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Open Access

Peer-reviewed

Research Article

Size matters: Predator outbreaks threaten foundation species in small Marine Protected Areas

Abstract

The unanticipated impacts of consumers in fragmented habitats are frequently a challenge for ecosystem management. On Indo-Pacific coral reefs, crown-of-thorns sea stars (Acanthaster spp.) are coral predators whose outbreaks cause precipitous coral decline. Across large spatial scales, Acanthaster densities are lower in large no-take Marine Protected Areas (MPAs) and reefs subject to limited human exploitation. However, using a combination of observational and manipulative experiments, we found that Acanthaster densities within a network of small, no-take MPAs on reef flats in Fiji were ~2–3.4 times greater inside MPAs than in adjacent fished areas and ~2–2.5 times greater than the upper threshold density indicative of an outbreak. This appeared to result from selective Acanthaster migration to the coral-rich MPAs from fished areas that are coral-poor and dominated by macroalgae. Small MPAs can dramatically increase the cover of foundation species like corals, but may selectively attract coral predators like Acanthaster due to greater food densities within MPAs or because the MPAs are too small to support Acanthaster enemies. As coral cover increases, their chemical and visual cues may concentrate Acanthaster to outbreak densities that cause coral demise, compromising the value of small MPAs. An understanding of predator dynamics as a function of habitat type, size, and fragmentation needs to be incorporated into MPA design and management.

Citation: Clements CS, Hay ME (2017) Size matters: Predator outbreaks threaten foundation species in small Marine Protected Areas. PLoS ONE 12(2): e0171569. https://doi.org/10.1371/journal.pone.0171569

Editor: Heather M. Patterson, Department of Agriculture and Water Resources, AUSTRALIA

Received: November 19, 2016; Accepted: January 22, 2017; Published: February 6, 2017

Copyright: © 2017 Clements, Hay. This is an open access article distributed under the terms of the Creative Commons Attribution License, which permits unrestricted use, distribution, and reproduction in any medium, provided the original author and source are credited.

Data Availability: Datasets used in this study will be available online from the BCO-DMO data system (http://www.bco-dmo.org/project/480717).

Funding: Financial support provided by NSF grant OCE- 0929119, NIH ICBG grant U19TW007401, and the Teasley Endowment to the Georgia Institute of Technology. The funders had no role in study design, data collection and analysis, decision to publish, or preparation of the manuscript.

Competing interests: The authors have declared that no competing interests exist.

Introduction

The increasing frequency and severity of anthropogenic impacts throughout the global ocean has led to habitat degradation, fragmentation, and trophic downgrading of marine ecosystems worldwide [1, 2]. To counter these trends and promote ecosystem recovery and resilience, Marine Protected Areas (MPAs) are increasingly being established–often with broadly defined goals oriented towards the protection of foundation species (e.g., coral, kelp, seagrass, mangroves, etc.) upon which a broad variety of other species depend [3]. Efforts to establish MPAs have been particularly urgent on tropical coral reefs, which have experienced dramatic declines in coral cover and coral-associated species [4, 5, 6, 7] and in numerous cases have transitioned from structurally complex systems dominated by corals to structurally simplified systems dominated by macrolagae [8, 9].

While the number of MPAs worldwide has steadily increased, MPA design and management strategies are variable, with many no-take MPAs being small habitat fragments embedded within a broader background of exploited, and often degraded, habitat [10]. Indeed, an explicit aim of many MPAs is to aid the rehabilitation of surrounding degraded areas via spillover of adults and export of larvae [11]. There is considerable debate over how size affects MPA performance, but much of this has focused on how size influences protection from human exploitation (e.g., incorporating species’ home ranges and migration) and replenishment of focal species populations (e.g., larval export, recruitment, and spillover) [12, 13, 14]. In contrast, the effects of reserve size on predator densities or behaviors have rarely been addressed, despite the ability of consumers to destabilize species and community-level dynamics–especially if they attack foundation species or ecosystem engineers [1, 15]. Because predators have dramatic direct and indirect impact on community structure and function [1, 16], predicting and mitigating predator-induced disturbances are necessary to safeguard ecosystem integrity and will be increasingly important as global-scale stressors continue to challenge the effectiveness of local management efforts [17, 18].

A major driver of the recent 50% loss in coral cover along the Great Barrier Reef and on reefs throughout the tropical Pacific is predation by the crown-of-thorns sea star (Acanthaster spp.) [5, 7], which exhibits population outbreaks that can reduce live coral over vast areas and can lead to the ecological collapse of entire reef systems [19]. Acanthaster outbreaks are hypothesized to occur via several mechanisms, including (i) reduced population constraints (e.g., predation) that contribute to one or successive mass recruitment events and/or (ii) concentrated aggregations of foraging adults (for review, see [20]). Acanthaster adults use a combination of chemical and visual sensory cues to navigate toward preferred corals [21, 22, 23], and during outbreaks, have been shown to aggregate on corals being eaten by conspecifics [22] and move en masse from areas of depleted coral to unexploited reef tracts in search of food [19, 24]. There is also correlative evidence across large spatial scales that limited or restricted fishing is associated with low densities of Acanthaster–hypothetically due to the maintenance of intact food webs that exert top-down control on Acanthaster populations [25, 26]. However, despite these correlations over large areas [25] and long time periods [26], the identity of critical predators and the life-stage of Acanthaster on which they feed remain unknown, and therefore speculative as a mechanism of population control.

