Revista Natureza & Conservação by Editora Cubo

ISSN 1679-0073

ISSN 1679-0073 Natureza & Conservação, 8(2) December 2010

Summary Essays and Perspectives Reservoir Fish Stocking: When One Plus One May Be Less Than Two Angelo Antonio Agostinho, Fernando Mayer Pelicice, Luiz Carlos Gomes & Horácio Ferreira Júlio Jr ................................... 103

Conservation Crossroads and the Role of Hierarchy in the Decision-Making Process Adrián Monjeau......................................................................................................................................................................... 112

Plasticity and Conservation Ulrich Lüttge ............................................................................................................................................................................. 120

Natureza & Conservação 8(2): December 2010

Brazilian Journal of Nature Conservation

Research Letters

Natureza & Conservação 8(2): December 2010

Predicting Patterns of Beta Diversity in Terrestrial Vertebrates Using Physiographic Classiﬁcations in the Brazilian Cerrado André Andrian Padial, Luis Mauricio Bini, José Alexandre Felizola Diniz-Filho, Nayara Pereira Resende de Souza & Ludgero Cardoso Galli Vieira ........................................................................................ 127

Regeneration and Colonization of an Invasive Macrophyte Grass in Response to Desiccation Thaisa Sala Michelan, Sidinei Magela Thomaz, Priscilla Carvalho, Roberta Becker Rodrigues & Márcio José Silveira ................................................................................................................... 133

Fish as Potential Controllers of Invasive Mollusks in a Neotropical Reservoir Camila Ribeiro Coutinho de Oliveira, Rosemara Fugi, Kelly Patrícia Brancalhão & Angelo Antonio Agostinho ...................... 140

Dealing with Data Uncertainty in Conservation Planning Kerrie Ann Wilson ..................................................................................................................................................................... 145

Reef Fisheries and Underwater Surveys Indicate Overﬁshing of a Brazilian Coastal Island Hudson Tercio Pinheiro, Jean-Christophe Joyeux & Agnaldo Silva Martins............................................................................. 151

Successional and Seasonal Changes in a Community of Dung Beetles (Coleoptera: Scarabaeinae) in a Brazilian Tropical Dry Forest Frederico de Siqueira Neves, Victor Hugo Fonseca Oliveira, Mário Marcos do Espírito-Santo, Fernando Zagury Vaz-de-Mello, Júlio Louzada, Arturo Sanchez-Azofeifa & Geraldo Wilson Fernandes............................... 160

How Can We Estimate Buﬀer Zones of Protected Areas? A Proposal Using Biological Data Brenda Alexandre, Renato Crouzeilles & Carlos Eduardo Viveiros Grelle ............................................................................... 165

Drafting a Blueprint for Functional and Phylogenetic Diversity Conservation in the Brazilian Cerrado Rodrigo Assis de Carvalho, Marcus Vinicius Cianciaruso, Joaquim Trindade-Filho, Maíra Dalia Sagnori, Rafael Dias Loyola .................................................................................................................................. 171

The Opportunity Cost of Conserving Amphibians and Mammals in Uganda Federica Chiozza, Luigi Boitani & Carlo Rondinini ................................................................................................................... 177

Forum Geoconservação em Áreas Protegidas: o Caso do GeoPark Araripe - CE Nájila Rejanne Alencar Julião Cabral & Teresa Lenice Nogueira da Gama Mota .................................................................... 184

Conhecimento Cientíﬁco Rogério Parentoni Martins & Francisco Ângelo Coutinho ........................................................................................................ 187

O Desaﬁo da Normatização de Informações de Biodiversidade para Gestão de Águas: Aproximando Cientistas e Gestores Tadeu Siqueira & Fabio de Oliveira Roque ............................................................................................................................... 190

Mudanças Climáticas e a Biodiversidade dos Biomas Brasileiros: Passado, Presente e Futuro Alexandre Aleixo, Ana Luisa Albernaz, Carlos Eduardo Viveiros Grelle, Mariana Moncassim Vale & Thiago Fernando Rangel.............................................................................................................. 194

O Protagonismo do Brasil no Histórico Acordo Global de Proteção à Biodiversidade Russell Mittermeier, Patrícia Carvalho Baião, Lina Barrera, Theresa Buppert, Jennifer McCullough, Olivier Langrand, Frank Wugt Larsen & Fabio Rubio Scarano.............................................................. 197 Sponsored by

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ISSN 1679-0073 Natureza & Conservação, 8(2) December 2010

Brazilian Journal of Nature Conservation

“Natureza & Conservação” (Brazilian Journal for Nature Conservation) is a peer-reviewed scientific journal devoted to improving theoretical, methodological and practical aspects of conservation science. Until 2009, Natureza & Conservação was edited by the Fundação “O Boticário de Proteção à Natureza”. From 2010 on, it became an official scientific journal of the new Brazilian Association for Ecological Science and Conservation (Associação Brasileira de Ciência Ecológica e Conservação - ABECO), with substantial support from the Boticário Foundation. The main goal of Natureza & Conservação is to communicate new research and conceptual advances in conservation science to different actors of society, including researchers, conservationists, technical officers and decision makers. Scientific papers should focus on new conceptual or methodological developments with practical implications, and case studies will be considered only if referred to more general contexts. Natureza & Conservação is currently indexed in Web of Science, Periodica, CABI International, Latindex and Hapi. Sections Natureza & Conservação publishes original papers in English, basically in two formats: Essays & Perspectives will be longer essays and reviews, updating recent topics of general interest in conservation science and highlighting new conceptual, practical or methodological advances. Papers in this section will usually be invited, but advance proposals can be submitted to the editors and are welcome. Original scientific research will be evaluated in a fast-track decision process in the format of Research Letters, which are concise manuscripts of about 3000 words (tied to a online supplementary material, if necessary). The other section, Forum, will comprise invited columns and forum, written in Portuguese, dealing with specific topics in conservation as well as correspondence, book reviews and highlights from the literature. The main aim of these sections is to make research results and methodological aspects more accessible and relevant to applied scientists and decision makers. Editorial Board

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ISSN 1679-0073

Brazilian Journal of Nature Conservation

Natureza & ConservaĂ§ĂŁo 8(2): December 2010

Natureza & Conservação / Associação Brasileira de Ciência Ecológica e Conservação (ABECO) – vol. 8, no. 2, (2010). São Carlos, SP: Editora Cubo, 2010. Brazilian Journal Nature Conservation Semestral ISSN 1679-0073 1. Ecologia. 2.Conservação da Natureza.

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From the cover: The Brazilian rocky reefs, sited mainly in southeast and south coast of Brazil, shelter a high abundance and richness of species, in some cases similar to coral reefs. Recently, in the State of Espírito Santo, a partnership among scientists, environmentalists and managers have collaborated to increase of research activities and elaboration of proposals of Marine Protected Areas involving these reef environments. Photo by Dr. Hudson Pinheiro.

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Essays & Perspectives

Brazilian Journal of Nature Conservation

Reservoir Fish Stocking: When One Plus One May Be Less Than Two Angelo Antonio Agostinho1,*, Fernando Mayer Pelicice2, Luiz Carlos Gomes1 & Horácio Ferreira Júlio Jr.3 1

Nupélia, Departamento de Biologia, Pós-graduação em Ecologia de Ambientes Aquáticos Continentais, Universidade Estadual de Maringá – UEM

Neamb, Pós-graduação em Ecologia de Ecótonos, Universidade Federal do Tocantins – UFT

Nupélia, Departamento de Biologia Celular e Genética, Pós-graduação em Ecologia de Ambientes Aquáticos Continentais, Universidade Estadual de Maringá – UEM

Abstract Fisheries management in Brazilian reservoirs is based (since the 1970’s) on stocking and construction of fish passes. Low landings of the fisheries and the precarious conservation status of native populations in the Upper Paraná River basin indicate how useless these practices were. Failures in most stocking programs conducted may be explained by the negligence of basic assumptions for implementation (clear goals, scientific foundation and evaluation of results). In spite of the common sense support, decision makers should consider that, for any management actions involving biomanipulation, there are relevant environmental risks related to the origin and selection of broodstock and production of fries, and to the releasing of reared fish. Among the latter should be mentioned introduction of associated non-native species (pathogens and parasites), genetic degradation of native stocks (bottleneck effects, loss of genetic variability and fitness, domestication), imbalances and changes in community structure. For an environmental friendly and economical and societal desirable stocking, the decision process should consider information on the receptor ecosystem, target species, uses and users of the resource, legislation and risks for biodiversity conservation. Therefore, the first aspect to be considered is the need for stocking and identification of environmental constrains to it. The ability to produce fish with genetic quality equivalent to native stock and with unaltered ability to spawn in nature (the main challenges in the stocking process) should also have decisive roles in determining whether a stocking program should be implemented. Size, quantity, season and site of releasing should be based on the life cycle, distribution and structure of natural populations, whereas evaluation and monitoring should be considered as integral and indissoluble parts of the stocking process. Habitat management and fishery control should be considered as alternatives or complements. Impoundments are sources of impacts on biodiversity and the success of stocking in such environments appears temporary. Ideally, the success should be quantified by the ability of stocked fish to reproduce in nature and to contribute to the genetic variability of the population. For ethical conservation reasons stocking cannot be only evaluated through fishery landings. Key words: Fishery Management, Exotic Species, Native Stocks, Fish Reproduction, Brazil.

Introduction The building of hydroelectric dams has profoundly changed the landscape in South American river systems by altering the quality and availability of habitats as well as the water dynamics. These changes exert strong selective pressure on pre-existing aquatic communities because not all species can colonize or maintain self-sustaining populations in *Send correspondence to: Angelo Antonio Agostinho Nupélia, Departamento de Biologia, Pós-graduação em Ecologia de Ambientes Aquáticos Continentais, Universidade Estadual de Maringá – UEM Av. Colombo 5790, bloco H-90, CEP 87020-900, Maringá, PR, Brazil E-mail: agostinhoaa@gmail.com

this new system (Agostinho et al. 2008). In addition, given the evolutionary past of South American fish, which have occurred in a predominantly lotic environment, species with pre-adaptation (sensu Fernando & Holcik 1991) in lacustrine or pelagic environments are rare (Agostinho et al. 2008). Then, several species that require running water (rheophilic species, Gomes & Miranda 2001) and extensive habitats (migratory species) are particularly vulnerable. These species are generally larger in size and have fishing value (Hoeinghaus et al. 2009). The perceptions of these impacts are old and it was already manifested by the efforts conducted to mitigate the impacts

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of the first large impoundments in Brazil (e.g., a fish ladder in the Itaipava Dam, Pardo River, in 1911). The mitigation attempts, which were generally focused on the impounded areas, involved the building of fish passes (e.g., ladders) and the installation of hatcheries for managing fish stocks and controlling fishing activity. However, these actions were only recently monitored and the results of the evaluations have indicated that several management initiatives, such as ladders, had adverse effects on the ichthyofauna targeted for conservation (Agostinho et al. 2007a; Pelicice & Agostinho, 2008; Volpato et al. 2009).

have self-sustaining populations anymore; that is, there is no natural recruitment, and the stock requires periodic and ongoing releases; (ii) supplementation stocking, when the natural stock has demographic or genetic restrictions due to a variety of reasons (e.g., habitat modification, fragmentation, excessive fishing, and natural failures in recruitment). In the first group, the interest is essentially fishing exploration. However, supplementation, if well-executed, can also serve to conservation purposes because it is implicit that natural recruitment is happening, but in low levels. In this case, the success of stocking for conservation can be measured by the proportion of individuals from the natural recruitment in the population or in the fishing, which indicates rehabilitation. There is, however, a continuum between supplementation and maintenance stocking, i.e., between interests that are strictly for conservation and those that are for fishing exploration. Because maintenance stocking is based on non-sustainable populations and is implicit continuity, it can be considered as ex-situ conservation in reservoir.

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In the context of conservation, stocking efforts are emblematic. Similar to the other management tools, stocking has been performed, in most of the cases, without clear objectives, scientific support or assessment of the results (Vieira & Pompeu 2001; Agostinho et al. 2004, 2007b); thus, most stocking efforts have been inadequate because this type of action demands a high level of knowledge and care to be successful (Blankenship & Leber 1995; Cowx 1999; Molony et al. 2003). Although stocking is intended to replenish stocks that have experienced genetic and/or demographic losses, the effect of previous stocking efforts has rarely been measured. Basic bio-ecological aspects, such as the size, quantity and genetic diversity of individuals to be stocked and the selection of target species, as well as the location and time of release have been frequently neglected or not properly addressed (Vieira & Pompeu 2001; Agostinho et al. 2004, 2007a). Ultimately, the absence or the inadequateness of monitoring has not allowed for improvements in stocking techniques, resulting in wasted efforts, resources and opportunities for over half a century. Therefore, the study briefly analyzes the theoretical concepts and assumptions that motivate the activity, presents associated potential risks, evaluates the Brazilian experience (emphasizing reservoirs) and provides basic technical recommendations for stocking to achieve greater success and environmental responsibility. We believe that a discussion of this topic is timely because stocking actions have strong popular support and, consequently, are used inappropriately and/or opportunistically (Agostinho et al. 2005). Furthermore, it is worrisome that people without any scientific knowledge, who are often motivated by laudable, albeit erroneous, reasons, can conduct stockings at any time.

Stocking Approaches Stocking is a method to manage fish populations management that involves the release of wild or cultivated organisms to replenish a specific stock with demographic and/or genetic restrictions (temporary stocking) or to increase the fishing yield above the one supported by natural recruitment (permanent stocking). Therefore, it can be motivated by interests in the conservation of stocks and/or biomass production for fishing. Stocking can be classified as the following: (i) maintenance stocking, when individuals of a species are released into an area where it historically occurs naturally, but it does not

In a broad sense, stocking can also be classified as (iii) addition or introduction, when it involves the release of a species into an area where it does not occur naturally but where it can, however, establish a self-sustaining population. Given the peculiarities of this stocking approach, it will not be discussed in this article (For more details on the subject, see Agostinho et al. 2007a). Other recurring terms for stocking are given by Cowx (1994, 1999), such as stocking for mitigation (voluntary or mandatory to attenuate or compensate for damage produced in the environment), stocking for enhancement (improve the fishing yield), stocking for restoration (complement other management actions designed to remove or reduce factors that limit stocks), creation of new fisheries (addition or introduction of new species). In the case of large impoundments, where the impact on ichthyofauna is relevant, even stocking that is performed to improve fishing should not, for ethical reasons, disregard conservation interests.

Potential Impacts As any management action, fish stocking can carry some environmental risks that can reach tragic proportions if conducted carelessly. The potential impacts of stocking include: the introduction of non-native species of fish, even by stockings that are not for that purpose; dissemination of pathogens and parasites; deleterious effects related to the genetic quality of matrices and fingerlings (bottleneck effects, loss of genetic variability and fitness, domestication); and impacts on the structure and functioning of communities (intra- and interspecific competition, predation) (Figure 1). Hydrological changes, which are inevitable in impoundments, along with introduced species, are currently the main threats to freshwater biota (Cambray 2003; Eby et al. 2006; Rahel 2007; Johnson et al. 2008; Vitule et al. 2009). Paradoxically,

Reservoir Fish Stocking

the stocking of exotic species was one of the strategies used to mitigate impacts arising from the impoundments. Because stocking with non-native species in public waters is illegal and blatantly contradicts the commitments made by countries that signed the Convention on Biodiversity, including Brazil, which promulgated it as law, this approach of fishing management will not be discussed in this article. In any case, the problems with predation, competition, parasitism, habitat changes and genetic degradation due to stocking with non-native species are widely discussed in the literature (Zaret & Paine 1973; Santos et al. 1994, 2001; Gabrielli & Orsi 2000; Vieira & Pompeu 2001; Gomieiro & Braga 2004; Canonico et al. 2005; Agostinho et al. 2007a; Resende et al. 2008; Latini & Petrere 2004; Fugi et al. 2008; Pelicice & Agostinho 2009). The introduction of pathogens and parasites during the process of stocking is not any less problematic for conservation; these invaders are introduced by the water used for transportation and through infected fish released in the environment (Molony et al. 2003 and citations therein). Genetic losses, although they have not been well researched in Brazilian reservoirs (for an exception, see Matsumoto & Hilsdorf 2009; Rodriguez-Rodriguez et al. 2010), have been considered among the most common and deleterious effects of the process of stocking (Hindar et al. 1991; Ford 2002; Araki et al. 2007) and can compromise the viability of wild populations in the short and long term (on ecological and evolutionary scales, respectively). These findings clash with public opinion, which regards the release of the fish as an inexorably positive event that can only help the recovery of the environment. In Brazil, the operational rules for the production of fingerlings for stocking are, frequently, the same used for their production for fish farming, and this production is, in general, implemented concurrently in the same hatchery. However, it should be noted that changes in gene frequencies are inevitable in the manipulation of breeding and rearing of wild animals in captivity. These changes, which are well documented for domestic livestock, arise from the breeding of individuals with a high degree of parentage and/or reduced effective population size (inbreeding), the crossing of genetically divergent fish (outbreeding), or artificial gene selection by favoring characteristics adapted to the rearing environment either during the maintenance of breeders or during the development of eggs, larvae and fingerlings (domestication; Flagg & Nash 1999). Therefore, the selection of breeders based on traits linked to the production in fish farms or maintenance of a reduced group for a prolonged time can contribute to the homogenization of the cultivated stock, distancing it from the wild gene pool (inbreeding). The genetic variability of the broodstock, which is usually very low (Calcagnotto & Toledo-Filho 2000), leads to reduction in the genetic variability of the wild population when continuous stocking is performed, because gene exchange between the groups is inevitable. In cultivation, it is also common the broodstock

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be close relatives of each other and this can increase the incidence of anomalies in the development of fingerlings, as well as interfere in the survival rate and growth rate of the fish. As the environmental pressures in captivity greatly differ from the pressures to which fish are subjected under natural conditions, the production of some generations in captivity is enough to alter the gene pool of the domesticated group and, consequently, to reduce its biological performance in the natural environment (Ford 2002; Caroffino et al. 2008). All of this indicates that, if there is gene exchange between domesticated and wild fish, the sustainability of the population will be negatively affected (Hansen 2002). The production and stocking of a given species in a basin from breeders obtained in another basin was a common procedure in Brazil until recently. Although the effects of this stocking have not been investigated, it is known that spatially isolated populations that are subjected to different selective pressures can, during the evolutionary process, have a gene pool that is adapted to local conditions, albeit with potentially poor performance in another basin, even if it is within the area of natural distribution of the species. The release of individuals from naturally distinct populations leads to the problem of outbreeding depression because it affects the viability and fertility of the receiving population, decreasing the fitness of individuals. Reproduction with individuals from remote populations can dilute the effects of locally advantageous alleles through the influx of new alleles; this possibility is especially critical if this event affects co-adapted “gene complexes” (combination of locally adapted genes). For example, hybrids of pink salmon Oncorhynchus gorbuscha were made between females from Auke Creek (Alaska) and Pillar Creek (Kodiak Island) using as control, crossing between males and females of the same creek. Parentage assignment from microsatellite analysis was used to improve estimates of survival. The hybridization reduced return rates of adults in the F1 generation and decreased the survival in F2 (Gilk et al., 2004). The crossing among populations of largemouth bass (Micropterus salmoides) from different basins (Mississippi River basin X Great lakes basin) also showed the effect of outbreeding depression. The increasing in the mortality rate was 3.6 times more in the generation F2 than in F1 and in native fishes, caused by susceptibility to infectious disease (Goldberg et al. 2005) Finally, it is worth noting that knowledge of the carrying capacity of the environment and of the size of the wild stock are basic concepts that should guide the need for stocking (Cowx 1999). However, these aspects have been systematically ignored by stocking programs, based on the rationale that the addition of fish to the system is always beneficial to fishing practices (Agostinho et al. 2008). In the case in which the environment does not support a surplus population, there is a risk of strong demographic changes (Molony et al. 2003; van Zyll de Jong et al. 2004). The added contingent can compete with the resident fish (Vehanen et al. 2009), increasing the mortality rate, decreasing the growth rate or leading to resource depletion. If the stocked species is a predator,

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there is risk of an exacerbated increase in the consumption of native invertebrates and fish, altering the organization of trophic webs (Skov et al. 2002). There is no evidence that these problems are occurring in Brazilian reservoirs due to a lack of relevant research and monitoring. Nevertheless, considering that reservoirs have their carrying capacities determined by the littoral zone, which is proportionally small when compared to the entire reservoir surface, (Agostinho et al. 1999), there is an increased risk that the ecological interactions intensify via stocking.

species were stocked showed, however, that there was no relationship between the stocking effort with these species and the fishing yield; greater yields had actually arisen from native species that had not been the object of stocking (Companhia de Energia do Estado de São Paulo 1996; AES Tiête 2007). These results caused the use of non-native species to be considered inadvisable in events promoted by the “Comitê Coordenador das Atividades de Meio Ambiente do Setor Elétrico Brasileiro” (Coordinating Committee for Environmental Activities of the Brazilian Electricity Sector; see a summary in Reuniões Temáticas: ações em http://www. eletrobras.com/elb/data/Pages/LUMIS187BD838PTBRIE. htm), while stocking with native species was recommended, albeit with reservations.

