Lobsters as keystone: Only in unfished ecosystems?
Introduction
Given the extent of worldwide fishing pressure on marine species, habitats, and entire ecosystems, studies that have compared current exploited states to historical or pristine states have invariably found that large-scale changes to species abundances and ecosystem structure and function have occurred as a result of fishing (Jackson et al., 2001, Pandolfi et al., 2003, Lotze et al., 2006). Traditional fisheries management practices have mostly focussed on single-species approaches to conduct stock assessments to determine the maximum sustainable yield (MSY) that can be harvested (Browman et al., 2004, Pikitch et al., 2004). However, more recently, the ecosystem-based fisheries management (EBFM) approach is increasingly being used by fisheries management agencies following a widespread call from the scientific and academic communities for its implementation (Browman et al., 2004, Pikitch et al., 2004, Pitcher et al., 2009). EBFM is broadly defined as the recognition of the need to move towards a management system that recognises the importance of food web linkages and an understanding of how human activity affects the integrity and sustainability of all components of marine ecosystems (Pitcher et al., 2009).
Implicit in this broader view of fisheries management is the need to quantify food web linkages, the flow of energy through the ecosystem, and the ecosystem effects of fisheries. Recent fisheries studies have applied ecosystem models to assess the impact of fisheries on marine ecosystems worldwide (Worm et al., 2009, Smith et al., 2011, Garcia et al., 2012). Results from such studies indicate that entire ecosystems are directly and indirectly impacted as a result of fishing activities (Worm et al., 2009, Smith et al., 2011, Garcia et al., 2012). Historical ecosystem reconstructions have been undertaken for northern British Columbia, Canada (Ainsworth et al., 2008), and for the North Adriatic (Coll et al., 2009a), South Catalan, (Coll et al., 2009b), and North Sea regions in Europe (Mackinson and Daskalov, 2007). These model reconstructions have documented large-scale ecosystem-wide changes that have occurred as a result of fishery harvest along with other human-mediated disturbances (Coll et al., 2009a, Coll et al., 2009b). Many ecosystem models have also been used to predict the ecosystem impacts of EBFM strategies for ecosystems (Worm et al., 2009, Smith et al., 2011, Garcia et al., 2012).
In New Zealand, Māori peoples settled approximately 760 years ago, about 600 years before European arrival (Wilmshurst et al., 2010). These first settlers had a high reliance on coastal marine resources (Leach, 2006, Smith, 2011a, Smith, 2011b), as evidenced by remains of lobster (Jasus edwardsii) and other invertebrates and vertebrates in middens located on Wellington's south coast and throughout New Zealand, which were harvested by diving, pots, and hoop nets (Leach, 2006, Booth, 2008). At the beginning of the 20th century, the commercial lobster fishery on Wellington's south coast was one of the first lobster fisheries in the country (Booth, 2008). In the late 1940s, most lobster were harvested from rocky inshore areas between depths of 5 and 25 m, but the late 1970s lobster were fished to depths of 50 m (Booth, 2008). In addition to this depth change, there is evidence that the average size of a lobster is smaller today than in the 1940s (Booth, 2008). Commercial fishing of lobster through the use of pots represents the main source of fishery revenue within the Wellington region and the fishery has been managed through the quota management system (QMS) since 1986. There is also a substantial recreational lobster fishery, taken by both potting and diving within the region. The lobster fishery in New Zealand is the country's most valuable export fishery, worth $229 million for 2.7 million kg of lobster landed in 2010 (Ministry of Fisheries, 2011). In addition to lobster fishing on Wellington's south coast, there are commercial and recreational fisheries for many finfish and shellfish species.