Retention of food-web connections, along with other fisheries and conservation benefits, have been touted in the literature and used to advocate for MPAs [3, 27], which are now one of the most widespread management tools used by coastal communities throughout the Pacific [28]. Despite their general success [29, 30], some MPAs appear ineffective and can even hasten degradation of remaining critical habitat if they lead to unexpected consumer impacts on foundation species [15]. Studies from terrestrial systems emphasize that habitat fragmentation can lead to mesopredator outbreaks via reduced top-down and bottom-up population constraints [31, 32], but these insights have received limited attention in planning and management of MPAs, especially as a function of size and of being embedded within increasingly fragmented and degraded marine ecosystems. Most MPAs are small (< 1.0 km2) [10]–with management focused almost solely on various forms of fishing restrictions (e.g., permanent, partial or periodic restrictions) [28]. Here, we provide evidence that small reserves can be at special risk for predator (Acanthaster spp.) outbreaks and suggest that the probability of outbreak densities may increase as conservation succeeds at increasing coral cover and thus food for, and attraction of, Acanthaster.

Materials and methods

Ethics statement

Research was approved by the Fijian government and the Korolevu-i-Wai Environment Committee, which oversees management of nearshore marine resources where we conducted our research. Fieldwork was performed in accordance with the ethical regulations of the Georgia Institute of Technology and Fijian Law.

Study area

This study was conducted within paired fished and no-take MPAs on reef flats (depth of ~0–2 m at low tide and ~1–3+ m at high tide) adjacent to Namada, Vatu-o-lalai, and Votua villages along the Coral Coast of Viti Levu, Fiji (18° 13.059’S, 177° 42.979’E) (Fig 1). Paired areas were located within an 11 km stretch of fringing reefs that are separated by a series of deep-water channels. MPAs within this reef system are small (0.45–0.78 km2) and separated by ~2.6–10 km. MPAs exhibited high coral cover (~38–56%) and low macroalgal cover (~1–3%) on hard substrates [33, 34], as well as higher biomass and diversity of herbivorous and piscivorous fishes often targeted by artisanal fishers [34, 35]. Conversely, adjacent fished areas were relatively degraded with low coral cover (4–16%), high macroalgal cover (~49–91%) [33, 34], and low biomass and diversity of herbivorous and piscivorous fishes [34, 35].

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Fig 1. Mean Acanthaster density is ~2–3.4 times greater within Marine Protected Areas (MPAs) than adjacent fished areas.

(Top panel) Village and MPA locations along the coast of Viti Levu, Fiji. Dark gray sections represent the MPAs at each site. (Bottom panel) Violin plots depicting the mean ± SE Acanthaster density (large black dots and error bars), the frequency of plots with differing densities of Acanthaster (the enclosed areas), and each individual plot as a function of Acanthaster counted in that 15 X 15 m plot (small black dots) within MPAs (dark gray) and adjacent fished areas (white) at each village (n = 15 quadrats reef-1 location-1). Data for each pairwise comparison were analyzed using a generalized linear model (GLM) with a Poisson distribution (Votua) or quasi-GLM models (Namada and Vatu-o-lalai).

https://doi.org/10.1371/journal.pone.0171569.g001

Acanthaster cf. planci density

Acanthaster density was quantified in paired MPAs and fished areas using 15×15 m quadrats (n = 15 reef-1 location-1) that were non-overlapping and distributed haphazardly within the reef flat of each area. Surveys entailed a single snorkeler carefully searching for Acanthaster within and under rock ledges and coral colonies within each quadrat for five minutes, especially near areas with obvious signs of Acanthaster feeding. Acanthaster abundance data violated parametric assumptions, so differences between paired MPAs and fished areas were evaluated using Wilcoxon rank-sum pair-wise comparisons.

Experimental tagging study

To evaluate how tagging might affect Acanthaster behavior, we conducted preliminary experiments comparing righting ability and feeding behavior of tagged and untagged Acanthaster (n = 10 individuals treatment-1) that were caged on the reef flat of Votua’s MPA. Ten individuals were each tagged by inserting five plastic tag fasteners at the base of individual arms near the oral disk (S1 Fig), and all individuals were held in individual cages on the reef flat for the 7-day duration of this experiment. Two days were allowed for tag acclimation among the treatment group before experiments were conducted. Righting ability was assessed on days 3 and 7 post-tagging by flipping individuals onto their aboral surface and measuring the time required to right themselves onto their oral surface. This was repeated three times for each individual with a 1-minute rest interval between trials. Prior to analysis, data were log transformed and tested for homogeneity of variance using Bartlett’s test. Mean righting times within and between days were compared using a two-way ANOVA. Individuals were also offered two small fragments of the coral Montipora hispida (~8–10 cm length) on days three and five post-tagging to assess the effects of tagging on feeding behavior. Comparisons of whether the corals offered were either both eaten or both not eaten within 24 h were conducted using a Fisher’s exact test (there were no cases of only one coral being eaten).

