INTRODUCTION TO THE CONCEPT OF PROTECTED AREAS
The benefits of marine protected area (MPA) establishment are readily recognised and accepted, but they are often set up with insufficient understanding of the fundamentals of conservation science. Many such reserves are consequently deemed ineffective because of poor management or weak establishment (Agardy et al. 2003). An MPA is “Any defined area within or adjacent to the marine environment, together with its overlying water and associated flora, fauna, historical and cultural features, which has been reserved by legislation or other effective means, including custom, with the effect that its marine and/or coastal biodiversity enjoys a higher level of protection than its surroundings” (Kelleher and Kenchington 1992; Agardy et al. 2003). This definition is specific in its boundaries and thus highlights a crucial flaw in MPA-based conservation. What about the migrating species that utilise the environment outside of the MPA? Marine habitats are dynamic and three-dimensional, so organisms can move and disperse across boundaries. This property has ramifications for reserve design and eventual species preservation (Carr 2003). Under current designs, marine protected areas can only provide protection during specific life stages, underscoring the importance of connectivity and ocean health as a whole.
The ever-increasing human population, with its associated range of activities, is the main driver of shifts in structure and ecological functioning of environments worldwide (Lubchenco et al. 2003). The establishment of marine protected areas and terrestrial nature reserves has become a common management tool in the effort to restore ecosystems’ health and functionality by excluding selected areas from human development and allowing them to return to a more natural state (Christianen et al. 2014). Humans depend on the environment for survival, so conserving biodiversity through the provision of protection for both marine and terrestrial species across the globe is an urgent matter. The process currently followed is simply to preserve habitats and increase population numbers in an attempt at species recovery. While the level of protection afforded by different MPAs is variable, two potentially conflicting objectives stand out in the establishment of most marine reserves. These are the conservation of the species and habitat within reserve boundaries and the facilitation of the sustainable use of their resources (Lubchenco et al. 2003; Roberts et al. 2003).
A HISTORICAL PERSPECTIVE OF PROTECTED AREAS
The design of reserve models has its roots in the terrestrial environment ,with modern protected areas having their origin in early nineteenth century England, while the first marine reserves were proclaimed in the early twentieth century (Agardy et al. 2003; Carr et al. 2003; Carroll 2003). The number of marine protected areas more than tripled from 1986 to 1995 (Agardy et al. 2003). Before the revolutions in fishing efficiency brought about by technological advances in storage, propulsion and equipment, many marine populations had temporal and spatial refugia, inaccessible to fishing pressures by virtue of their depth or their isolation during specific climatic events (Bohnsack 1998; Dayton et al. 2000). When stocks elsewhere were depleted, these refuge sites began to be exploited and the need for areas of protection was realised (Dayton et al. 2000).
Through the last century, there has been a progression from simple regulation and management of marine resources and activities, to the establishment of small MPAs located in economically valuable areas, and finally to the formation of large multi-use MPAs governed by an integration of conservation and fisheries policies in second half of the twentieth century (Kelleher & Kenchington 1992). The Geneva Conventions of the Law of the Sea was developed in 1958, forming a legal framework within which the establishment of MPAs and the conservation of marine resources could take place (Kelleher & Kenchington 1992). Certain key environmental and ecological differences between terrestrial and marine ecosystems have implications for reserve design and management objectives (Carr et al. 2003). While marine reserves exhibit varying degrees of protection in an attempt to balance the sustainable use of resources while allowing populations to recover and promoting biodiversity, their terrestrial counterparts have a greater focus on biodiversity. Without careful monitoring and effective management, unintentional consequences can arise through a combination of increasing densities and limited knowledge of the interactions between species and their environment.
LOCAL OVERABUNDANCE OF SPECIES IN PROTECTED AREAS
Elephants are threatened with extinction in some regions of their range but are able to successfully persist and reach a degree of overabundance in conserved pockets, resulting in the management of Africa’s elephants being a complex endeavour (Blake & Hedges 2004; van Aarde & Jackson 2007). High local abundance in elephant populations within conservation areas, along with their subsequent and often detrimental impacts on vegetation and ecological processes, has led to the formation of the “elephant problem” concept (Caughley 1976; van Aarde & Jackson 2007). This concept is one explanation for how a species achieves local overabundance in a designated preservation area, with cascading effects on community and habitat structure, often to the detriment of the species afforded protection (van Aarde & Jackson 2007).
