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Keywords:

  • Evidence-based conservation;
  • marine protected areas;
  • meta-analysis;
  • no-take zones;
  • systematic review

Abstract

  1. Top of page
  2. Abstract
  3. Introduction
  4. Methods
  5. Results
  6. Discussion
  7. Acknowledgments
  8. References
  9. Supporting Information

Marine reserves, areas closed to all fishing and other extractive activities, provide a refuge for species of commercial and conservation importance. Given the considerable resources committed to designing temperate reserve networks, we synthesized data from temperate reserves worldwide to determine their ecological effects. In common with other studies, we found higher density, biomass, and species richness in temperate marine reserves compared to adjacent exploited areas. However, there was considerable heterogeneity in magnitude of effect among reserves, variability which was largely unexplained by species or reserve characteristics. Our analytical approach allowed for formal power analyses, indicating that detection of large reserve effects in temperate systems globally requires monitoring at least 37 reserves. These results must be qualified by the limitations of data available and will undoubtedly vary at different spatio-temporal scales and for different focal species, but provide guidance for the design and monitoring of future marine conservations plans. International commitments toward establishment of multiple reserves offer a unique opportunity to assess reserve effectiveness; this opportunity can only be realized if reserves are designed to achieve clear and quantifiable objectives and are adequately monitored before and after establishment, based on appropriate power analyses, to assess how well those objectives are achieved.


Introduction

  1. Top of page
  2. Abstract
  3. Introduction
  4. Methods
  5. Results
  6. Discussion
  7. Acknowledgments
  8. References
  9. Supporting Information

Marine reserves, also known as “no-take zones,” are areas of the sea typically closed to all fishing and other extractive activities. Their use as conservation and fishery management tools in temperate regions is a subject of debate within the policy community (Jones 2007), and has been propelled high up on the agenda of many governments that have agreed to the commitments of the World Summit on Sustainable Development (e.g., http://ec.europa.eu/maritimeaffairs). This period of active policy development and implementation creates a pressing need to evaluate the current evidence on effectiveness of temperate marine reserves and an opportunity to put in place a coordinated research and monitoring program to fill important knowledge gaps. Therefore, we have undertaken the first systematic review focusing solely on the effects of temperate marine reserves worldwide. Our synthesis extends existing meta-analyses by using a more rigorous meta-analytical approach (Gates 2002; Pullin & Stewart 2006; Roberts et al. 2006; Sutton & Higgins 2008), including all published studies, examining species-level data, and assessing the number of reserves required to detect an effect when evaluating effectiveness across multiple temperate reserves.

Many marine reserves have been established around the world (Wood et al. 2008), offering an opportunity to ascertain the magnitude of effects resulting from the removal of fisheries and other extractive activities, e.g., hydrocarbon and aggregate extraction. Several synthetic studies reveal that despite variation in the magnitude and statistical evidence of observed effects, marine reserves tend to sustain higher densities and larger sizes of fish and invertebrates compared with adjacent fished areas (Gell & Roberts 2003; Halpern 2003; Micheli et al. 2004; Lester et al. 2009). On the other hand, some authors have suggested that there may be considerable heterogeneity in reserve performance, varying with the reserve management objectives, whether the reserve is part of a network of reserves, the location, size, and protection duration of the reserve, and the characteristics of the species under consideration (Jennings 2000; Mosqueira et al. 2000; Côtéet al. 2001; Micheli et al. 2004; Kaiser 2005; Claudet et al. 2008).

Our focus on temperate marine reserves is driven in part by the current political agenda in the European Union that requires member states to designate networks of marine reserves by 2012. Despite the fact that tropical realms have a greater number of MPAs than temperature realms, there are adequate numbers of temperature marine reserves, with study results published in peer-reviewed journals, to justify a detailed, synthetic examination of their effects (Spalding et al. 2008). We have excluded the many studies of tropical marine reserves to prevent any confounding effects of nonrelevant studies. In addition, there may be ecological reasons to suppose that marine reserves in tropical systems respond differently to those in temperate systems. For instance, temperate species tend to be more mobile as adults and have longer distance dispersal as larvae (O’Connor et al. 2007), leading to suggestions that reserves in temperate systems might need to be larger than those in tropical settings in order to achieve comparable effects (Shipp 2003; Laurel & Bradbury 2006). However, such a supposition remains speculative and was not supported by the most recent global synthesis of ecological effects within marine reserves (Lester et al. 2009).