To test whether Acanthaster selectively migrated into the MPAs versus the fished areas, 120 adults of 36 ± 2 cm diameter (from the tips of opposite arms) were collected from the MPAs and adjacent fished areas of reefs flats near Votua, Vatu-o-lalai, and Namada villages, with 20 individuals collected from within and 20 from outside the MPAs at each village site (40 individuals village-1 site-1). Each individual was tagged with five plastic tag fasteners between the base of individual arms, and labeled flagging tape was attached to the end of each tag fastener to aid in location and identification (S1 Fig). Individuals were then enclosed within cages located along the MPA border perpendicular to the coastline at each site (20 individuals border-1 location-1) for 48 h to allow for tag acclimation. Upon release, individuals’ movements were monitored at 24 h intervals for four to eight days by physically locating each individual and recording its location via GPS (Garmin GPS 76CSX). GPS coordinates of individual Acanthaster positions were imported into ArcMAP (Version 10.3.1), and the Geospatial Modeling Environment extension (Version 0.7.4.0) was used to calculate individuals’ initial and final directions of movement relative to their release point along their respective MPA border, as well as each individual’s net displacement between consecutive days (S1 Fig). The angular directions of individuals’ positions relative to the MPA border were plotted as circular data and together tested for circular uniformity against an alternative that presumes a specified angle (e.g., 90°) using Batschelet’s modified Hodges-Ajne test. This analysis was conducted for both the first and final relocation of each individual because initial orientations are more suitable for evaluating patch detection capabilities [36, 37].

To determine whether an individual’s origin influenced their movement direction, we compared relocations, both pooled across all villages and individually for each MPA border, of Acanthaster collected from the MPAs and fished areas. We also characterized the path directionality of individual sea star movements at each border where data from two or more consecutive movements (relocations on two+ days in a row) were available using the ratio of D (the net displacement from initial to final position in the path) to W (total distance traveled between days) [38, 39]. A D:W ratio of 1 represents an individual exhibiting uniformly directional movement (i.e., straight-line path). Values > 0.7 are considered highly directional, > 0.5 partially directional, and < 0.5 undirected [38, 39].

MPA border benthic surveys

Surveys of benthic community composition were conducted to assess habitat differences inside and outside of each MPA border and the relationship between coral cover and Acanthaster displacement at each border. Surveys used 40 m point intercept transects (n = 20 transects border-1 MPA-1, points at 0.5 m intervals, 1,600 points border -1) that were non-overlapping (mean distance between transects = ~12 m) and oriented parallel to the coastline, with the midpoint (20 m) of each transect positioned on the MPA border (20 m within the MPA and 20 m within the fished area) (S1 Fig). Benthic data from within and outside each MPA border were square root transformed if needed, and analyzed using t-tests. When benthic data could not be transformed to meet parametric assumptions, the original count data were used and analyzed with quasi-GLM models. To test for correlations between coral cover and Acanthaster movement among sites, coral cover along the transect at the site of each individual’s release as well as pooled coral cover by MPA border were each, separately, linearly regressed against the displacement between consecutive days exhibited by the individual sea star at that location and the mean displacement along each individual MPA border, respectively.

Results

We found that Acanthaster densities within MPAs (~80–98 ha-1) were ~2–3.4 times greater than within fished areas (~23–47 ha-1; p ≤ 0.030, Fig 1), as well as ~2–2.5 times greater than the upper threshold density indicative of an outbreak (40 individuals per hectare [24]). Our tagging methods affected neither righting times (p = 0.190) nor frequencies of feeding (p = 1.000) for Acanthaster; there also was no effect of assessing these behaviors on days three or seven post tagging (p ≥ 0.719, S1 Table).

When Acanthaster were released along MPA borders, their directions of initial movement were significantly biased toward the MPA for five of the six borders (p < 0.050, Fig 2), and suggestive of an MPA preference in the remaining contrast. Approximately 73% of all individuals released and relocated (85 of 116) moved to the MPA, a pattern that was consistent regardless of whether Acanthaster were originally collected from the MPAs or fished areas (p > 0.656, S2 Table). Similarly, final movement positions were significantly biased toward MPAs for all six contrasts (p < 0.050, Fig 2). The ratio of net displacement (D) to total displacement between consecutive days (W) indicated that Acanthaster movement paths exhibited considerable directionality at five of the six MPA borders (D:W = 0.453–0.717, S3 Table).