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The Brazilian Experience The first stockings in public waters in Brazil were conducted in the Northeast by the “Comissão Técnica de Piscicultura do Nordeste” (Technical Commission on Pisciculture of the Northeast), which was later called the “Diretoria de Pesca e Piscicultura do Departamento Nacional de Obras Contra as Secas” – DNOCS (Board of Fishing and Pisciculture of the National Department of Works Against Drought – DNOCS), more specifically on August 14, 1933, in the Campos da Sementeira Dam, in the town of Arcoverde, in Pernambuco State (Gurgel & Nepomuceno 1988). Repeated stockings in this region obtained self-sustaining populations of non-native species, and improved fish yield, especially for tilapias (Paiva et al. 1994). The Northeast strategy, especially with non-native species, has disseminated to other regions in Brazil and became the main fishing management method practiced by fishery development agencies and hydroelectric power companies throughout the 20th century. However, many of these were inefficient or constituted an additional threat to biodiversity (Agostinho et al. 2004, 2007b). In the other regions of Brazil, stockings were usually considered as part of initiatives to mitigate the impacts of dams on fishing resources, and they were supported by positive public opinion (common sense) and had a strong political-electoral appeal (Agostinho et al. 2005). This same common sense explains the popular acceptance of stocking as a compensatory measure for impacts and the compulsory character of its use as a punishment in cases of environmental law offenses. Stocking activities are also common as part of holiday celebrations because they are considered an important strategy for environmental education. Until 1990, non-native species were predominant in stocking programs in the South and Southeast regions of Brazil. Under the claim that urgent measures should be taken and that they could not wait for the results of the study with native species, more than one dozen species of fish from other basins, including other continents, were stocked in reservoirs (Agostinho et al. 2007a). Some of them achieved success in colonization and are currently disseminated in these basins (e.g. Plagioscion squamosissimus and Cichla spp.), while others, although not disseminated, are locally abundant (e.g., tilapias, Astronotus and Triportheus). Monitoring data on commercial fishing landings in reservoirs where exotic

Stocking actions in Brazil were historically performed based on limited knowledge both of the system to be managed and of the species to be stocked, as well as the need for this action. Furthermore, inexperience in conducting the stocking (species and quantity needed, as well as the location, size and time of release, etc.) led to the practice of “trial and error”; however, without monitoring it was not possible to learn from these practices (Gomes et al. 2004; Agostinho et al. 2004, 2007a; Pelicice et al. 2009). Therefore, neglecting the genetic quality of breeders and the possibility of negative impacts on natural populations has made this management activity a potential and constant threat to local populations and to fishing itself - although such consequences have never been assessed by empirical studies. Data from Quirós (1999) on stocking rates and yields in more than 700 ponds, reservoirs and lakes around the world show that, for large reservoirs, where stocking is generally supplementary, the yield is naturally low and the response to this effort is of little relevance. Although this author does not emphasize this fact, the comparison of large and small reservoirs from the Upper Paraná River basin shows similar tendencies (AES Tietê 2007). According to Quirós (1999), small and medium sized reservoirs that presented high yields were stocked with densities between 500 and 800 ind.ha–1.year–1. In reservoirs from the Paraná River basin, this value varied between 6 and 30 ind.ha–1.year–1, that is well below what is needed. Therefore, a large reservoir, like Água Vermelha (64,700 ha), in the Grande River basin, should have a carrying capacity proportionally similar to that of a medium or small reservoir, and should be stocked with 32 to 52 million fingerlings per year, to show a satisfactory response to the stocking effort. The number of fingerlings from three species released in this reservoir for a decade was approximately one million (AES Tietê 2007). Considering that Brazil has 3,400,000 ha of impounded water (Agostinho et al. 2007a), the quantity of fingerlings to be produced would be unreachable, and attempts to reach this number could have severe environmental consequences. In the Upper Paraná River, more promising results have been registered in reservoirs with smaller areas, with a relationship between the quantity of stocked individuals and the fishing yield, as observed for Piaractus mesopotamicus

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Reservoir Fish Stocking

(Pacu), Prochilodus lineatus and Leporinus elongatus, in some reservoirs located in the Pardo River (AES Tiête 2007). Nevertheless, positive responses to the stocking effort were only apparent when the weight of released fingerlings increased from 8 to 25 g for P. mesopotamicus and from 6 to 18 g for P. lineatus (Belmont et al. 2004). However, the carelessness in or lack of assessment of stocking results is the main factor that enabled innocuous programs to persist for more than 50 years and perpetuate the waste of resources, efforts and opportunities, as shown by the low fishing yields in Upper Paraná River reservoirs and the precarious conservation state of stocks in the main tributaries of this stretch of the river.

Strategies For Stocking Given the irreversible character of many of the effects of impoundment on the ichthyofauna (Agostinho et al. 2008), it is expected that the discussions about the need for stocking in reservoirs will extrapolate the economic and social

dimensions, and it will be necessary to effectively consider the ecological dimension. Special care is expected in the prevention of potential impacts that the stocking actions may promote. An environmentally friendly, economic and socially desirable stocking strategy should consider a sequence of procedures in which prerequisites cannot be ignored and reliable prior knowledge is essential. The minimum background information needed to the stocking process is shown in Figure 2.

Deciding process The first requirement to decide to use stocking must be the evaluation of need, followed by economic, social and environmental viability. This decision requires comprehensive knowledge of all system components (environment, fish population, fishermen, risks, etc. - Figure 2), including environmental or artificial factors that lead to stock or population depletion. During the discussions that precede the stocking decision, clear and quantifiable objectives must be outlined to provide

Figure 1. Summarized representation of the entire stocking process, showing the main impacts (dashed diagrams on the right) discussed in the text.

Figure 2. A conceptual model showing the complex information needed to conduct sound stocking.

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Agostinho et al.

criteria for the evaluation of its effectiveness. Thus, it is of fundamental importance to consider other management strategies (e.g., fisheries control, habitat management or doing nothing; Agostinho et al. 2007a) and to consistently establish the reasons for stocking. Although the depletion of a given population or stock can require management measures, the option for stocking may be necessitated in particular cases, such as events of increased mortality due to fishing (overfishing) or failures/insufficiencies in recruitment (due to climate issues, degradation in spawning and initial development habitats). The increase in carrying capacity due to input of nutrients (eutrophication) can also indicate a situation in which stocking may be recommended. The assessment of stocking viability requires, besides the knowledge of the factors that limit the stock, the recognition that only some stocks and environments have the potential to respond to the stocking effort and that the impacts of this activity on the target stock and the ecosystem can be high and, sometimes, irreversible (changes in community structure, disease dissemination and losses of genetic integrity; Cowx 1999). Therefore, decisions on conducting stocking are not easy. Below, we present a diagram to subsidize these decisions (Figure 3).

Hatchery Ideally, hatcheries should be defined as recommended by Flagg & Nash (1999), that is, facilities designed to breed and disseminate a stock of fish with genetic resources equivalent to the native stock and with an unaltered ability to naturally reproduce in its original habitat. These authors believe that this notion of a hatchery still does not exist in the world and suggest that the great challenge is to match fish production strategies with those that reduce the risk of supplementary stocking. Therefore, the success of stocking and its ecological viability are intricately linked to how the fish in this method of management are produced. The genetic effects of domestication on the reduction of the reproductive capacity of species has been analyzed by Araki et al. (2007), who estimated this reduction to be around 40% per captive-reared generation when fish are released in the natural environment. Therefore, the environmental responsibility that must guide stocking programs recommends that some guidelines should be adopted by managers to operate hatcheries, especially those indicated by Flagg & Nash (1999), namely: •

to provide fish with minimal genetic divergence from their natural counterparts to maintain the long-term

Figure 3. A decision diagram showing the links among all types of stocking, which may serve has guidelines for decision on conducting or not conducting stocking.

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adaptive characteristics, using a group of regional breeders that is sufficiently diversified; •

to manage the broodstock to maintain the natural seasonality of gonadal maturation events, ensuring high quality gametes and minimizing the early maturation of males;

•

to manage the incubation process and the characteristics of the incubator, there must be options to comply with the complexity of the habitat to produce fish without selection, and with natural appearances and behaviors, as well as high survival rates;

•

to establish specific targets for growth patterns that are similar to the natural patterns;

•

to use low densities of fish in the production process to improve survival and avoid selection; and

•

to have the option of applying anti-predatory conditioning methods during the production of fingerlings and juveniles.

We do recognize that it is difficult to follow all the guidelines presented above, but managers must pursue achieving them.

Releasing Among the factors that have received little attention in stocking programs are those related to the quantity and size of the fish to be released, as well as the location and time of release. In an ideal situation, these variables would be defined based on knowledge of the life cycle, distribution and structure of the natural population, ideally after pilot studies (Molony et al. 2003). Therefore, the size of the fish to be released must be defined within the series of sizes registered in the wild population, except for cases in which the wild population is facing imminent extinction and requires higher survival rates (Flagg & Nash 1999). Larger fish are more costly to produce; this expense could be compensated for by the higher survival rate after release. However, the greater time commitment required for the cultivation increased the probability of domestication selection and the development of behaviors that are not adequate for the natural environment, such as schooling, increases in naivety or loss of competiveness (Molony et al. 2003). The choice of the release location, on the other hand, must consider information about the type of habitat in which fish from the same ontogenetic phase occur naturally, which will eventually lead to releases in distant locations from where adults occur. This also implies that release locations must not be chosen based on accessibility, a criterion that has guided stockings in reservoirs (e.g. close to bridges, beaches and margins). The time of release, which like other variables, depends on the species and its life cycle, is influenced by the size or phase in which the fish must be released. Other

factors to be considered are the environmental conditions and availability of adequate food. The quantity of released fish should, ideally, be established based on the carrying capacity of the receiving body of water and avoid exceeding it. It is necessary to recognize that productivity in natural systems has limits (Wiley 1995) and that the biogenic capacity in reservoirs is usually restricted to the littoral zone. Furthermore, ecosystems are not static and instead show considerable variations in their biogenic capacities, suggesting the need to consider the fluctuations in carrying capacity to better estimate stocking effort, because these fluctuations could allow the substitution of the wild stock with the stock produced in the hatcheries (Pearsons 2010).

Monitoring The stocking programs must be evaluated and monitored so that changes and improvements are incorporated or so that the need for abandonment is detected. Therefore, monitoring is an integral and inseparable part of the stocking action and must be clearly and consistently outlined in the decision-making stage of this mode of management. Stocking should not be considered without clearly defining the assessment method for its effectiveness and possible impacts. In this evaluation, it is crucial to distinguish and quantify the capture rate for fish from stockings and from natural recruitment, using some of the many available marking techniques (Molony et al. 2003). The monitoring of fish in hatcheries (e.g., monitoring of their genetic, morphophysiological, behavioral and health traits, in addition to their origins) and the detailed recording of the locations and dates of releases, as well as the quantity and size of released fish, are indispensable variables for explaining the results of stocking programs.

Final Considerations Stocking has wide popular acceptance as the most complete solution for the recovery of depleted stocks. In addition to how easily understood stocking is, the interest of the media in this topic and the existence of technology for the production of a large number of fingerlings of several species contributes to this perception (Molony et al. 2003). However, studies on fish stockings in tropical reservoirs are rare in the literature and are generally restricted to cases of massive releases of fish, often including non-native species. Although this can be due to a lack of assessment of stocking for supplementation, the fact that recurring failures are not recorded needs to be considered. Given the risks associated with stocking, before this mode of management is adopted, an in-depth assessment of other alternatives is recommended (e.g., habitat management and fishing control) based on the factors that lead the stock to depletion. When stocking method is considered appropriate and is believed to have acceptable impact levels, this strategy should be adopted in combination with improvements in the habitats and regulation of exploratory activity. The same

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way that a broad initial knowledge of the system is necessary for the decision to pursue stocking for supplementation, a detailed record and assessment of procedures are vital after implementation. Increases in fishing landings, though they may be adequate to managers and the public, cannot be used as an indication of the effectiveness of these measures from the perspective of resource conservation.

Cambray JA, 2003. Impact on indigenous species biodiversity caused by globalization of alien recreational freshwater fisheries. Hydrobiologia, 500:217-230.

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Acknowledgments AAA and LCG thank CNPq for the Research Grant (“Bolsa Produtividade em Pesquisa”) provided.

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Hansen MM, 2002. Estimating the long-term effects of stocking domesticated trout into wild brown trout (Salmo trutta) populations: an approach using microsatellite DNA analysis of historical and contemporary samples. Molecular Ecology, 11:1003-1015. Hindar K, Ryman N & Utter F, 1991. Genetic effects of cultured fish on natural fish populations. Canadian Journal of Fisheries and Aquatic Sciences, 38:1867-1876. Hoeinghaus DJ, Agostinho AA, Gomes LC, Pelicice FM, Okada EK, Latini JD, Kashiwaqui, EAL & Winemiller KO, 2009. Effects of river impoundment on ecosystem services of large tropical rivers: embodied energy and market value of artisanal fisheries. Conservation Biology, 23:1222-1231. Johnson PTJ, Olden JD & Vander Zanden MJ, 2008. Dam invaders: Impoundments facilitate biological invasions into freshwaters. Frontiers in Ecology and the Environment, 6:357-363. Latini AO & Petrere Jr M, 2004. Reduction of a native fish fauna by alien species: an example from Brazilian freshwater tropical lakes. Fisheries Management and Ecology, 11:71-79. Matsumoto CK & Hilsdorf AWS, 2009. Microsatellite variation and population genetic structure of a neotropical endangered Bryconinae species Brycon insignis Steindachner, 1877: implications for its conservation and sustainable management. Neotropical Ichthyology, 7:395-402. Molony BW, Lenanton R, Jackson G & Norriss J, 2003. Stock enhancement as a fisheries management tool. Fish Biology and Fisheries, 13:409-432. Paiva MP, Petrere Jr M, Petenate AJ, Nepomuceno FH & Vasconcelos EA, 1994. Relationship between the number of predatory fish species and fish yield in large northeastern Brazilian reservoirs. In: Cowx I (Ed.). Rehabilitation of freshwater fisheries. Oxford: Fishing New Books. p. 120-129. Pearsons TN, 2010. Operating hatcheries within an ecosystem context using the adaptative stocking concept. Fisheries, 35:23-31. Pelicice FM & Agostinho AA, 2008. Fish-passage facilities as ecological traps in large neotropical rivers. Conservation Biology, 22:180-188. Pelicice FM & Agostinho AA, 2009. Fish fauna destruction after the introduction of a non - native predator (Cichla kelberi) in a Neotropical reservoir. Biological Invasions, 11:1789-1801. Pelicice FM, Agostinho AA & Agostinho CS, 2009. Conservação da ictiofauna na área de influência de Peixe Angical e recomendações ao manejo. In: Agostinho CS, Pelicice FM & Marques EE (Ed.). Reservatório de Peixe Angical:

Rahel FJ, 2007. Biogeographic barriers, connectivity and homogenization of freshwater faunas: it’s a small world after all. Freshwater Biology, 52:696-710. Resende EK, Marques DKS & Ferreira LKSG, 2008. A successful case of biological invasion: the fish Cichla piquiti, an Amazonian species introduced into the Pantanal, Brazil. Brazilian Journal of Biology, 68:799-805. Rodriguez-Rodriguez MP, Lopera-Barrero NM, Ribeiro RP, Povh JA, Vargas L, Sirol RN & Jacometo CP, 2010. Diversidad genética de piracanjuba usada en programas de repoblación con marcadores microsatélites. Pesquisa Agropecuária Brasileira, 45:56-63. Santos GB, Maia-Barbosa PM, Vieira F & López CM, 1994. Fish and zooplankton communities structure in reservoirs of southeastern Brazil: Effects of the introduction of exotic predatory fish. In: Pinto-Coelho RM, Giani A & von Sperling E (Ed.). Ecology and human impact on lakes and reservoirs in Minas Gerais with special reference to future development and management strategies. Belo Horizonte: Segrac. p. 77-83. Santos LN, Gonzalez AF & Araujo FG, 2001. Dieta do tucunaréamarelo Cichla monoculus (Bloch & Schneider) (Osteichthyes, Cichlidae), no Reservatório de Lajes, Rio de Janeiro, Brasil. Revista Brasileira de Zoologia,18:191-204. Skov C, Perrow MR, Berg S & Skovgaard H, 2002. Changes in the fish community and water quality during seven years of stocking piscivorous fish in a shallow lake. Freshwater Biology, 47:2388-2400. Van Zyll de Jong MC, Gibson RJ & Cowx IG, 2004. Impacts of stocking and introductions on freshwater fisheries of Newfoundland and Labrador, Canada. Fisheries Management and Ecology, 11:183-193. Vehanen T, Huusko A & Hokki R, 2009. Competition between hatchery-raised and wild trout Salmo Trutta in enclosures – do hatchery releases have negatives effects on wild populations? Ecology of Freshwater Fish, 18:261-268. Vieira F & Pompeu OS, 2001. Peixamentos - uma alternativa eficiente? Ciência Hoje, 30:28-33. Vitule JRS, Freire CA & Simberloff D, 2009. Introduction of non native freshwater fish can certainly be bad. Fish and Fisheries, 10:98-108. Volpato GL, Barreto RE, Marcondes AL, Moreira PSA & Ferreira MFB, 2009. Fish ladder select fish traits on migration – still a growing problem for natural fish populations. Marine and Freshwater Behaviour and Physiology, 42:307-313. Wiley RW, 1995. A common sense protocol for the use of hatchery-reared trout. American Fisheries Society Symposium, 15:465-471. Zaret TM & Paine RT, 1973. Species introduction in a tropical lake. Science, 182:449-455.

Received: August 2010 First Decision: August 2010 Accepted: September 2010

Essays & Perspectives

Brazilian Journal of Nature Conservation

Conservation Crossroads and the Role of Hierarchy in the Decision-Making Process Adrián Monjeau Instituto de Análisis de Recursos Naturales, Universidad Atlántida Argentina & CONICET, Mar del Plata,Argentina

Abstract This essay examines several crossroads, paradoxes and controversies in conservation politics in Latin America: populated and non-populated protected areas, local versus global forces, and the role of the national government in making long-term, ecologically correct decisions versus short-term politically correct local decisions. Ecologically sound projects at a global scale, such as the maintenance of the ecosystems working order, exceed the lifetime of the present generation. In addition to this, as decisions in a particular area may have ecological consequences that go beyond the sphere of that area, the responsibility cannot be delegated to the local management level. Local consensus is essential to implement conservation goals on the ground, but it should never be opposed to global priorities, especially because this antagonism puts the ecosystem working order at risk. In this ranking, the hierarchical organization of the decision making process, from global to local, is crucial, so that the State retains its organizational role while working along with the local forces in their effort to implement conservation. Key words: Protected Areas, People & Parks, Politics, Biodiversity, Global versus Local.

Introduction The huge demographic growth and burst of western techno-capitalism has given place to a world map in which the human footprint has reached and domesticated almost every single patch of fertile land (see Sanderson et al. 2002). This devastating force has homogenized and simplified landscapes across the globe and made them accessible to the language of multiplication of capital. This landscape transformation is contributing to ever faster declines in species and the system they depend on (Agrawal & Redford 2007). Most of the world’s diversity, be it in biology, wildlife, ethno-linguistic groups and other minorities that still survive landscape transformation, is fenced into the last corners of the planet, where the global market routes have not reached yet. For this reason, a great part of the areas with the highest biological diversity also enclose the largest linguistic diversity (Toledo 2001, 2005). In other words, several non-protected wilderness areas in developing countries still survive thanks to their inaccessibility (Monjeau 2007). As a result, whether those wildernesses will survive or not depends on road construction. On behalf of the progress, soon remoteness will be a thing of the uncivilized past. Resources of all *Send correspondence to: Adrián Monjeau Instituto de Análisis de Recursos Naturales, Universidad Atlántida Argentina y CONICET, Arenales 2740 (7600) Mar del Plata, Argentina E-mail: amonjeau@parkswatch.org

kinds will be exploited to exhaustion unless restrictions are vigorously imposed (Terborgh 1999). Natural landscapes in arable land will disappear. There are many examples of what has happened in remote areas where the global market has been able to extend its reach. It is in these places that non-protected ecosystems are now strongly altered, while strictly protected areas have demonstrated their capacity to resist landscape transformation (Bruner et al. 2001). If it were not for the pioneering efforts of a few people capable of understanding the future, today we would be unable to enjoy the national parks that harbor the last bastions of wilderness on Earth (Soulé & Terborgh 1999).

Are Protected Areas Enough to Stop Extinction? Even though protected areas are acknowledged as the most efficient and necessary tools to safeguard the last remnants of wildlife (Bruner et al. 2001, Brooks et al. 2006 ; Kramer et al. 1997), they are not enough to stop global losses in biodiversity and wild places. Conservation scientists have realized that protected areas, if isolated, cannot achieve their objective of sustaining natural functioning and diversity (Newmark 1987; Belovsky 1987; Soulé & Terborgh 1999). In order to achieve this objective, it is necessary to manage large interconnected extensions dedicated to conservation, managing ecosystems as integers (Soulé & Terborgh 1999; Toledo 2005; Interagency Ecosystem Management Task Force

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1995). Consequently, effective biodiversity conservation must include conservation outside the boundaries of protected areas (Allison et al. 1998; Soulé & Terborgh 1999). These goals are coupled to problems that exceed the reach of biology, since the management of entire ecosystems implies the accommodation of the best possible mosaic between stakeholders of diverse interests that overlap in a very complex way the objectives of conservation (Brandon 1998, 2002; Brandon & Wells 1992; Brandon et al. 1998; Ritcher & Redford 1999; Salafsky & Wollenberg 2000; Wondolleck & Yaffee 2000). Consensus is the most desirable attitude to achieve these conservation objectives at an ecosystem scale. This does not mean, however, that a consensus will coincide exactly with the ecologically correct decisions in every single grid of the mosaic, since the consensus finds its limits where the diffuse collective interests of stakeholders imply restrictions on individual ones (Peterson et al. 2005). This is a commons problem. Paraphrasing Garrett Hardin (1972), we could argue that the ‘I-here-now” has a vast preference over the weaker “All-there-after”. This tragedy of the commons has ultimately structured the present world map and its tendencies. The current biodiversity crisis is caused in part by the accumulation in time and space of zillions of little local decisions that disregard further global consequences. Taking this statement into account, we believe that the State, that represents us all everywhere in the long term, should be the integrating tool for the above-mentioned mosaic, considering that its main objective is the regulation of individuals for coexistence to be possible. It has been demonstrated that internal and external pressures will eventually transform a natural area which is under no conservation management of any sort. The main issue would be how much of the wild areas is actually safe from the pressures of transformation. In a recent study (Monjeau 2007) it was calculated that the percentage of protected areas under strict conservation management without pressure from their surroundings is in the order of 8% (a total of 1511 protected areas were studied) in South America. This is obviously in short supply to carry out functions such as preservation of biological diversity, endemisms and ecosystem processes crucial to human and natural economy: soil fertility, global temperature regulation and water provision. If the remaining 92% depend – to any degree – on achieving a consensus between conservationist agendas and the surrounding social entities and villagers, one core question of this essay would be to what extent the State can uphold a set of correct ecological decisions vis-à-vis such a consensus.