Here, we address three key research issues with direct relevance to policy formulation: (1) We evaluate the effect of reserve protection on density, biomass, and species richness of marine biota within reserve borders, (2) we assess how reserve effects vary in relation to taxon, reserve parameters, and species characteristics, and (3) we use this information to estimate the number of reserves required to detect specified levels of effectiveness.

Methods

  1. Top of page
  2. Abstract
  3. Introduction
  4. Methods
  5. Results
  6. Discussion
  7. Acknowledgments
  8. References
  9. Supporting Information

Systematic review

We used systematic review methods (Pullin & Stewart 2006; Higgins & Green 2008) developed from the health sciences for use in ecology to test the effectiveness of policy or management interventions (The review protocol and full review are available at http://www.environmentalevidence.org/SR23.htm). Such evidence-based approaches are considered the most robust methods on which to base decision making (Gates 2002; Roberts et al. 2006; Sutton & Higgins 2008). We searched the literature using repeatable methods to identify and select relevant studies. We extracted data reporting the mean and variance of density, biomass, and species richness, inside and outside reserves to derive species level effect sizes, along with covariates of interest. We synthesized these data and explored reasons for variation using standard meta-analytical techniques, and undertook sensitivity analyses to assess the robustness of the results. Details of the methods are provided in Table S1.

Results

  1. Top of page
  2. Abstract
  3. Introduction
  4. Methods
  5. Results
  6. Discussion
  7. Acknowledgments
  8. References
  9. Supporting Information

Thirty-four studies presented data on 30 temperate no-take areas that fulfill our inclusion criteria (Table 1). Many studies were excluded because they had no control (fished site or “before” measurement) or because the marine reserve conferred only partial protection (e.g., Wolff 1992; Murawski et al. 2000; Bradshaw et al. 2001; Blyth-Skyrme et al. 2006). Of those studies included in the review, only one (Claudet et al. 2006) used a Before/After/Control/Impact (BACI) design; the rest were based on comparisons of no-take areas and adjacent control areas (i.e., fished areas). In addition, all studies, with the exception of Tuya et al. (2000), Paddack & Estes (2001), and Willis & Millar (2005), were based on repeated (pseudo-replicated) observations in single no-take areas. Studies measuring fish abundance most often utilized visual dive transects (19 reserves). Eight of the nine remaining reserves studied had at least one author in common. This is a potential source of performance (measurement) bias. Twenty studies (17 reserves) were conducted in the Mediterranean, with the remaining 14 (13 reserves) from Australasia, Africa, or America. The median age of reserves (at the time the study was conducted) was 23 years, the median length of monitoring was 14 years and median period from reserve establishment to detection of effect was 9.5 years. Median values for other reported characteristics are: latitude 37.4 (39° S to 23° 26′ 21″ S and 49.2° N to 23° 26′ 21″ N), year of establishment 1986 (1963–1998), size 3 km2 (0.01–300), depth 14.5 m (3–230), and number of taxa studied 7 (1–202) (Table 1).

