Marine reserve networks conserve biodiversity ... | Wiley Online Library

Marine reserve networks conserve biodiversity by stabilizing communities and maintaining food web structure STEPHEN R. WING

AND

LUCY JACK

Department of Marine Science, University of Otago, 310 Castle Street, Dunedin, New Zealand 9054 Citation: Wing, S. R., and L. Jack. 2013. Marine reserve networks conserve biodiversity by stabilizing communities and maintaining food web structure. Ecosphere 4(11):135. http://dx.doi.org/10.1890/ES13-00257.1

Abstract. Theory predicts that networks of fully protected marine reserves conserve biodiversity by stabilizing communities and maintaining food web structure in the face of inadequately constrained fishery exploitation. To test these ideas we examine trends in species incidence, community and trophic structure of temperate reef fishes over an eight year period within the Fiordland no-take marine reserve network, at management zones subject to commercial fishing and at those closed to commercial exploitation but open to recreational fishers. We use information from extensive stratified subtidal surveys of the reef fish community and abundance of macroalgae, as well as oceanographic data collected in 2002, 2006 and 2010. Our analyses indicate a regional decline in species richness of exploited reef fish in areas open to fishing between 2002 and 2010. Following implementation of spatial management (2006–2010), richness of ‘exploited’ species increased within marine reserves, but remained unchanged in areas open to fishing. Further, analysis of differences in community structure in this time period (2006–2010) indicate that both ‘exploited’ and ‘non-target’ groups were more stable within marine reserves than were those within fished areas. Consequentially average trophic level of the community remained stable within marine reserves but declined sharply in areas open to fishing, indicating both declines in large omnivorous species and increases in forage fish within exploited regions. These analyses offer an important test of the direct and indirect effects of marine reserve networks on the dynamics of reef fish communities at the landscape scale. We demonstrate the potential for multiple no-take reserves spread over a heterogeneous marine landscape to maintain biodiversity by stabilizing community structure and preserving intact food webs on a regional scale. Key words: conservation; diversity; fishing; marine reserve network; reef fish community; species richness; trophic level. Received 19 August 2013; accepted 16 September 2013; published 18 November 2013. Corresponding Editor: S. Cox. Copyright: Ó 2013 Wing and Jack. 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. http://creativecommons.org/licenses/by/3.0/ E-mail: [email protected]

INTRODUCTION

genic perturbation. Early theoretical studies predicted that restricted diet within a food web lowers community stability (MacArthur 1955). Indeed virtual food webs with large numbers of weak, ephemeral linkages are more stable, and thus more persistent, than those with few strong, requisite linkages (Polis 1994, McCann and Hastings 1997). Omnivory, where a consumer feeds at multiple trophic levels (Pim and Lawton 1978), results in a fractional trophic position of

Successful ecosystem management puts measures in place that guard against the erosion of ecological processes and promotes the maintenance of biodiversity. Ecologists have long been fascinated by the interaction between the complexity of food web architecture and community stability, where more stable communities have greater resistance to environmental or anthropov www.esajournals.org

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the consumer (Levine 1980), calculated as an average of the constituent trophic levels of the prey base. The degree of omnivory of a species may change during ontogeny (Abrams 2011), vary spatially within a population (Parsons and LeBrasseur 1970) or vary according to competitive interactions within linked food webs (Holt and Polis 1997). Food webs with a high degree of omnivory support fewer strong interactions and have a reduced likelihood of dramatic indirect effects such as trophic cascades (Polis 1994). Consequentially, complex communities with high trophic level omnivores can maintain more stable temporal and spatial dynamics (Polis and Strong 1996). These studies form the basis for asking a question of great importance for successful management of exploited systems: What happens to a community subject to selective removal of top trophic level omnivores? Omnivory is widespread in marine systems where large bodied, generalist predators have been heavily exploited by fishing, in many cases resulting in a simplification of food webs (Pauly et al. 1998) and in changes to the trophic dynamics of communities (Frank et al. 2011). Temperate reef fish communities contain a wide diversity of omnivorous species. Here selective removal of high trophic level omnivores has provided an explicit test of system stability in their absence. The effects of localized fishing mortality on community dynamics may be relatively elusory in open systems where subsidies and rescue effects from nearby refuge populations can dampen local exploitation-driven changes in spatial dynamics (Wing and Wing 2001, Kritzer and Sale 2006). However in more insular systems, the effects of exploitation may be locally enhanced (Roberts 1995). In theory, networks of no-take marine reserves may stabilize communities by providing a direct refuge to the exploited component of a community, typically omnivorous and top predatory fishes (Fogarty 1999). The present study investigated this idea by examining patterns in reef fish community and food web structure within a new marine reserve network established in Fiordland, southwest New Zealand. The network of marine protected areas in Fiordland offers a model system for testing the effects of selective removal of high trophic level omnivores in nature and for testing the ability of v www.esajournals.org