Politically Correct or Ecologically Wise? The problem is not simple. There seems to be an inverse relation between what is politically correct and what is ecologically acceptable. In today’s world, nature conservation movements find themselves at a crossroads, where two

extreme myths or beliefs are dialectically opposed (Monjeau & Solari 2008; Terborgh 1999; Terborgh et al. 2002; Schwartzman et al. 2000; Wilshushen et al. 2002): 1) it is impossible to establish effective conservation if people are involved; 2) it is impossible to establish effective conservation if people are not involved. Both assumptions coexist under pressure inside the core of the conservation paradigm because they are only partly correct. Arguments in favor of myth 1 are built upon the belief that human presence invariably impacts nature negatively. Arguments in favor of myth 2 suggest that a harmonious and sustainable relationship is the only way that conservation goals can be achieved. It is not hard to guess that the coexistence of these two approaches generates a great amount of confusion in conservation policies. This gives way to the bewildering fact that the engine driving the history of conservation has become dialectical (sensu Hegel 1966) between abuse and inexperience.

The Abuse Factor The abuse factor is clearly exemplified by the first blunders committed during the periods of land occupation and subsequent territorial designation (Redford & Fearn 2007). Protected areas – preceded by a history of usurpation and even genocide at a regional scale – find themselves in conflict with the rights of the original inhabitants who were forcedly displaced from them. At this end of the spectrum, the unviability put forward by myth 2 is understandable, since in the memories of the usurped there is a “karma” against conservation as a symbol of imposition by a foreign enemy who expelled them from their native lands. To make the issue even more complex, due to the fact that just a few elite or rich people have been displaced as the result of the creation of protected areas, “the usurped linked conservation with a concern of the wealthy and the powerful” (Brosius 2007). Given that perception, displaced people or people with restricted access to local resources may have strong incentives to destroy wildlife and resources in protected areas (Brandon 2002; Delrio 2005; Brosius 2007). As Brosius (2007, p. 106) puts it: [...] In several forums, local community advocates have put forward propositions for organizing reconciliation commissions which inquire into past injustices promulgated in the name of conservation, and the idea of restitution is receiving increasing attention.

The pressures to return to the use of the natural resources in their native forests, valleys and prairies (also understandable) swell up in the public consciousness to a point where myth 2 becomes valid. Recently, WWF (Worldwide Fund for Nature 2006) published an assessment of the case in which humans and wildlife species can coexist and prosper. In the same avenue of pressures against myth 1, conservation has become the target of indigenous movement. In their discourses “they have come to equate conservation with the extinguishing of rights, the latest in a long line of attempts to dispossess them”, as Brosius (2007) puts it. In most of

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the wilderness settings where many important protected areas and peripheral settlements are located, conservation success is likely dependent on local acceptance or resistance (Redford & Fearn 2007). This is another argument to validate myth 2. However, when the State grants the lands, conservation goals are not part of their subsistence practices (again, understandable as well as justifiable). The empirical evidence shows that a free-for-all human use is generally not compatible with biodiversity conservation. Bottom-up decision making processes will not improve the security of parks if they rely a hundred percent on voluntary compliance (Terborgh 1999). There are many examples of the failure of voluntary compliance to produce the desired results. For instance, Agrawal & Redford (2006) survey of 37 projects which attempted joint achievement of biodiversity conservation and poverty alleviation found little evidence in favor of synergies between human use and conservation. These facts tend to re-validate myth 1, and so the dialectic continues.

alleviation, conservation organizations are shifting the message persuading donors that conservation can be accomplished together with poverty alleviation (Wells & McShane 2004) and that biodiversity conservation is important in ecosystem services for human well being in the long run (Burton et al. 1992). Under these new banners of “conservation and human needs”, conservation groups have had impressive rates of financial growth (Terborgh 1999).

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The Inexperience Factor The inexperience factor consists in presuming that the debts of abuse can be paid merely by putting the conservation goals of protected areas in the hands of their original inhabitants or local communities (the “positive-historical” conservation action in Agrawal & Redford 2007). This is a short-term demagogic approach, politically naïve, that compromises the public’s right to live in a healthy environment, endangers the planet’s inhabitability and – worst of all – does not repay the debt of a shameful injustice of continental dimensions! To cover up errors with more errors does not seem to be the wisest road to a fair, ethnically integrated society that is in concord with its ecosystem; it is more like a road to the political positioning of a few opportunistic individuals who take advantage of the dynamics of consensus. In this way they favor the process of deconstruction of the State towards territorial tribalization. The confluence of critiques that blunt the ethical focus of biodiversity conservation highlights the distress that conservation goals impose to human populations. This ethical side of the criticism have gained a great deal of attention in international forums such as the World Park Congress (see Terborgh 2004, 2005 versus Andrade 2005), World Conservation Congress and Convention on Biological Diversity Conference of the Parties, and in international conservation organizations like WWF, CI, TNC, WCS, and IUCN. This is one of the most critical problems confronting conservation worldwide. However, whether the restrictions imposed to local populations are justified or not on behalf of global, longterm conservation goals is perhaps alongside this point: the vociferous human rights criticism regarding a lack of synergy between conservation and social goals has undeniable negative impact on conservation funds. Because available funds for conservation have declined in comparison to the increasing emphasis on social equity and poverty

Likewise, a confluence of the major international donors (under the pressure to adopt greener policies) and conservation organizations (wanted to be seen by these donors as “social sensitive”) converged in the form of integrated conservation and development projects, called ICDPs. The contribution to conservation claimed by the ICDPs is to reduce external threats to parks by promoting sustainable development outside the protected areas. However, under the “socio-green” shield, the ICDPs contain a mixture of inexperience working together. On one hand, big international donors, like the USAID, have little experience in conservation. On the other hand, conservation organizations are inexperienced in managing international development projects. Conservation projects have been spending millions of dollars in economic assistance to local villagers and promoted development with little benefits for conservation. On the contrary, due to the synergy between two processes, the inexperience factor under discussion may be the cause of an “eutrophication effect” negative to conservation: 1) Improvements in the livelihood of villagers inside or near parks stimulate migration from other places, population density increases, and consequently, pressure on the park´s resources increases. 2) Local villagers might voluntarily respect park boundaries and other restrictions if they could raise their standards of living through means other than exploitation of the park´s resources. Nevertheless, alternative economic activities supported by external sources of money may stop when the source of money is discontinued. Then, the hope of a better life on behalf of conservation disappears in no more than the blink of an eye. Millions of dollars later, the situation is worse than it used to be: a much bigger population which is unwilling to believe in new promises and claiming for the use of the park´s resources as they did before. As far as we can see on the ground, there is a false link between ICDPs and conservation. If the goal is conservation of biodiversity, investment in local people is the wrong target.

Who Will Advocate for the Defense of Nature? This uncertainty between the right way and the convenient way is another tectonic movement in the philosophical roots of conservationism. Unfortunately, if conservation organizations are reluctant to go where criticism wishes them to go, the financial forces will push them into putting more emphasis on the social issues than on biodiversity

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conservation priorities. The question is: If most of the conservation organizations are resigning their conservation goals on behalf of social issues (the sphere of the wellfunded aid organizations together with governments), in whose hands will the future of biodiversity be? Who will advocate for the defense of nature? Should the leading conservation organizations change their oxymoronic names? This trend is continually gaining traction and if conservation communities do not address this negative momentum with concrete proposals, the consequences for global conservation could be devastating. Unquestionably, conservation has entered a new age. As a consequence, it is imperative to clarify which are and which are not real conservation goals in order to invest wisely only on projects which are beneficial for conservation. Nowadays, a “Hegelian end to history” is being pursued, resolving the dialectic with a paradigm shift on the priority of functions that a protected area must perform. The protectionist paradigm, which defends the necessity of strict conservation, is being replaced by a paradigm that is more socio-environmentalist and prioritizes the social functions of conservation under the main banner of sustainable development (Brechin et al. 2002). Alongside these lines, the idea has therefore been to put forward that strict protected areas are not viable in situations of high social pressures because people do not want them. This argument is as dangerous as stating that traffic lights are not viable at crossings where traffic is excessive and consensus is that they should be removed because people do not want to stop at a red light. As a consequence of this predisposition, the promises of the so-called sustainable development are being carried out in areas under the category of “managed resources”. In most cases, buffer or reserve areas are open to human exploitation. As a result, they cannot be differentiated (as pertains to conservation objectives) from purely productive establishments outside the parks. This is an important point: if sustainable development is useful to correct a pre-existent development inside the boundaries of protected areas mitigating the environmental damages and makes the achievement of sustainability more efficient than before, then… congratulations. But if sustainable development is masked as a new panacea to be able to dig into previously unavailable natural resources, then the ally becomes the traitor. If the socioenvironmentalists were complying with their work outside the boundaries of protected areas ever since the first protectionists began to comply with theirs in the 1890’s, a significant percentage of the Earth’s surface would be now sustainable. Given that, the socioenvironmentalists would not be by now knocking on the door of the protected areas and claiming the use of resources that have survived the debacle. The debacle, of course, is the result of the fact that the natural resources preserved into the parks are surrounded by millions of cases of failed natural resources management.

The conflict between local people and conservation is intrinsically a spatial problem, say, a problem of territorial ordination of human uses: Protected areas should sustain the functions they were created for, and the supposed objectives of sustainable development should be implemented across the rest of the planet. But this is not the case: the rest of the planet is being devastated at an alarming rate and protected areas are gradually relaxing their limits. The decay of wildlife is not a necessary sacrifice for society’s wellbeing; let us not kid ourselves: the devastation of natural resources advances as swiftly as hunger, inequality and poverty. After billions of dollars poured on behalf of poverty alleviation, the current world shows little progress on that matter. If the world was, is and will be full of injustice, we must then ask ourselves why effective protected areas -a tiny fraction of the Earth surface- are now linked with the cause of human misery.

Conservation Crossroads and the Vagueness of the Language Conservation crossroads cannot be solved in part because of the vagueness of the language. As Wittgenstein (1999) puts it, the interface between words is a source of unquiet thinking, full of false meanings, oxymorons and ambiguity that cause false interactions among concepts. As in the times of “evangelization”, nature is again being pushed towards the unproductive corners of the world on behalf of discourses that are camouflaged with beautiful words such as equality, welfare, democracy, humanity. They are in fact an elegant justification for the scavenging and greed of the globalized market. It is largely in “areas of freedom” that the human footprint has advanced the most. If natural areas become “freed” from the care of the State, how long they will survive is directly proportional to the time taken by the “free” markets to reach their resources (and the ethnic groups that inhabit them). With the tremendous confusion that nowadays exists, the daft idea (false link) that indigenous people are enemies of conservation has become quite trendy. On the contrary, both, cultural and natural diversity are threatened by the same enemy: it is a giant that advances at a huge step. It is so gigantic that it cannot be visible at a local scale. If it were up to this giant, there would not be a single square meter of land that would not have been seized. A striking example of a false link between words is the oxymoronic concept of “sustainable development”. Taking into account that the planet’s sustainable production capacity to feed the human species was reached and exceeded ca. 1978 (Wackernagel et al. 2002) – let alone the energetic demands of other species- rather than to “sustainable development” we should be referring to a “reasonable reversal”. Reversal is reasonable because it seeks to replace the impending catastrophic and irrational scenario typical of shipwrecks. It is essential to implement a program of reduction of the human footprint, both, in quantity and in intensity. This applies to us all.

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Are We “the World”? An environmentally friendly plausible human society must be much smaller and have a completely different attitude towards nature (Næss 1989). This consists in removing humankind from the epistemological centre (as Foucault 1985 requests) in conservationist philosophy. An exclusive concern with human development often leads to undesirable impacts on biodiversity conservation (Redford et al. 2006). There are portions of ecosystems where nature must be complete in order to function, and that either do not need or actually resist human presence (Terborgh 1999; Terborgh et al. 2001). We humans need space to sustain our wellbeing (and protected areas have much to offer), but it is unviable to presume that we need the entire planet for productive activities. It has already become evident (Wackernagel et al. 2002) that if we do not limit our humongous appetite, even an entire planet will not be enough. Human economy will not be possible either if we do not allow nature to function correctly (Sachs 2008), the dialectic is resolved when understanding that the only way to adapting to a plausible future is to cede spaces in a mosaic of coexistence (Toledo 2005).

The Green Leviatan At these crossroads that we are analyzing, we see that one of the main problems to carry out effective conservation in a mosaic of coexistence is the tension that exists between public and individual interests. Individuals prioritize their interests in the short term and in the reduced environment for the convenience of their group (Velázquez Delgado 2006). As we put it above, this – the Hardin´s tragedy of the commons – is one of the great problems in conservation biology. Hobbes’ assertion in the 17th century “homo homini lupus” is hard to refute when looking back into historical evidences (Hobbes 2006). If humans destroy each other in defense of their own interests (“bellum omnium contra omnes”), it would be ingenuous to expect a majority’s consensus to be concerned about other species for generations still to come or for what – roughly – might occur to the planet in future scenarios. It is very hard to get present generations to give up even a small amount of its present welfare for the sake of a better future. This behavior is “reasonable” if we base it on a demolishing argument: as the bearing of time as a variable is unidirectional, the present is in no way dependent on what might come to be in the future. In other words, what has the future done for us for us to have to do something for it? Given that the inverse relationship is opposite (in a cause-and-effect sense), the irresponsible behavior of past generations has led to “that future” and become present in our every day life (Cecchetto 2007). Ecologically sound projects at a global scale, such as the maintenance of the ecosystems working order, exceed the lifetime of the present generation (Green 1977). In addition to this, as decisions in a particular area may have ecological consequences that go beyond the sphere of that area, the responsibility cannot be

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delegated to the local management level. Local consensus is necessary to accomplish conservation goals, but it is not enough. It is evident that the State (in the democratic sense of res publica) must make decisions for the common good, at a global scale and in the long term, because future survival depends on this. Nevertheless, as Ortega y Gasset states in The Revolt of the Masses, the government lives the day hidden in the present, and avoids solving current conflicts imposed upon by the urgent circumstances of the moment, neither delving into the future nor constructing anything for the long-term survival, even if its opportunities to do so are vast. In addition, there are many governments subordinated to the power of multinational corporations and, therefore, conservation goals cannot be warranted in the hands of these countries. At this point we have another crossroads in conservation politics: How can responsibility be in the hands of the national government if this is unable to manage the resources in an efficient way? This argument - together with the recognition of the local people’s rights to use their natural resources and the indigenous people´s rising claims on the property of the territories they inhabit - is used to support the bottom-up against the top-down decision making. This is a wrong argument. At local level, decisions are made ad hoc, on a case-by-case basis, often under political pressure and driven by the demands of the moment. The ultimate challenge is counteracting the tragedy of the commons. Experience teaches that restraint in the use of renewable resources will not spring up from the bottom. The selfish demands of the clamorous few are promoted over the diffuse interests of a passive majority. Instead of looking downwards to find the solution, when the State is not reliable we have to look higher up: global and long-term conservation correct decisions must be in the hands of a partnership of national governments to take priority over the sovereignty of the member nations.

Where do We Go from Here? Local versus global and people versus nature crossroads are not necessarily antagonist forces if each State –in a global partnership- is able to rank the federal territory following priorities for long-term biodiversity conservation and ecosystem services without trespassing the primordial function of protected areas as a tool for conservation. As stated in the Principle of Subsidiarity in many national laws (Esain 2009), local consensus is essential to implement conservation goals on the ground, but it should never be opposed to global priorities, especially because this opposition puts the functioning of the ecosystem at risk. Given the hierarchical structure of nature (Bailey 1987; Klijn & Udo de Haes 1994; Monjeau et al. 1998), in which upper components (climate, geology) influence the features and functioning of the lower components (soil, vegetation), most of the ecological systems and their interactions are not

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Conservation Crossroads and Hierarchy

Table 1. Proposed scheme of decision-making that matches the hierarchical structure of natural components and processes with the hierarchy of political decisions. This scheme proposes the integration of local and regional forces following a top-down cascade of enforcement.

Scale

Natural components and processes potentially affected

Biotic division potentially affected

Suggested hierarchical level of decision making

World Continental National Sub-national Local

Climate Geology, geomorphology Geomorphology, soils, hydrology Soils, hydrology Hydrology, vegetation, fauna

Biosphere Biomes Biomes and ecoregions Ecoregions and habitats Habitats and microhabitats

Global, International environmental authority Multinational at continental level Multinational at regional or national level State, department, province Municipality, local

politically divisible at a local level. Therefore, it is crucial to structure the decision making process hierarchically, from global to local, so that the State aligns to the local forces in their effort to implement conservation, without losing its organizational role (Table 1). The alliances global-local must be negotiated in a participative manner with an unequivocal priority on conservation goals. In addition to the international conventions like the Convention on Biological Diversity (CBD), some initiatives are tested in search of allies for conservation at larger scale. For instance, in Brazil, the tax on the circulation of goods and services (ICMS) is collected by the government and is distributed among towns (municipios) mainly following the criteria established in the federal constitution (75% of the total) combined with the criteria of each state legislative congress (remaining 25%) in proportion to environmental conservation units and watershed protection areas (Young & Roncisvale 2002). This is a good example of a top-down initiative where the large-scale conservation goals are negotiated between the national and state governments and the local level. This incentive is successful in encouraging municipios to increase the total area under conservation inside their boundaries, both public and private, since this represents a higher budget. Hierarchy is the key of success in this decision making process. Firstly, the decision making process at local level should be subordinated to global goals. Secondly, it should always reinforce the levels of protection. As the ICMS example, environmental legislation and economic instruments are developing favorably in this direction in several South American countries, although many of these legislative solutions are still in cradle. As a concluding remark, a possible way-out from these conservation crossroads might consist in being able to distinguish friends from dangerous enemies (with their affable words) for both, biodiversity conservation and local livelihoods, especially from a Latin American perspective. The ecological history of this part of the world is full of Trojan horses.

Acknowledgments I would like to thank John Terborgh, Eduardo Rapoport, Magdalena Stanulescu, Alicia Bugallo, Fabián Gonzalez, Katrina Brandon, Fernando Fernandez, Jim Barborak, Amanda Paulos, Leonardo Cermelo and Ricardo Rozzi for their influence in the ideas expressed in this essay. Their opinions do not necessarily convey an agreement with my viewpoints. This work is an updated version of a conference given on behalf of The Conservation Land Trust. I am grateful to the Consejo Nacional de Investigaciones Científicas y Técnicas and Instituto de Análisis de Recursos Naturales for their financial support.

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Received: August 2010 First Decision: September 2010 Accepted: October 2010

Essays & Perspectives

Brazilian Journal of Nature Conservation

Plasticity and Conservation Ulrich Lüttge Institut fuer Botanik, Technische Universitaet Darmstadt, Darmstadt, Germany

Abstract Plasticity is a feature of phenotypes and genotypes. Genetic variation is reflected in phenotypic plasticity. When conservation protects genomes it protects complexity and with complexity it protects the beauty of life. However, plasticity is not to be considered as a stationary distinction of the genome. Phenotypic and genetic plasticity are interrelated via plasticity of gene regulation by intrinsic networks with non-linear dynamics. In the “evo-devo” (evolution – development) perspective of plasticity epigenetic dynamics must be essential, and thus, plasticity may also be considered to be a process. Closely related to plasticity is diversity and this is of enormous concern in conservation. As a feature of space, ecological niches are shaped by spatiotemporal-functional dynamics of plasticity. Plasticity of plants in growth and development is essential for their occupation of space. There is a dual plasticity, i.e. plasticity of species and plasticity of niches. Conservation of protected areas sustains space as a resource of life. Plasticity may either attenuate or accelerate speciation. In the latter case it is a source of biodiversity. The genotypes but also the epigenetic properties of DNA and histone methylation are products of evolution. With the essential role of diversity in the regulation of ecosystem functions and stability this evidently merges evolutionary and ecological aspects. We need to extend the evo-devo to an evo-devo-eco concept. It underlines the prominent role of conservation for developing ethics of respect of past evolutionary heritage and present ecological treasure. Conservation protects biodiversity as an important natural heritage of the biosphere of which man is a biological part by evolution and an operator by culture, on which man is totally dependent, and which by ethical imperative man must conserve. Key words: Conservation, Ecophysiology, Plasticity, Plants, Phenotypes, Genotypes.

