore Serranidae Kopuku mangara Inverti vore Kopuku kava Inverti vore Priacanthidae Mata uira, Matahuira Invertivore+planktivore Carangidae po'opo'o Invertivore+piscivore Mullidae 'a'avere Invertivore Kyphosidae Mahaki, Nanue (c) Omnivore Nanue Herbivore Chaetodontidae Tipi tipi Omnivore Tipi tipi uri Invertivore Tipi tipi Omnivore Tipi tipi hoe Invertivore Multispine Planktivore butterflyfish Pomacanthidae Lemonpeel angelfish Herbivore Kntoti para Herbivore Cirrhitidae Pilikoa Piscivore+invertivore Pomacentridae Maniata Planktivore Maniata Planktivore Kototi Omnivore Labridae Mori, Marari Invertivore Pahika Invertivore Matuku Invertivore Ure ure Invertivore Teteme Invertivore Kntea Invertivore Mori vaihi Invertivore Kakaka, Ra'emea Invertivore Gobiidae Pascua Goby Invertivore Scaledcheek goby Invertivore Acanthuridae Ma'ito Herbivore Soleidae Rahai Invertivore+piscivore Monacanthidae Paoa Omnivore Koreva Omnivore Koreva Omnivore Balistidae Kokiri Planktivore Ostraciidae Momo tara Invertivore Tetraodontidae Titeve kapovai, Invertivore Pez globo Diodontidae Titeve tara tara Invertivore Total (%) Family Video Census Endemic Muraenidae + + + + Congridae + + Synodontidae + + Holocentridae + + Fistulariidae + + Aulostomidae + + Scorpaenidae + + Serranidae + + + + Priacanthidae + Carangidae + + Mullidae + + Kyphosidae + + + + Chaetodontidae + + + + + + + + Pomacanthidae + + + + Cirrhitidae + Pomacentridae + + + + + + + Labridae + + + + + + + + + + - + + + + Gobiidae + + + + Acanthuridae + + Soleidae + Monacanthidae + + + + + * + Balistidae + + Ostraciidae + + Tetraodontidae + + Diodontidae + + Total (%) 87% 80% 24%
|Printer friendly Cite/link Email Feedback|
|Title Annotation:||articulo en ingles|
|Author:||Wieters, Evie A.; Medrano, Alba; Perez-Matus, Alejandro|
|Publication:||Latin American Journal of Aquatic Research|
|Date:||Oct 1, 2014|
|Previous Article:||Estructura comunitaria del macrobentos y peces costeros submareales del archipielago Juan Fernandez, Chile.|
|Next Article:||La pesca en Isla de Pascua, una historia reciente (1950-2010).|