spatial closures to conserve or stimulate recovery of intact communities. The Fiordland Marine Area (Te Moana o Atawhenua) (FMA) is a region of globally significant natural and cultural heritage and of great economic importance to New Zealand. In recognition of this, two no-take marine reserves were established in 1993, in Doubtful and Milford Sounds. Furthermore, in 2005, the Fiordland Marine Management Act closed the inner regions of eleven fjords to commercial fishing (46,002 ha; 59% of the FMA) and established a network of eight new marine reserves nested within commercial exclusion zones, bringing the total no-take areas to 10, and covering 10,421 ha or 13.11% of the FMA (Fig. 1). To meet fisheries and biodiversity conservation objectives in this highly subdivided habitat (Lubchenco et al. 2003) efforts were taken to distribute fully protected marine reserves widely across the region among fjords and within representative habitat types (Wing and Jack 2010). All except one of the no-take marine reserves were nested in a buffer zone, where commercial fishing was closed and recreational fishing limits were reduced. Testing the ecological consequences of a protected area network is implicit to a successful adaptive management strategy whose goal is to optimize regional biodiversity patterns. Commonly employed impact-control studies measure spatial management outcomes by comparing regions that are under different management regimes (no-take reserve, commercial exclusion zones, open fished areas) but that are otherwise comparable, before and after implementation. However the New Zealand fjords are characterized by strong physical environmental gradients containing the extremes of salinity, wave action and irradiance that together influence patterns of pelagic productivity and density of habitatproviding macroalgae along the axis of each fjord (Goebel et al. 2005, Wing et al. 2007). The resulting gradient in composition of organic matter source pools can be directly linked to diet, sub-population structure, growth and fecundity of reef fish (Wing et al. 2012, Beer and Wing 2013). Thus comparable control sites in differing management zones are rare or nonexistent. To overcome this limitation, in the present study we make use of a time series from three 2


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Fig. 1. Map of the Fiordland Marine Area in southwestern New Zealand with marine reserves (dark gray), commercial exclusion zones (light gray) and open areas (white) indicated. Positions of long term monitoring sites surveyed in 2002, 2006 and 2010 are indicated by black circles. Additional sites surveyed only in 2006 and 2010 are indicated by open circles.

changes in diversity, community composition and food web structure are coincident with spatial management of fishing.

Fiordland-wide subtidal surveys conducted over an eight-year period to compare changes in the species composition and trophic structure of the fish community within regions under different management regimes. We use information on surface salinity and density of the habitatforming kelp Ecklonia radiata to partition out variability associated with the fjordic environmental gradients in our models. We focus on resolving both direct and indirect effects of fishing among management zones by dividing the reef fish community into exploited and nontarget components. In the resulting analysis we are able to explicitly test whether observed v www.esajournals.org

MATERIALS

AND

METHODS

Study sites The study was conducted using data collected in 2002, 2006 and 2010 at a core set of 21 longterm monitoring sites and in 2006 and 2010 at a further 9 monitoring sites in Fiordland, New Zealand (Fig. 1). The core data from 21 sites provided the opportunity to analyze variability in the temporal trajectories of community com3


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position in the different management zones, before and after implementation of the FMA. The larger data set of 30 sites provided greater statistical power to resolve changes in the fish community between 2006–2010 (after FMA implementation). All surveys were conducted during austral summer (2002: December; 2006 and 2010: January–February). We characterized the position of the sites along the major environmental gradients at each survey event using two summarizing metrics. 1. Surface salinity.—Water column structure varies from highly stratified at the fjord head to well mixed at the outer coast (Stanton 1984). We collected salinity readings every 0.5 sec from the surface to depth at each study site using a Seabird SBE-19 conductivity, temperature and depth profiler (CTD). Data were post-processed to 0.5 m bins using standard Seabird processing algorithms for the pumped SBE-19. Surface salinity at each site was calculated as the mean salinity at 0–2.5 m depth, averaged among years. 2. Kelp density.—Biogenic habitat type varies from fragile encrusting invertebrate communities that colonize rock walls of the inner fjord, to kelp forest on the open coast (Smith and Witman 1999, Wing and Jack 2012). To characterize the biogenic habitat at each site, the density of the dominant stipitate kelp Ecklonia radiata was measured during each survey event. Six replicate paired 2-m2 quadrats were haphazardly sampled within depth strata centered at 5, 10 and 15 m. Mean densities of E. radiata were calculated as the hierarchical mean of means, and standard errors were calculated by depth, survey and site.