Plasticity and Conservation: Setting the Scene Plastic is anything – mostly any material –, that is capable of being moulded. An encyclopedic dictionary says that in biology plastic means exhibiting adaptability to environmental change. However, this definition is too narrow. Plasticity includes responses to internal conditions. Most plastic is the human brain and with it the human mind. In reactions to external and internal stimuli and even with repair of some serious injuries the brain’s plasticity is enormous (La Recherche 2010). One outcome of plasticity is flexibility. However, what has plasticity to do with conservation? Closely related to plasticity is diversity and this indeed is of enormous concern in conservation. The fashionable key word is “biodiversity”. Plasticity may endow organisms with functional flexibility. This may then lead to functional diversity of organisms within populations or of species within larger units such as habitats or ecosystems. There is diversity of species, or floristic diversity when we talk of plants, and plasticity may support speciation as we *Send correspondence to: Prof. Dr. Ulrich Luettge Institut fuer Botanik, Technische Universitaet Darmstadt, Schnittspahnstrasse 3-5, D-64287 Darmstadt, Germany E-mail: Luettge@bio.tu-darmstadt.de

shall see below. There is diversity of niches, sites, habitats and ecosystems. Although globally there is still a very large gap between practice and basic consent it is generally accepted that diversity is important for and in need of conservation. This centers plasticity into the focus of conservation. Evidently, with this said, a note of caution is also required. Some environments are characterized by often single dominating factors of stress or “stressors”. Such sites, habitats or ecosystems can be highly unique and valuable natural heritage although under extreme single factor stress there will be little biodiversity. For acclimation specialization and rigidity will be more in demand than plasticity. Illustrative examples are extreme deserts where water is the dominant stressor or Antarctic territories and high altitude alpine systems (outside the tropics) where temperature is the dominant stressor. Only a small number of species that have evolved specific traits for dealing with the specific stress will occupy such sites. Nevertheless they can be just as urgently worth protection and conservation than diversity hot spots. Conversely in the tropics it is normally multi-factor and not single-factor stress that

Plasticity and Conservation

characterizes the various ecosystems, including the great ecosystems of Brazil, namely its forests (mainly moist or rain forests, flood forests), its flood plains, its cerrados, its inselbergs and high altitude vegetation. Naturally, there are also unique special sites in the tropics with single factor stress. The rushes of water in rivers and water-falls with their specialized rheophytes, e.g. Podostemonaceae, come to my mind. Generally, however, multi-factor stress rather than single-factor stress is determining plant life and this requires plasticity. Thus, it is the intention of this brief essay to evaluate various aspects of plasticity and why they need to be considered in strategies of conservation.

Plasticity: Phenotypes and Genotypes It is one of the most basic concerns of conservation to protect genotypes. The genotype comprises the complete genetic information of an organism. It is the product of evolution. Evolutionary selection produces adaptation, i.e. genetically adapted individuals. Such evolutionary adaptation is a long term process. In development during ontogeny the genotype generates a certain phenotype. For the phenotypes we distinguish morphotypes and physiotypes. The complete set of phenotypical traits generated by a genotype in the morphological domain is the morphotype and in the physiological domain is the physiotype (Kinzel 1972, 1982). The morphotypes are structural life forms as delineated by comparative morphology and anatomy. The physiotypes are physiological life forms as delineated by comparative physiology, biochemistry and molecular biology. In ecology and physiological ecology the phenotype is always the direct receiver of environmental input. This reception of external information may require acclimation by short-term and often also reversible phenotype expression. By feed-back or feed-forward via the phenotype there will be a regulatory demand on the genotype with selective gene regulation potentially generating differently acclimated phenotypes. Thus, due to feed-back or feed-forward the phenotypes are both receivers and products of environmental input. They are generated by the genotype under the pressure of external cues. This implies that the various genotypes may be capable of generating phenotypic plasticity to a smaller or larger extent. “Phenotypic plasticity is defined as the ability of one genotype to produce more than one phenotype when exposed to different environments” (Garland & Kelly 2006). Individual genotypes can express numerous phenotypes (Vogt et al. 2008). In this sense we may speak of plasticity of individual organisms. Conversely, if we consider phenotypic plasticity in relation to co-occurrence of different genotypes within a population which are each adapted to a slightly different environment we have the plasticity of a population. In this case it is genetic variation which is reflected in phenotypic plasticity (Booy et al. 2000). The individual plasticity given by one individual genotype evidently implies that plasticity is due to not only genotypic

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variation (in populations!) but also gene regulation (in individuals!). This involves hierarchies of genes as well as net works. We realize that phenotypic and genetic plasticity may be interrelated. Phenotypes and genotypes are integrated in common networks (Hütt & Lüttge 2005; Lüttge 2005). Plasticity may be considered as a dynamic process rather than a fixed character. A plethora of biochemical, metabolic and physiological processes are involved in generating plasticity at the level of individual organisms. Since Darwinian selection acts on the individual it can be stated that the capacity of plasticity itself will be subject to evolutionary selection. One may ask the question if there may be particular concrete plasticity genes. Indeed, it is often argued in the life science literature that for complex inclinations including physiological and psychical diseases and aberrations of man specific genes are responsible and research is searching for them. Thus, why should there not be specific plasticity genes? With the multitude of functions involved in plasticity I very much doubt that this can be a complete and satisfactory explanation. There are no single distinct paraphernalia making up plasticity. Major traits are complexes and comprise complements of subordinate traits (Garland & Kelly 2006). Generally genomes appear to be too small to have specific genes governing most complex complements of functionality. The more likely explanation is epigenetic regulation of gene expression. The concept of epigenetics dates back to Johann Friedrich Blumenbach (1752-1840), and the term epigenetics then has been specified at the beginning of the 1930s by Conrad Hal Waddington (1905-1975) (Gierer 1988). It is now strongly revived by the evo-devo-concept (evolution – development). For among other reasons the concept is much needed because we realize that genomes are often astonishingly small. An outstanding example is the genome of man who has 20 to 30 thousand genes compared to the 15 to 20 thousand genes of a nematode and the fly Drosophila or the 27 thousand genes of the higher plant Arabidopsis thaliana, and who shares 98.7% of his genes with the chimpanzee, i.e. differs by barely a few hundred genes from the chimpanzee. Evidently we need much more than genomics for explaining high complexity. Biology must abandon its genome centered views. This is precisely what the epigenetics concept achieves. It aims at explaining the external manifestation or overt output of genetic activity and with that also the appearance of new structures during embryonic development or ontogeny. We can consider epigenetics broadly as gene regulation with the complex and integrated functional networks of the entire developing or developed organism. The feed-back and feed-forward links (edges) between the various nodes of networks bear out non-linear dynamics which intrinsically comprise a high degree of complexity. This also should assist us in understanding the complexity of plasticity. The expression of different phenotypes by one given genotype in the same environment demonstrates the power of epigenetics (Vogt et al. 2008). Under the influence of

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epigenetics phenotypical plasticity is part of a developmental process. There will be developmental variation (Vogt et al. 2008). Phenotypical plasticity is the target of evolutionary natural selection. Followed by genetic assimilation where a novel phenotype is genetically fixed, genetically stable populations or ecotypes will be the result (Kinzel 1982; Turesson 1992; Pigliucci et al. 2006).

increasing r. At medium r there are bifurcations first resulting in two oscillating states between two branches. After a further bifurcation on each of the two branches there are four oscillating states. We certainly recognize a certain degree of order in these oscillations. However, when r increases further new branches are erratically generated even with infinitesimally small changes of r. The system determined by the mathematical equation (3) has moved into chaos and its performance has become increasingly unpredictable.

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Evidently, the genetics of plasticity and the evolution of plasticity are far from being explained by any specific one gene target. Epigenetic regulation involves methylation of DNA and nucleosomal histones. It is an intriguing question and remains to be understood how selection might affect evolution of methylation properties of these macromolecules. Here we witness very strong interrelations of evolution, development and ecology. We should advance towards developing an “evo-devo-eco” concept.

Plasticity: Non-Linear Dynamics We have seen that when conservation protects genomes it also protects complexity. The complexity is maintained by plasticity. I alluded to this by the initial remark on the human brain. This is the most complex organ with high plasticity which we know in the entire biosphere. Technical examples are systems of routes for traffic or electricity lines (Watts 1999). When certain nodes and edges break down alternatives are available and the punctually perturbed system may not even show any overt symptoms. Complexity of feedback systems becomes evident in deterministic chaos (Schuster 1995). Deterministic chaos is borne out by population dynamics. This is also an aspect of conservation when it aims at protecting populations. We may be deeply submerged into complexity if we even only consider an apparently very simplistic case of a population, x, of a single organism with its resources, r, although this certainly is a highly reductionist approach vis-à-vis reality in nature. The size of the population x after a time interval of t1 from the start t0 will be proportional to the amount of resources, r: xt1 ~ r . xt0

(1)

A negative feedback is that the population will generate competition between individuals and it will produce waste and pollution. Therefore, xt1 will also be negatively related to the size at t0, i.e. it will be also proportional to 1 – xt0: xt1 ~ (1 – xt0)

(2)

The result is the logistic equation: xt1 = r . xt0 . (1 – xt0)

(3)

Robert May (1976) has studied this equation in detail and called it “simple mathematical model with very complicated dynamics”. At low r, xt1 increases monotonically with

The interest of conservationists must lie at first sight in the relevance of this to population dynamics (Hastings et al. 1993). However, we can advance it if we extend the meaning of r from simple resource and subsume in it any kind of effective external cues (Lüttge 2008). Then we return to very general concerns of conservation. Let us consider biodiversity. I think conservationists may argue and the view is widely accepted that in a functional and highly complex ecosystem network like that of a tropical rainforest the biodiversity is its very state of order. Even if we take “order” here only metaphorically for the oscillations obtained with equation (3) we can relate this well to our observations of biodiversity. With very high stress or low r there is low diversity. Only highly specialized and stress acclimated species will survive. With very low stress or high abundances or high r there is also low diversity with the monopoly of only a few robust and outcompeting species. At medium stress or medium r there is the “order” of diversity. High plasticity allows a wealth of solutions of acclimation and adaptation occurring together. Grime et al. (1987) have actually demonstrated this in an experiment with a number of microcosms where they applied different degrees of stress. High diversity was given only within a rather narrow window of stress intensities allowing – for the conditions of the British Isles – a dry biomass of no less than 350 and no more than 750 g.m–2. A further aspect of nonlinear dynamics in nature is the ordering power of stochastic noise. Is that not a paradox, the ordering power of noise? Why should it be of interest for conservation? It is our daily experience that there is noise everywhere. There is noise in all living systems and subsystems. There is noise in our brain and it has been found that a certain level of noise is important for sustaining its functions via the plasticity of neuronal connections. For explaining the contribution of noise to the generation of overt ordered functional structures let us consider a simple mathematical geometric problem. Let us assume that there is a regular sinoidal oscillation. The inherent rhythm of the oscillation becomes only visible if the peaks of the oscillation pass a certain upper threshold. However, the peaks always remain below the threshold, and hence, there is no overt rhythm. Let us now superimpose some noise to the oscillation. If the noise is rather small nothing new shall happen. If the noise is rather high the regular oscillation will disappear in the noise. A medium degree of noise, however, will lift the peaks of the oscillation just

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above the upper threshold and now the rhythm will be seen. This is called stochastic resonance or stochastic coherence. In this way noise can establish regular structures and sustain functions. There are even medicinal applications of this phenomenon. In conclusion of this section, we see that when conservation protects complexity with its nonlinear dynamics it also protects the beauty of life with its unlimited fantasy of creating fantastic variety. The beauty of life is maintained by plasticity.

Plasticity: Acquisition of Space Conservation has a lot to do with space. It needs to protect sites, habitats and ecosystems which occupy space. We can consider space per se as a resource (Grams & Lüttge 2010). Acquisition and defense of space by organisms under strong environmental abiotic stress as well as under competition, i.e. strong biotic stress, is an important facet of their ecological performance. Plants are basically immobile. Therefore unlike animals, they cannot acquire space by moving around. Plant’s plasticity in growth and development is essential for their occupation of space. With respect to partitioning of resources within the whole plant organisms, i.e. allocation and distribution of nutrients and assimilates to the various organs during growth, we may distinguish between isometry and allometry (Pretzsch 2009). In isometry all linear dimensions of the plant change proportionally to each other. In this way geometric similarity is maintained during growth and development. Shape is determined by a rigid developmental program. In allometry linear dimensions do not change proportionally. There is plasticity in the development of shape in response to environmental cues. Allometry comprises both the internal allometric partitioning processes driven by the development of size and external factors which determine biomass allocation. Allometry can be well exemplified by considering the relationship between stem diameter and crown width of European beech (Fagus sylvatica L.) and Norway spruce (Picea abies (L.) Karst.). Crown width and with it the growing space requirement of the trees increases with increasing stem diameter. For the same increase in diameter, beech has a higher demand for growing space than spruce. There is typically more lateral crown spread in the former in comparison with the more vertical and pyramidal growth of the latter (Pretzsch 2009). This is particularly important for trees competing for space in mixed dense forest stands. Hence, plasticity determines the occupation of space by plants. A modular concept of phenotypic plasticity in plants has been proposed by De Kroon et al. (2005). It suggests that plasticity is mainly operating at the sub individual level, i.e. plasticity is effective via the expression of modules. These authors propose that “plasticity of whole plants is a by-product of modular responses, shaped by hierarchical

selection” and they even quote a very extreme view where the modular concept culminates in the caricature that “A tree is not a tightly integrated organism but a by-product of its parts” (Hankioja 1991, quoted by De Kroon et al. 2005). Conversely, systems biology progressively unravels highly interactive regulatory networks in development and functioning of plants. Allometry reveals coordinated plastic development of the various parts of plants. Although they are non-animated living beings we must not deny plants to be integrated organisms.

Plasticity: Ecological Niches We may consider space per se in a rather abstract way as something that can be occupied but does not in itself provide material resources that can be utilized. Thus, “empty” space in itself can be a resource (Grams & Lüttge 2010). This concept of space differs fundamentally from the concept of ecological niche. Traditionally niche is not considered purely as space but most importantly also implies functional aspects. Recently concepts of species niches have been extended and became more sophisticated. There are two main components (Chase & Leibold 2003): requirements of species for their survival in their environment, and impacts of species on their environment. It is important for conservation to realize this distinction. First, the species which one wants to protect with conservation of an environment evidently have specific requirements which conservation must sustain. Second, however, quite clearly all life is always affecting its own environment. We see this not only very directly in the feedback from individuals within populations as illustrated by the logistic equation discussed above (Equation 3) and in the most dramatic way by the consequences of the increasingly growing population of man on the biosphere of our planet of which man is a biological part. In a more subtle way we observe that any life will shape its environment so that niches will change. If species arrive and get established in “empty” space, e.g. as pioneer species on bare rocks, on open sand or gravel of deserts or on newly formed volcanic islands of oceans, their activity will create new niches. We understand that niches have very complex spatiotemporal/ functional dynamics, certainly nonlinear dynamics. Niches show plasticity. With respect to the central theme of this essay, namely the plasticity of species, it is comprehensible that the two plasticities, that of niches and that of species, must match. In other words species plasticity must bear out niche plasticity. If a species has high genotypic and phenotypic plasticity it will have a larger niche width than a species with low plasticity. The niche of the species with high plasticity may overlap with the niche of the species with low plasticity and even comprise it entirely but not vice versa, of course. I shall briefly illustrate this using two examples from tropical sites and ecosystems. The first example is the comparison of two species, i.e. species of the shrub and tree genus Clusia. Clusia multiflora H.B.K.

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is a species obligatorily performing C3-photosynthesis (C3). Clusia minor L. is a species which can switch between C3-photosynthesis and crassulacean acid metabolism (CAM). Focus and space of this essay do not allow to explain these modes of photosynthesis in detail (see Lüttge 2007a, b). It is sufficient for understanding the example to say as much as this: In C3-photosynthesis carbon dioxide, CO2, can only be taken up and fixed in the light, i.e. during the day, with simultaneous assimilation to carbohydrate. CAM plants can also fix CO2 in the dark, i.e. during the night, and store it in the form of malic acid (nota bene: “acid metabolism”) from which it can be remobilized behind closed stomata during the day and be assimilated in the light. In addition some direct CO2 fixation can also be performed by CAM plants during the day. C3-plants must open their stomata for CO2 uptake during the day subjecting them simultaneously to pronounced respiratory loss of water. Stomatal opening for CO2 uptake during the night is associated with much less loss of water. CAM has an enormous intrinsic flexibility of expressing night time and day time CO2 uptake. In addition the C3/CAM intermediate species C. minor can reversibly switch completely between the C3 and the CAM-mode of photosynthesis. Evidently C. minor has much more ecophysiological plasticity than C. multiflora. Observations in a secondary savanna in Venezuela have shown that C. multiflora occupies open sun exposed sites, while C. minor grows in semi-shade of a deciduous forest. However, C. minor can also intrude into the open sites of C. multiflora, where both species can be found growing side by side. Conversely, C. multiflora was not observed in the semi-deciduous forest. At first sight this appears paradox, because with CAM as a water saving mode of photosynthesis C. minor should be better acclimated to open exposed sites. However, plasticity and niche width explain the distribution of both species. With acclimation but low adaptability the low-plasticity species C. multiflora is restricted to its smaller niche, the open sites. With high adaptability the high-plasticity species C. minor has a much larger niche width, comprising both the space of the forest and that of the open savanna (Lüttge 2007b).

river in some areas marine sandy deposits date from the Pleistocene (120,000 year BP) and remained acquiring their final shape after a series of invasions and regressions of the sea during the Holocene (Martin et al. 1993). The plant species we find in the young restingas to a large extent were coming from the old Atlantic rainforest. Approximately 80% of the species occurring in the restingas of the State of Rio de Janeiro are also found in montane rainforests (Araujo 2000). Plasticity of the rainforest species must have been important for their successful migration from the multi-factor stress environment of the rainforest to the more extreme and seasonal conditions of the restinga habitats (Scarano 2002). In other words, with high plasticity the newly arrived species in the restingas had large niche widths. However, more speciation obviously did not occur. There are hardly any endemic species in the restingas (Rizzini 1979, Araujo 2000).

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The second example is the comparison of two ecosystems, i.e. the Atlantic rainforest and the coastal restingas of Brazil. The Atlantic rainforest is an old ecosystem. Tropical rainforests are characterized as multi-factor stress environments (Lüttge 2010). This implies that there is a particularly large variety and plasticity of niches. This in turn supports speciation as we shall consider a bit more in the following section of this essay. Hence, conditions and time have allowed speciation to occur in the Atlantic rainforest. Indeed it is rich of species including many endemic species. Conversely, the restingas are young ecosystems (Scarano 2002; Scarano et al. 2005). They stand on quaternary terrains, i.e. sandy coastal, plains (Martin et al. 1993, Scarano et al. 1997). Along the coastline of Rio de Janeiro the quaternary sandy deposits and dunes date mostly from the Holocene, having been established and re-established from 5,000 to 3,000 BP, but further north under the influence of the Paraiba do Sul

The two examples, i.e. that of two species and that of two ecosystems, are from very different scales. Evidently plasticity occurs at the individual level but then has implications at higher hierarchical levels, such as species and ecosystems. The relevance for conservation at the different scalar levels appears similar. The performance of the two Clusia species with different degrees of plasticity and niche width is important in attempts of reforestation of the secondary savanna, which has in fact been initiated at the respective site in Venezuela but unfortunately was abandoned due to lack of dedicated manpower. The plasticity and niche width of Atlantic rainforest species allowing them to conquer the young restinga habitat and to create a new ecosystem with new niches there underlines the uniqueness of both Atlantic rainforest and restinga and their perplexing interrelations. In different ways as explained, plasticity affects them both. Both are most remarkable and wonderful natural heritage of Brazil with the highest priority for conservation.

Plasticity and Speciation Species diversity is the product of speciation. There is a controversy of whether plasticity accelerates or attenuates speciation. Various genotypes may be capable of generating phenotypic plasticity to a smaller or larger extent. On this background it can indeed be debated whether plasticity may support development of new species diversity and evolutionary selection of new genotypes. A key question is if plasticity is adaptive or not. This may be different from case to case. In any way, it is evident that plasticity may be associated with high costs in terms of resources of a plant. By this way plasticity may reduce fitness (van Kleunen & Fischer 2005). However, this also depends on how we define and especially how we quantify fitness. This is a difficult problem and much more complex than measuring sexual reproduction and seed production (Lüttge & Scarano 2007). Plasticity may also hinder speciation by stabilizing genotypes. Flexible

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adaptations by plasticity could protect given genotypes from selection under environmental pressure. On the other hand plasticity may enhance speciation by allowing large ecological amplitudes (Grime et al. 1986; West-Eberhard 1986, 1989, 2003; Solbrig 1994; Lüttge 1995a,b, 2000, 2005; Gehrig et al. 2001; Lüttge & Scarano 2004). Thus, plastic variation of phenotypes generates biodiversity. Ecological amplitudes may separate populations with reduced sets of genotypes which are specially adapted to particular sites. This may lead to new genetically stable populations which we call ecotypes (Turesson 1992; Kinzel 1982; De Jong et al. 2005). Then adaptive radiation and segregation may lead to speciation in the path of the wanderings of populations in the space of genotypes so that the development of phenotypes from genotypes is the real origin of complexity in biology (Schuster 1998). Important opportunities to witness the traces of speciation are given on oceanic islands of recent volcanic origin such as the islands of Hawaii or of the Galápagos archipelago. The Galápagos islands are about 4 million years old. A comparison of three cactus genera, namely Opuntia, Brachycereus and Jasminocereus (all of them CAM-species), of these islands is most interesting (McMullen 1999). There are 6 species of Opuntia which together with their varieties make 14 endemic members of the genus occurring on different islands of the archipelago. Unfortunately there do not appear to be any ecophysiological studies so that we have no information about plasticity. Evidently, however, speciation must be quite active in this genus. By contrast, in another endemic cactus genus Brachycereus there is only one species, Brachycereus nesioticus (K.Schum.) Backbg., the so called lava cactus. It occurs on five islands and shows very little variability. No trends towards speciation are indicated. This endemic stem succulent cactus, grows directly on completely sun exposed and highly irradiance-absorbent black surfaces of lava fields and may be so very specifically acclimated to the extreme stress of its site that selection would immediately eliminate any variations. The third example Jasminocereus also expresses only one species on the islands, i.e. Jasminocereus thoursaii (Weber) Backbg. This species, however, represents an intermediate case as it is very variable and occurs with three varieties. It appears to be in an early stage of differentiation. Studies of the ecophysiological performance and expression of plasticity in these cacti might promise important insights into the relation between plasticity and speciation.