The factors behind the local overabundance problem are a result of human modifications of the environment. Artificial water provision concentrates animals in small areas around water sources, while the reduction and fragmentation of home ranges through the destruction of habitats and formation of artificial barriers prevent natural dispersal. Elephants, once widely distributed across sub-Saharan Africa, are now restricted to isolated pockets within a fragmented mosaic of habitats accounting for only 16% of their historical distribution range (Blanc et al. 2005; van Aarde & Jackson 2007).
If the establishment of MPAs provides effective protection, sustainable development of target organism populations will be promoted with their increased density signifying the “successfulness” of the model (Halpern 2003; Christianen et al. 2014). However, especially in smaller reserves, MPA zonation has sometimes resulted in an accumulation of target species far exceeding their historic densities, which is detrimental to ecological processes and their habitat as a whole (Halpern 2003; Caughley 1976; Berger et al. 2001; Christianen 2014). These effects are rarely considered when constructing conservation policies (Christianen et al. 2014). Marine reserves are a relatively recent phenomenon when compared to terrestrial systems, so few documented examples exist of a population’s density increasing to such an extent that it becomes locally overabundant.
A well depicted case in a terrestrial context is the destructive effect of the African elephant (Loxodonta africana) on an environment supporting dense populations. Elephants are migratory and historically traversed great distances across Africa seeking out water sources and foraging for seasonally-available food. As numbers dwindled in the face of hunting and habitat loss, protection of the species became paramount and so fences were erected and reserves formed. Elephant numbers in these enclosed areas subsequently increased but the landscapes within the reserve boundaries have become severely degraded with detrimental consequences for other species (Blanc et al. 2005; Young et al. 2009).
African Elephants and Landscape Transformation
In a terrestrial context, large herbivores are landscape engineers in that they modify habitat structure spatially and temporally (Gonnet 2001; Landman et al. 2012). Being megaherbivores, African elephants transform the landscape and, especially when in high densities, directly influence the structure and composition of the vegetation, thus modifying an array of ecological processes (Ogada et al. 2008; Landman et al. 2014).
The distribution of resources across a landscape and thus its transformation as a result of herbivore-induced impacts, such as foraging and trampling, is spatially heterogeneous (Coughenour 1991; Washington-Allen et al. 2004). The intensity and diversity of these impacts are dynamic and determined by a spatiotemporal hierarchy of factors resulting in a phenomenon known as the piosphere effect, a radial pattern of diminishing impacts from a patch centre (Chamaillé-Jammes et al. 2007; Landman et al. 2012). A familiar example is the radial vegetation gradient found around water sources (Coughenour 1991; Chamaillé-Jammes et al. 2009). The effects on vegetation dynamics cascade through the system reverberating across the full spectrum of biological elements; from individual species to the overall health and functionality of the ecosystem and culminating in shifts in biodiversity patterns (Owen-Smith 1996; Landman et al. 2012).
Landscape transformation, as documented for the high density elephant populations in both the Kruger National Park (KNP) and the Addo Elephant National Park (AENP) of South Africa, has caused reserve managers to revise conservation goals (Owen-Smith et al. 2006; Landman et al. 2012). Focus has transferred from species-specific conservation to the state of biological diversity, with the minimisation of single species impacts being top priority to ensure system resilience (Rogers 2003; Owen-Smith et al. 2006; Chamaillé-Jammes et al. 2007). The move from species-specific conservation to the preservation of biological diversity acknowledges the influence that high densities of elephants have on ecological patterns and processes, spanning biotic and abiotic system components (Kerley et al. 2008; Landman et al. 2014). The location and availability of key resources such as surface water and the quality and quantity of food are fundamental in determining the variation of the intensity and heterogeneity of elephant-induced impacts (Chamaillé-Jammes et al. 2007). Through the adoption of the metapopulation concept and allowing dispersal-induced fluctuation of elephants between preserved areas, not treated as isolated conservation islands, the impacts on vegetation and community structure could be reduced (van Aarde & Jackson 2007). By way of conservation corridors or transfrontier reserves, this model would take into account the spatiotemporal shifts in quality and quantity of resources, allowing elephants to migrate before severely degrading the landscape of a particular region.
