Marine Reserve Networks Conserve Biodiversity by Stabilizing Communities and Maintaining Food Web Structure

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(10):XXX. http://dx.doi.org/10.1890/ES13-00257.1

Abstract. Theory predicts that networks of fully protected marine reserves conserve biodiversity by stabilising 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 00 Month 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 Successful ecosystem management puts mea- lowers community stability (MacArthur 1955). sures in place that guard against the erosion of Indeed virtual food webs with large numbers of ecological processes and promotes the mainte- weak, ephemeral linkages are more stable, and nance of biodiversity. Ecologists have long been thus more persistent, than those with few strong, fascinated by the interaction between the com- requisite linkages (Polis 1994, McCann and plexity of food web architecture and community Hastings 1997). Omnivory, where a consumer stability, where more stable communities have feeds at multiple trophic levels (Pim and Lawton greater resistance to environmental or anthropo- 1978), results in a fractional trophic position of

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the consumer (Levine 1980), calculated as an spatial closures to conserve or stimulate recovery average of the constituent trophic levels of the of intact communities. The Fiordland Marine prey base. The degree of omnivory of a species Area (Te Moana o Atawhenua) (FMA) is a region may change during ontogeny (Abrams 2011), of globally significant natural and cultural vary spatially within a population (Parsons and heritage and of great economic importance to LeBrasseur 1970) or vary according to competi- New Zealand. In recognition of this, two no-take tive interactions within linked food webs (Holt marine reserves were established in 1993, in and Polis 1997). Food webs with a high degree of Doubtful and Milford Sounds. Furthermore, in omnivory support fewer strong interactions and 2005, the Fiordland Marine Management Act have a reduced likelihood of dramatic indirect closed the inner regions of eleven fjords to effects such as trophic cascades (Polis 1994). commercial fishing (46,002 ha; 59% of the FMA) Consequentially, complex communities with high and established a network of eight new marine trophic level omnivores can maintain more stable reserves nested within commercial exclusion temporal and spatial dynamics (Polis and Strong zones, bringing the total no-take areas to 10, 1996). These studies form the basis for asking a and covering 10,421 ha or 13.11% of the FMA question of great importance for successful (Fig. 1). To meet fisheries and biodiversity management of exploited systems: What hap- conservation objectives in this highly subdivided pens to a community subject to selective removal habitat (Lubchenco et al. 2003) efforts were taken of top trophic level omnivores? to distribute fully protected marine reserves Omnivory is widespread in marine systems widely across the region among fjords and where large bodied, generalist predators have within representative habitat types (Wing and been heavily exploited by fishing, in many cases Jack 2010). All except one of the no-take marine resulting in a simplification of food webs (Pauly reserves were nested in a buffer zone, where et al. 1998) and in changes to the trophic commercial fishing was closed and recreational dynamics of communities (Frank et al. 2011). fishing limits were reduced. Temperate reef fish communities contain a wide Testing the ecological consequences of a diversity of omnivorous species. Here selective protected area network is implicit to a successful removal of high trophic level omnivores has adaptive management strategy whose goal is to provided an explicit test of system stability in optimize regional biodiversity patterns. Com- their absence. The effects of localized fishing monly employed impact-control studies measure mortality on community dynamics may be spatial management outcomes by comparing relatively elusory in open systems where subsi- regions that are under different management dies and rescue effects from nearby refuge regimes (no-take reserve, commercial exclusion populations can dampen local exploitation-driv- zones, open fished areas) but that are otherwise en changes in spatial dynamics (Wing and Wing comparable, before and after implementation. 2001, Kritzer and Sale 2006). However in more However the New Zealand fjords are character- insular systems, the effects of exploitation may be ized by strong physical environmental gradients locally enhanced (Roberts 1995). In theory, containing the extremes of salinity, wave action networks of no-take marine reserves may stabi- and irradiance that together influence patterns of lize communities by providing a direct refuge to pelagic productivity and density of habitat- the exploited component of a community, typi- providing macroalgae along the axis of each cally omnivorous and top predatory fishes fjord (Goebel et al. 2005, Wing et al. 2007). The (Fogarty 1999). The present study investigated resulting gradient in composition of organic this idea by examining patterns in reef fish matter source pools can be directly linked to community and food web structure within a diet, sub-population structure, growth and fe- new marine reserve network established in cundity of reef fish (Wing et al. 2012, Beer and Fiordland, southwest New Zealand. Wing 2013). Thus comparable control sites in The network of marine protected areas in differing management zones are rare or nonex- Fiordland offers a model system for testing the istent. effects of selective removal of high trophic level To overcome this limitation, in the present omnivores in nature and for testing the ability of study we make use of a time series from three

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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.