Conclusions Plasticity is a feature of all living organisms including man. Degrees of plasticity may vary. There may be low plasticity and high rigidity or high plasticity and low rigidity. Plasticity is a complex trait of organisms but it does not materialise in distinct paraphernalia. It is a quality expressing spatiotemporal and functional nonlinear dynamics. Therefore plasticity may even be considered

as a process. Summarizing the relevance of plasticity for conservation we may distil two central roles of plasticity from this essay: First, it is the role of plasticity in shaping ecological niches, which is essential in the acquisition and defense of the resource space by life. Areas protected by conservation are the space of life. Man is endowed with a particularly high plasticity. By evolution man is biologically part of the biosphere on which he depends. By culture man also is operator on the biosphere. Second, it is the role plasticity can play in speciation creating biodiversity. Areas protected by conservation are the holding of biodiversity.

References Araujo DSD, 2000. Análise florística e fitogeografíca das restingas do estado do Rio de Janeiro. [D.Sc.Thesis]. Rio de Janeiro: Universidae Federal do Rio de Janeiro. Booy G et al., 2000. Genetic diversity and the survival of populations. Plant Biology, 2:379-395. Chase M & Leibold MA, 2003. Ecological niches. Chicago: The University of Chicago Press. De Jong G, 2005. Evolution of phenotypic plasticity: patterns of plasticity and the emergence of ecotypes. New Phytologist, 166:101-118. De Kroon H et al., 2005. Constraints on the evolution of adaptive phenotypic plasticity in plants. New Phytologist, 166:49-60. Garland T & Kelly SA, 2006. Phenotypic plasticity and experimental evolution. Journal of Experimental Biology, 209:2344-2361. Gehrig H et al., 2001. Molecular phylogeny of the genus Kalanchoë (Crassulaceae) inferred from nucleotide sequences of the IST-1 and IST-2 regions. Plant Science, 160:827-835. Gierer A, 1998. Im Spiegel der Natur erkennen wir uns selbst. Wissenschaft und Menschenbild. Reinbeck: Rowohlt. Grams TEE & Lüttge U, 2010. Space as a resource. Progress in Botany, 72:349-370. Grime JP et al., 1987. Floristic diversity in a model system using experimental microcosms. Nature, 328:420-422. Grime JP, Crick JC & Rincon JE, 1986. The ecological significance of plasticity. In: Jennings DH and Trewawas AJ (Ed.). Plasticity in plants. Cambridge: Cambridge University Press. p. 5-29. Hankioja E, 1991. The influence of grazing on the evolution, morphology and physiology of plants as modular organisms. Philosophical Transactions of the Royal Society B: Biological Sciences, 333:241-247. Hastings A et al., 1993. Chaos in ecology: is mother nature a strange attractor? Annual Review of Ecology & Systematics, 24:1-33. Hütt M-T & Lüttge U, 2005. Network dynamics in plant biology: current progress in historical perspective. Progress in Botany, 66:277-310.

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Kinzel H, 1972. Biochemische Pflanzenökologie. Schriften des Vereins zur Verleihung naturwissenschaftlicher Kenntnisse in Wien, 112:77-98.

McMullen CK, 1999. Flowering plants of the Galápagos. Ithaca, London: Cornell University Press.

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Kinzel H, 1982. Pflanzenökologie und Mineralstoffwechsel. Stuttgart: Ulmer. La Recherche, 2010. Le cerveau comment il se réorganise sans cesse. Les dossiers de LaRecherche, 40:2010. Special isssue. Lüttge U, 1995a. Ecophysiological basis of the diversity of tropical plants: the example of the genus Clusia. In: Heinen HD, San José JJ & Caballero-Arias H (Eds.). Nature and human ecology in the neotropics. Venezuela. p. 23-36. (Scientia Guaianae, 5). Lüttge U, 1995b. Clusia: Ein Modellfall der ökophysiologischen Plastizität in einer tropischen Gattung. In: Rundgespräche der Kommission für Ökologie. Bayerische Tropenforschung - Einst und jetzt. vol. 10. München: Akademie der Wissenschaften. p. 173-186. Lüttge U, 2000. Photosynthese-Physiotypen unter gleichen Morphotypen, Species und bei Klonen: Kann ökophysiologische Plastizität zur Entstehung von Diversität beitragen? Berichte der Reinhold-Tüxen-Gesellschaft, 12:319-334. Lüttge U, 2005. Genotypes - phenotypes - ecotypes: relations to crassulacean acid metabolism. Nova Acta Leopoldina, NF 92:177-193. Lüttge U, 2007a. Photosynthesis. In: Lüttge U (Ed.). Clusia: a woody neotropical genus of remarkable plasticity and diversity. Berlin: Springer. p. 135-186. (Ecological Studies, 194). Lüttge U, 2007b. Physiological ecology. In Lüttge U (Ed.). Clusia: a woody neotropical genus of remarkable plasticity and diversity. Berlin: Springer. p. 187-234. (Ecological Studies, 194). Lüttge U, 2008. Physiological ecology of tropical plants. Berlin: Springer. Lüttge U, 2010. Struggle of plants with crassulacean acid metabolism (CAM) in tropical environments under the action of dynamic networks of stressors. Aob Plants, published electronically. Lüttge U & Scarano FC, 2004. Ecophysiology. Revista Brasileira de Botânica, 27:1-10. Lüttge U & Scarano FR, 2007. Synecological comparisons sustained by ecophysiological fingerprinting of intrinsic photosynthetic capacity of plants as assessed by measurements of light response curves. Brasilian Journal of Botany, 30:355-364. Martin L, Suguio K & Flexor JM, 1993. As fluctuações de nível do mar durante o quaternário superior e a evolução geológica de “deltas” Brasileiros. Boletim do Instituto de Geografia da Universidade de São Paulo, Publicação Especial, 15:1-186. May RM, 1976. Simple mathematical models with very complicated dynamics. Nature, 261:459-467.

Pigliucci M, Murren CJ & Schlichting CD, 2006. Phenotypic plasticity and evolution by genetic assimilation. Journal of Experimental Biology, 209:2362-2367. Pretzsch H, 2009. Re-evaluation of allometry: state-of-the-art and perspective regarding individuals and stands of woody plants. Progress in Botany, 71:339-369. Rizzini CT, 1979. Tratado de Fitogeográfia do Brasil. vol. 2. São Paulo: Edusp. Scarano FR, 2002. Structure, function and floristic relationships of plant communities in stressful habitats marginal to the Brazilian Atlantic rain forest. Annals of Botany, 90:517-524. Scarano FR et al. 1997. Plant establishment on flooded and un-flooded patches of a fresh water swamp forest in southeastern Brazil. Journal of Tropical Ecology, 14:793-803. Scarano FR et al. 2005. Physiological synecology of tree species in relation to geographic distribution and ecophysiological parameters at the Atlantic forest periphery in Brazil: an overview. Trees, 19:493-496. Schuster HG, 1995. Deterministic chaos. Weinheim: Verlag Chemie. Schuster P, 1998. Evolution in molekularer Auflösung. Berichte und Abhandlungen. Berlin-Brandenburgische Akademie der Wissenschaften: Akademie Verlag, 6:187- 215. Solbrig OT, 1994. Plant traits and adaptive strategies: their role in ecosystem function. Berlin: Springer. p. 97-116. (Ecological Studies, 99). Turesson G, 1922. The genotypic response of the plant species to the habitat. Hereditas, 3:211-350. van Kleunen M & Fischer M, 2005. Constraints on the evolution of adaptive phenotypic plasticity in plants. New Phytologist, 166:49-60. Vogt G et al., 2008. Production of different phenotypes from the same genotype in the same environment by developmental variation. Journal of Experimental Biology, 211:510- 513. Watts DJ, 1999. Small worlds. Princeton: Princeton University Press. West-Eberhard MJ, 1986. Alternative adaptations, speciation, and phylogeny (a review). Proceedings of the National Academy of Sciences of the United States of America, 83:1388-1392. West-Eberhard MJ, 1989. Phenotypic plasticity and the origins of diversity. Annual Review of Ecology & Systematics, 20:249-278. West-Eberhard MJ, 2003. Developmental plasticity and evolution. Oxford: Oxford University Press.

Received: September 2010 First Decision: October 2010 Accepted: October 2010

Research Letters

Brazilian Journal of Nature Conservation

Predicting Patterns of Beta Diversity in Terrestrial Vertebrates Using Physiographic Classifications in the Brazilian Cerrado André Andrian Padial1,2,*, Luis Mauricio Bini1, José Alexandre Felizola Diniz-Filho1, Nayara Pereira Resende de Souza1 & Ludgero Cardoso Galli Vieira3 1

Programa de Pós-graduação em Ecologia e Evolução, Departamento de Ecologia, ICB, Universidade Federal de Goiás, Goiânia, GO, Brazil

Instituto Federal de Educação, Ciência e Tecnologia de Mato Grosso do Sul, Campus Nova Andradina, Nova Andradina, MS, Brazil

Universidade de Brasília, Planaltina, DF, Brazil

Abstract The effectiveness that environmental classifications have in representing terrestrial vertebrates inhabiting the Brazilian Cerrado was tested. Data on species composition of vertebrates were mapped on a cell’s grid covering the Cerrado. These cells were classified according to environmental classification schemes and the classification strength (CS) of these schemes was tested. Classification schemes based on biological data were also generated. Moreover, the classification scheme generated using the biological data of one taxonomic group was used as criteria to calculate the CS values using the similarity matrix generated by a second taxonomic group, thus generating a cross-taxa CS. The environmental classification scheme at the fine spatial scale significantly represented the variability of terrestrial vertebrates found in the Cerrado; however, CS values were generally low. Cross-taxa CS values, although higher than those obtained by the environmental classifications, were also generally low and indicate that biological surrogates as predictors of overall biota have little effectiveness. Key words: Biological Surrogates, Classification Strength, Environmental Surrogates, Physiographic Classification, Terrestrial Vertebrates.

Introduction Conservation planning strategies are based recurrently on surrogates as a means to defy the Linnean (Raven & Wilson 1992) and Wallacean shortfalls (Lomolino 2004). For instance, besides using specific taxonomic groups as surrogates for conservation targets (Rodrigues & Brooks 2007) broad-scale biodiversity attributes have been used (e.g., land types, Harding & Winterbourn 1997; Margules & Pressey 2000). Several features (e.g., landform types, drainage basins, and dominant vegetation types) are used to generate groups of homogeneous areas that become the targets for conservation (Harding & Winterbourn 1997; Silva et al. 2006). Heino & Mykrä (2006) suggested that areas from different groups of a classification scheme should be preserved to maximize the conservation of the regional biodiversity.

*Send correspondence to: Andre Andrian Padial Instituto Federal de Educação, Ciência e Tecnologia de Mato Grosso do Sul, Campus Nova Andradina, MS-473, Km 23, CEP 79750-000, Nova Andradina, MS, Brazil E-mail: aapadial@gmail.com

Silva et al. (2006) proposed a regionalization scheme for the Cerrado biome (Central Brazil). Five “landscape units” and 15 “ecological units” (nested within the landscape units) were delineated as important schemes in characterizing the spatial heterogeneity of the Cerrado biome. These schemes classify the biome in areas with similar general ecological conditions, and they can be used by managers as an objective criterion to delineate conservation areas (Silva et al. 2006). If the biota corresponds closely to the environmental classifications, each class of the physical classification scheme will encompass a portion of the overall variability of biodiversity (Heino & Mykrä 2006). Nevertheless, the effectiveness of the landscape or ecological units proposed by Silva et al. (2006) as surrogates for conservation was not properly evaluated for any biological group. We evaluated whether there was a correspondence between the classification schemes described above (hereafter referred to as a priori classifications) and the patterns of beta diversity generated by terrestrial vertebrates in the Cerrado biome (Brazil). We used data collected from 1199 species of mammals, birds, reptiles, and amphibians (see Diniz-Filho et al. 2008 for details about the biological

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data). We also evaluated whether there was a correspondence between the a priori classifications and the discrete units generated a posteriori using the biological data. The role of a given taxon as a biological surrogate of another taxon was evaluated with a cross-taxa analysis. For this analysis the classification generated by one taxonomic group was used to describe the patterns of beta diversity of a second taxonomic group.

landform and vegetations types (see Figure 1 in Silva et al. 2006). A second classification scheme (with 15 classes) was formed by splitting the five landscape units into ecological units: five units of well-drained plains and plateaus dominated by savannas, three units of hilly terrains dominated by savannas, four units of plains dominated by deciduous and semi-deciduous forests, two units of plains with evergreen, semi-deciduous, and other forests and one unit of poorly drained lowlands dominated by seasonally flooding savannas (see Silva et al. 2006 for a detailed description of the different landscape and ecological units). Each cell in our grid was classified according to the categories provided by Silva et al. (2006) (Figure 1).

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Methods Species’ geographic ranges were mapped on a grid with a spatial resolution of 1o that covered the area of the biome used to delineate the a priori classifications (see Silva et al. 2006 for detailed maps showing the a priori classifications). With this grid we created a matrix of species presence-absence by cells (N = 133 cells; see Diniz-Filho et al. 2008 for a partition of the Cerrado in 181 cells). The Jaccard similarity coefficient was used to calculate faunal similarity between cells in the biome for each taxonomic group. We used the two classification schemes of the Cerrado biome (i.e., landscape units and ecological units) proposed by Silva et al. (2006). They used 41 environmental variables (see Silva et al. 2006 Table 1) in a cluster analysis to identify five landscape units that were characterized by dominant

The grid system we used slightly differed from the area Silva et al. (2006) used to delineate the a priori classifications. As a consequence, three ecological units were excluded from our analysis: one ecological unit (3B) was not located in any of the cells we used, and two ecological units (2C and 3C) were only found in one cell. Thus, environmental classification at the fine spatial scale was comprised of 12 classes (Figure 1). We calculated the classification strength (CS) of the a priori classifications for each taxonomic group (van Sickle & Hughes 2000). The CS analysis measures the correspondence strength between a classification scheme and a matrix

Figure 1. A map of Brazil showing the area of the Cerrado biome we analyzed (upper panel; divided into 133 grid cells of 1° × 1°) and the a priori classifications (lower panel; LU = Landscape units; EU = Ecological units) proposed by Silva et al. (2006).

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of biological data by calculating the difference between the average within-group similarity (W) and the average between-group similarity (B). A large CS value (CS values vary from 0 to 1, van Sickle & Hughes 2000) indicates a high within-group and a low between-group similarity. In theory, this result indicates that the classes generated by a given classification scheme are valid to reproduce the main patterns of vertebrate beta-diversity. Significance levels of CS values were assessed by a Monte Carlo procedure with 10,000 permutations (Smith et al. 1990). The largest CS value found by comparing the two criteria indicates the classification scheme considered to be the best and most preferred scheme for conservation purposes. We also used the k-means method to generate typologies of the Cerrado using the biological data (i.e., the a posteriori classifications). We reduced the dimensionality of the biological data with a non-metric multidimensional scaling (NMDS) technique to transform the binary data (presence/ absence) into a set of quantitative vectors more suitable for the Euclidian metrics used in the k-means method. We used the k-means clustering method to divide the landscape into

the most appropriate and meaningful groups. As a method of determining when to stop adding groups to the analysis we used the following R statistic (Hartigan 1975): ⎛ tr ( W )(k ) ⎞ − 1 (n − k − 1) R=⎜ ⎝ tr ( W )(k + 1) ⎟⎠

(1)

where tr(W) is the trace of the dispersion matrix within groups, n is the number of samples, and k is the number of groups. This statistic was calculated for different values of k, and the results were compared to an F distribution (df = p; the number of variables, which was equal to two because we always used the first two axes of the NMDS for all biological groups). We calculated the CS values for the groups generated by the k-means method to compare the a priori and a posteriori classifications. Finally, we used clusters generated by a given taxonomic group (in the surrogate role) as a classification criterion to calculate the CS values with the similarity matrix generated by a second taxonomic group (in the target role; see Pinto et al. 2007). Figure 2 summarizes our analytical approach.

Figure 2. Our analytical approach. Step 1: Classification strength of the a priori classification schemes; Step 2: Classification strength of the a posteriori classification schemes; Step 3: Classification strength of the classification schemes based on a surrogate taxa.

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classifications were substantially higher than the CS values of the a priori classifications (Table 1); however, all CS were low (Table 1).

Results CS values for the classification scheme based on the five landscape units were significant for bird and reptile species. The correspondences between the classifications based on the ecological units and the biological data were greater than those based on landscape units and significant for all vertebrate taxa (Table 1).

When the classification criterion generated by a given taxonomic group was used as surrogate for the other three taxonomic groups CS values, albeit low, the values were always significant and substantially higher than the values derived from the environmental data alone (Tables 1 and 2). The typology generated by birds received the highest CS values for the three other taxonomic groups.

The number of classes generated by the k-means method varied for taxonomic groups (ranging from five for mammals to 12 for birds). As expected, the CS values of the a posteriori

Table 1. Classification strength for each a priori classification scheme. B = between group mean similarity; W = within-group mean similarity; CS = classification strength; p = significance level based on 10,000 random permutations. The a posteriori classifications were not tested because they were generated using the biological data, and thus they are a good and significant classification scheme.

Classification scheme

Taxa

Number of classes

Mammals

0.604

0.608

0.003

0.356

a priori

Birds

0.648

0.669

0.021

0.014

Landscape units

Reptiles

0.415

0.432

0.017

0.025

Amphibians

0.592

0.591

–0.001

0.551

Mammals

0.600

0.661

0.061

<0.001

a priori

Birds

0.652

0.709

0.058

<0.001

Ecological units

Reptiles

0.411

0.501

0.090

<0.001

Amphibians

0.585

0.647

0.063

<0.001

Mammals

0.569

0.747

0.178

Birds

0.635

0.851

0.215

Reptiles

0.394

0.686

0.291

Amphibians

0.565

0.812

0.247

a posteriori

Table 2. Results of the classification strength of a given taxonomic group (in the surrogate role; generated by a k-means method) in relation to other taxonomic groups (in the target role). B = between group mean similarity; W = within-group mean similarity; CS = classification strengths; p = significance based on 10,000 random permutations.

Target taxa

Mammals

Birds

Reptiles

Amphibians

Surrogate classification

Birds (12 classes)

0.583

0.828

0.245

<0.001

Reptiles (11 classes)

0.588

0.774

0.186

<0.001

Amphibians (9 classes)

0.580

0.808

0.228

<0.001

Mammals (5 classes)

0.631

0.751

0.120

<0.001

Reptiles (11 classes)

0.644

0.760

0.115

<0.001

Amphibians (9 classes)

0.637

0.799

0.162

<0.001

Mammals (5 classes)

0.390

0.544

0.154

<0.001

Birds (12 classes)

0.395

0.672

0.277

<0.001

Amphibians (9 classes)

0.393

0.648

0.255

<0.001

Mammals (5 classes)

0.556

0.734

0.178

<0.001

Birds (12 classes)

0.571

0.790

0.218

<0.001

Reptiles (11 classes)

0.577

0.734

0.157

<0.001

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Correspondence Between Classifications and Biota

Discussion We found that the patterns of beta diversity in terrestrial vertebrates could not be predicted by the classifications based on the environmental and vegetation variability of the Brazilian Cerrado (Silva et al. 2006). Thus, these classification schemes cannot be considered reliable surrogates for biodiversity in conservation planning and cannot be used to establish a regionalization plan for monitoring terrestrial fauna in the Cerrado biome. Weak CS values have been previously reported (e.g., Hawkins et al. 2000; Heino & Mykrä 2006). In these studies, coarser environmental discontinuities, which determine the groups of a typology, may not have been predictive of beta diversity patterns (Brooks et al. 2004). The low correspondence found between the biological and environmental classifications may also relate to the means in which different taxonomic groups respond to local environmental factors (Heino et al. 2004). Correspondence between biota and climatic factors are evident and well-known at a global scale; however, local factors may be important for species within a biome. At a scale of landscape units (the largest spatial scale we used), low levels of beta diversity may have been found due to the ubiquity of common species of the biome (Higgins et al. 2005), which resulted in a high between-group similarity compared to the within-group similarity. In this case, the local environmental differences between cells found in each landscape unit may be ignored (Faith & Walker 1996). One could argue that the number of optimal a posteriori classes could differ from the number of optimal a priori classes and thus the CS values could not be compared. However, the “relative” CS may be calculated to circumvent such limitations (Snelder et al. 2005). For this calculation, the a posteriori classification scheme must be generated with the same number of classes as the a priori classification scheme. For instance, we can calculate the relative CS for mammals considering landscape units (0.003/0.178) and for birds considering ecological units (0.058/0.215; note that the a priori and a posteriori classification schemes have the same number of classes in these cases; Table 1). The relative CS for mammals considering landscape units is 1.68%, and the relative CS for birds considering ecological units is 26.97%. These percentages indicate how good the a priori classification is in relation to the best classification possible (i.e., the a posteriori one, see also Snelder et al. 2005). It is clear that the a priori classifications have a low correspondence to the diversity of terrestrial vertebrates found in the Cerrado. Congruent spatial patterns of different taxa would indicate the reliability of the use of biological surrogates of the overall biodiversity (Heino et al. 2005). This would be true if the classification scheme generated by a given taxonomic group (in the surrogate role) was valid for a second taxonomic group (in the target role). However, our results do not support the use of biological surrogates; the cross-taxa classifications strengths we found were also low. Effective surrogates

cannot be considered as an alternative for a biodiversity analysis as they are difficult to find. Thus, if a taxonomic group or biodiversity pattern is considered important for conservation, the conservation targets and goals should be based directly on the available data for the specific group or pattern without trying to reach other indirect benefits (e.g., for other groups). Despite being necessary, surrogates for biodiversity conservation planning (Margules & Pressey 2000) are generally unreliable. Currently, most ecologists agree that using species data in systematic conservation planning is the best alternative method available in selecting a network of protected areas that reaches a given conservation goal at the lowest cost (Brooks et al. 2004). However, it is important to highlight that we are analyzing a biodiversity hotspot, the Cerrado, which has a high level of endemism and a high rate of habitat loss (Myers et al. 2000). We cannot wait for the proposition of conservation plans based solely on reliable data concerning species’ occurrences (Pressey 2004). We can circumvent the paucity of species data using species distribution modeling, especially for rare species (Brooks et al. 2004; Guisan et al. 2006). One can assume that considering environmental heterogeneity as an additional criterion in selecting priority areas for conservation would be a way to simultaneously encompass ecological processes and ecosystem services worthy of preservation (Higgins et al. 2004) and minimize biases caused by the paucity of biological data. Land types have been considered as entities relevant to conservation because they represent unique ecological processes and ecosystem services (Higgins et al. 2004; Molnar et al. 2004). Moreover, an area with idiosyncratic habitats and environmental conditions may have high conservation value because it may support endemic and/or rare species (Trakhtenbrot & Kadmon 2006). However, assuming that the environmental variability properly represents the variability of the biota can lead to incorrect interpretations. At best, a multiple array of surrogates to circumvent idiosyncrasies of one approach may be a more comprehensive and stronger strategy for conservation planning purposes (Pressey 2004). Nevertheless, the use of either environmental or biological surrogates for conservation strategies does not exclude the urgent need for collecting baseline data on species distribution.