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Fig 2. Acanthaster move selectively to Marine Protected Areas (MPAs).

Movement directions of initial (solid dots) and final (open dots) Acanthaster relocations from release points at MPA/non-MPA borders on each side of the MPA at each of the three villages. Arrows represent the resultant vector (R) for initial (black) and final (gray) relocations. * and * indicate a significant difference between Acanthaster movement towards MPAs (white region) rather than fished areas (shaded region) (Modified Hodges-Ajne test, p < 0.050).

https://doi.org/10.1371/journal.pone.0171569.g002

Benthic community composition commonly differed immediately within versus outside MPAs, with coral and macroalgal cover exhibiting the most frequent significant differences across MPA borders (Fig 3). Coral cover 20 m within MPA borders was 80–440% greater than in the 20 m outside MPA borders, while macroalgal cover was 20–610% greater immediately outside versus inside the MPAs; differences were even more pronounced toward the centers of each area [34]. Acanthaster rates of displacement were negatively correlated with mean coral cover along each border, both when plotted by individual Acanthaster (R2 = 0.209, p < 0.001) and when pooled by MPA border (R2 = 0.756, p = 0.030; Fig 4).

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Fig 3. Reef habitat differs immediately inside vs. outside MPAs.

Comparisons of benthic cover (mean % ± SE) 20 m inside (black) and 20 m outside (gray) of MPA borders perpendicular to the coastline at Namada, Vatu-o-lalai, and Votua villages (n = 20 transects border-1 location-1). The category “Other” includes dead coral, rock, rubble/sand, and uncommon benthic organisms (e.g., zooanthids, soft coral). Asterisks after p-values indicate comparisons analyzed with quasi-GLM models.

https://doi.org/10.1371/journal.pone.0171569.g003

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Fig 4. Acanthaster displacement is negatively correlated with local coral cover.

(a) Relationship between individual Acanthaster displacement between consecutive days (m day-1) and coral cover (%) at each individual’s release location along MPA borders. (b) Relationship between coral cover (mean % ± SE) and Acanthaster displacement between consecutive days (m day-1; mean ± SE) when pooled by MPA border. See Fig 1 for village site names. Coefficients of regression (R2) and p-values are indicated in the graph. Two data points with extreme Acanthaster displacement values (y1 = 42.65 m, y2 = 34.39 m) at low coral cover (x1 = 0%, x2 = 11.25%) were excluded from analyses after performing an outlier analysis (Jackknife distances) using JMP (Version 11.0.0).

https://doi.org/10.1371/journal.pone.0171569.g004

Discussion

Our findings suggest that at small scales, common MPA benefits (e.g., increased coral cover) may attract predators such as Acanthaster. Acanthaster were 2–3.4 times as abundant within the coral-rich MPAs, exhibiting densities similar to those that have caused extensive coral decline (e.g., >50% [40]), and lead to cascading effects on reef structure and associated species [19]. This unanticipated pattern may provide an important lesson for the management of MPAs across the Pacific, as the overwhelming majority of tropical Pacific MPAs are small (<0.5 km2 [28]) and like those in this study, are situated within a background of increasingly degraded reef habitat [5]. Given the widespread use of small MPAs as management tools [10, 28] and the destructive impacts that Acanthaster feeding can have on coral reefs at the densities documented here [19, 40], our study highlights the need to consider how the size and placement of MPAs influence their susceptibility to Acanthaster outbreaks, and whether the probability of outbreaks increases with MPA success (enhanced coral cover, the foundation species for this system). Despite these high densities of Acanthaster, corals are still abundant in the MPAs we investigated [33, 34]. This may be due to Acanthaster densities increasing recently and not yet strongly suppressing coral cover or due to coral growth rates on these shallow, turbulent, and well-lit platforms being high enough to generate positive net growth despite high rates of consumption.

Our findings contrast with previous studies where Acanthaster densities were reduced in large MPAs or areas subject to limited fishing pressure [25, 26]. Our patterns may differ from these earlier studies due to (i) small MPAs having greater perimeter to area ratios that facilitate increased movement of Acanthaster into coral-rich MPAs, (ii) habitat disparities between coral-rich MPAs and surrounding degraded reefs that enhance Acanthaster recruitment and immigration to coral-rich MPAs, (iii) differences in critical consumers or processes between MPAs located on shallow (~1–3+ m) reef flats like those we studied vs. reefs from previous studies (~7m in depth or greater), or (iv) large MPAs supporting critical consumers or processes that are not sustainable in the small MPAs we studied. That said, it is critical to note that previous studies assumed that predation suppressed Acanthaster densities in large MPAs or areas with reduced fishing pressure [25, 26], but this assumption was not directly tested. Neither the identities of the critical predators of Acanthaster nor the life stage at which predation could control Acanthaster have been determined. Regardless, it is evident from our tagging and density data that Acanthaster can selectively migrate into the coral-rich MPAs vs. the coral-poor fished areas and that predation within these reef systems is insufficient to reduce Acanthaster numbers to densities below those capable of causing considerable damage to coral communities. These findings highlight an important risk for the many small MPAs embedded within increasingly fragmented and degraded reef ecosystems.