Fiordland-wide subtidal surveys conducted over changes in diversity, community composition an eight-year period to compare changes in the and food web structure are coincident with species composition and trophic structure of the spatial management of fishing. fish community within regions under different management regimes. We use information on MATERIALS AND METHODS surface salinity and density of the habitat- forming kelp Ecklonia radiata to partition out Study sites variability associated with the fjordic environ- The study was conducted using data collected mental gradients in our models. We focus on in 2002, 2006 and 2010 at a core set of 21 long- resolving both direct and indirect effects of term monitoring sites and in 2006 and 2010 at 30 fishing among management zones by dividing monitoring sites in Fiordland, New Zealand (Fig. the reef fish community into exploited and non- 1). The core data from 21 sites provided the target components. In the resulting analysis we opportunity to analyze variability in the tempo- areabletoexplicitlytestwhetherobserved ral trajectories of community composition in the

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different management zones, before and after Statistical analysis implementation of the FMA. The larger data set Differences in habitat variables among management of 30 sites provided greater statistical power to zones.—To test whether management zones were resolve changes in the fish community between distributed evenly across the Fiordland marine 2006–2010 (after FMA implementation). All environment, we use data collected at 30 sites to surveys were conducted during austral summer construct ANOVAs (JMP Pro 10, SAS) to test for (2002: December; 2006 and 2010: January–Febru- differences in the two habitat variables: (1) ary). We characterized the position of the sites density of Ecklonia radiata and (2) surface along the major environmental gradients at each salinity among management zones (fixed, three survey event using two summarizing metrics. levels) and fjord (random, 12 levels). Including 1. Surface salinity.—Water column structure fjord in the model accounted for inherent varies from highly stratified at the fjord head to variability among individual catchments and well mixed at the outer coast (Stanton 1984). To basins. Variance components were estimated characterize this at each site, we collected salinity using restricted maximum likelihood (REML). readings every 0.5 sec from the surface to depth The results of this model (see Results) influenced at each study site using a Seabird SBE-19 our decision to include the effects of (1) Ecklonia conductivity, temperature and depth profiler radiata density (2) surface salinity in further (CTD). Data were post-processed to 0.5 m bins models. using standard Seabird processing algorithms for Biodiversity and change in species richness.—To the pumped SBE-19. Surface salinity at each site test for linear changes in biodiversity over time was calculated as the mean salinity at 0–2.5 m we used data collected at 21 sites to fit general depth, averaged among years. linear models (JMP Pro 10, SAS) and test the 2. Kelp density.—Biogenic habitat type varies relationship between species richness of exploit- from fragile encrusting invertebrate communities ed and non-target reef fish and (1) Ecklonia radiata that colonize rock walls of the inner fjord, to kelp density, (2) surface salinity, and (3) year (contin- forest on the open coast (Smith and Witman 1999, uous; 2002, 2006, 2010), in each management Wing and Jack 2012). To characterize the biogenic zone type (no-take reserves, commercial exclu- habitat at each site, the density of the dominant sion zones and regions open to fishing). Second- stipitate kelp Ecklonia radiata was measured ly, to detect changes in species richness since during each survey event. Six replicate paired FMA implementation, we used data collected at 2-m2 quadrats were haphazardly sampled within 30 sites but in only two years (providing greater depth strata centred at 5, 10 and 15 m. Mean statistical power but over a shorter timeframe), to densities of E. radiata were calculated as the fit ANOVAs (JMP Pro 10, SAS) to test the hierarchical mean of means, and standard errors relationship between species richness of exploit- were calculated by depth, survey and site. ed and non-target reef fish by (1) management zone (fixed, three levels), (2) year (fixed, two Reef fish community structure and composition levels) and (3) the interaction management zone We conducted scuba surveys of reef fish at 30 3 year. In each model Site (30 levels) was fitted as sites in 2006 and 2010 and at a subset of 21 core a random effect, and variance components were sites in 2002. A pair of divers counted conspic- estimated using restricted maximum likelihood uous reef fish on a series of 5 m wide, 2.5 m high (REML). and 50 m long belt transects centered at 5, 10 and Community stability and change in community 15 m depth respectively (for a full description of structure.—To test for changes in community the methods see Wing et al. 2006, Wing and Jack structure since FMA implementation, using data 2010). Data were averaged by site, which collected at 30 sites we constructed resemblance reduced replication at the lowest level (and matrices based on (1) presence–absence data therefore power) but decreased the noise associ- and (2) relative abundance data using the Bray- ated with high levels of variation among tran- Curtis similarity index (Cherel et al. 2007). sects. To investigate the direct and indirect effects Relativeabundancedataweresquare-root of fishing, we separated the fish into exploited transformed and a dummy variable of 0.001 and non-target communities (Table 1). was added before analysis. Analysis of pres-