The exploitation of coastal marine resources affects not only the targeted species, but also other species and habitats in the ecosystem (Jackson et al., 2001, Pandolfi et al., 2003, Lotze et al., 2006). By studying trophic dynamics in areas protected by no-take MRs in comparison to exploited areas, it is possible to understand the ecosystem effects of fishing and how exploited ecosystems recover. Keystoneness, an indicator for identifying keystone speices, is one of many useful indicators for understanding how ecosystems respond to changes in abundance of certain species (Paine, 1966, Paine, 1969, Power et al., 1996, Libralato et al., 2006, Libralato et al., 2010, Link et al., 2010a, Link et al., 2010b). A keystone species is defined as a species whose effect on an ecosystem is disproportionately large relative to its abundance and is important for understanding how individual species affect ecosystems (Power et al., 1996). In New Zealand, a top-down trophic cascade has been observed at a MR, where urchin (Evechinus chloroticus) grazed areas have been reduced in spatial extent through top-down predation on the urchin population by recovering populations of protected predators such as lobster (J. edwardsii) and fish (snapper – Chrysophrys auratus) (Cole and Keuskamp, 1998, Shears and Babcock, 2002, Shears and Babcock, 2003).
In 2008, the Taputeranga MR was established on Wellington's south coast (Pande and Gardner, 2009). This full no-take MR protects 854.79 hectares of coastal waters, including habitats suitable for lobster and other harvested reef species. In order to understand the ecosystem effects of fishing and ecosystem response to MR protection on the south coast of Wellington, we have constructed ecosystem models for three time frames: historical past, pre-MR establishment, and distant future. Using fisheries catch records and stock assessments, we constructed a historical ecosystem model for 1940 prior to large-scale commercial exploitation. Using extensive field observations, a model was constructed to represent the ecosystem prior to implementation of the Taputeranga MR in 2008 (exploited state). This model was used to simulate the future ecosystem in 2050 following 42 years of protection by the Taputeranga MR. These three models were analysed to determine the ecosystem effects of fisheries, and how the ecosystem responds to MR protection. We then compared our results to other ecosystems protected by MRs in New Zealand (Shears and Babcock, 2003, Pinkerton et al., 2008), and ecosystem responses to lobster fisheries worldwide.
Section snippets
Study area
The study area on Wellington's south coast includes the Taputeranga MR (41°20 S, 174°45 E). This full no-take reserve extends from Princess Bay on the eastern boundary to Quarry Bay on the western boundary (Fig. 1) and was officially designated in August 2008. We conducted research in the Taputeranga MR in collaboration with, and permission from, the Department of Conservation that manages the MR. The marine environment that the Taputeranga MR protects is representative of the temperate Cook
Ecosystem structure and function
The pre-MR, exploited ecosystem model is described by 24 trophic groups linked by 77 predator–prey interactions, and approximately five trophic levels (Fig. 2 and Table 1), with the majority of biomass within the ecosystem being made up of primary producers (Fig. 2). Macroalgal trophic groups accounted for 78% of the biomass in the ecosystem, being made up of 51% canopy, 25% foliose and 2% crustose species. Microphytes accounted for 10% of ecosystem biomass. The benthic invertebrate trophic
Historical, pre-MR establishment, and future ecosystem states
The degree of ecosystem exploitation and change that has taken place during the last 70 years of fishing activity along the Wellington south coast is not as severe as has been documented at European locations, which have in some cases been subjected to 2500 years of exploitation (Coll et al., 2009a, Coll et al., 2009b). The degree of exploitation that we have observed is more similar to marine ecosystems with shorter and less intense exploitation histories, such as in northern British Columbia,
Acknowledgements
We are grateful to Matt Pinkerton and Carolyn Lundquist of NIWA New Zealand for providing support with ecosystem modelling, and to Villy Christensen and Divya Varkey at the University of British Columbia's Fisheries Centre for assisting with Ecopath. Helen Kettles, Daniel Boyce, and Ben Knight provided assistance with data collation. Andrew Rae and Benjamin Magana Rodriguez provided technical support with GIS procedures. Tyler Eddy, Jamie Tam, and Tim Jones were supported by Victoria University
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Present address: Department of Conservation, Conservation House, 18 Manners Street, Wellington 6011, New Zealand.