Predator outbreaks may occur in small MPAs due to increased resource availability as the MPAs become effective and enhance the abundance of foundation species that serve as attractive foods for consumers [15]. Our findings build on a small, but growing, body of evidence that consumer attraction may be a critical vulnerability for effective management, as similar scenarios have been documented in other systems, including attraction and overgrazing of seagrass MPAs by sea turtles [15] and plant community regime shifts due to elephant aggregations in African reserves [41, 42]. On coral reefs, this phenomenon may be especially problematic if degraded areas near MPAs serve as nurseries for predators such as Acanthaster. For Acanthaster, degraded areas surrounding reserves have abundant coral rubble (into which juvenile Acanthaster selectively recruit [43]) and considerable abundance of crustose coralline algae, a favored food of juvenile Acanthaster (for review, see [20]). Increased juvenile survival in these degraded reef areas followed by selective migration to coral-rich MPAs could contribute to the high Acanthaster densities we documented within MPAs.

A second possibility, or additional contributor, to the density difference we noted is that small, fragmented systems may lack top predators, sometimes allowing mesopredators like Acanthaster to escape consumer control [31, 32, 44]. Predatory fish biomass was low in both the small MPAs and the fished areas we investigated [35, 45, 46] and is comparable to, or lower than, the biomass of predatory fishes on reefs previously associated with high Acanthaster population densities [25]. Lower coral abundance in fished areas may also reduce predation on larval and juvenile Acanthaster by coral-associated planktivorous fishes, which have been shown in laboratory trials to prey upon Acanthaster larvae [47]. While the identity and roles of predatory fishes controlling Acanthaster densities in the wild are largely unknown [20, 25, 26], it is plausible that our predator-depauperate reefs are incapable of exerting top-down control on Acanthaster (e.g., predation during vulnerable pre-reproductive stages [26]).

Regardless of what processes normally control Acanthaster densities, our tagging data show that migration of adult sea stars from degraded areas could lead to outbreak densities within the coral-rich MPAs. Acanthaster consistently moved towards the MPAs at rates proportional to local coral density; a behavior consistent with outbreak scenarios where sea stars migrate from areas of low coral abundance and aggregate on remaining coral patches [19, 48, 49]. However, rather than aggregative behavior induced by recent coral decline, data from the MPAs we studied suggest that increases in live coral following MPA establishment [33, 34, 45] are producing “food hotspots” that attract sea stars from surrounding overfished areas to form ‘spot’ outbreaks [50]. Greater herbivore control of macroalgae within MPAs [34] may further exacerbate this hotspot effect because macroalgae suppress Acanthaster feeding on adjacent corals [51], resulting in corals within the MPAs being not only more abundant and more attractive, but also more accessible to Acanthaster than corals in the degraded, seaweed-dominated areas surrounding the MPAs. Thus, common benefits of MPAs may become liabilities if reef spatial dynamics, consumer movements, and species interaction networks are not considered in a community context that extends beyond reserve borders.

While many outbreak densities of Acanthaster appear to occur following massive recruitment events [20, 52], this did not appear to be the process generating outbreak densities in our sites. We did not note high densities of Acanthaster in the fished areas or on deeper portions of adjacent reefs. Rather than resulting from boom and bust cycles, the high densities noted in the MPAs we studied appeared to result from lower chronic densities of Acanthaster aggregating in the food hot-spots generated within MPAs. Thus, these localized outbreak densities seem to be generated by different processes [48, 53] and to occur on different temporal and spatial scales than outbreaks noted in many previous investigations (for review, see [20]).

Optimizing local-scale management can provide a critical buffer for ecosystems subject to an increasing array of local and global disturbances [54]. Our study highlights a shortcoming of basic extraction restrictions if these are not integrated with issues of scale, migration, and food web dynamics. Across the Pacific where customary ownership and governance of marine resources occurs at a local scale, small MPAs are among the most common strategies used to manage coral reef ecosystems [28]. When enforced, they can produce remarkably positive effects [27, 29, 30], but as positive outcomes accumulate, this success may concentrate coral predators and endanger MPA resilience. An appreciation for mechanisms generating predator outbreaks needs to be included in the conceptual toolkit of MPA managers. This is particularly relevant to small, locally-managed MPAs where control of Acanthaster by physical removal, injections, or other means is likely feasible [55]. However, most MPA management strategies are limited to fishing restrictions that vary in scope and duration (e.g., permanent, partial, or periodic restrictions) [28] and are likely incapable of facilitating adequate biological control of Acanthaster. While protection from extraction may be conferring other benefits commonly expected from MPAs, the concern is that without active management of predators like Acanthaster, current schemes may promote situations where predation threatens the foundation species upon which MPA success is built. This could compromise gains that have been made since reserve establishment, as well as those expected for the future.