Statistical analysis Differences in habitat variables among management zones.—To test whether management zones were distributed evenly across the Fiordland marine environment, we use data collected at 30 sites to construct ANOVAs (JMP Pro 10, SAS) to test for differences in the two habitat variables: (1) density of Ecklonia radiata and (2) surface salinity among management zones (fixed, three levels) and fjord (random, 12 levels). Including fjord in the model accounted for inherent variability among individual catchments and basins. Variance components were estimated using restricted maximum likelihood (REML). The results of this model (see Results) influenced our decision to include the effects of (1) Ecklonia radiata density (2) surface salinity in further models. Biodiversity and change in species richness.—To test for linear changes in biodiversity over time we used data collected at 21 sites to fit general linear models (JMP Pro 10, SAS) and test the relationship between species richness of exploited and non-target reef fish and (1) Ecklonia radiata density, (2) surface salinity, and (3) year (continuous; 2002, 2006, 2010), in each management zone type (no-take reserves, commercial exclusion zones and regions open to fishing). Secondly, to detect changes in species richness since FMA implementation, we used data collected at 30 sites but in only two years (providing greater statistical power but over a shorter timeframe), to fit ANOVAs (JMP Pro 10, SAS) to test the relationship between species richness of exploited and non-target reef fish by (1) management zone (fixed, three levels), (2) year (fixed, two levels) and (3) the interaction management zone 3 year. In each model Site (30 levels) was fitted as a random effect, and variance components were estimated using restricted maximum likelihood (REML). Community stability and change in community structure.—To test for changes in community structure since FMA implementation, using data collected at 30 sites we constructed resemblance matrices based on (1) presence–absence data and (2) relative abundance data using the BrayCurtis similarity index (Cherel et al. 2007). Relative abundance data were square-root transformed and a dummy variable of 0.001 was added before analysis. Analysis of presence–absence data alongside relative abundance

Reef fish community structure and composition We conducted scuba surveys of reef fish at 30 sites in 2006 and 2010 and at a subset of 21 core sites in 2002. A pair of divers counted conspicuous reef fish on a series of 5 m wide, 2.5 m high and 50 m long belt transects centered at 5, 10 and 15 m depth respectively (for a full description of the methods see Wing et al. 2006, Wing and Jack 2010). Data were averaged by site, which reduced replication at the lowest level (and therefore power) but decreased the noise associated with high levels of variation among transects. To investigate the direct and indirect effects of fishing, we separated the fish into exploited and non-target communities (Table 1). v www.esajournals.org

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WING AND JACK Table 1. Reef fish species in Fiordland arranged by frequency of occurrence. Exploited species are indicated by (e). Information on habitat, maximum total length (ML, cm) and diet is from Francis (2001); information on trophic level (TL, with SE in parentheses) is from Fishbase.org and Wing et al. (2012). Species

Common name

Habitat

ML

Diet

Notolabrus celidotus Caesioperca lepidoptera Pseudolabrus miles Notolabrus fucicola Parapercis colias Helicolenus percoides Notolabrus cinctus Hypoplectrodes huntii Aplodactylus arctidens Nemadactylus macropterus Mendosoma lineatum Scorpaena papillosus Parika scaber Lepidoperca tasmanica Squalus acanthias Callanthias allporti Latris lineata Odax pullus Latridopsis ciliaris Pseudophycis barbata Cephaloscyllium isabellum Latridopsis forsteri Parapercis gilliesi

spotty butterfly perch scarlet wrasse banded wrasse blue cod (e) sea perch (e) girdled wrasse redbanded perch marblefish tarakihi (e) telescopefish scorpionfish leatherjacket red lined perch spiny dogfish splendid perch trumpeter (e) greenbone (e) blue moki (e) bastard cod carpet shark copper moki (e) yellow weaver

estuarine - reef water column-reef deep reef wave-exposed reef gravel, shell debris deep reef reef associated benthic reef, caves wave-exposed reef sand-mud, reef water column-reef benthic reef reef associated deep reef reef associated water column-reef water column-reef kelp beds sand-mud, reef benthic reef, caves reef associated reef associated gravel, shell