Table 1.  Reserve characteristics and measured outcomes of no-take zones included in this study (full references provided in supplementary material)
ReferencesReserve nameLatitudeLongitudeYear of establishmentSize (km2)Depth (m)TaxonOutcome measures
nAlgaeInvertebrataFishDensityBiomassSpecies richness
Macpherson et al. (1997)Banyuls38.20–0.481974277  XX  
Bell (1983)Banyuls-Cerbere42.303.07197421574  XX X
Tuya et al. (2000)Bell Island and Lime Kiln No take areas48.33122.00199715 203 X X  
Paddack & Estes (2001)Big Creek Marine Ecological Reserve35.68–121.30199471410  XXX 
Harmelin et al. (1995)Carry-le-Rouet43.155.101982  0.01149  XX X
Claudet et al. (2006)Courome no take zone in Cote Bleue Marine Park43.304.501995 0.2162  XX X
Branch & Odendaal (2003), Lasiak (1999)Dwesa−32.1828.5019635103 X X X
Duran & Castilla (1989)Estacion costera de investigaciones marinas−33.30−71.381982436XX X  
Borja et al. (2006)Gaztelugatxe43.44−2.781998252 X XX 
Paddack & Estes (2001)Hopkins36.60−121.9019843910  XXX 
Tuya et al. (2006)Isla La Graciosa e islotes del norte de lanzarote (Chinijo)28.0615.2419951308X XXX 
Shears & Babcock (2003), Willis & Anderson (2003), Willis & Millar (2005), Willis et al. (2003)Leigh−36.16174.48197751721XXXXX 
Garcia-Rubies & Zabala (1990), Sabates et al. (2003), Sala et al. (1998)Medes Island41.602.901983115202XXXX  
Moreno et al. (1986)Mehuin−39.24−73.13197815 88 X X  
Guidetti et al. (2005)Miramare mpa45.5013.3019861713  XX X
Lasiak (1999)Mkambati−31.1830.00?15 51 X X X
Schroeder & Love (2002)Platform Gail−34.00−119.501988  0.022303  XX  
Paddack & Estes (2001)Point Lobos36.50−121.90196331210  XXX 
Bevilacqua et al. (2006)Punta Campanella MPA40.3414.23199712 616XX X X
Castilla & Bustamante (1988)Punta El Lacho−33.30−71.3819823103XX XX 
Tuya et al. (2006)Punta la Restinga-Mar de las Calmas (El Hierro)28.0615.2419961308X XXX 
Francour (1996)Scandola42.298.4019756173  XXX 
Bordehore et al. (2003)Tabarca Island Marine Reserve38.500.601989  13932XX X  
Babcock et al. (1999), Willis & Millar (2005)Tawharanui Marine Park−36.22174.501981  4224 XXXX 
Willis & Millar (2005)Te Whanganui a Hei (Hahei)−36.49175.471992  8202  XXX 
Guidetti (2006)Torre Guaceto Marine Reserve40.7117.791992 221026XXXX  
Buxton & Smale (1989)Tsitsikamma−34.5923.341964300253  XX  
Michelli et al. (2005)Tuscan archipelago (Capraia and Giannutri)43.0010.301989  3101  XX  
Vacchi et al. (1988)Urtica38.713.181996  130?  X  X
Martell et al. (2000)Whytecliff/Porteau cove no take areas49.20−123.301993  1331  XX  

Magnitude of effects of marine reserve protection

Overall, there was evidence that temperate marine reserves harbored higher density, biomass, and species richness of all biota combined (i.e., fish, invertebrates, and algae) than areas outside of reserves (Figure 1). Combining species within reserves prior to combining effects across reserves increased the apparent effectiveness of protection (Figure 1) and thus, reserve-level analyses will overestimate effectiveness for some species. There was heterogeneity in response to protection among species and reserves in all cases despite the low power of the tests (Table 2). Failsafe numbers suggest that species-level biomass and reserve-level density analyses may overestimate the effectiveness of no-take areas, despite systematic data retrieval. By contrast, the species-level richness and density analyses are robust (Table 2).

image

Figure 1. Pooled Hedges g effect sizes based on two analyses (species level, reserve level) of three outcome parameters (density, biomass, and species richness). Error bars are 95% CIs.

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Table 2.  Summary of effects of protection on density, biomass, and species richness of all species
ModelnOverall effect sizesHeterogeneityFail safe n
Hedge's d95% CIPln rr95% CIQP
  1. Two measures of effect size are given: Hedge's d and ln rr, with associated 95% confidence intervals. Overall effect sizes were obtained by either combining all species-level or reserve-level effect sizes. Effect sizes are considered significant when their confidence intervals do not overlap zero. N in parentheses indicates sample sizes for ln rr analysis. *failsafe number > 5n+ 10, which indicates a statistically robust result.