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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 TL Notolabrus celidotus spotty estuarine - reef 24 benthic invertebrates 3.3 (0.5) Caesioperca lepidoptera butterfly perch water column-reef 30 planktivore 3.1 (0.2) Pseudolabrus miles scarlet wrasse deep reef 27 benthic invertebrates 3.6 (0.6) Notolabrus fucicola banded wrasse wave-exposed reef 38 benthic invertebrates 3.5 (0.6) Parapercis colias blue cod (e) gravel, shell debris 45 piscivore, omnivore 4.9 (0.7) Helicolenus percoides sea perch (e) deep reef 47 piscivore, omnivore 4.0 (0.7) Notolabrus cinctus girdled wrasse reef associated 30 benthic invertebrates 3.6 (0.5) Hypoplectrodes huntii redbanded perch benthic reef, caves 20 benthic invertebrates 4.1 (0.7) Aplodactylus arctidens marblefish wave-exposed reef 65 herbivore 2.0 (0.1) Nemadactylus macropterus tarakihi (e) sand-mud, reef 70 omnivore, infauna 3.4 (0.4) Mendosoma lineatum telescopefish water column-reef 40 planktivore 3.1 (0.6) Scorpaena papillosus scorpionfish benthic reef 30 benthic crustaceans 4.0 (0.6) Parika scaber leatherjacket reef associated 31 sessile invertebrates 3.0 (0.2) Lepidoperca tasmanica red lined perch deep reef 20 planktivore 3.1 (0.2)à Squalus acanthias spiny dogfish reef associated 160 piscivore, omnivore 4.3 (0.7) Callanthias allporti splendid perch water column-reef 30 planktivore 3.4 (0.5) Latris lineata trumpeter (e) water column-reef 120 piscivore, omnivore 3.8 (0.6) Odax pullus greenbone (e) kelp beds 40 herbivore 2.1 (0.1) Latridopsis ciliaris blue moki (e) sand-mud, reef 80 infauna 3.2 (0.5) Pseudophycis barbata bastard cod benthic reef, caves 63 benthic invertebrates 3.5 (0.5) Cephaloscyllium isabellum carpet shark reef associated 100 omnivore 4.2 (0.6) Latridopsis forsteri copper moki (e) reef associated 65 piscivore, omnivore 3.7 (0.6) Parapercis gilliesi yellow weaver gravel, shell 37 benthic invertebrates 3.5 (0.6) Deep water emergent species. à TL estimated from closest feeding guild member.

ence–absence data alongside relative abundance estimates were then used to calculate the data allowed us to compare the effects of species average trophic level of the entire fish commu- loss with those from changes in relative abun- nity at each site (hereafter ‘‘trophic level’’). To dance. We used permutational multivariate test for linear changes in trophic level since analysis of variance (PERMANOVAþ,Primer-e 2002, we used data collected at 21 sites to version 6) to test for changes in structure of the construct general linear models (JMP Pro 10, exploited and the non-target reef fish commu- SAS) to test for relationships between trophic nity from 2006 to 2010. We used sequentially level and (1) density of Ecklonia radiata,(2) fitted terms to first partition variance associated surface salinity, and (3) year (continuous, 2002, with (1) Ecklonia radiata density (2) surface 2006, 2010), within each management zone salinity and then to assess the importance of type (no-take reserves, commercial exclusion the remaining variability due to (3) year (fixed, zones and regions open to fishing). To assess two levels), within each management zone type. changes in trophic level between 2006 and We use principal coordinates analysis 2010, we used data collected at 30 sites, to (PERMANOVAþ,Primer-e version 6) to calcu- construct an ANOVA to test for differences in late the distance between the centroids for each trophic level of the reef fish community among pair of years (2006 and 2010) in each manage- (1) management zones (fixed, three levels), (2) ment zone. This distance is in essence the years (fixed, two levels) and (3) the interaction difference between the mean positions of the management zone 3 year. In each model site community in multivariate space in each year (30 levels) was included as a random effect, and so is a metric for change in community and variance components were estimated using structure (Anderson et al. 2008). restricted maximum likelihood (REML). Addi- Ecosystem functioning and change in trophic tionally,toassessinwhichportionofthefish level.—Estimates of fish trophic level were community the greatest changes occurred, we sourced from Fishbase.org and from Wing et compared plots of the cumulative density of al. (2012) for Parapercis colias (Table 1). These reef fish by trophic level in each management