24 30 27 38 45 47 30 20 65 70 40 30 31 20 160 30 120 40 80 63 100 65 37

benthic invertebrates planktivore benthic invertebrates benthic invertebrates piscivore, omnivore piscivore, omnivore benthic invertebrates benthic invertebrates herbivore omnivore, infauna planktivore benthic crustaceans sessile invertebrates planktivore piscivore, omnivore planktivore piscivore, omnivore herbivore infauna benthic invertebrates omnivore piscivore, omnivore benthic invertebrates

TL 3.3 3.1 3.6 3.5 4.9 4.0 3.6 4.1 2.0 3.4 3.1 4.0 3.0 3.1 4.3 3.4 3.8 2.1 3.2 3.5 4.2 3.7 3.5

(0.5) (0.2) (0.6) (0.6) (0.7) (0.7) (0.5) (0.7) (0.1) (0.4) (0.6) (0.6) (0.2) (0.2)à (0.7) (0.5) (0.6) (0.1) (0.5) (0.5) (0.6) (0.6) (0.6)

Deep water emergent species. à TL estimated from closest feeding guild member.

data allowed us to compare the effects of species loss with those from changes in relative abundance. We used permutational multivariate analysis of variance (PERMANOVAþ, Primer-e version 6) to test for changes in structure of the exploited and the non-target reef fish community from 2006 to 2010. We used sequentially fitted terms to first partition variance associated with (1) Ecklonia radiata density (2) surface salinity and then to assess the importance of the remaining variability due to (3) year (fixed, two levels), within each management zone type. We use p rincipal coordinates a nalysis (PERMANOVAþ, Primer-e version 6) to calculate the distance between the centroids for each pair of years (2006 and 2010) in each management zone. This distance is in essence the difference between the mean positions of the community in multivariate space in each year and so is a metric for change in community structure (Anderson et al. 2008). Ecosystem functioning and change in trophic level.—Estimates of fish trophic level were sourced from Fishbase.org and from Wing et al. (2012) for Parapercis colias (Table 1). These estimates were then used to calculate the v www.esajournals.org

average trophic level of the entire fish community at each site (hereafter ‘‘trophic level’’). To test for linear changes in trophic level since 2002, we used data collected at 21 sites to construct general linear models (JMP Pro 10, SAS) to test for relationships between trophic level and (1) density of Ecklonia radiata, (2) surface salinity, and (3) year (continuous, 2002, 2006, 2010), within each management zone type (no-take reserves, commercial exclusion zones and regions open to fishing). To assess changes in trophic level between 2006 and 2010, we used data collected at 30 sites, to construct an ANOVA to test for differences in trophic level of the reef fish community among (1) management zones (fixed, three levels), (2) years (fixed, two levels) and (3) the interaction management zone 3 year. In each model site (30 levels) was included as a random effect, and variance components were estimated using restricted maximum likelihood (REML). Additionally, to assess in which portion of the fish community the greatest changes occurred, we compared plots of the cumulative density of reef fish by trophic level in each management zone, for both the exploited and non-target reef 5


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Fig. 2. Average salinity (grey bars) and average density of Ecklonia radiata (black bars) among management zones. Letters not connected by same letter within each variable are significantly different as indicated by a Tukey post hoc test. Error bars indicate 61 SE.

fish communities.

Between 2006 and 2010, species richness of exploited reef fish increased in marine reserves but remained constant in commercial exclusion zones and open fished areas (r 2 ¼ 0.38, RMSE ¼ 0.92, p , 0.0001; Fig. 4a). Non-target reef fish.—We did not detect any changes in species richness between 2002 and 2010 or 2006 and 2010 in the three management zones ( p . 0.05, Fig. 3b; p . 0.05, Fig. 4b).

RESULTS Habitat variables Open fished areas incorporated higher density kelp forest (r 2 ¼ 0.71, RMSE ¼ 0.89, p ¼ 0.0001; Fig. 2) and higher surface salinity conditions (r 2 ¼ 0.85, RMSE ¼ 2.97, p ¼ 0.0001; Fig. 2) than areas protected by commercial exclusion zones or marine reserves, which incorporated more estuarine habitats.