Green Sea Turtles and Sea Grass Pastures
Landscape degradation is not unique to terrestrial habitats, as documented in various studies concerning turtle populations and their effects on sea grass habitats (Murdoch et al. 2007; Fourqurean et al. 2010; Christianen et al. 2014). The Atlantic and Pacific Ocean populations of green sea turtles (Chelonia mydas) have been recovering as a result of conservation efforts aimed at minimising the overexploitation of this species (Chaloupka et al. 2008). The focus of protection is on breeding and feeding aggregations, which has substantially contributed to the increase in populations during recent times, but is inadequate in its protection of the species as a whole. Sea turtles play key ecological roles in their environment; including that of consumer, ecosystem engineer, transporter of nutrients, prey, and facilitator (Heithaus 2013). Limited research into these roles has resulted in little knowledge on which to base conservation plans. The complex life history of C. mydas has huge implications for conservation efforts. Turtle move between a variety of habitats during their lifetime; ranging from the terrestrial, during egg-laying, to both neritic and oceanic environments depending on the requirements of each life stage (Bolten 2003).
Other than the little understood early juvenile stage, two broad assertions can be made for sea turtles; juveniles tend to be resident in certain neritic environments that support sufficient grazing habitats and adults carry out migrations between breeding and feeding grounds (Carr 1986; Bolten 2003). Juveniles have the ability to attain high densities if their feeding grounds fall within the protection of an MPA while adults are exposed to various threats, from the fishing industry among others, during their long-distance migrations. The environments in which turtles aggregate for either feeding or breeding, tend to be “static” in their boundaries with turtles often returning to the same sites year after year. This remigration behaviour makes it easier to protect these habitats than the oceanic environments utilised by migrating individuals. C. mydas is primarily herbivorous and is capable of consuming vast amounts of sea grass during its resident life stage. Green turtles are essential for the maintenance of healthy sea grass pastures, reducing canopy structure and preventing build-up of organic matter within this habitat (Thayer et al. 1984; Fourqurean et al. 2010).
Sea grass pastures may have formed the prime habitat and food source for green turtles for as long as fifty million years (Christianen et al. 2014). In an MPA on the Bermuda Platform, the sea grass habitat within its protection supports a juvenile green sea turtle population that utilises the pastures for up to fourteen years before returning to their natal beaches to breed (Godley et al. 2004). The Caribbean Coastal Marine Productivity Program has observed a decline in sea grass meadows in this region as a result of the intense grazing pressure exerted by turtles on a habitat which has yet to recover (Fourqurean et al. 2010). The inability of the sea grass to recover from intense levels of grazing could be detrimental to the juvenile population inhabiting the area, as the nearest suitable sea grass habitat is over 1000 km away (Murdoch et al. 2007; Fourqurean et al. 2010).
A four year research study monitored populations and habitat in a ten-year-old Indonesian MPA where C. mydas receives full protection (Christianen et al. 2014). Direct habitat degradation occurred as a result of increased green sea turtle density and subsequent intensified grazing pressure on sea grass pastures. This population not only achieved the highest global density, surpassing even the density estimates of populations preceding human exploitation, but was also at least four times the density of those outside the protection of the MPA (McClenachan et al. 2006; Christianen et al. 2014). The grazing pressure of large numbers of turtles congregating in this area, attributed to immigration, has been detrimental to resources. In addition to removing 100 % of the total daily leaf production, turtles used their flippers to dig in the sediment for rhizomes and roots, which is a previously undescribed feeding strategy (Christianen et al. 2012; Christianen et al. 2014). This behaviour lead to barren tracts in the sea grass pastures, enhancing erosion and reducing rejuvenation potential, causing severe degradation of sea grass habitat (Christianen et al. 2014), analogous to the effects of large elephant populations on terrestrial vegetation structure and composition.