Density-all species481 (360)0.2930.21–0.376<0.0010.3350.213–0.4571366.50  <0.0013848*
Density-all reserves290.5290.262– 0.796<0.0010.6450.206–1.08576.96<0.00164
Biomass-all species29 (23)0.5710.277–0.865<0.001 1.03270.413–1.65392.74<0.00170
Biomass-all reserves10 (8) 0.5850.125–1.044<0.0130.7270.184–1.27 40.35<0.00118
Species richness-all species130.8040.311–1.296<0.0010.2420.094–0.39059.16<0.00187*
Species richness-all reserves81.1090.308–1.91<0.0070.5210.208–0.83441.91<0.00150*

When broad taxonomic groupings (algae, invertebrates, fish) were considered separately, there was evidence that both algae (Hedge's g, 0.302, 95% CI: 0.021–0.584, P= 0.04) and fish (Hedge's g, 0.317, 95% CI: 0.229–0.406, P < 0.001) were more abundant within reserves, but invertebrates were not (Hedge's g, 0.161, 95% CI: −0.124–0.447, P= 0.27). Heterogeneity was present within each group (P < 0.001). However, when density was aggregated at the reserve level, no effect of protection was apparent for either algae (Hedge's g, 0.103, 95% CI: −0.349–0.555, P= 0.65) or invertebrates (Hedge's g, 0.271, 95% CI: −0.14–0.683, P= 0.19). Fish density remained higher within reserves (Hedge's g, 0.311, 95% CI: 0.107–0.513, P= 0.004). The effect size for fish density is equivalent to ln rr= 0.448 (95% CI: 0.303–0.593, n= 260, failsafe = 3,340), or a 57% increase in fish density within no-take areas compared with outside. There were too few data available to examine the effects of protection on biomass and species richness of algae and invertebrates.

Reasons for variation in the effects of marine reserves

Although there are differences in density within and outside reserves, there are no differences in the magnitude of effect among fish with different mobility, habitat associations, or fishery status (Figure 2). The positive response of fish density to protection increased weakly with maximum body length (median: 44.5 cm, range: 4.3–200 cm, coefficient = 0.003, P= 0.02). By contrast, there was no evidence that fish density response to protection varied with resilience (measured as K, median: 0.28, range: 0.05–2.9, power > 0.8). Therefore, within the limitation of current studies of temperate marine reserves, variation among fish species in response to protection remains largely unexplained in terms of life-history characteristics.

image

Figure 2. Variation in fish density associated with migration habits (A); environment (B); and exploitation level (C). Error bars are upper 95% CI limits.

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Reserve-level density effect sizes increased with reserve size (coefficient = 0.004, 95% CI = 0.001–0.007, z= 2.82, n= 29, P= 0.006) but the goodness of fit is low (R2= 0.037). Variation in reserve effects on density was not related to other reserve characteristics, such as reserve age, latitude, or depth (Table 3). Biomass effect sizes increased with latitude (coefficient =−0.0102, z=−2.11, P= 0.035) but again, the goodness of fit (R2= 0.082) was low. Mediterranean and non-Mediterranean studies could not be distinguished in post hoc subgroup analyses of fish density, biomass, or species richness effect sizes given that in all cases, 95% confidence intervals overlapped.

Table 3.  Meta-regression coefficients for reserve level covariates where P > 0.05
 CoefficientStandard errorP
Latitude−0.0080.0050.087
Age 0.0030.0140.817
Size 0.0030.0030.318
Maximum depth−0.0020.0040.640
Time from establishment to monitoring 0.0170.0150.246

Power analysis

Classical power analysis based on the pooled effect from all available studies suggests that 37 reserves would be the threshold required to detect a moderate to large effect size of 0.6 even at low power, if examining temperate reserves globally (1 –ß= 0.6) (Table 4). This effect equates to a 90% difference in density inside versus outside the reserve (Table 2). More than 100 reserves would be required to detect effects as small as 0.38 equating to approximately 50% difference (Figure 3). The Bayesian analysis suggests that these estimates are low and that as many as 240 reserves may require monitoring to detect global effects as large as 0.6 (Figure 4).

Table 4.  Changes in detectable effect size in relation to number of marine reserves and power
Number of reservesPower 0.6Power 0.8Power 0.9
101.3141.7242.033
190.9081.1891.4  
280.7370.9651.136
370.6370.8340.981
image

Figure 3. Power curves illustrating the relationships between effect size (D) and number of reserves at 1 –ß= 0.6, 0.8, 0.9.

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image

Figure 4. Predictive distributions from WinBUGS for necessary sample size n to achieve 80% power.