Community stability and change in community structure The changes observed in the structure of the reef fish community between 2006 and 2010 were coincident with the level of exploitation from fishing whereby communities in fished regions were more dissimilar between time periods than were those within marine reserves. Exploited reef fish.—In commercial exclusion zones, we detected a change in the community structure of the exploited reef fish between 2006

Biodiversity and change in species richness Exploited reef fish.—We did not detect linear changes in the species richness of exploited reef fish between 2002–2010 in marine reserves or commercial exclusion zones ( p . 0.05, Fig. 3a). In areas open to commercial fishing, species richness of exploited reef fish declined between 2002–2010 (F2,5 ¼ 6.43, p ¼ 0.05; Fig. 3a). v www.esajournals.org

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Fig. 3. Time series from 2002, 2006 and 2010 surveys of average species richness among sites for (a) seven exploited species and (b) 17 non-target species from marine reserves (open circles), commercial exclusion zones (closed triangles), and areas open to fishing (closed circles). Time of implementation of the Fiordland Marine Management Act 2005 (April 20, 2005) is indicated by a vertical line. Error bars indicate 61 SE.

and 2010 that was mainly due to changes in species relative abundance rather than species gain or loss (Bray Curtis on relative abundance, pseudo F ¼ 3.18, p ¼ 0.01, Fig. 5a–c; Bray Curtis on presence-absence, pseudo F ¼ 2.64, p ¼ 0.08). v www.esajournals.org

These metrics also changed considerably in areas open to commercial fishing (Fig. 5a–c), although this change was not statistically significant as variability among sites was high (Bray Curtis on relative abundance; pseudo F ¼ 2.29, p ¼ 0.12; 7


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abundance, pseudo F ¼ 0.98, p ¼ 0.40; Bray Curtis on presence–absence, pseudo F ¼ 0.41, p ¼ 0.69; Fig. 5a, c). Non-target reef fish.—Non-target fish community structure remained most stable in marine reserves (Bray Curtis on relative abundance, pseudo F ¼ 0.50, p ¼ 0.73; Bray Curtis on presence–absence, pseudo F ¼ 0.26, p ¼ 0.85; Fig. 5b–d), followed by in commercial exclusion zones, where the relative abundance of fishes changed but species incidence did not (Bray Curtis on relative abundance, pseudo F ¼ 1.5, p ¼ 0.16; Bray Curtis on presence–absence, pseudo F ¼ 2.84, p ¼ 0.04). In open fished areas, the structure of the non-target fish community changed between 2006 and 2010 mostly due to species loss (Bray Curtis on relative abundance, pseudo F ¼ 5.53, p ¼ 0.03; Bray Curtis on presence–absence, pseudo F ¼ 5.59, p ¼ 0.01; Fig. 5b–d).

Food web structure and change in trophic level We observed linear declines in trophic level between 2002 and 2010 in commercial exclusion zones (F3,29 ¼ 6.26, p ¼ 0.02; Fig. 6), but not in open areas or marine reserves ( p . 0.05; Fig. 6). Trophic level was lower in 2010 than in 2006 at sites in commercial exclusion zones and in open fished regions but remained stable at sites in marine reserves (r 2 ¼ 0.60, RMSE ¼ 0.102, p , 0.0001; Fig. 7). We detected marked shifts in the trophic structure of the fish community between 2006 and 2010 that differed among management zones (Fig. 8). In marine reserves, the density of higher trophic level (4–5) species increased and to a lesser extent the density of lower trophic level (3–4) non-target forage fish also increased (Fig. 8a, d). In commercial exclusion zones and open fished regions, higher trophic level (4–5) exploited species decreased and a lower trophic level (3–4) forage fish greatly increased (Fig. 8b, e). These changes were most pronounced in open fished areas (Fig. 8c, f ), where higher trophic level (4 –5) exploited species decreased and lower trophic level (3–4) forage fish greatly increased, so that the amount of change in both the exploited and non-target portions of the community was relative to the level of exploitation.

Fig. 4. Average species richness among sites from 2006 (gray) and 2010 (black) surveys for (a) seven exploited species and (b) 17 non-target species from marine reserves, commercial exclusion zones, and areas open to fishing. Letters not connected by same letter within each variable are significantly different as indicated by a Student t post hoc test. Error bars indicate 61 SE.