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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.

zone, for both the exploited and non-target reef 2002–2010 (F2,5 ¼ 6.43, p ¼ 0.05; Fig. 3a). fish communities. Between 2006 and 2010, species richness of exploited reef fish increased in marine reserves RESULTS but remained constant in commercial exclusion zones and open fished areas (r2 ¼ 0.38, RMSE ¼ Habitat variables 0.92, p , 0.0001; Fig. 4a). Open fished areas incorporated higher density Non-target reef fish.—We did not detect any 2 kelp forest (r ¼ 0.71, RMSE ¼ 0.89, p ¼ 0.0001; changes in species richness between 2002 and 2 Fig. 2) and higher surface salinity conditions (r ¼ 2010 or 2006 and 2010 in the three management 0.85, RMSE ¼ 2.97, p ¼ 0.0001; Fig. 2) than areas zones ( p . 0.05, Fig. 3b; p . 0.05, Fig. 4b). protected by commercial exclusion zones or marine reserves, which incorporated more estu- Community stability and change in arine habitats. community structure The changes observed in the structure of the Biodiversity and change in species richness reef fish community between 2006 and 2010 were Exploited reef fish.—We did not detect linear coincident with the level of exploitation from changes in the species richness of exploited reef fishing whereby communities in fished regions fish between 2002–2010 in marine reserves or were more dissimilar between time periods than commercial exclusion zones ( p . 0.05, Fig. 3a). In were those within marine reserves. areas open to commercial fishing, species rich- Exploited reef fish.—In commercial exclusion ness of exploited reef fish declined between zones, we detected a change in the community

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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.

structure of the exploited reef fish between 2006 on presence-absence, pseudo F ¼ 2.64, p ¼ 0.08). and 2010 that was mainly due to changes in These metrics also changed considerably in areas species relative abundance rather than species open to commercial fishing (Fig. 5a–c), although gain or loss (Bray Curtis on relative abundance, this change was not statistically significant as pseudo F ¼ 3.18, p ¼ 0.01, Fig. 5a–c; Bray Curtis variability among sites was high (Bray Curtis on

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between centroids (Bray Curtis on relative 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 (r2 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 Fig. 4. Average species richness among sites from zones (Fig. 8). In marine reserves, the density 2006 (gray) and 2010 (black) surveys for (a) seven of higher trophic level (4–5) species increased exploited species and (b) 17 non-target species from and to a lesser extent the density of lower marine reserves, commercial exclusion zones, and trophic level (3–4) non-target forage fish also areas open to fishing. Letters not connected by same increased (Fig. 8a, d). In commercial exclusion letter within each variable are significantly different as zones and open fished regions, higher trophic indicated by a Student t post hoc test. Error bars level (4–5) exploited species decreased and a indicate 61 SE. lower trophic level (3–4) forage fish greatly increased (Fig. 8b, e). These changes were most pronounced in open fished areas (Fig. 8c, f ), relative abundance; pseudo F ¼ 2.29, p ¼ 0.12; where higher trophic level (4–5) exploited Bray Curtis on presence-absence, pseudo F ¼ species decreased and lower trophic level (3–4) forage fish greatly increased, so that the amount 1.65, p ¼ 0.21). The structure of the exploited fish of change in both the exploited and non-target community remained most stable at sites within portions of the community was relative to the marine reserves, as shown by the least distance level of exploitation.

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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.