Supporting information

S1 Fig. Experimental tagging and monitoring of Acanthaster.

(a) Diagram of Acanthaster tagging and (b) photograph of a tagged Acanthaster. (c) Diagram of experimental design for tagged Acanthaster released along each MPA border and benthic surveys conducted along each MPA border. See key below diagram for symbol identification.

https://doi.org/10.1371/journal.pone.0171569.s001

(TIF)

S1 Table. Experimental tagging did not affect Acanthaster righting ability.

Two-way ANOVA on the effect of tagging on righting time of Acanthaster after 2 and 7 days. All data were log transformed. Bartlett test for homogeneity of variances (F = 0.883, p = 0.347).

https://doi.org/10.1371/journal.pone.0171569.s002

(TIF)

S2 Table. Movements of Acanthaster originating from the MPAs or fished areas into the MPA or fished areas at each MPA border.

Comparisons between Acanthaster of different origins tested with Fisher’s exact test.

https://doi.org/10.1371/journal.pone.0171569.s003

(TIF)

S3 Table. Acanthaster movement paths exhibited considerable directionality at five of the six MPA borders.

Acanthaster net displacement (m day-1; mean ± SE), displacement between consecutive days (m day-1; mean ± SE), and D:W ratio (mean ± SD) at MPA border locations.

https://doi.org/10.1371/journal.pone.0171569.s004

(TIFF)

Acknowledgments

We thank the Fijian government and the Korolevu-i- wai district elders for collection and research permissions, and V. Bonito for scientific and cultural support.

Author Contributions

  1. Conceptualization: CSC MEH.
  2. Data curation: CSC.
  3. Formal analysis: CSC.
  4. Funding acquisition: MEH.
  5. Investigation: CSC.
  6. Methodology: CSC MEH.
  7. Project administration: CSC MEH.
  8. Resources: CSC MEH.
  9. Supervision: CSC MEH.
  10. Validation: CSC MEH.
  11. Visualization: CSC MEH.
  12. Writing – original draft: CSC MEH.
  13. Writing – review & editing: CSC MEH.