Bray Curtis on presence-absence, pseudo F ¼ 1.65, p ¼ 0.21). The structure of the exploited fish community remained most stable at sites within marine reserves, as shown by the least distance between centroids (Bray Curtis on relative v www.esajournals.org

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Fig. 5. Community disimilarity between 2006 and 2010 at 30 sites as indicated by multivariate centroid distances calculated using the Bray-Curtis index based on square root of relative abundances for (a) exploited species and (b) non-target species, and for species incidence (presence-absence) of (c) exploited species and (d) non-target species. Significant differences between 2006 and 2010 from a PERMANOVA on communities in marine reserves, commercial exclusion zones and areas open to fishing are indicated by an asterisk.

8-year time frame indicate that fishing pressure corresponded with a general trend toward greater heterogeneity in occurrence across the landscape. In contrast, within areas designated as no-take reserves, this trend was reversed during the four-year time period after the implementation of the marine reserve network. These patterns indicate that a network of spatial closures can combat the erosion of biodiversity and rebuild a more homogeneous, species-rich community on a regional scale over a relatively short time frame. Our examination of changes in community structure expanded upon these results. The observed patterns of change between 2006 and 2010 indicated an important difference in the stability of communities within different management zones. In this analysis we use multi-

DISCUSSION The data and analyses presented here provided an important replicated test of the effects of fishing mortality on reef fish community dynamics. Our results are consistent with the idea that networks of marine protected areas act to stabilize reef fish communities and to preserve intact food webs by providing a direct refuge for the large, high trophic level omnivores that make up the exploited component of the community. In a distinct bioregion such as Fiordland, increases in site level (alpha) species richness indicate that sites, on average, contain a greater compliment of the regional species pool and that communities have increased in homogeneity across the region. Declines in local richness of exploited species in areas open to fishing over an v www.esajournals.org

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Fig. 6. Time series from 2002, 2006 and 2010 surveys of average trophic level among sites from marine reserves (open circles), commercial exclusion zones (closed triangles), and areas open to fishing (closed circles). Error bars indicate 61 SE.

Fig. 7. Average trophic level within marine reserves, commercial exclusion zones and open regions (data from Fishbase.org and Wing et al. 2012). Gray bars indicate trophic level in 2006 and black bars indicate trophic level in 2010. Levels not connected by same letter within each variable are significantly different as indicated by a Student t post hoc test. Error bars indicate 61 SE.

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Fig. 8. Cumulative density for surveys in 2006 (open circles) and 2010 (black circles) versus trophic level for exploited reef fish in (a) marine reserves, (b) commercial exclusion zones, (c) open regions, and for non-target reef fish in (d) marine reserves, (e) commercial exclusion zones, and (f ) open regions. Error bars indicate 61 SE.

variate centroid distances to measure the relative similarity of communities as they change through time. Because we considered an orthogonal data set with equal sample sizes among years and timed during austral summer months, distributed across replicate management zones, and containing a full range of habitat types, the effects we observe are likely associated with management. We observed that within no take marine reserves, the structure of the exploited reef fish community was more stable between 2006 and 2010, than in areas open to commercial or recreational exploitation. In the same time period, areas subject to recreational fishing but not commercial fishing changed an intermediate amount whilst communities in commercially exploited areas changed the most. These results are consistent with the prediction that the cessation of fishing mortality in marine reserves would result in v www.esajournals.org

stability in the exploited reef fish community (Micheli et al. 2004, Lester et al. 2009). These communities are predominately made up of large, high trophic level, omnivorous species such as blue cod (Parapercis colias), sea perch (Helicolenus percoides), tarakihi (Nemadactylus macropterus) and trumpeter (Latris lineata). A significant portion of the diet in these species comprises smaller reef fish species such as wrasses (Notolabrus celidotus, Pseudolabrus miles, Notolabrus fucicola) and smaller schooling planktivorous species such as butterfly perch (Caesioperca lepidoptera) and telescopefish (Mendosoma lineatum) (Beer 2011, Wing et al. 2012). Accordingly, we present evidence that relative changes in the composition of the exploited reef fish community are coupled with changes in the composition of the non-target community. Patterns of stability in non-target reef fish community structure were also relative to the level of 11