While above-ground grazing of sea grass sustains the turtle population, below-ground biomass contributes to continued regeneration, so the foraging of this particular biomass substantially decreases the resilience of the pastures to the expanding turtle density (Moran & Bjorndal 2005). This foraging behaviour triggers a collapse in the vegetation, converting a landscape abundant in sea grass to an erosion-driven wasteland. In order for sea grass to effectively recover the turtle population would have to be reduced to close to zero individuals (Christianen et al. 2014). This unintended consequence renders the MPA model ineffective in its goal of providing protection for the endangered turtles (Christianen et al. 2014). The overabundance of turtles in protected areas could be a result of destruction of the sea grass habitat of surrounding areas through anthropogenic activities influencing productivity. Promoting the health of these surrounding habitats would raise MPAs from isolated islands of protection to valuable components in a network geared towards sustainability.
Other turtle species might have to utilise sea grass habitats as a result of the decline in availability and quality of their preferred habitats, exerting additional pressure on the system. The hawksbill sea turtle (Eretmochelys imbricata) is another circumtropical species: it spends the first few years of its life in oceanic habitats before recruiting to neritic habitats, primarily coral reefs (Bjorndal & Bolten 2010). A decline in reef habitats may endanger the health and persistence of hawksbill populations, driving them to utilise peripheral habitats such as sea grass beds in order to survive (Jones et al. 2004; Bjorndal & Bolten 2010). The potential sharing of this habitat type by two bulk grazing species highlights the importance of protecting foraging habitats both inside and outside of protected areas.
Monk Seal Translocation Program within the Hawaiian Archipelago
Many diverse habitats are required for ensuring the health and resilience of an ecosystem. The effectiveness of a protected area or network of areas at safeguarding the recovery of a species depends not only on reserve size but also on the condition of habitats outside reserve boundaries that are essential for the survival of threatened species. So why can one of the world’s largest marine protected areas not support a successfully recovering endemic Hawaiian monk seal (Monachus schauinslandi) population? The 360 000 km2 Papahānaumokuākea Marine National Monument, established in 2006 to incorporate the Northwestern Hawaiian Islands and surrounding waters, can only protect a fraction of the distribution range of larger migratory species (Gerber et al. 2011).
The establishment of larger protected areas with stringent prohibitions and no-take policies is an attempt to encompass sufficient habitat diversity and space to go some way towards protecting migratory animals. Despite conservation efforts on this scale, the Hawaiian monk seal population residing within the reserve boundaries along the Hawaiian Archipelago is declining towards extinction while the population inhabiting the largely unprotected Main Hawaiian Islands is growing (Gerber et al. 2011). The failure of this MPA to enhance monk seal population numbers may be a result of low juvenile survival rates under increased interspecific competition and predation by jacks (Carangidae) and Galapagos sharks (Carcharhinus galapagensis). These are target species harvested outside of the protection of the reserve (Baker & Thompson 2007; Baker et al. 2011; Gerber et al. 2011). Various conservation efforts are currently under way in order to recover the declining seal population. One proposed intervention programme involves the translocation of weaned pups from the low juvenile survival areas within the reserve to areas in the rest of the Hawaiian island system that exhibit higher survival rates (Baker et al. 2011; Gerber et al. 2011). Once old enough, these pups will be returned to their original population with an expected increase in their probability of survival (Gerber et al. 2011).
Although heralded as an adaptive management strategy, the translocation of monk seals between populations inside and outside of the reserve poses several complications. The condition of the foraging habitat, and the densities of seal competitors and predators, could potentially invalidate the effectiveness of the program: individuals returning to the unfamiliar environment of the reserve would face reduced food availability and increased risk of predation (Gerber et al. 2011). The survival of the Hawaiian monk seal species now depends on a two phase management strategy that supports the increasing population in the Main Hawaiian Islands region while maintaining viable subpopulations within the boundaries of Papahānaumokuākea (Baker et al. 2011; Gerber et al. 2011). This example highlights the importance of understanding not just the patterns observed in the environment but also the ecological processes driving them. Ecosystem interactions should be of chief importance when planning and managing a protected area, with a focus wider than the alleviation of pressure on certain species, inclusive of the determination and mitigation of underlying ecological drivers and the integrity of whole ecosystems (Christianen et al. 2014).