DISCUSSION 8-year time frame indicate that fishing pressure corresponded with a general trend toward The data and analyses presented here provid- greater heterogeneity in occurrence across the ed an important replicated test of the effects of landscape. In contrast, within areas designated as fishing mortality on reef fish community dynam- no-take reserves, this trend was reversed during ics. Our results are consistent with the idea that the four-year time period after the implementa- networks of marine protected areas act to tion of the marine reserve network. These stabilize reef fish communities and to preserve patterns indicate that a network of spatial intact food webs by providing a direct refuge for closures can combat the erosion of biodiversity the large, high trophic level omnivores that make and rebuild a more homogeneous, species-rich up the exploited component of the community. community on a regional scale over a relatively In a distinct bioregion such as Fiordland, short time frame. increases in site level (alpha) species richness Our examination of changes in community indicate that sites, on average, contain a greater structure expanded upon these results. The compliment of the regional species pool and that observed patterns of change between 2006 and communities have increased in homogeneity 2010 indicated an important difference in the across the region. Declines in local richness of stability of communities within different man- exploited species in areas open to fishing over an agement zones. In this analysis we use multi-

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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 rela- stability in the exploited reef fish community tive similarity of communities as they change (Micheli et al. 2004, Lester et al. 2009). These throughtime.Becauseweconsideredanor- communities are predominately made up of thogonal data set with equal sample sizes large, high trophic level, omnivorous species among years and timed during austral summer such as blue cod (Parapercis colias), sea perch months, distributed across replicate manage- (Helicolenus percoides), tarakihi (Nemadactylus ment zones, and containing a full range of macropterus) and trumpeter (Latris lineata). A habitat types, the effects we observe are likely significant portion of the diet in these species associated with management. We observed that comprises smaller reef fish species such as within no take marine reserves, the structure of wrasses (Notolabrus celidotus, Pseudolabrus miles, the exploited reef fish community was more Notolabrus fucicola) and smaller schooling plank- stable between 2006 and 2010, than in areas tivorous species such as butterfly perch (Caesio- open to commercial or recreational exploitation. perca lepidoptera)andtelescopefish(Mendosoma Inthesametimeperiod,areassubjectto lineatum) (Beer 2011, Wing et al. 2012). Accord- recreational fishing but not commercial fishing ingly, we present evidence that relative changes changed an intermediate amount whilst com- in the composition of the exploited reef fish munities in commercially exploited areas community are coupled with changes in the changed the most. These results are consistent composition of the non-target community. Pat- with the prediction that the cessation of fishing terns of stability in non-target reef fish commu- mortality in marine reserves would result in nitystructurewerealsorelativetothelevelof

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protection afforded by fisheries closures. regions support more abundant kelp and oceanic Finally, we present evidence that the marine salinity regimes (Fig. 2). The consequence is some reserve network was able to conserve food web loss of balance in direct comparisons among structure across the region. We observed that zones, and highlights the importance of includ- within no take areas, higher trophic level ing site and habitat variables in our statistical omnivores were maintained at greater densities. models. We also note that areas with higher kelp In areas open to commercial and recreational density and more oceanic salinity harbour higher exploitation, the structure of the community had species richness of reef fish, which underscores changed since implementation of the FMA. the importance of examining temporal changes Fewer high trophic level omnivores and greater within management zones in the present analy- numbers of lower trophic level forage fish now sis. Further, this result highlights the under- make up the community in these regions. This representation of marine reserves or commercial pattern is consistent with incidental effects of exclusion zones in the outer coastal habitats in fishing mortality on non-target species or by Fiordland. indirect effects of removing large omnivores We conclude that in areas with high fishing from fished regions. In this case we see a greater mortality, community dynamics and associated prevalence of small forage fish in the commer- trophic interactions changed more between 2006 cially fished regions, consistent with a release and 2010 than within marine reserves, which had from predation pressure. more persistent populations of large piscivorous We use several statistical tools to elucidate omnivores. Our results are consistent with this these trends, each with important considerations. idea, however several caveats accompany this First we make use of general linear models to conclusion. Our data are based on natural resolve trends over an eight-year time series that populations subject to environmental forcing of includes a period of three years before imple- recruitment, survival and migration. We expect mentation of spatial management. Here a sample that there would be some coupling of these size of 21 sites gave us moderate resolution of processes and associated dynamics among spe- trends over time. Nevertheless some important cies. For example, conditions favouring recruit- patterns emerge among management zones and ment might be common across species and our species groupings (Figs. 3 and 6). We then observation that communities of both exploited increase our statistical resolution of the period and non-target reef fish are more dynamic in after implementation of the marine reserve fished areas could be explained by extrinsic network (2006–2010) by increasing sample size environmental forcing on both groups rather by 30%, including a more balanced representa- than the effects of management. We cannot tion of management zones. Here we are able to completely reject this alternative explanation. resolve differences in species richness and tro- Nevertheless because we considered an orthog- phic level both among and between management onal data set with replication of multiple zones (Figs. 4 and 7). We extended this analysis management zones across a large region, and using permutational multivariate analysis of because we have statistically accounted for much variance on community composition to resolve of the variability associated with site and differences in the change of whole communities environmental differences among sites, the most between 2006 and 2010 (Fig. 5). Our result that parsimonious explanation for the observed pat- communities within marine reserves were more terns is that they are a result of spatial manage- stable over this time period provides evidence of ment. A second caveat is that observed changes success for an important management objective, in the non-target reef fish community could be to increase community stability and persistence. partially or wholly driven by incidental effects of Our analysis of the density of Ecklonia radiata and fishing. Several of the non-target species suffer surface salinity underscore important differences incidental mortality, which might have direct in these indicators of habitat quality among effects on community composition. However the management zones. We demonstrate that com- largest observed changes in this community mercial exclusion zones and marine reserves are occurred in small, lower trophic level forage fish, similar in terms of habitat quality, while open which are generally not subject to direct fishing