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Discussion

  1. Top of page
  2. Abstract
  3. Introduction
  4. Methods
  5. Results
  6. Discussion
  7. Acknowledgments
  8. References
  9. Supporting Information

Available evidence indicates that temperate no-take marine reserves have higher densities, biomass, and species richness of marine biota within their boundaries compared to areas outside reserves. Fish density and biomass in particular are higher in no-take than in adjacent areas, which is consistent with previous analyses (Mosqueira et al. 2000; Côtéet al. 2001; Halpern 2003; Micheli et al. 2004; Claudet et al. 2008). We also found that algal densities are higher inside reserves, which may indicate a trophic effect. In temperate systems, the common grazers such as sea urchins often reach high densities in the presence of fishing because their predators are removed from the system, and thus this result suggests a return to a more intact ecosystem state inside reserves (Behrens & Lafferty 2004).

However, some of the effects of reserves on exploited ecosystems are uncertain because of the nature of the available data. For example, only one of the studies included in the analysis used a Before-After/Control-Impact design. As a result, it is difficult to attribute definitively the differences between no-take areas and adjacent controls to the establishment of reserves. Confounded baselines have been identified as barriers to the interpretation of previous studies, with some authors arguing that the apparent effect of protection is instead due to higher-quality habitat or site-specific features in reserves before the onset of protection or changing patterns of fishing pressure after protection is put into place (Willis et al. 2003; Edgar et al. 2004). However, other analyses suggest that the reserve effects observed in this subset were not attributable to habitat differences (Halpern et al. 2004; Lester et al. 2009). Until more studies using high-quality designs are done in temperate reserves, it will remain challenging to disentangle the effects of protection from those of differences in habitat or fishing pressure.

There was considerable heterogeneity in the magnitude of effect among reserves, which remained largely unexplained by species or reserve characteristics, although protection may be more effective for larger fish and in larger reserves. Power for these analyses was low (<0.6), which is not surprising given the low sample sizes. Potential variation in the life-history parameters of the species over their lifespan, or the existence of complex nonlinear relationships are other possible explanations for the results. For comparison, the only other synthesis of wholly temperate studies documented effects of both reserve size and age (Claudet et al. 2008). It is not uncommon to find that two or more meta-analyses reach different conclusions (Kerlikowske et al. 1995; Lortie & Callaway 2006). Here the discrepancy could be methodological (i.e., analysis of raw data from 12 reserves using a weighted generalized linear mixed model (Claudet et al. 2008) versus analyzing summary metrics from 30 reserves using a meta-regression approach [this study]). Alternatively, species response within a reserve may be modified by reserve size and age, but variation in species composition between reserves may obscure this relationship. Note that Claudet et al. (2008) included only Mediterranean reserves that were likely to be more homogeneous in species composition than the global set considered here. Latitude and reserve size do not appear to have a strong influence (both had small coefficients) on reserve effectiveness and there are no differences between the pooled effects of Mediterranean reserves and other temperate reserves. It is therefore difficult to forecast the impact of proposed marine reserves in temperature latitudes using physical, geographic, or biological attributes as predictors. Within Europe, the current legal requirement to design coherent networks of marine reserves provides a unique opportunity to address this knowledge gap provided that monitoring is implemented with exploration of between-reserve heterogeneity as an objective.

One important aspect which could not be examined owing to the lack of available information was the intensity of resource exploitation, either before reserve implementation, in the surrounding waters during the period of protection, or within the reserve when enforcement was not complete. Intense resource exploitation outside and adjacent to reserve boundaries clearly increases the likelihood of finding large effects of protection, while poaching in reserves will decrease reserve effectiveness (Samoilys et al. 2007). Socioeconomic factors have also explained heterogeneity in marine reserve effectiveness in the tropics (McClanahan et al. 2006) and were not examined here due to the lack of consistent data across studies. Spatially and temporally explicit monitoring of fishing effort would be required to provide these data.

Although we have shown that temperate marine reserves might provide benefits for biodiversity conservation, we did not examine whether reserves also benefit fisheries through spillover of adults and/or larval export. Only one study included in our analysis (Guidetti 2007) explicitly stratified the sampling to analyze change in fish density with increasing distance from the reserve. However, there is growing empirical and theoretical evidence that reserves may be able to benefit unprotected areas outside through one or both of these mechanisms (e.g., Abesamis & Russ 2005; Tupper 2007; Goni et al. 2008). It is a more challenging but important step to demonstrate whether these impacts on outside areas can offset the losses to fisheries resulting from reserve establishment.