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protection afforded by fisheries closures. Finally, we present evidence that the marine reserve network was able to conserve food web structure across the region. We observed that within no take areas, higher trophic level omnivores were maintained at greater densities. In areas open to commercial and recreational exploitation, the structure of the community had changed since implementation of the FMA. Fewer high trophic level omnivores and greater numbers of lower trophic level forage fish now make up the community in these regions. This pattern is consistent with incidental effects of fishing mortality on non-target species or by indirect effects of removing large omnivores from fished regions. In this case we see a greater prevalence of small forage fish in the commercially fished regions, consistent with a release from predation pressure. We use several statistical tools to elucidate these trends, each with important considerations. First we make use of general linear models to resolve trends over an eight-year time series that includes a period of three years before implementation of spatial management. Here a sample size of 21 sites gave us moderate resolution of trends over time. Nevertheless some important patterns emerge among management zones and species groupings (Figs. 3 and 6). We then increase our statistical resolution of the period after implementation of the marine reserve network (2006–2010) by increasing sample size by 30%, including a more balanced representation of management zones. Here we are able to resolve differences in species richness and trophic level both among and between management zones (Figs. 4 and 7). We extended this analysis using permutational multivariate analysis of variance on community composition to resolve differences in the change of whole communities between 2006 and 2010 (Fig. 5). Our result that communities within marine reserves were more stable over this time period provides evidence of success for an important management objective, to increase community stability and persistence. Our analysis of the density of Ecklonia radiata and surface salinity underscore important differences in these indicators of habitat quality among management zones. We demonstrate that commercial exclusion zones and marine reserves are similar in terms of habitat quality, while open v www.esajournals.org

regions support more abundant kelp and oceanic salinity regimes (Fig. 2). The consequence is some loss of balance in direct comparisons among zones, and highlights the importance of including site and habitat variables in our statistical models. We also note that areas with higher kelp density and more oceanic salinity harbor higher species richness of reef fish, which underscores the importance of examining temporal changes within management zones in the present analysis. Further, this result highlights the underrepresentation of marine reserves or commercial exclusion zones in the outer coastal habitats in Fiordland. We conclude that in areas with high fishing mortality, community dynamics and associated trophic interactions changed more between 2006 and 2010 than within marine reserves, which had more persistent populations of large piscivorous omnivores. Our results are consistent with this idea, however several caveats accompany this conclusion. Our data are based on natural populations subject to environmental forcing of recruitment, survival and migration. We expect that there would be some coupling of these processes and associated dynamics among species. For example, conditions favoring recruitment might be common across species and our observation that communities of both exploited and non-target reef fish are more dynamic in fished areas could be explained by extrinsic environmental forcing on both groups rather than the effects of management. We cannot completely reject this alternative explanation. Nevertheless because we considered an orthogonal data set with replication of multiple management zones across a large region, and because we have statistically accounted for much of the variability associated with site and environmental differences among sites, the most parsimonious explanation for the observed patterns is that they are a result of spatial management. A second caveat is that observed changes in the non-target reef fish community could be partially or wholly driven by incidental effects of fishing. Several of the non-target species suffer incidental mortality, which might have direct effects on community composition. However the largest observed changes in this community occurred in small, lower trophic level forage fish, which are generally not subject to direct fishing 12


WING AND JACK University of Otago Research Committee and from the Royal Society of New Zealand’s Marsden Fund (UOO038).

pressure. The patterns observed here have important implications for the application of spatial management to temperate reef fish communities. Over a relatively short period of four years we observed differences in community dynamics within regions subject to fishing versus those under two levels of protection: no-take marine reserves and commercial exclusion. Importantly, we observe the strongest positive effects of management in fully no-take reserves and very little benefit of partial reserves or effort-based management. Though many investigations have focussed on local changes in biomass and age structure associated with single no-take marine reserves (Halpern 2003, Lester et al. 2009), relatively few have resolved trajectories of change for whole communities, or shifts in food web structure within networks of reserves (e.g., Micheli et al. 2004, Babcock et al. 2010). The Fiordland marine area provides an important opportunity to resolve patterns in community dynamics across a replicated system. Based on the evidence presented here, we suggest that networks of fully protected areas are a powerful tool, capable of conserving biodiversity, and maintaining stable and fully functional marine communities across a landscape. Our evidence also suggests that even in remote areas such as the New Zealand fjords, these benefits may not be reaped purely by exclusion of commercial activities if the number of recreational fishers is left unconstrained. The observed stabilizing effects of marine protection on the dynamics of whole communities and maintenance of food web structure are consistent with theoretical predictions of the stabilizing effect of top predators and omnivores on community structure. These results are an important example of the value of spatial networks of no-take marine reserves for the regional maintenance of intact and fully functional ecosystems at the landscape scale.

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