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pressure. University of Otago Research Committee and from the The patterns observed here have important Royal Society of New Zealand’sMarsdenFund implications for the application of spatial man- (UOO038). agement to temperate reef fish communities. Over a relatively short period of four years we LITERATURE CITED observed differences in community dynamics Abrams, P. 2011. Simple life-history omnivory: re- within regions subject to fishing versus those sponses to enrichment and harvesting in systems under two levels of protection: no-take marine with intraguild predation. American Naturalist reserves and commercial exclusion. Importantly, 178:305–319. we observe the strongest positive effects of Anderson, M., R. Gorley, and K. Clarke. 2008. management in fully no-take reserves and very PERMANOVAþ for PRIMER: guide to software little benefit of partial reserves or effort-based and statistical methods. University of Auckland, management. Though many investigations have Auckland, New Zealand. focussed on local changes in biomass and age Babcock, R., N. Shears, A. Alcala, N. Barrett, G. Edgar, structure associated with single no-take marine K. Lafferty, T. McClanahan, and G. Russ. 2010. Dacadal trends in marine reserves reveal differen- reserves (Halpern 2003, Lester et al. 2009), tial rates of change in direct and indirect effects. relatively few have resolved trajectories of Proceedings of the National Academy of Sciences change for whole communities, or shifts in food 107:18256–18261. web structure within networks of reserves (e.g., Beer, N. 2011. Blue cod population structure and Micheli et al. 2004, Babcock et al. 2010). The connectivity in the New Zealand fjords. Disserta- Fiordland marine area provides an important tion. University of Otago, Dunedin, New Zealand. opportunity to resolve patterns in community Beer, N. and S. Wing. 2013. Trophic ecology drives dynamics across a replicated system. Based on spatial variability in growth among subpopulations of an exploited temperate reef fish. New Zealand the evidence presented here, we suggest that Journal of Marine and Freshwater Research 47:73– networks of fully protected areas are a powerful 89. tool, capable of conserving biodiversity, and Cherel, Y., K. Hobson, C. Guinet, and C. Vanpe. 2007. maintaining stable and fully functional marine Stable isotopes document seasonal changes in communities across a landscape. Our evidence trophic niches and winter foraging individual also suggests that even in remote areas such as specialization in diving predators from the South- the New Zealand fjords, these benefits may not ern Ocean. Journal of Animal Ecology 76:826–836. be reaped purely by exclusion of commercial Fogarty, M. 1999. Essential habitat, marine reserves activities if the number of recreational fishers is and fishery management. Trends in Ecology and Evolution 14:133–134. left unconstrained. The observed stabilising Francis, M. 2001. Coastal fishes of New Zealand: an effects of marine protection on the dynamics of identification guide. Third edition. Reed Books, whole communities and maintenance of food Auckland, New Zealand. web structure are consistent with theoretical Frank, K., B. Petrie, J. Fisher, and W. Leggett. 2011. predictions of the stabilising effect of top Transient dynamics of an altered large marine predators and omnivores on community struc- ecosystem. Nature 477:86–89. ture. These results are an important example of Goebel, N., S. Wing, and P. Boyd. 2005. A mechanism the value of spatial networks of no-take marine for onset of diatom blooms in a fjord with persistent salinity stratification. Estuarine Coastal reserves for the regional maintenance of intact and Shelf Science 64:546–560. and fully functional ecosystems at the landscape Halpern, B. 2003. The impact of marine reserves: Do scale. they work and does size matter? Ecological Applications 13:S117–137. ACKNOWLEDGMENTS Holt, R. and G. Polis. 1997. A theoretical framework for intraguild predation. American Naturalist 149:745– We thank M. Francis, J. Davis, N. Beer, C. Archibald, 764. K. Clarke, K. Rodgers, D. Neale, K. Blakemore, G. Kritzer, J. and P. Sale. 2006. Marine metapopulations. Funnel, A. Smith, E. Green and M. McArthur for Elsevier Academic Press, London, UK. valuable contributions to this research. Support was Lester, S., B. Halpern, K. Grorud-Colvert, J. Lubchenco, provided from New Zealand’s Ministry for the B. Ruttenberg, S. Gaines, S. Airame, and R. Warner. Environment and Department of Conservation, the 2009. Biological effects within no-take marine