***On the 19 August 2015, the National Marine Fisheries Service announced that a further 7 000 square miles of critical habitat in the Main Hawaiian Islands has been identified as a region requiring special considerations for the minimisation of coastal destruction and degradation (Marine Conservation Institute).
SHIFTING BASELINES IN A TRANSFORMED ENVIRONMENT
A stumbling block with regards to conservation and the promotion of population growth is the phenomenon of shifting baselines. Some studies of the fishing industry failed to identify an appropriate baseline against which to measure the change in populations for exploited fish species (Pauly 1995). Cognitive bias causes people to compare the current size of a population to that at the beginning of their experience, rather than objectively analysing historical evidence. In sea turtle populations, the identification of accurate baselines is challenging because populations were already diminished before recording of numbers and quantifying of trends (Bjorndal & Bolten 2003). The World Conservation Union (IUCN) uses the standard of ten years or three generations, whichever is longer, as a baseline in the formation of recovery goals. Using this criterion for the green turtle proves to be a problem because in the late 1800s the species was already declining under excessive exploitation (Bjorndal & Bolten 2003). Conservation studies should aim at determining accurate baselines from archaeological evidence and verified historical accounts (Bjorndal & Bolten 2003). These estimates are often set in a pristine and pre-exploitation environment, so the question of sustainability arises. Is it possible for the environment, in its current state, to support these historical densities? Green turtle foraging impacts aside, seagrass meadows have been declining globally as a result of anthropogenic activities (Waycott et al. 2009; Christianen et al. 2014). Would it not then be detrimental to increase the green turtle population to prehistoric densities without attempting to ensure the health of the environment on which they depend? A reduction in the size of a healthy habitat implies that fewer turtles would be required to fulfil their ecological roles (Bjorndal & Bolten 2003).
Returning to the terrestrial analogy, the same holds for the African elephant. Providing protection to restore populations to historic densities without taking into account the historic migratory behaviour that mitigates its impact on the landscape would surely be a short-sighted conservation strategy. In the past, Kruger National Park managers used cull events in an effort to reduce the effect of a growing population of elephants on the environment. In more recent times, translocation programs, conservation corridors and transfrontier parks have become the favoured methods of impact mitigation. In marine systems however, where there are no fences and species are able to freely disperse and aggregate in search of resources, population control becomes more complex. The recovery goals set for sea turtles should be focused on the fulfilment of their ecological roles within the environment, as this strategy includes the health of a single species in a broader, functioning ecosystem (Bjorndal and Bolten 2003). In protected habitats exhibiting local overabundance, the zoning of the MPA could be such that limited exploitation would be permitted for sustainable use. Many aboriginal people of oceanic islands and coastal regions have traditionally relied on the sea for sustenance and under specific zoning and special decrees, taking their traditional methods into account, could continue to do so (Campbell 2003). The challenge however is in local versus global population health. While some turtle populations are stable or even increasing in certain regions (Nel et al. 2013), global populations are still declining (Campbell 2003) and therefore allowing exploitation at this stage may be counterproductive. Focus should rather be on minimising the drivers of population decline while restoring functioning ecosystems. This approach supports the potential recovery of populations, allowing for a more even distribution of the numbers.
OPTIMAL DESIGN FOR EFFECTIVE MARINE PROTECTED AREAS
The establishment of marine protected areas is an essential component in the conservation of biodiversity and the sustainable management of fisheries. Conditions specific to regions and ecosystems determine the objectives and subsequent effectiveness of protected areas (Halpern 2003; Roberts et al. 2003; Christianen et al. 2014). An understanding of the ecology of target species, and the anthropogenic and climate change pressures exerted on them, will help in the protection of optimal habitats and refuge areas in the face of shifting suitability (Bjorndal & Bolten 2010). The sustainability of local populations under the protection of MPAs depends on several key elements. Minimising local overabundance requires the preservation of habitats inside and outside reserve boundaries in order to provide sufficient alternative foraging habitat. Alternative habitats in an interconnected network of protected areas should promote migration and alleviate density-dependent pressures on the resources within the MPA (Christianen et al. 2014). In marine reserves of a larger scale, protecting natural predators may prevent local overabundance and stimulate the dispersal of certain species over a larger area (Bjorndal & Bolten 2010; Christianen et al. 2014).