References

  1. 1. Estes JA, Terborgh J, Brashares JS, Power ME, Berger J, Bond WJ, et al. Trophic downgrading of planet Earth. Science. 2011;333(6040):301–306. pmid:21764740
  2. 2. McCauley DJ, Pinsky ML, Palumbi SR, Estes JA, Joyce FH, Warner RR. Marine defaunation: animal loss in the global ocean. Science. 2015;347(6219):1255641. pmid:25593191
  3. 3. Lubchenco J., Palumbi S.R., Gaines S.D. & Andelman S. (2003). Plugging a hole in the ocean: the emerging science of marine reserves. Ecological Applications, 13, 3–7.
  4. 4. Gardner TA, Côté IM, Gill JA, Grant A, Watkinson AR. Long-term region-wide declines in Caribbean corals. Science. 2003;301(5635):958–960. pmid:12869698
  5. 5. Bruno JF, Selig ER. Regional decline of coral cover in the Indo-Pacific: timing, extent, and subregional comparisons. PLoS One. 2007;2(8):e711. pmid:17684557
  6. 6. Carpenter KE, Abrar M, Aeby G, Aronson RB, Banks S, Bruckner A, et al. One-third of reef-building corals face elevated extinction risk from climate change and local impacts. Science. 2008;321(5888):560–563. pmid:18653892
  7. 7. De'ath G, Fabricius KE, Sweatman H, Puotinen M. The 27-year decline of coral cover on the Great Barrier Reef and its causes. Proc Natl Acad Sci USA. 2012;109(44):17995–17999. pmid:23027961
  8. 8. Mumby PJ, Steneck RS. Coral reef management and conservation in light of rapidly evolving ecological paradigms. Trends Ecol Evol. 2008;23(10):555–563. pmid:18722687
  9. 9. Hughes TP, Graham NAJ, Jackson JBC, Mumby PJ, Steneck RS. Rising to the challenge of sustaining coral reef resilience. Trends Ecol Evol. 2010;25(11):633–642. pmid:20800316
  10. 10. Costello MJ, Ballantine B. Biodiversity conservation should focus on no-take Marine Reserves: 94% of Marine Protected Areas allow fishing. Trends Ecol Evol. 2015;30(9):507–509. pmid:26321055
  11. 11. Russ GR, Alcala AC. Enhanced biodiversity beyond marine reserve boundaries: the cup spillith over. Ecol Appl. 2011;21(1):241–250. pmid:21516901
  12. 12. Roberts CM, Andelman S, Branch G, Bustamante RH, Carlos Castilla J, Dugan J, et al. Ecological criteria for evaluating candidate sites for marine reserves. Ecol Appl. 2003;13(sp1):199–214.
  13. 13. Claudet J, Osenberg CW, Benedetti-Cecchi L, Domenici P, García-Charton J-A, Pérez-Ruzafa Á, et al. Marine reserves: size and age do matter. Ecol Lett. 2008;11(5):481–489. pmid:18294212
  14. 14. Gaines SD, White C, Carr MH, Palumbi SR. Designing marine reserve networks for both conservation and fisheries management. Proc Natl Acad Sci USA. 2010;107(43):18286–18293. pmid:20200311
  15. 15. Christianen MJA, Herman PMJ, Bouma TJ, Lamers LPM, van Katwijk MM, van der Heide T, et al. Habitat collapse due to overgrazing threatens turtle conservation in marine protected areas. Proc R Soc B. 2014;281(1777):7.
  16. 16. Ohgushi T, Schmitz OJ, Holt RD. Trait-mediated indirect interactions: ecological and evolutionary perspectives. Ohgushi T, Schmitz OJ, Holt RD, editors. New York: Cambridge University Press; 2012.
    • 17. Scheffer M, Barrett S, Carpenter SR, Folke C, Green AJ, Holmgren M, et al. Creating a safe operating space for iconic ecosystems. Science. 2015;347(6228):1317–1319. pmid:25792318
    • 18. Rocha JC, Peterson GD, Biggs R. Regime shifts in the Anthropocene: drivers, risks, and resilience. PLoS One. 2015;10(8):e0134639. pmid:26267896
    • 19. Kayal M, Vercelloni J, de Loma TL, Bosserelle P, Chancerelle Y, Geoffroy S, et al. Predator crown-of-thorns starfish (Acanthaster planci) outbreak, mass mortality of corals, and cascading effects on reef fish and benthic communities. PLoS One. 2012;7(10):e47363. pmid:23056635
    • 20. Pratchett MS, Caballes CF, Rivera-Posada JA, Sweatman HPA. Limits to understanding and managing outbreaks of crown-of-thorns starfish (Acanthaster spp.). Oceanogr Mar Biol Ann Rev. 2014;52:133–199.
    • 21. Barnes DJ, Brauer RW, Jordan MR. Field and laboratory observations of the crown-of-thorns starfish, Acanthaster planci: locomotory response of Acanthaster planci to various species of coral. Nature. 1970;228(5269):342–4. pmid:4394407
    • 22. Ormond RFG, Campbell AC, Head SH, Moore RJ, Rainbow PR, Saunders AP. Formation and breakdown of aggregations of crown-of-thorns starfish, Acanthaster planci (L). Nature. 1973;246(5429):167–9.
    • 23. Sigl R, Steibl S, Laforsch C. The role of vision for navigation in the crown-of-thorns seastar, Acanthaster planci. Scientific Reports. 2016;6:30834. pmid:27476750
    • 24. Moran PJ, Death G. Estimates of the abundance of the crown-of-thorns starfish Acanthaster planci in outbreaking and non-outbreaking populations on reefs within the Great Barrier Reef. Mar Biol. 1992;113(3):509–515.
    • 25. Dulvy NK, Freckleton RP, Polunin NVC. Coral reef cascades and the indirect effects of predator removal by exploitation. Ecol Lett. 2004;7(5):410–416.
    • 26. Sweatman H. No-take reserves protect coral reefs from predatory starfish. Curr Biol. 2008;18(14):R598–R599. pmid:18644332
    • 27. Graham NAJ, Ainsworth TD, Baird AH, Ban NC, Bay LK, Cinner JE, et al. From microbes to people: tractable benefits of no-take areas for coral reefs. Oceanogr Mar Biol Ann Rev. 2011;49:105–135.