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reserves: a global synthesis. Marine Ecology Pro- Roberts, C. 1995. Rapid buildup of fish biomass in a gress Series 384:33–46. Caribbean Marine Reserve. Conservation Biology Levine, S. 1980. Several measures of trophic level 9:815–826. applicable to complex food webs. Journal of Smith, F. and J. Witman. 1999. Species diversity in Theoretical Biology 83:195–207. subtidal landscapes: maintenance by physical Lubchenco, J., S. Palumbi, S. Gaines, and S. Andelman. processes and larval recruitment. Ecology 80:51–69. 2003. Plugging a hole in the ocean: the emerging Stanton, B. 1984. Some oceanographic observations in science of marine reserves. Ecological Applications the New Zealand fiords. Estuarine Coastal and 13:S3–S7. Shelf Science 19:89–104. MacArthur, R. 1955. Fluctuations of animal popula- Wing, S., N. Beer, and J. Jack. 2012. The resource base tions and a measure of community stability. of blue cod (Parapercis colias) subpopulations in Ecology 36:533–536. marginal fjordic habitats is linked to chemoauto- McCann, K. and A. Hastings. 1997. Re-evaluating the trophic production. Marine Ecology Progress Series omnivory-stability relationship in food webs. Pro- 466:205–214. ceedings of the Royal Society B 264:1249–1254. Wing, S. and L. Jack. 2010. Biological monitoring of the Micheli, F., B. Halpern, L. Botsford, and R. Warner. Fiordland Marine Area and Fiordland’s marine 2004. Trajectories and correlates of community reserves: 2010. Department of Conservation, Te change in no-take marine reserves. Ecological Anau, New Zealand. Applications 14:1709–1723. Wing, S. and L. Jack. 2012. Resource specialisation Parsons, T., and R. LeBrasseur. 1970. The availability of among suspension-feeding invertebrates on rock food to different trophic levels in the marine food walls in Fiordland, New Zealand, is driven by chain. In J. Steele, editor. Marine food chains. water column structure and feeding mode. Marine University of California Press, Berkeley, California, Ecology Progress Series 452:109–118. USA. Wing, S., L. Jack, and M. H. Bowman. 2006. Biological Pauly, D., V. Christensen, J. Dalsgaard, R. Froese, and monitoring of the Fiordland Marine Area and F. Torres. 1998. Fishing down marine food webs. Fiordland’s marine reserves: 2006. Department of Science 279:860–863. Conservation, Te Anau, New Zealand. Pim, S. and J. Lawton. 1978. Feeding on more than one Wing, S., J. Leichter, C. Perrin, S. Rutger, M. Bowman, trophic level. Nature 275:542–544. and C. Cornelisen. 2007. Topographic shading and Polis, G. 1994. Food webs, trophic cascades and wave exposure influence morphology and eco- community structure. Australian Journal of Ecolo- physiology of Ecklonia radiata (C. Agardh 1817) in gy 19:121–136. Fiordland, New Zealand. Limnology and Ocean- Polis, G. A. and D. R. Strong. 1996. Food web ography 52:1853–1864. complexity and community dynamics. American Wing, S. and E. Wing. 2001. Prehistoric fisheries in the Naturalist 147:813–846. Caribbean. Coral Reefs 20:1–8.

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