Connectivity between populations through a series of protected areas enhances genetic diversity and allows replenishment of depleted stocks in areas not under protection (Carr et al. 2003). The strategic placement of reserves along migratory routes and in spawning grounds, with adult reserves being located in close proximity to nursery habitats, ensures the viability of protected populations (Carr et al. 2003; Halpern 2003). In the terrestrial context, static conservation corridors are sufficient to ensure population viability but in the more dynamic marine environment, an understanding of the oceanographic elements influencing larval dispersal and species movements is necessary for determining MPA locations (Carr et al. 2003). Marine protected areas should maintain ecosystems that are functionally intact at regional scales, to ensure the persistence of viable populations both inside and outside of reserve boundaries (Roberts et al. 2003).
Explore. Dream. Discover.
Agardy MT, Bridgewater P, Crosby MP, Day J, Dayton PK, Kenchington R, Laffoley D, McConney P, Murray PA, Parks JE, Peau L. 2003. Dangerous targets? Unresolved issues and ideological clashes around marine protected areas. Aquatic Conservation: Marine and Freshwater Ecosystems 13: 353-367.
Baker JD, Harting AL, Wurth TA, Johanos TC. 2011. Dramatic shifts in Hawaiian monk seal distribution predicted from divergent regional trends. Marine Mammal Science 27: 78-93.
Baker JD, Thompson PM. 2007. Temporal and spatial variation in age-specific survival rates of a long-lived mammal, the Hawaiian monk seal. Proceedings of the Royal Society of Biological Sciences 274: 407-415.
Berger J, Swenson JE, Persson IL. 2001. Recolonising carnivores and naïve prey: Conservation lessons from Pleistocene extinctions. Science 291: 1036-1039.
Bjorndal KA, Bolten AB. 2003. From ghosts to key species: Restoring sea turtle populations to fulfil their ecological roles. Marine Turtle Newsletter 100: 16-21.
Bjorndal KA, Bolten AB. 2010. Hawksbill sea turtles in seagrass pastures: success in a peripheral habitat. Marine Biology 157: 135-145.
Blake S, Hedges S. 2004. Sinking the flagship: The case of forest elephants in Asia and Africa. Conservation Biology 18: 1191-1202.
Blanc JJ, Barnes RFW, Craig CG, Douglas-Hamilton I, Dublin HT, Hart JA, Thouless CR. 2005. Changes in elephant numbers in major savanna populations in eastern and southern Africa. Pachyderm 38: 19-28.
Bohnsack JA. 1998. Application of marine reserves to reef fisheries management. Australian Journal of Ecology 23: 298-304.
Bolten AB. 2003. Variation in sea turtle life history patterns: Neritic vs. oceanic developmental stages. In: The biology of sea turtles: Volume II, (eds) P.L. Lutz, J.A. Musick & J. Wyneken, CRC Press, Florida.
Carr AF. 1986. Rips, FADS, and little loggerheads. Bioscience 36: 92-100.
Carr MH, Neigel JE, Estes JA, Andelman S, Warner RR, Largier JL. 2003. Comparing marine and terrestrial ecosystems: implications for the design of coastal marine reserves. Ecological Applications 13 (1): 90-107.
Carroll V. 2003. The natural history of visiting: Responses to Charles Waterton and Walton Hall. Studies in History and Philosophy of Biological and Biomedical Sciences 35 (1): 31-64.
Caughley G. 1976. Elephant problem – Alternative Hypothesis. East African Wildlife Journal 14: 265-283.