Acknowledgements A.A. Padial and N.P.R. Souza received scholarships and acknowledge CAPES and CNPq for the financial support. L.M. Bini and J.A.F. Diniz-Filho also acknowledge CNPq for research grants. Financial support for this study came from a PRONEX program of CNPq/SECTEC-GO for establishing conservation priorities in Cerrado area.

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References Brooks T, Fonseca GAB & Rodrigues ASL, 2004. Species, data, and conservation planning. Conservation Biology, 18:1682-1688. Diniz-Filho JAF et al., 2008. Spatial patterns of terrestrial vertebrate species richness in Brazilian Cerrado. Zoological Studies, 47:146-157. Faith DP & Walker PA, 1996. Environmental diversity: on the best-possible use of surrogate data for assessing the relative biodiversity of sets of areas. Biodiversity and Conservation, 5:399-415. Guisan A et al., 2006. Using niche-based models to improve the sampling of rare species. Conservation Biology, 20:501-511. Harding JS & Winterbourn MJ, 1997. An ecoregion classification of the South Island, New Zealand. Journal of Environmental Management, 51:275-287. Hartigan JA, 1975. Clustering Algorithms, New York: Wiley. Hawkins CP et al., 2000. Evaluation of the use of landscape classifications for the prediction of freshwater biota: synthesis and recommendations. Journal of the North American Benthological Society, 19:541-556. Heino J, Louhi P & Muotka T, 2004. Identifying the scales of variability in stream macroinvertebrate abundance, functional composition and assemblage structure. Freshwater Biology, 49:1230-1239. Heino J et al. 2005. Searching for biodiversity indicators in running waters: do bryophytes, macroinvertebrates, and fish show congruent diversity patterns? Biodiversity Conservation, 14:415-428. Heino J & Mykrä H, 2006. Assessing physical surrogates for biodiversity: do tributary and stream type classifications reflect macroinvertebrate assemblage diversity in running waters? Biological Conservation, 129:418-426. Higgins JV et al., 2004. Beyond Noah: saving species is not enough. Conservation Biology, 18:1672-1673. Higgins JV et al., 2005. A freshwater classification approach for biodiversity conservation planning. Conservation Biology, 19:432-445. Lomolino MV, 2004. Conservation biogeography. In Lomolino MV & Heaney LR (eds.), Frontiers of Biogeography: new

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directions in the geography of nature, Sunderland: Sinauer Associates, p. 293-296. Margules CR & Pressey RL, 2000. Systematic conservation planning. Nature, 405:243-253. Molnar J, Marvier M & Kareiva P, 2004. The sum is greater than the parts. Conservation Biology, 18:1670-1671. Myers N et al. 2000. Biodiversity hotspots for conservation priorities. Nature, 403: 853-858. Pinto MP et al., 2007. Biodiversity surrogate groups and conservation priority areas: birds of the Brazilian Cerrado. Diversity and Distributions, 14:78-86. Pressey RL, 2004. Conservation planning and biodiversity: assembling the best data for the job. Conservation Biology, 18:1677-1681. Raven PH & Wilson EO, 1992. A fifty-year plan for biodiversity surveys. Science, 258:1099-1100. Rodrigues ASL & Brooks TM, 2007. Shortcuts for biodiversity conservation planning: the effectiveness of surrogates. Annual Review of Ecology, Evolution, and Systematics, 38:713-737. Silva JF et al., 2006. Spatial heterogeneity, land use and conservation in the Cerrado region of Brazil. Journal of Biogeography, 33:536-548. Smith EP, Pontasch KW & Cairns J, 1990. Community similarity and the analysis of multi- species environmental data: a unified statistical approach. Water Research, 24:507-514. Snelder TH, Biggs BJF & Woods RA, 2005. Improved eco-hydrological classification of rivers. River Research and Applications, 21:609-628. Trakhtenbrot A & Kadmon R, 2006. Effectiveness of environmental cluster analysis in representing regional species diversity. Conservation Biology, 20:1087-1098. van Sickle J & Hughes RM, 2000. Classification strengths of ecoregions, catchments, and geographic clusters for aquatic vertebrates in Oregon. Journal of the North American Benthological Society, 19:370-384.

Received: April 2010 First Decision: June 2010 Accepted: August 2010

Research Letters

Brazilian Journal of Nature Conservation

Regeneration and Colonization of an Invasive Macrophyte Grass in Response to Desiccation Thaisa Sala Michelan1, Sidinei Magela Thomaz2,*, Priscilla Carvalho3, Roberta Becker Rodrigues3 & Márcio José Silveira4 1

Graduate Program in Ecology and Evolution, Universidade Federal de Goiás – UFG, Campus II, CEP 74001-970, Goiânia, GO, Brazil

Biological Science Department/Nupélia, Universidade Estadual de Maringá – UEM, Av. Colombo 5790, bloco H-90, CEP 87020-900, Maringá, PR, Brazil

Graduate Program in Ecology of Continental Aquatic Environments, Universidade Estadual de Maringá – UEM

Graduate Program in Comparative Biology, Universidade Estadual de Maringá – UEM

Abstract The emergent macrophyte Urochloa subquadripara, an exotic and invasive species, causes extensive damage to aquatic plant assemblages. Regeneration and colonization by fragments of U. subquadripara may be affected by desiccation and may differ according to the fragment portion (apical, intermediate or basal). We tested the hypotheses that the ability of U. subquadripara fragments to regenerate and colonize declines with increasing time of exposure to desiccation, and that apical portions regenerate and colonize more quickly than intermediate and basal ones. Apical, intermediate and basal portions were exposed to different desiccation periods and left to grow in a greenhouse. Sprout and root growth were negatively and significantly affected by the desiccation period; fragments exposed to intermediate levels of desiccation regenerated and colonized the sediment at rates similar to those of the control. In addition, apical fragments showed greater development and sediment colonization than intermediate and basal fragments. Thus, the data supported our two hypotheses. Our results show that U. subquadripara has a high regeneration potential, indicating that the use of water drawdown to control its spread may be ineffective, and new strategies are required. Key words: Non-Native Species, Urochloa subquadripara, Vegetative Propagation, Plant Dispersal, Tropical Signalgrass.

Introduction Alien species are commonly introduced in freshwater ecosystems, where they threaten several natives. This phenomenon has become a central concern for freshwater ecologists and managers because inland aquatic ecosystems are often more diverse compared to terrestrial and marine ecosystems (Balian et al. 2008) and are experiencing high extinction rates (Jenkins 2003). Several species of aquatic macrophytes are well suited to be successful invaders because they reproduce prolifically, disperse easily and are adapted to a myriad of habitats (Santamaria 2002). Understanding how alien plant species become invasive is a fundamental goal in invasion ecology (Rejmánek et al. 2005) and an essential component of vegetation dynamics (including alien species) is the *Send correspondence to: Sidinei Magela Thomaz Biological Science Department/Nupélia, Universidade Estadual de Maringá – UEM, Av. Colombo 5790, bloco H-90, CEP 87020-900, Maringá, PR, Brazil E-mail: smthomaz@nupelia.uem.br

ability of plants to colonize available habitats (Riis 2008). Spatial patterns of abundance may be determined by the biological characteristics of a species (e.g., its life form) and physical features of its environment (e.g., water and sediment physico-chemistry) (e.g., Henry-Silva et al. 2008). Dispersal by plant fragments is also an important mechanism contributing to plant regeneration and spatial patterns of species abundance (Riis et al. 2009). Regional recovery after disturbances should occur more rapidly in species with widespread dispersal owing to their ability to disseminate propagules over relatively large areas (Reed et al. 2000). Vegetative propagation is a common dispersal strategy of many aquatic plants, and dispersal and colonization can take place through a variety of vegetative plant portions, including entire plants, rhizomes, stolons, tubers, turions and simple stem fragments (Barrat-Segretain & Cellot 2007; Riis et al. 2009). Macrophyte fragments are released by several types of disturbance, such as sediment mobility, herbivory and changes in water flow (Riis & Sand-Jensen

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Natureza & Conservação 8(2):133-139, December 2010

2006). Fragments can travel considerable distances via water flow, water birds and boats, and they can initiate colonization in distant habitats (Santamaria 2002; Thomaz et al. 2009).

The mean daily photoperiod was 13 hours ± 0.9 SD, the mean daily maximum temperature was 31.18 °C ± 2.33 SD and the mean daily minimum temperature was 21.02 °C ± 1.48 SD.

In wetlands, drying generally leads to biomass losses in the established vegetation (Combroux & Bornette 2004). This process also occurs when reservoirs are subject to water drawdown (Thomaz et al. 2009). Thus, the water regime is a major determinant of aquatic-plant community development, and the time of exposure to desiccation of plant fragments is an important determinant of their growth and success in freshwater ecosystems (Barrat-Segretain & Cellot 2007; Silveira et al. 2009).

Here, we consider regeneration to be the emergence of new sprouts and colonization to be rooting in sediment (sensu Barrat-Segretain & Bornette 2000). To assess the period over which tropical signalgrass can resist desiccation, plants were kept in a tray without water for 2, 5, 9, 17 and 26 days. We did not choose these periods in advance, but they were chosen over the course of the experiment by analyzing regeneration. For example, fragments regenerated quickly after the first three desiccation periods, so the intervals between periods were short during the early portion of the experiment. However, regeneration was substantially lower after 17 days, so we extended the final period to 26 days of desiccation before finishing the experiment. After 26 days of desiccation, almost no regeneration occurred (see Results). Fragments not submitted to desiccation were used as controls (zero days of desiccation).

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Recent studies indicate that some troublesome exotic species can accumulate high biomasses (Pott VJ & Pott A 2003; Thomaz et al. 2009; Michelan et al. 2010). One example is the Poaceae Urochloa subquadripara (Trin.) R.D. Webster [syn. Brachiaria subquadripara (Trin.) Hitchc., Brachiaria arrecta (Hack.) Stent.], commonly known as tropical signalgrass (Teuton et al. 2004). It is native to Africa, exhibits high invasion potential (Kissman 1997) and has colonized several natural and artificial aquatic ecosystems in Brazil (Pott VJ & Pott A 2003; Thomaz et al. 2009) and in the United States (Teuton et al. 2004). According to Lorenzi (2000), this species propagates exclusively by stolons, although regeneration via seeds has also been reported (Teuton et al. 2004). Reservoirs and floodplains suffer water-level drawdown, which causes mortality of the emergent parts of tropical signalgrass (Thomaz et al. 2009; Michelan et al. 2010). Investigating the vegetative regeneration of this species is important to understand how it colonizes sediment and recovers from drought stress. In this study, we investigated how vegetative fragments of U. subquadripara recover from desiccation. We hypothesized that (i) the ability of tropical signalgrass fragments to regenerate new sprouts and colonize sediment declines with increasing time of exposure to desiccation and (ii) younger portions (apical) of tropical signalgrass stems submitted to desiccation regenerate and colonize more quickly than older (basal) ones. According to these hypotheses, we tested the predictions that sprout and root biomass, sprout length and sprout and root RGR (Relative Growth Rates) are greater in younger fragments submitted to shorter desiccation periods. The rationale for this hypothesis is that macrophytes, particularly grasses, usually regenerate most quickly from apical fragments where meristematic activity is higher (Kissman 1997; Riis et al. 2009), and that desiccation represents a stress that may reduce regeneration (Silveira et al. 2009).

Materials and Methods We collected tropical signalgrass and sediment from a floodplain lake (Upper Paraná River floodplain; 22° 45’ 24’’ S and 53° 23’ 28’’ W) in January 2009. The material was transported to the laboratory on the same day. The experiment was carried out from January to March 2009.

To evaluate differences in regeneration and colonization between different plant fragments, we used fragments from close to the tip (apical) and from the intermediate and basal portions of tropical signalgrass. From each portion, we selected a fragment with two nodes. Basal (close to the root) fragments are the oldest and apical (close to the tip) fragments are the youngest. We added sediment and water to plastic trays (45 × 38 × 9 cm). The water layer in the trays was about 5 cm deep and was maintained by adding tap water every day. Three fragments from the same individual (apical, intermediate and basal portions) in each desiccation treatment (control, 2, 5, 9, 17 and 26 days of desiccation) were added to the trays. Thus, there were five trays (five replicates), each containing 18 fragments (three plant portions × six desiccation treatments). For each fragment, the following biological attributes were measured after 26 days: dry weights (DW; g) of sprouts and roots, sprout length (cm), relative growth rates (RGR) of sprout length and biomass, number of sprouts and number of leaves per fragment. We chose a maximum of 26 days of desiccation because this period was long enough to almost completely eliminate fragment regeneration (see Results). We measured DW after drying the plant material in an oven at 80 °C until a constant weight was reached. The RGRs were calculated for each sprout (following Radford 1967) based on its increase in length, according to the equation: RGR = (ln Xt - ln Xt-1)/Δt

(1)

where Xt = final length (26 days), Xt-1 = length at the beginning of the experiment and Δt = days of desiccation. Because sprout lengths were zero at the beginning, the equation became: RGR = ln Xt/Δt

(2)

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Regeneration of an Invasive Macrophyte

We used the same equation to estimate the RGR of sprout biomass, but with sprout dry weight substituted for length. Because each fragment had, in general, more than one sprout, the mean RGR per fragment was considered in the analyses. Because there were two nodes in each fragment, we summed the sprout and root dry weights and the number of sprouts and leaves; thus, these response variables became a single number per fragment. An analysis of covariance (ANCOVA) was performed to test the homogeneity of slopes derived from the regression analysis, considering the relationship between dependent variables (sprout and roots biomass, mean sprout length, sprout length RGR and sprout biomass RGR) and the time of exposure to desiccation (in days) across the three different plant segments (apical, intermediate and basal) (Crawley 1993). In this analysis, the time of exposure to desiccation (in days) was considered to be the covariate. The nonparametric Kruskal-Wallis test was used to analyze the effect of time of exposure to desiccation and plant segment (apical, intermediate and basal) on the numbers of sprouts and leaves. We used this nonparametric analysis because these response variables are discrete. We used the program STATISTICA 7.0 (StatSoft Inc. 2007) to perform the ANCOVA and the Kruskal-Wallis analysis.

Results Desiccation negatively affected the regeneration and colonization of tropical signalgrass, and the apical, intermediate and basal portions responded differently to desiccation stress. Fragments not submitted to desiccation (control) or submitted to 2 days of desiccation had 100% regeneration (i.e., all fragments emitted sprouts). A lesser but still high degree of regeneration was found after the 5 and 9 days of desiccation (93%). Much lower regeneration occurred after 17 and 26 days of desiccation (40 and 13%, respectively). Among plants submitted to 5 and 9 days of desiccation, only one basal fragment did not regenerate. Among plants submitted to 17 days of desiccation, regeneration did not occur in four basal fragments, three intermediate fragments and two apical fragments. However, only two intermediate fragments regenerated in the treatment submitted to 26 days of desiccation. In general, all attributes were relatively stable from zero to 9 days of desiccation in fragments from the apical and intermediate segments, but a steep decrease occurred after 26 days of desiccation. In contrast, generally lower and constantly decreasing values with increasing time of exposure to desiccation were observed in the basal segments of tropical signalgrass (Figures 1 and 2). After 26 days of growth, control fragments from the apical portion had the highest mean values of sprout (0.26 ± 0.08 SE g DW) and root biomass (0.32 ± 0.15 SE g DW). Desiccation period and fragment position negatively and significantly affected sprout and root

biomasses, but the interaction between desiccation period and position was not significant (Table 1; Figure 1a, b). The greatest mean sprout length was also recorded for the apical portion in the control (18.02 ± 3.73 SE cm); however, at least one intermediate fragment exhibited a tall sprout, even after 17 days of desiccation (25.6 cm). The mean sprout length was also negatively and significantly affected by the desiccation period and fragment position (Table 1; Figure 1c). The interaction was not significant (Table 1). The sprout-length RGR and sprout-biomass RGR decreased significantly with days of desiccation. These variables were also affected significantly by fragment position (Table 1; Figure 1d, e). For both variables, the interactions were not significant (Table 1). The highest mean sprout length RGR (0.11 ± 0.009 SE d–1) and sprout biomass RGR (0.008 ± 0.002 SE d–1) were found for the apical fragment in the control treatment, corresponding to doubling times of 6 and 86 days, respectively. However, the lowest mean sprout length and biomass RGR values were close to zero in the three types of fragments after 26 days of desiccation. In contrast to the other variables, the mean number of sprouts was highest in the intermediate fragments after 5 days of desiccation (2.2 ± 0.2 SE), and the highest number of leaves was found in apical fragments in the same treatment (9.6 ± 0.7 SE). Both variables were negatively and significantly affected by desiccation period and fragment position (Table 2; Figure 2a, b).

Table 1. Results of an ANCOVA for effects of fragment position (apical, intermediate and basal) and desiccation (0, 2, 5, 9, 17 and 26 days) and their interactions on sprout biomass, root biomass, sprout lenght, sprout-lenght RGR and sprout-biomass RGR.

Sprout biomass (g) Fragment position Days of desiccation Interaction Root biomass (g) Fragment position Days of desiccation Interaction Mean sprout length (cm) Fragment position Days of desiccation Interaction Sprout-length RGR Fragment position Days of desiccation Interaction Sprout-biomass RGR Fragment position Days of desiccation Interaction

2 1 2

14.71 41.26 2.62

<0.001 <0.001 0.079

2 1 2

9.41 26.21 1.81

<0.001 <0.001 0.169

2 1 2

8.25 77.81 0.83

0.001 <0.001 0.44

2 1 2 2 1 2

3.4 128.6 0.1 14.73 45.6 2.5

0.038 <0.001 0.868 <0.001 <0.001 0.089

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Figure 1. Sprout biomass a), root biomass b), sprout length c), and RGRs measured for sprout length d) and biomass e) of fragments of Urochloa subquadripara taken from apical (closer to the tip), intermediate and basal parts of the plant. Attributes were measured across different desiccation periods. Mean values and standard errors of five replicates are shown. Lines represent ordinary leastsquares fits.

Discussion Several species of the family Poaceae are invasive in terrestrial (Kissman 1997; Petenon & Pivello 2008) or freshwater aquatic ecosystems (Douglas & O’Connor 2003; Michelan et al. 2010). The success of plant invaders can be explained by several features, such as their ability to reproduce vegetatively, high growth rates, high seed production, high seed germination rates, easy dispersal, short life cycles and production of allelopathic agents (Elton 1958; Pyšek et al. 2008). These features are common in several species of the family Poaceae (e.g., Urochloa mutica and Urochloa subquadripara, our target species).

Vegetative reproduction via stolons is common in Poaceae (Kisssman 1997). This ability to reproduce by stolons is due in part to the presence of culms (i.e., grass stems), which are segmented by nodes. Nodes in Poaceae are the regions with the highest meristematic activity and where roots, branches and other culms originate (Kissman 1997). Thus, nodes have the potential to originate new individuals. Our results show a high degree of regeneration (emission and growth of sprouts and roots) and colonization (rooting in sediment) from the nodes of tropical signalgrass. Although we are not aware of studies addressing the growth and regeneration of tropical signalgrass, high growth rates (Bianchini Jr. et al. 2010) and regeneration rates (greater

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Regeneration of an Invasive Macrophyte

Figure 2. Numbers of sprouts a) and leaves b) per fragment of Urochloa subquadripara taken from apical (closer to the tip), intermediate and basal parts of the plant. Attributes associated with regeneration were measured across different desiccation periods. Mean values and standard errors of five replicates are shown. Table 2. Results of a Kruskal Wallis test for effects of fragment position (apical, intermediate and basal) and desiccation (0, 2, 5, 9, 17 and 26 days) on the numbers of sprouts and leaves.