    • 28. Jupiter SD, Cohen PJ, Weeks R, Tawake A, Govan H. Locally-managed marine areas: multiple objectives and diverse strategies. Pacific Conservation Biology. 2014;20(2):165–179.
    • 29. Lester SE, Halpern BS, Grorud-Colvert K, Lubchenco J, Ruttenberg BI, Gaines SD, et al. Biological effects within no-take marine reserves: a global synthesis. Mar Ecol Prog Ser. 2009;384:33–46.
    • 30. Selig ER, Bruno JF. A global analysis of the effectiveness of Marine Protected Areas in preventing coral loss. PLoS One. 2010;5(2):e9278. pmid:20174644
    • 31. Prugh LR, Stoner CJ, Epps CW, Bean WT, Ripple WJ, Laliberte AS, et al. The rise of the mesopredator. Bioscience. 2009;59(9):779–791.
    • 32. Crooks KR, Soule ME. Mesopredator release and avifaunal extinctions in a fragmented system. Nature. 1999;400(6744):563–566.
    • 33. Bonaldo RM, Hay ME. Seaweed-coral interactions: variance in seaweed allelopathy, coral susceptibility, and potential effects on coral resilience. PLoS One. 2014;9(1):e85786. pmid:24465707
    • 34. Rasher DB, Hoey AS, Hay ME. Consumer diversity interacts with prey defenses to drive ecosystem function. Ecology. 2013;94(6):1347–1358. pmid:23923498
    • 35. Clements C, Bonito V, Grober-Dunsmore R, Sobey M. Effects of small, Fijian community-based marine protected areas on exploited reef fishes. Mar Ecol Prog Ser. 2012;449:233–243.
    • 36. Goodwin BJ, Bender DJ, Contreras TA, Fahrig L, Wegner JF. Testing for habitat detection distances using orientation data. Oikos. 1999;84(1):160–163.
    • 37. Zollner PA, Lima SL. Orientational data and perceptual range: real mice aren't blind. Oikos. 1999;84(1):164–166.
    • 38. Ferlin V. The mode of dislocation of Astropecten aranciacus. Helgoland Wiss Meer. 1973;24(1–4):151–156.
    • 39. Scheibling RE. Optimal foraging movements of Oreaster reticulatus (L.) (Echinodermata: Asteroidea). J Ex Mar Biol Ecol. 1981;51(2–3):173–185.
    • 40. Pratchett MS, Schenk TJ, Baine M, Syms C, Baird AH. Selective coral mortality associated with outbreaks of Acanthaster planci L. in Bootless Bay, Papua New Guinea. Mar Environ Res. 2009;67(4–5):230–236. Epub 2009/03/31. pmid:19327821
    • 41. Dublin HT, Sinclair ARE, McGlade J. Elephants and fire as causes of multiple stable states in the Serengeti-Mara woodlands. J Anim Ecol. 1990;59(3):1147–1164.
    • 42. Landman M, Schoeman DS, Hall-Martin AJ, Kerley GIH. Long-term monitoring reveals differing impacts of elephants on elements of a canopy shrub community. Ecol Appl. 2014;24(8):2002–2012.
    • 43. Zann L, Brodie J, Berryman C, Naqasima M. Recruitment, ecology, growth, and behaviour of juvenile Acanthaster planci (L.) (Echinodermata: Asteroidea). Bull Mar Sci. 1987;41(2):561–575.
    • 44. Brashares JS, Prugh LR, Stoner CJ, Epps CW. Ecological and conservation implications of mesopredator release. Trophic cascades: predators, prey, and the changing dynamics of nature. 2010:221–240.
    • 45. Bonito V, Simpson R, Waqairagata F. Evaluating the performance of LMMAs in the districts of Korolevu-i-wai, Dawasamu, and Nakorotubu. NOAA Technical Report NA09NOS4630017. 2014. Available from: http://data.nodc.noaa.gov/coris/library/NOAA/CRCP/other/grants/International_FY09_Products/Final_Report_NA09NOS4630017.pdf
      • 46. Bonaldo RM, Guimarães PR, Pires MM, Hay ME. Even small marine protected areas provide refuge to reef fish assemblages, feeding groups, and coral recruits: evidence from Fijian coral reefs. PLoS One. In press.
        • 47. Cowan Z-L, Dworjanyn SA, Caballes CF, Pratchett MS. Predation on crown-of-thorns starfish larvae by damselfishes. Coral Reefs. 2016;35(4):1253–62.
        • 48. Dana TF, Newman WA, Fager EW. Acanthaster aggregations: interpreted as primarily responses to natural phenomena. Pacific Science. 1972;26(4):355–372.
        • 49. Kenyon JC, Aeby GS. Localized outbreak and feeding preferences of the crown-of-thorns seastar Acanthaster planci (echinodermata, asteroidea) on reefs off Oahu, Hawaii. Bulletin of Marine Science. 2009;84(2):199–209.
        • 50. Engelhardt U. Fine-scale surveys of crown-of-thorns starfish (Acanthaster planci) in the central Great Barrier Reef region. CRC Reef Research Centre technical report no. 30. Townsville, Australia: CRC Reef Research Centre. 1999.
          • 51. Clements CS, Hay ME. Competitors as accomplices: seaweed competitors hide corals from predatory sea stars. Proc R Soc B. 2015;282(1814):221–229.
          • 52. Birkeland C, Lucas JS. Acanthaster planci: major management problem of coral reefs. Boca Raton, Florida: CRC Press. 1990.
            • 53. Pratchett MS. Dynamics of an outbreak population of Acanthaster planci at Lizard Island, northern Great Barrier Reef (1995–1999). Coral Reefs. 2005;24(3):453–62.
            • 54. Anthony KRN, Marshall PA, Abdulla A, Beeden R, Bergh C, Black R, et al. Operationalizing resilience for adaptive coral reef management under global environmental change. Glob Change Biol. 2015;21(1):48–61.
            • 55. Moutardier G, Gereva S, Mills SC, Adjeroud M, Beldade R, Ham J, et al. Lime juice and vinegar injections as a cheap and natural alternative to control COTS outbreaks. PLoS One. 2015;10(9):e0137605. pmid:26356840