Chaloupka M, Bjorndal KA, Balazs GH, Bolten AB, Ehrhart LM, Limpus CJ, Suganuma H, Troëng S, Yamaguchi M. 2008. Encouraging outlook for recovery of a once severely exploited marine megaherbivore. Global Ecology and Biogeography 17: 297-304.
Chamaillé-Jammes S, Fritz H, Madzikanda H. 2009. Piosphere contribution to landscape heterogeneity: A case study of remote-sensed woody cover in a high elephant density landscape. Ecography 32: 871-880.
Chamaillé-Jammes S, Valeix M, Fritz H. 2007. Managing heterogeneity in elephant distribution: Interactions between elephant population density and surface-water availability. Journal of Applied Ecology 44: 625-633.
Christianen MJA, Govers LL, Bouma TJ, Kiswara W, Roelofs JGM, Lamers LPM, van Katwijk MM. 2012. Marine megaherbivore grazing may increase seagrass tolerance to high nutrient loads. Journal of Ecology 100: 546-560.
Christianen MJA, Herman PMJ, Bouma TJ, Lamers LPM, van Katwijk MM, van der Heide T, Mumby PJ, Silliman BR, Engelhard SL, van de Kerk M, Kiswara W, van de Koppel J. 2014. Habitat collapse due to overgrazing threatens turtle conservation in marine protected areas. Proceedings for the Royal Society of Biological Sciences 281: 20132890.
Coughenour MB. 1991. Spatial components of plant-herbivore interactions in pastoral, ranching, and native ungulate ecosystems. Journal of Range Management 44: 530-542.
Dayton PK, Sala E, Tegner MJ, Thrush S. 2000. Marine reserves: Parks, baselines, and fishery enhancement. Bulletin of Marine Science 66(3): 617-634.
Fourqurean JW, Manuel S, Coates KA, Kenworthy WJ, Smith SR. 2010. Effects of excluding sea turtle herbivores from a seagrass bed: Overgrazing may have led to loss of seagrass meadows in Bermuda. Marine Ecology Progress Series 419: 223-232.
Gerber LR, Estes J, Crawford TG, Peavey LE, Read AJ. 2011. Managing for extinction? Conflicting conservation objectives in a large marine reserve. Conservation Letters 4: 417-422.
Godley BJ, Broderick AC, Campbell LM, Ranger S, Richardson PB. 2004. An assessment of the status and exploitation of marine turtles in Bermuda. In: An assessment of the status and exploitation of marine turtle in the UK overseas territories in the wider Caribbean. Final Project Report for the Department of Environment. London: Food and Rural Affairs and the Foreign and Commonwealth Office.
Gonnet JA. 2001. Influence of cattle grazing on population density and species richness of granivorous birds (Emberizidae) in the arid plain of the Monte, Argentina. Journal of Arid Environments 48: 569-579.
Halpern BS. 2003. The impact of marine reserves: Do reserves work and does reserve size matter? Ecological Applications 13: 117-137.
Heithaus MR. 2013. Predators, prey, and the ecological roles of sea turtles. In: The biology of sea turtles: Volume III, (eds) J. Wyneken, K.J. Lohmann & J.A. Musick, CRC Press, Florida.
Jones GP, McCormick MI, Srinivasan M, Eagle JV. 2004. Coral decline threatens fish biodiversity in marine reserves. Proceedings of the Natural Academy of Sciences 101: 8251-8253.
Kelleher G, Kenchington R. 1992. Guidelines for establishing Marine Protected Areas: A Marine Conservation and Development Report. International Union for Conservation of Nature (IUCN): Switzerland.
Kerley GIH, Landman M, Kruger L, Owen-Smith N, Balfour D, Boer WF, Gaylard A, Lindsay K Slotow R. 2008. Effects of elephants on ecosystems and biodiversity. In: The 2007 scientific assessment of elephant management in South Africa, (eds) K.G. Mennell & R.J. Scholes, Council for Scientific and Industrial Research, Pretoria.
Landman M, Schoeman DS, Hall-Martin AJ, Kerley GIH. 2012. Understanding long-term variations in an elephant piosphere effect to manage impacts. PLoS ONE 7 (9): e45334.