Number of sprouts Fragment position Days of desiccation Number of leaves Fragment position Days of desiccation

2 5

6.91 42.73

0.031 <0.001

2 5

12.68 35.05

0.001 <0.001

than 60%) are common in other invasive macrophytes (Barrat-Segretain & Bornette 2000; Barrat-Segretain & Cellot 2007; Riis et al. 2009). However, regeneration of some macrophyte species is higher when fragments contain the apical tips (Riis et al. 2009). For tropical signalgrass, regeneration does not depend on the presence of apical tips; we obtained 100% regeneration from all fragment positions in the control treatment. This high regeneration rate may partly explain the high frequency of occurrence of tropical signalgrass (Pott VJ & Pott A 2003; Thomaz et al. 2009) and its impacts on natives (Michelan et al. 2010). However, even though intermediate and basal fragments were capable of regenerating, both regeneration and colonization were more likely in fragments taken from closer to the apical portion of the parent plants, which are younger than the basal fragments. Younger tissues of macrophytes such as Potamogeton perfoliatus L. and Elodea canadensis Michx. exhibit higher meristematic activity than older tissues (Riis et al. 2009); this pattern is consistent with our results. Notably, our results indicate that tropical signalgrass is highly resistant to desiccation stress: some regeneration occurred even after 26 days of desiccation. These results have practical implications. For example, water level drawdown is a strategy suggested to control aquatic macrophytes in

some situations (Wade 1990). However, our results indicate that plans meant to control and manage tropical signalgrass will not succeed if short-term drawdowns are used. The responses of regeneration and colonization to desiccation stress vary in different macrophyte species. Following intermediate desiccation periods (i.e., a few days of drought), for example, some species may grow and develop faster, leading to successful colonization (Riss et al. 2009) and changing the composition and structure of macrophyte assemblages (Hill et al. 1998). Results similar to ours (i.e., high rates of regeneration and colonization even after 9 days of desiccation) have been found for other species of aquatic macrophytes, such as Elodea nuttallii (Planchon) St. John, Hydrilla verticillata (L. f.) Royle, Egeria najas Planch. and Egeria densa Planch., whose survival rates and sprout and root lengths did not differ between various desiccation periods and controls (Barrat-Segretain & Cellot 2007; Silveira et al. 2009). In an experiment using herbicide, Carbonari et al. (2003) have shown that tropical signalgrass is difficult to control and that the regeneration of this species is even higher when an intermediate level of herbicide is used. Although herbicide and desiccation affect plant physiology differently, both are sources of plant of stress; thus, it seems that intermediate levels of stress may actually increase the regeneration of tropical signalgrass. In summary, our data support the hypothesis that regeneration of new sprouts and colonization of the sediment decline with increasing time of exposure to desiccation in tropical signalgrass fragments. Desiccation negatively and significantly affected all attributes that we measured (sprout and root biomasses, mean sprout length, sprout length RGR, sprout biomass RGR and the numbers of sprouts and leaves). However, these results are valid only for longer periods of desiccation (>10 days); at intermediate levels of desiccation (up to 9 days), fragments regenerated and

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Natureza & Conservação 8(2):133-139, December 2010

colonized sediments quite well. Our hypothesis that younger portions of this species are more likely to regenerate and colonize than older ones was also supported; fragments taken from closer to the apical portions grew and colonized sediments more often than those taken from farther from the apical portions.

Elton CS, 1958. The Ecology of Invasions by Animals and Plants, Methuen: The University of Chicago Press.

138

Finally, we contend that the presence of U. subquadripara in reservoirs with different surface areas, and in natural ecosystems such as streams and lakes, is a matter of concern because this species significantly affects the richness and composition of native macrophyte assemblages (Michelan et al. 2010). The easy regeneration of this species via fragments, as shown by our experiments, increases this concern. Any mechanical management strategy that produces fragments of U. subquadripara may not be efficient. Our results emphasize that controlling this species is difficult and that if water drawdown is applied, it will be effective only if long desiccation periods are used.

Acknowledgments P. Carvalho and M. J. Silveira thank the Parque Tecnológico da Itaipu (PDTA/FPTI-BR) and T. S. Michelan acknowledges the Brazilian Council of Research (CNPq) for providing scholarships. S. M. Thomaz is especially thankful to CNPq for continuous funding through a Research Productivity Grant. This research was funded by CNPq/Ministério da Ciência e Tecnologia through the Long-Term Ecological Research Program (site number 6) and Itaipu Binacional.

References Balian JH et al., 2008. The Freshwater Animal Diversity Assessment: an overview of the results. Hydrobiologia, 595:627-637. Barrat-Segretain MH & Bornette G, 2000. Regeneration and colonization abilities of aquatic plant fragments: effect of disturbance seasonality. Hydrobiologia, 421:31-39. Barrat-Segretain MH & Cellot B, 2007. Response of invasive macrophyte species to drawdown: The case of Elodea sp. Aquatic Botany, 87:255-261. Bianchini Jr. I et al., 2010. Growth of Hydrilla verticillata (L.f.) Royle under controlled conditions. Hydrobiologia, 644:301-312. Carbonari CA et al., 2003. Controle de Brachiaria subquadripara e Brachiaria mutica através de diferentes herbicidas aplicados em pós-emergência. Planta Daninha, 21:79-84. Combroux ICS & Bornette G, 2004. Propagule banks and regenerative strategies of aquatic plants. Journal of Vegetable Science, 15:13-20. Crawley MJ, 1993. GLIM for ecologists. Oxford: Blackwell Scientific Publications. Douglas MM & O’Connor RA, 2003. Effects of the exotic macrophyte, para grass (Urochloa mutica), on benthic and epiphytic macroinvertebrates of a tropical floodplain. Freshwater Biology, 48:962-971.

Henry-Silva GG et al., 2008. Growth of free-floating aquatic macrophytes in different concentrations of nutrients. Hydrobiologia, 610:153-160. Hill NM et al., 1998. A hydrological model for predicting the effects of dams on the shoreline vegetation of lakes and reservoirs. Environmental Management, 22:723-736. Jenkins M, 2003. Prospects for biodiversity. Science, 302:1175-1177. Kissman KG, 1997. Plantas infestantes e Nocivas. São Paulo: BASF. Tomo I. Lorenzi H, 2000. Plantas daninhas do Brasil: terrestres, aquáticas, parasitas e tóxicas. São Paulo: Nova Odessa. Michelan TS et al., 2010. Effects of an exotic-invasive macrophyte (tropical signalgrass) on native plant community composition, species richness and functional diversity. Freshwater Biology, 55:1315-1326. Petenon D & Pivello VR, 2008. Plantas invasoras: representatividade da pesquisa dos países tropicais no contexto mundial. Natureza & Conservação, 6:65-77. Pott VJ & Pott A, 2003. Dinâmica da vegetação aquática do Pantanal. In: Thomaz SM & Bini LM (Ed.). Ecologia e manejo de macrófitas aquáticas. Maringá: Eduem. p. 145-162. Pyšek P et al., 2008. Geographical and taxonomic biases in invasion ecology. Trends in Ecology & Evolution, 23:237-244. Radford PJ, 1967. Growth analysis formulae: their use and abuse. Crop Science, 7:171-175. Reed DC et al., 2000. The role of dispersal and disturbance in determining spatial heterogeneity in sedentary organisms. Ecology, 81:2011-2026. Rejmánek M et al., 2005. Ecology of invasive plants: State of the art. In: Moody HA et al. (Ed.). Invasive alien species: a new synthesis. Washington: Island Press. p. 104-162. Riis T & Sand-Jensen K, 2006. Dispersal of plant fragments in small streams. Freshwater Biology, 51:274-286. Riis T, 2008. Dispersal and colonisation of plants in lowland streams: success rates and bottlenecks. Hydrobiologia, 596:341-351. Riis T et al. 2009. Regeneration, colonisation and growth rates of allofragments in four common stream plants. Aquatic Botany, 90:209-212. Santamaria L, 2002. Why are most aquatic plants widely distributed? Dispersal, clonal growth and small-scale heterogeneity in a stressful environment. Acta Oecologica, 23:137-154. Silveira MJ et al., 2009. Effects of desiccation and sediment type on early regeneration of plant fragments of three species of aquatic macrophytes. International Review of Hydrobiology, 94:169-178. StatSoft Inc, 2007. STATISTICA. Data analysis software system. Version 8.0. Available from: <www.statsoft.com>.

Regeneration of an Invasive Macrophyte

Teuton TC et al., 2004. Factors affecting seed germination of tropical signalgrass (Urochloa subquadripara). Weed Science, 52:376-381. Thomaz SM et al., 2009. Temporal trends and effects of diversity on occurrence of exotic macrophytes in a large reservoir. Acta Oecologica, 35:614-620.

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Wade PM, 1990. Physical control of aquatic weeds. In: Pieterse AH & Murphy KJ (Ed.). Aquatic weeds. The ecology and management of nuisance aquatic vegetation. Oxford: Oxford Science Publications. p. 93-135.

Received: August 2010 First Decision: August 2010 Accepted: September 2010

Research Letters

Brazilian Journal of Nature Conservation

Fish as Potential Controllers of Invasive Mollusks in a Neotropical Reservoir Camila Ribeiro Coutinho de Oliveira1, Rosemara Fugi2, Kelly Patrícia Brancalhão2 & Angelo Antonio Agostinho2,* 1

Limnobios Consultoria em Ambientes Aquáticos, Maringá, PR, Brazil

Universidade Estadual de Maringá – Nupelia/DBI/PEA, Maringá, PR, Brazil

Abstract We evaluated the importance of the invasive mollusks (Corbicula fluminea and Limnoperna fortunei) in the diets of the fish species through analysis of stomach contents of fish caught in the commercial fishery in the years 2005 and 2006 in the Itaipu Reservoir (Upper Paraná River). The degree that these bivalves are processed in the digestive tract was used to identify potential controllers. Food items were evaluated by using the methods of occurrence, volume, and Feeding Index (IAi). Out of the 36 fish species present in the fishery landings, 24 consumed L. fortunei and 12 consumed C. fluminea. In order to evaluate the possibility that the bivalves pass through the digestive tract of the predator without any damage, their degree of digestion was evaluated. Only species with morphological pre-adaptations for a malacophagous diet, or that managed to crush the shells of these bivalves were very successful with this kind of food. In particular, Megalancistrus parananus, Leporinus obtusidens, and Leporinus macrocephalus may be regarded as potential biological control agents for these invasive populations. Key words: Introduced Species, Biological Control, Golden Mussel, Corbicula, Limnoperna.

Introduction The introduction of species is widely recognized as a serious problem that involves huge economic losses and constitutes one of the most important global threats to biodiversity and ecosystem functioning (Pimentel et al. 2000). Currently, a large number of organisms are released outside their native range, a process that involves multiple taxonomic groups and many types of environments in different parts of the world (Vitousek et al. 1997). Given the current ease of transport, innumerable introductions can take place within a short space of time. The impacts on native wildlife resulting from the proliferation of an introduced species can be categorized as: i) competition for resources, ii) exacerbation of predation, iii) changes in the habitat and functioning of the system, iv) introduction of pathogens and parasites, and v) genetic changes (Agostinho et al. 2006). An important aggravating factor is that other environmental disturbances often occur in conjunction with introductions (reservoir construction, fishing, pollution, habitat alterations), increasing the chances of colonization *Send correspondence to: Angelo Antonio Agostinho Universidade Estadual de Maringá – Nupelia/DBI/PEA, Av. Colombo, 5790, CEP 87020-900, Maringá, PR, Brazil E-mail: agostinhoaa@gmail.com

of the invader and intensifying its negative effects, including causing extinctions (Gurevitch & Padilha 2004). Two recent cases of species introduced into inland waters of Brazil have drawn the attention of biologists, engineers, and economists. This is the arrival and dispersal of two species of Asian bivalves, the Asian clam Corbicula fluminea and the golden mussel Limnoperna fortunei (Darrigran 2002). These species were introduced into the mouth of the La Plata River basin through the release of ballast water from ships arriving from Southeast Asia. The first record of C. fluminea in South America occurred in the late 1960s (Ituarte 1981), and of L. fortunei in the early 1990s (Darrigran & Pastorino 1993). Since the 1990s, C. fluminea has been found in the floodplain of the Upper Paraná River (Takeda et al. 2004), and in 1994 a population boom of this species occurred throughout the length of the Itaipu Reservoir (Okada 2001). The first record of L. fortunei in this reservoir was in August 2001, and it was found in the floodplain located upstream of this reservoir in December 2002 (Takeda et al. 2004). Since then, these species have been causing negative effects on the native fauna, dam operations, and water-supply uptake (Takeda et al. 2007; Suriani et al. 2007).

141

Fish as Potential Controllers of Invasive Mollusks

Fish have been reported to consume these mollusks, particularly the Doradidae Pterodoras granulosus, whose diet contains large amounts of C. fluminae and L. fortunei (Ferriz et al. 2000; Gaspar da Luz et al. 2002; Cantanhêde et al. 2008). Although the eradication of aquatic invasive species is a difficult task (Vitule et al. 2009), it is expected that intensive consumption can have an important role in their control. In this case, however, it is necessary to gain a comprehensive understanding of ecosystem processes and elements that operate in the reduction of the invasive populations. In this context, the main objective of the present study was to evaluate the components of the fish fauna that may have the conditions to control these invasive mollusks in Itaipu Reservoir, through analysis of stomach contents during two annual periods. Specifically we sought to determine the importance of the invasive mollusks in the diets of the fish species, and the degree that these bivalves are processed in the fish digestive tract.

Materials and Methods Fish were obtained from the landings of the commercial fishery in Itaipu Reservoir (Upper Paraná River), during the years 2005 and 2006. All fishes sampled were measured and weighed, and stomachs and intestines containing food were fixed in 4% formalin. The gastro-intestinal contents of 3,752 specimens were analyzed (Table 1), and the food items were identified under optical and stereoscope microscope. The analyses were performed by using the methods of frequency of occurrence (FO) and volumetric frequency (FV) (Hyslop 1980). The importance of food items was inferred from the combination of these methods in the Feeding Index (IAi) (Kawakami & Vazzoler 1980).

The IAi values were ordered from highest to lowest, and cumulative summations were made based on which were classified as preferred – up to 50%; secondary – between 50 and 75%; and accessory – all remaining items (Rosecchi & Nouaze 1987). For species in which the invasive bivalves were prominent in the diet, the degree of digestion (DD) of mollusks found in the stomach and intestine was evaluated according to the degree of integrity, using the following scale: 0 = intact, 1 = almost intact, 2 = valves fragmented and 3 = muscle digested.

Results Diet composition The general analysis of the diet of 36 fish species sampled during 2005 (31 species) and 2006 (32) revealed the presence of 53 food items. These were grouped into nine categories, namely algae, basidiomycetes, crustaceans, detritus/sediment, insects, mollusks, other invertebrates, fish, and higher plants. Taking as a criterion the occurrence and the values of the Feeding Index for the different categories of food items analyzed in 2005 and 2006, we noticed that algae, higher plants, detritus/sediment, mollusks, and insects occurred widely among the fish species. These were taken by at least 65% of the 31 species analyzed in 2005, and 71% of the 32 in 2006 (Table 2). Mollusks were recorded in the stomach of 71 and 84% of the species in 2005 and 2006, respectively. However, mollusks proved to be dominant items for only 13 and 15% of the species during these years, respectively.

Table 1. Number of stomach contents analyzed (N), and standard length range (cm) (SL).

Species

A. osteomystax

118

17.1-28.6

M. lippincottianus

210

8.2-34.2

A. ucayalensis

18.7-23.6

P. anisitsi

19.3-43.4

C. jenynsii

15.4-17.4

P. galeatus

195

11.8-19.5

C. kelberi

19.3-35.2

P. granulosus

1368

16.3-57.3

H. commersoni

19.9-33.7

P. maculatus

9.0-32.5

H. regani

20.3-32.8

P. mesopotamicus

12.4-38.0

H. ternetzi

16.5-30.5

P. squamosissimus

480

12.5-55.9

H. aurogutatus

16.5-25.3

P. lineatus

15.6-43.2

H. malabaricus

16.2-34.7

P. motoro

22.5-33.7

Hypostomus sp.

13.3-28.8

P. ornatus

18.9-32.6

101

14.7-26.3

P. pirinampu

20.3-55.3

I. labrosus L. friderici

18.0-28.5

R. aspera

15.4-40.5

L. macrocephalus

21.0-53.0

R.. vulpinus

20.8-52.7

L. obtusidens

19.5-25.8

S. borellii

15.5-31.5

15.1-21.4

S. marginatus

11.7-38.0

Loricaria sp.

13.8-43.6

S. pappaterra

211

9.2-20.5

L. platymetopon

21.5-28.7

S. nasutus

18.4-31.2

165

18.0-44.7

S. maculatus

10.3-24.6

L.lacustris

M. parananus

Oliveira et al.

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Natureza & ConservaĂ§ĂŁo, 8(2):140-144, December 2010

Table 2. Number of species in which different food categories were recorded (Occur) and in which they were preferred (Preferred).

Item

2005

2006

Occur Preferred

Algae

Basidiomycetes

Crustaceans

Detritus/sediment

Occur Preferred

Insects

Mollusks

Other invertebrates

Fish

Higher plants

Total species

2 32

Diets of the potential control species for invasive mollusks Among the fish that consumed mollusks, 24 species ate L.Â fortunei and 12 ate C. fluminea. However, the importance of these items in the diet showed high interspecific and interannual variation. Limnoperna fortunei was recorded in the digestive tract of 15 out of the 31 fish species analyzed in 2005, and in 24 of 32 examined in 2006 (Figure 1). However, this item was prominent in the diet of only 6 species analyzed in 2005, i.e., Megalancistrus parananus (84%), Leporinus macrocephalus (79%), Pterodoras granulosus (47%), Hypostomus ternetzi (37%), Serrasalmus marginatus (17%), and Piaractus mesopotamicus (15%). In 2006, although more species used this invasive bivalve in their diet, in only three of them the golden mussel reached meaningful proportions: M. parananus (82%), Leporinus obtusidens (60%), and P. granulosus (29%).

Figure 1. Importance of the invasive golden mussel Limnoperna fortunei (a) and Asian clam Corbicula fluminea (b) in the diet of fish species from the commercial fishery landings in Itaipu Reservoir, during the years 2005 and 2006.

The clam C. fluminea was present in the diet of 10 species in 2005, and of 9 species in 2006. Notably, however, this clam did not comprise more than 10% of the diet of any of them (Figure 1).

Degree of digestion of the bivalves Of the seven species that consumed meaningful amounts of L. fortunei, L. obtusidens, L. macrocephalus and M. parananus consumed the highest percentages of this mollusk, and with the maximum degree of digestion (83, 80 and 44%, respectively) (Figure 2). Approximately 70% of the digestive tracts of L. obtusidens and M. parananus contained mussels showing evidence of at least the beginning of digestion. For L. macrocephalus this percentage was 100%. In the digestive tract of P. mesopotamicus, about half of the mussels were moderately digested (Degree of Digestion = 2). The highest incidences of intact mussels (DD = 0) were observed for P. granulosus (78%) and S. marginatus (89%) (Figure 2).

Figure 2. Percentage of occurrence of the invasive bivalves Limnoperna fortunei and Corbicula fluminea in different degrees of digestion (DD; DD-0, intact; DD-1, almost intact; DD-2, valves fragmented; DD-3, muscle digested) in the digestive tract of malacophagous fish species in the Itaipu Reservoir.

Fish as Potential Controllers of Invasive Mollusks

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Regarding the proportion of clams C. fluminea that were digested to some degree by the three species that consumed meaningful amounts of it, S. marginatus and I. labrosus were inefficient in digesting clams, and for P. granulosus, although the majority of shellfish found in the gut were not digested (DD = 0), a percentage of this item was found in advanced stages of digestion (Figure 2).

In this study, M. parananus, Leporinus obtusidens and L. macrocephalus showed to be important mollusk predators. Among the species considered here as malacophagous, M. parananus seems to act most effectively in the control of L. fortunei because it crushes the mussel. Characterized as a detritivore, this species has a ventral mouth equipped with strong maxillar and pharyngeal teeth compared to other species of loricariids (Delariva & Agostinho 2001).

Discussion

Leporinus obtusidens, an omnivorous species, profoundly altered its diet in the Middle Paraná River after the arrival of L. fortunei (Garcia & Montalto 2006). This species possesses a small terminal mouth provided with large strong incisiform oral teeth, in addition to pharyngeal teeth. The small mouth opening means that ingested organisms must be small in size, or be fragmented by the oral teeth. The conformation of their premaxillary and mandibular teeth affords a wide grinding surface that allows the fragmentation of hard organisms (Occhi & Oliveros 1974), which explains the presence of broken bivalve shells in their stomach contents.