Our systematic review has revealed clear gaps in the evidence base regarding the effectiveness of temperate marine reserves for either biodiversity conservation or sustainable fisheries management. Small sample sizes result in high uncertainty about the responses of specific taxa. In particular, algae and invertebrates are understudied, which is perhaps understandable as they represent a smaller portion (in biomass and revenue) of fisheries. There are also insufficient data on biomass, species richness, deep-water no-take areas, pelagic fish species, and soft sediment systems.

Our results also emphasize the challenges of establishing adequate monitoring to detect and assess reserve effects. Using the median period from reserve establishment to detection of effect of 9.5 years, our power analyses suggest that 37 marine reserves distributed worldwide would be the threshold sample required to detect a moderate to large effect of 0.6, unless heterogeneous data are pooled across studies (DerSimonian & Laird 1986; Cooper & Hedges 1994; Gurevitch & Hedges 1999). Many of the covariate categories examined here had smaller sample sizes than this, preventing our examination of specific questions about detecting effects for individual taxa or specific habitats. Additionally, large effects may take more time to develop than small effects. Thus, sampling to detect large effects may be best achieved by establishing fewer spatial replicates but longer-term protection and monitoring. However, from an adaptive management perspective, more widespread sampling and early detection of smaller effects may be more desirable.

Given that we used a global sample, there are likely instances for which fewer than 37 reserves require monitoring to detect an effect of reserve protection. For example, specific regions could exhibit large effects of protection or low variance, such as the Channel Islands, California, USA, where evidence for impacts was detected across 15 reserves (http://www.dfg.ca.gov/marine/channel_islands/fiveyears.asp) after only 5 years of protection. Where prior knowledge of variance is available for the region of concern, a targeted power analysis can be conducted and may suggest the need for fewer reserves than the number suggested by our global power analysis. However, lower variance at the regional scale is likely not universal and must be investigated on a case-by-case basis. There may also be a case for managers of individual reserves to monitor effectiveness with power analyses based on within-reserve sampling to inform local decision making.

We argue that the deficiencies in the current evidence base can be overcome by a large-scale coordinated experimental program incorporating Before-After-Control-Impact-paired replicated reserve/comparator designs, standardized methods for parameter estimation, and agreed protocols for assessing spillover effects and fishing effort. The establishment of multiple reserves, such as those proposed in the United Kingdom (DEFRA 2008), provides a unique opportunity to overcome many of these deficiencies and more rigorously assess reserve effectiveness. The global scientific community should impress upon policy makers the need for a coordinated approach for providing the data required to inform the conservation and sustainable use of marine resources. Pre-baseline monitoring is an essential component of this approach, and the more intense the monitoring, the greater the ability to detect subsequent changes. While policy makers may wince at the costs of collecting the appropriate evidence to assess the performance of marine reserves, the costs associated with establishing inadequate networks of reserves could be disproportionately higher.

Editor : Prof. Chris Thomas

Acknowledgments

  1. Top of page
  2. Abstract
  3. Introduction
  4. Methods
  5. Results
  6. Discussion
  7. Acknowledgments
  8. References
  9. Supporting Information

This work was funded by the award of U.K. Natural Environment Research Council Knowledge Exchange grant NE/C508734/2 to ASP.

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  2. Abstract
  3. Introduction
  4. Methods
  5. Results
  6. Discussion
  7. Acknowledgments
  8. References
  9. Supporting Information
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Supporting Information

  1. Top of page
  2. Abstract
  3. Introduction
  4. Methods
  5. Results
  6. Discussion
  7. Acknowledgments
  8. References
  9. Supporting Information

Table S1 Details of the general sources, web resources, and search terms used to access information. The Search terms are expressed in Boolean terms; *denotes a wildcard. Note that marine AND no take was modified in some searches as ``no'' and ``take'' are considered stop words in many resources. Where this occurred we used ``no take'' or ``no take zone.''

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CONL_74_sm_SuppMat.doc76KSupporting info item

Please note: Wiley Blackwell is not responsible for the content or functionality of any supporting information supplied by the authors. Any queries (other than missing content) should be directed to the corresponding author for the article.