Landman M, Schoeman DS, Hall-Martin AJ, Kerley GIH. 2014. Long-term monitoring reveals differing impacts of elephants on elements of a canopy shrub community. Ecological Applications 24 (8): 2002-2012.
Lubchenco J, Palumbi SR, Gaines SD, Andelman S. 2003. Plugging a hole in the ocean: The emerging science of marine reserves. Ecological Applications 13 (1): 3-7.
McClenachan L, Jackson JBC, Newman MJH. 2006 Conservation implications of historic sea turtle nesting beach loss. Frontiers in Ecology and the Environment 4: 290-296.
Moran KL, Bjorndal KA. 2005. Simulated green turtle grazing affects structure and productivity of seagrass pastures. Marine Ecology Progress Series 305: 235-247.
Murdoch TJT, Glasspool AF, Outerbridge M, Ward J, Manuel S, Gray J, Nash A, Coates KA Pitt J, Fourqurean JW, Barnes PA, Vierros M, Holzer K, Smith SR. 2007. Large-scale decline in offshore seagrass meadows in Bermuda. Marine Ecology Progress Series 339: 123-130.
Nel R, Punt AE, Hughes GR. 2013. Are coastal protected areas always effective in achieving population recovery for nesting sea turtles? PLoS ONE 8(5): e63525
Ogada DL, Gadd ME, Ostfeld RS, Young TP, Keesing F. 2008. Impacts of large herbivorous mammals on bird diversity and abundance in an African savannah. Oecologia 156: 387-397.
Owen-Smith N, Kerley GIH, Page B, Slotow R, van Aarde RJ. 2006. A scientific perspective on the management of elephants in the Kruger National Park and elsewhere. South African Journal of Science 102: 389-394.
Owen-Smith RN. 1996. Ecological guidelines for waterpoints in extensive protected areas. South African Journal of Wildlife Research 26: 107-112.
Pauly D. 1995. Anecdotes and the shifting baseline syndrome of fisheries. Trends in Ecology and Evolution 10: 430.
Roberts CM Branch G, Bustamante RH, Castilla JC, Dugan J, Halpern BS, Lafferty KD, Leslie H, Lubchenco J, McArdle D, Ruckelshaus M, Warner RP. 2003. Application of ecological criteria in selecting marine reserves and developing reserve networks. Ecological Applications 13(1): 215-228.
Rogers KH. 2003. Adopting a heterogeneity paradigm. Implications for the management of protected savannas. The Kruger Experience. In: Ecology and Management of Savanna Heterogeneity, (eds) J.T. Du Toit, K.H. Rogers & H.C. Biggs, Island Press, Washington DC.
Soutullo A. 2010. Extent of the global network of terrestrial protected areas. Conservation Biology 24 (2): 362-365.
Thayer GW, Bjorndal KA, Ogden JC, Williams SL, Zieman JC. 1984. Role of larger herbivores in seagrass communities. Estuaries 7: 351-376.
United Nations Environment Programme (UNEP): World Conservation Monitoring Centre (WCMC). 2014. Global statistics from the World Database on Protected Areas, UNEP- WCMC: Cambridge.
Van Aarde RJ, Jackson TP. 2007. Mega-parks for metapopulations: Addressing the causes of locally high elephant numbers in southern Africa. Biological Conservation 134 (3): 289-297.
Washington-Allen RA, Van Niel TG, Ramsey RD, West NE. 2004. Remote sensing-based piosphere analysis. GIScience and Remote Sensing 41 (2): 136-154.
Waycott M, Duarte CM, Carruthers TJB, Orth RJ, Dennison WC, Olyarnik S, Calladine A, Fourqurean JW, Heck KL, Hughes AR, Kendrick GA, Kenworthy WJ, Short FT, Williams SL. 2009 Accelerating loss of seagrasses across the globe threatens coastal ecosystems. Proceedings of the National Academy of Sciences 106 (30) 12377-12381.
Young KD, Ferreira SM, van Aarde RJ. 2009. Elephant spatial use in wet and dry savannas of southern Africa. Journal of Zoology 278: 189-205.