The ingestion of mollusks by about 2/3 of the species examined can be explained by the massive proliferation of the invading bivalves C. fluminea and L. fortunei in the Itaipu Reservoir, as reported by Takeda et al. (2004, 2007). These invasive bivalves were found in the diet of several fish species, and comprised a significant share of the gastric contents of some of them. The presence of these bivalves attests to the high feeding plasticity, and especially the trophic opportunism of the fish, which exploited an unusual resource, invasive species that have recently become established locally, but are now easily available. Of the two invasive species, L. fortunei was the more preyed upon, and was recorded in more than 65% of the species analyzed. This indicates that this mussel may be more available than the clam C. fluminea. Unlike the latter, L. fortunei colonizes hard substrates, and the lack of competitors for this type of substrate in the Itaipu Reservoir may have led to increased density of this mollusk (Takeda et al. 2002). The shell of L. fortunei is more fragile than that of C. fluminae, and it is easier to break and to access its tissues (Cantanhêde et al. 2008), which may also account for the fishes preference for the mussels. Different groups of fish incorporated the invasive mollusk into their diet, especially species with omnivorous or detritivorous habits (Garcia & Montalto 2006). However, ingestion of these organisms does not necessarily imply a successful strategy for the predator, as evidenced by the high incidence of intact (undigested) prey in the digestive tract. The presence of intact mollusks in several fish species suggests that the mollusks cannot be digested and probably pass through the fish alive. In this case, these fish species may, rather than controlling these invasive organisms, instead contribute to their dispersal. This possibility was suggested for P. granulosus by Cantanhêde et al. (2008), who recorded a large number of closed shells of C. fluminae present in the end of the intestine of this species. Only species with morphological preadaptations for a malacophagous diet, or that can break the valves of these bivalves, are very successful with this type of food and can be considered predators. Garcia & Montalto (2006) noticed that fish that predate these exotic mollusks generally have mouths adapted for suction or are provided with strong incisiform or molariform teeth, and may have pharyngeal teeth with different degrees of development.

Leporinus macrocephalus, a species previously absent from the Upper Paraná River, has, like L. obtusidens, a terminal mouth, a small oral cleft, and a relatively large buccopharyngeal cavity, with incisiform oral teeth and hooked pharyngeal teeth, responsible for the holding and maceration of the prey (Rodrigues et al. 2006). The catfish P. granulosus is the species most often taken in the commercial fishery in Itaipu Reservoir (Okada et al. 2005). Its consumption of L. fortunei and C. fluminae has been documented for the Paraná River basin (Ferriz et al. 2000; Gaspar da Luz et al. 2002; Cantanhêde et al. 2008). The pharyngeal teeth, arranged in plates, are small, and the species’ diet has been described as omnivorous, opportunistically taking the most available food (Gaspar da Luz et al. 2002). In contrast to the observations for Leporinus spp. and M. parananus, P. granulosus ingests the whole bivalves, and although it can digest a portion of them, the integrity of many individuals present in the digestive tract suggests that this great migrator can contribute to the dispersal of these invaders, especially C. fluminea. In summary, our findings show that these bivalve invaders are widely consumed by the components of the ichthyofauna of the Itaipu Reservoir, as a result of the feeding plasticity of the fish and the high availability of this new food resource in the basin. The effectiveness of the fish species in using this resource and the role of this group of animals in the control of these invaders depends, however, on the degree of pre-adaptation of the species to a malacophagous diet. In discussions about the role of fish in the biological control of invasive species of bivalves, there is a consensus that, even though the fish cannot exterminate them, they contribute to maintaining their populations at lower levels of density (Garcia & Montalto 2006; Cantanhêde et al. 2008). However, this contribution is naturally species-specific, and experimental investigations on the possible disperser capacity of some of the fish species are needed to clarify this role.

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Oliveira et al.

Acknowledgements The authors are grateful to the Núcleo de Pesquisas em Limnologia, Ictiologia e Aqüicultura of the Universidade Estadual de Maringá, in the persons of Edson K. Okada and Maria de Lourdes B. Nunes who kindly provided the biological material for the development of this study. Thanks are also due to CNPq for a scientific initiation grant to Camila R. C. Oliveira, and a researcher grant to Angelo A. Agostinho.

References Agostinho AA, Pelecice FM & Júlio Jr HF, 2006. Biodiversidade e Introdução de Espécies de Peixes: unidades de conservação. In: Campos JB, Tossulino MGP & Muller CRC (Ed.). Unidades de Conservação: ações para valorização da biodiversidade. Curitiba: Instituto Ambiental da Paraná. p. 95-117. Cantanhêde G et al., 2008. Invasive molluscs in the diet of Pterodoras granulosus (Valenciennes, 1821) (Pisces, Doradidae) in the Upper Paraná River floodplain, Brazil. Ecology of Freshwater Fish, 17:47-53. Darrigran G & Pastorino G, 1993. Bivalvos invasores en el Rio de la Plata, Argentina. Comunicaciones de la Sociedad Malacologica del Uruguay, 7:309-313. Darrigran G, 2002. Potencial impact of filter-feeding invaders on temperate inland freshwalter environments. Biological Invasions, 4:145-156.

Natureza & Conservação, 8(2):140-144, December 2010

Kawakami E & Vazzoler G, 1980. Método gráfico e estimativa de índice alimentar aplicado no estudo de alimentação de peixes. Boletim do Instituto Oceonográfico, 29:205-207. Occhi RN & Oliveros OB, 1974. Estudio anátomo-histológico de la cavidad bucofaringea de Leporinus obtusidens Valenciennes y su relación com el régimen alimentario (Pisces, Tetragonopteridae). Physis, 33:77-90. Okada EK, 2001. Gradientes espaço-temporais na pesca artesanal do reservatório de Itaipu. [Tese]. Maringá: Universidade Estadual de Maringá. Okada EK, Agostinho AA & Gomes LC, 2005. Spatial and temporal gradients in artisanal fisheries: a case study of the Itaipu Reservoir, Brazil. Canadian Journal of Fisheries and Aquatic Sciences, 62:714-724. Pimentel D et al., 2000. Environmental and economic costs of nonindigenous species in the United States. Bioscience, 50:53-65. Rodrigues SS, Navarro RD & Menin E, 2006. Adaptações anatômicas da cavidade bucofaringiana de Leporinus macrocephalus Garavello & Britiski, 1988 (Pisces, Characiformes, Anostomidae) em relação ao hábito alimentar. Biotemas, 19:51-58. Rosecchi L & Nouaze Y, 1987. Comparaison de cinq indices alimentaries utilisés dans l’ analyse des contenus stomacaux. Revuedes Travaux de I’Institute de Peches Maritimes, 49:111-123.

Delariva RL & Agostinho AA, 2001. Relationship between morphology and diets of six neotropical loricariids. Journal of Fish Biology, 58:832-847.

Suriani AL, Franca RS & Rocha O, 2007. A malacofauna bentônica das represas do médio rio Tietê (São Paulo, Brasil) e uma avaliação ecológica das espécies exóticas invasoras, Melanoides tuberculata (Müller) e Corbicula fluminea (Müller). Revista Brasileira de Zoologia, 24:21-32.

Ferriz RA et al., 2000. Alimentacion del Pterodoras granulosus (Valenciennes), (Pisces, Doradidae) em la baja cuenca del Plata. Revista del Museo Argentino de Ciências Naturales, 2:151-156.

Takeda AM et al., 2002. Limnoperna fortunei dispersion – a new macrobenthic population in the Itaipu Reservoir (Brazil). In: Anais do Symposium on Reservoir Management in Tropical and Sub-tropical Regions. Foz do Iguaçu, PR, Brazil. p. 1-7.

Garcia M & Montalto L, 2006. Los peces depredadores de Limnoperna fortunei en los ambientes colonizados. In: Darrigran G and Damborenea C (Ed.). Bio-invasion del mejillón dorado en el continente americano. La Plata: Edulp. p. 111-127.

Takeda AM, Fujita DS & Fontes Jr. HM, 2004. Perspectives on exotic bivalves proliferation in the Upper Paraná River floodplain. In: Agostinho AA et al. (Ed.). Structure and functioning of the Paraná River and its floodplain. Maringá: EDUEM. p.97-100.

Gaspar da Luz KD et al., 2002. Alterations in the Pterodoras granulosus (Valenciennes, 1833) (Osteichthyes, Doradidae) diet due to the abundance variation of a bivalve invader species in the Itaipu Reservoir, Brazil. Acta Scientiarum, 24:427-432. Gurevitch J & Padilha DK, 2004. Are invasive species a major cause of extinctions? Trends in Ecology and Evolution, 19:470-474. Hyslop EJ, 1980. Stomach contents analysis review of methods and their applications. Journal of Fish Biology, 17:411-429. Ituarte CF, 1981. Primera noticia acerca de la presencia de pelecípodos asiáticos en el área rioplatense. Neotropica, 27:79-82.

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Received: June 2010 First Decision: July 2010 Accepted: August 2010

Research Letters

Brazilian Journal of Nature Conservation

Dealing with Data Uncertainty in Conservation Planning Kerrie Ann Wilson The University of Queensland, School of Biological Sciences, Brisbane, Queensland, Australia

Abstract Conservation planning analyses often employ data on biodiversity and sometimes vulnerability and these data are generally assumed to be accurate and correct. Here, different ways of exploring uncertainty associated with typical input data used for conservation planning are illustrated. First the uncertainty associated with predicted species distribution data is measured, summarised, and visualised. Second the uncertainty associated with the choice of vulnerability model is evaluated using Bayesian Model Averaging and the implication of this uncertainty on inference about key relationships associated with native forest conversion is assessed. The approaches used to assess uncertainty are applicable to any conservation planning exercise and such assessments will increase confidence in the products developed and reduce the risk that conservation effort is misdirected. Key words: Conservation Planning, Bayesian Model Averaging, Species Distribution Models, Uncertainty, Vulnerability.

Introduction Conservation planning is the process of locating and designing networks of terrestrial and marine protected areas to protect biodiversity in situ. The aim is to efficiently meet quantitative targets within a system of representative and complementary areas. Such analyses often rely upon data on biodiversity values and the vulnerability of these values to threatening processes (Moilanen et al. 2009). A commonly disregarded source of uncertainty in the planning process is the uncertainty associated with the input data. In the face of uncertainty, conservation planners could adopt a risk-averse attitude. Being risk-averse is to avoid risks associated with uncertainty and preferentially seek circumstances in which the risk is minimised. One riskaverse response to errors and uncertainty is to undertake further data collection. In a global and dynamic economic climate, however, such an adjournment entails uncertainty and risks of its own: ecosystems may be degraded and species may go extinct. For this reason, and in line with the precautionary principle, conservation decisions must be made and actions must be initiated in the face of uncertainty (Ludwig et al. 1993; Moilanen et al. 2006). Conservation planners should therefore identify the errors and uncertainties in the planning process and where necessary, evaluate the sensitivity of conservation planning outcomes to these. *Send correspondence to: Kerrie Ann Wilson The University of Queensland, School of Biological Sciences, Brisbane, Queensland, 4072, Australia E-mail: k.wilson2@uq.edu.au

Errors and uncertainty associated with biodiversity data Species locality data are commonly used for conservation planning. However, these are often biased to parts of a region or towards particular species, are incomplete, or contain errors. The selection of conservation areas cannot generally be delayed pending acquisition of improved species locality data so predicted species distribution data are increasingly relied upon (Elith & Leathwick 2009). Predicted species distributions can be derived by modelling the relationship between species locality data and mapped environmental information, such as climate, terrain, and soil (Guisan & Zimmermann 2000). Predicted species distribution data can contain errors and exhibit uncertainty, due to errors in the species locality data and mapped environmental information (McKelvey & Noon 2001). In addition, errors might be introduced due to decisions made during the modelling process (Diniz-Filho et al. 2009; Wilson et al. 2005c). Often, however, predicted species distribution data are presented as accurate digital representations without measures of precision. While measures of model accuracy based on misclassifications can be useful to identify spatial locations where errors occur, most model evaluation statistics assess overall model performance and do not provide information about the spatial distribution of prediction uncertainties (Fielding & Bell 1997). Confidence intervals, which express the uncertainty associated with parameter estimation, can be generated for probabilities of species occurrence. These intervals can

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be used to summarise and present spatially the uncertainty associated with predicted species distribution data. For example, conservation planners might be interested in predictions that have wide confidence intervals, as these might be the most unreliable sources of data for that species and therefore require further field sampling. Alternatively, conservation planners might be most interested in areas where the certainty is greatest as these might represent lower risk options for investment. The first objective of this paper is to measure, summarise, and visualise the uncertainty associated with predicted species distribution data.

Errors and uncertainty associated with vulnerability data Areas of the landscape that are priorities for conservation should be those that are both vulnerable to threats and that if lost or degraded, will result in conservation targets being compromised (Araújo et al. 2002). The vulnerability of sites can be predicted using quantitative models that relate the extent of a past threat to characteristics believed to have predisposed areas to the threat (for example, proximity to roads or soil type). Presently unaffected areas that share these characteristics are identified as vulnerable (Wilson et al. 2005b). An assessment of the uncertainty associated with vulnerability information is important in order to minimise the misallocation of conservation effort. For example, if vulnerability is overestimated, scarce resources could be allocated to areas that are not in urgent need of protection. Conversely, if vulnerability is underestimated, areas that are threatened could be overlooked possibly resulting in their biodiversity values being reduced or eliminated. Estimating and exploring the uncertainty associated with vulnerability assessments could determine the sensitivity of predictions to input data and assumptions. In particular, model structural uncertainty, which arises when predictions are based on a single, ‘best’ model of a particular structure could be investigated (Burnham & Anderson 2002, page 154). An approach to dealing with model uncertainty is to avoid selecting a single, ‘best’ model but rather average over a number of possible models. Model averaging can be applied in both Frequentist and Bayesian frameworks (Araujo & New 2007). Bayesian approaches to model averaging weight each model according to its posterior probability, as determined by the support it receives from the observed data and prior knowledge. Madigan & Raftery (1994) and Raftery et al. (1997) found that model-averaged predictions are more accurate than those obtained from a single, ‘best’ model. The second objective of this paper is to use Bayesian Model Averaging to assess the uncertainty associated with the choice of vulnerability model describing the conversion of native forest to plantations in Southern Chile. The effect of model uncertainty on inference about native forest conversion and on the vulnerability predictions is investigated.

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Materials and Methods Errors and uncertainty associated with biodiversity data Predictions of occurrence of Acacia ausfeldii, a rare plant endemic to the Box-Ironbark region of Victoria (Australia) were generated using logistic regression by modelling the relationship between survey data for this species and environmental variables (Wilson et al. 2005c). In order to quantify the uncertainty associated with the probabilities of species occurrence, Wald statistic confidence intervals for the logit were calculated (Hosmer & Lemeshow 1995). This uncertainty was then summarised and visualised by (1) mapping the upper and lower bounds on the probabilities independently of the probabilities of occurrence, (2) mapping the width of the confidence intervals, where a large width indicates high uncertainty, and (3) depicting the probabilities of occurrence and the uncertainty associated with these simultaneously. These methods to summarise and visualise the uncertainty associated with predictions of species occurrence could be applied to the data from any species distribution modeling method that generates predictions of occurrence and associated measures of uncertainty (Elith et al. 2006).

Errors and uncertainty associated with vulnerability data The Bayesian approach to model averaging involves calculating predictions under each possible model. These predictions are then weighted by the posterior probability (degree of belief) of each model (Hoeting et al. 1999). A model-averaged prediction for a particular outcome (Δ) is obtained via: k

P ( Δ | D ) == ∑ P ( Δ | Sk , D ) P (Sk | D )

(1)

k =1

where P(Δ|Sk, D) is a posterior prediction of the outcome (Δ) according to model Sk and the data (D) and P(Sk|D) is the posterior probability of model Sk, given the data (D) and prior knowledge. A single, “best” model describing the conversion of native forest to plantation in south central Chile was developed in order to identify areas of native forest vulnerable to conversion (Wilson et al. 2005a). The explanatory variables available to calibrate the model were soil type, annual rainfall, minimum annual temperature, slope, altitude, latitude, distance to towns, distance to roads, and distance to timber mills (with distance to towns, distance to timber mills and minimum annual temperature found to be correlated). While BMA may appear an intuitively attractive means to account for model uncertainty, there are three main difficulties associated with its implementation. First, when the number of models is large the direct evaluation of

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model-averaged predictions P(Δ|D) is computationally infeasible. For example, if the number of explanatory variables (v) is 20, the number of possible models is 2v (approximately 1 million). However, usually only a small number of these models will receive support from the data. Identifying a subset of parsimonious, data-supported models (termed Occam’s Window, Madigan & Raftery 1994) greatly reduces the number of models requiring summation. This subset can be identified by removing any model that has a posterior probability far less than the best model and then removing any model that has a lower posterior probability than any simpler sub-model. The leaps and bounds algorithm (Furnival & Wilson 1974) provides a tool for searching for this subset. The second obstacle to BMA is that the higher order integrals implicit in calculating posterior model probabilities can be analytically intractable, but approximate methods of integration have been developed (Kass & Raftery 1995). One asymptotic approximation is the Schwarz criterion (or BIC approximation, Schwarz 1978), which is reasonably accurate for large samples and computationally efficient (Volinsky & Raftery 2000). Specification of the prior belief that model Sk is the true model presents the third challenge. Clyde (2000) presented objective prior distributions for BMA of GLMs. These distributions, referred to as the Calibrated Information Criterion prior distributions (referred to herein as Clyde’s CIC prior distributions) include standard model selection criteria such as BIC, AIC (Akaike Information Criterion), and RIC (Risk Inflation Criterion). The use of Clyde’s CIC

prior distributions permits model inference based on maximum likelihood theory. To perform the BMA analysis, the S-PLUS function [BMA.GLM] (available from http://www.research.att. com/~volinsky/bma.html) was used. Three model spaces were evaluated. Each model space contains one of the correlated explanatory variables. To assess the fit of the models, the deviance was converted into an estimated D2 with values between 0.2 and 0.4 representing a very good model fit (Wrigley 1985). The Receiver Operating Characteristic (ROC) curve was employed to measure the discrimination ability of the models (Pontius & Schneider 2001), with areas greater than 0.8 indicating good discrimination. A random selection of the data was withheld to validate the models.

Results The probability of occurrence of A. ausfeldii was mapped along with the upper and lower bounds on the predictions (Figure 1). The range of the confidence interval width was from 0.00 to 0.42, with the upper end of this range indicating greatest uncertainty (Figure 2). The data and its uncertainty were displayed simultaneously (Figure 3). No areas were identified as having high probability of occurrence and low uncertainty. This is likely to be partially related to the greater certainty associated with areas predicted to have a high probability of occurrence, but also because confidence intervals around predicted probabilities of 0.5 tend to be wider than those closer to one or zero. This effect is due to the binomial nature of the response variable.

Figure 1. a-c) The probability of occurrence of A. ausfeldii a) expected probability of occurrence b) upper probability of occurrence c) lower probability of occurrence. The darker areas indicate a higher probability of occurrence.

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Figure 2. Width of the confidence interval associated with the probabilities of occurrence of A. ausfeldii. Darker areas depict a greater confidence interval width.

Figure 3. The probabilities of occurrence of A. ausfeldii and the associated uncertainty displayed simultaneously. A high probability of occurrence was anything greater than 0.5. A threshold of 0.2 was chosen for the high uncertainty values, as the range of the confidence interval width for the probabilities was from 0 to 0.42.

When the three model spaces were submitted to BMA, no model subsets were identified. That is, for each of the three model spaces, the model with all the variables included was identified as the favoured model. The model which included distance to towns was identified to be the best model in terms of deviance explained. However, the three models performed equally well (deviance explained ranged from 42.1 to 42.7%). The models also had equal discrimination ability as measured by the area under the ROC curve of 0.95 for each. Under the three models, the relationships between each of the explanatory variables and the presence or absence of native forest conversion were consistently positive or negative and the parameter coefficients were almost identical. Generally, the model that included minimum annual temperature gave odds that were slightly more rational (for example, greater odds of conversion occurring at low elevations).

been made in this regard, within the field of conservation planning there has been limited explicit assessment of the uncertainty associated with input data.

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Discussion Research within the field of conservation planning has focused on the development of theories and tools to design reserve networks that protect biodiversity in an efficient and representative manner. Whilst much progress has

Here, the uncertainty associated with predicted species distribution data is summarised and visualised in order to highlight information that is important to conservation planners. Displaying separately, the predicted probability of species occurrence and the upper and lower bounds on these predictions was the simplest means employed to do this. This mapping procedure will be cumbersome when there are many species of interest. Further, little information is provided on how the probabilities of occurrence and their associated uncertainty correspond. Combining the probabilities of occurrence and the estimates of their uncertainty using Boolean operators enabled the predicted species distribution data and its associated uncertainty to be depicted on a single map. This approach takes advantage of subjective associations by which people deal with a sequence of hues or particular hues. For example, red might represent areas that are important but have high risk (high probability and high uncertainty). The difficulty associated with this procedure is choosing an appropriate threshold for the display of the different categories.

Uncertainty in Conservation Planning

An alternative procedure would involve using a continuum of values to represent the probabilities of occurrence and their associated uncertainty by employing varying colour hues and saturations respectively (Davis & Keller 1997). For example, high probability of occurrence could be represented by green and low probability by red. The amount of uncertainty associated with the predictions could be represented by the colour saturation, with hazy colours representing greater uncertainty and vivid colours representing greater certainty. This approach provides the most information on how the probabilities of occurrence and their associated uncertainty correspond and avoids the need to define thresholds, but produces a complex interpretation key. Whilst a single, “best” model of native forest conversion was obtained (Wilson et al. 2005a), it was recognised that this might be only one of many possible models that perform equally well, but result in divergent predictions. Here, the single, “best” model of vulnerability was extended to incorporate model structural uncertainty, specifically that associated with the choice of explanatory variables included in the model. While there is uncertainty associated with model choice, each of the models performed well and the impact on inference and prediction was negligible. The approaches used here to assess the uncertainty associated with vulnerability data and and to visualise the uncertainty associated with predicted species distribution data are broadly applicable to conservation planning exercises. Other sources of uncertainty in conservation planning include those associated with the scale, resolution, and accuracy of input data and these also require consideration. The assessment of uncertainty in conservation planning can be used to increase confidence in the use of species and vulnerability data in conservation planning, help reduce the risk that conservation effort is misdirected, and increase the likelihood that conservation decisions are made that are optimal for biodiversity conservation.

References Araujo MB & New M, 2007. Ensemble forecasting of species distributions. Trends in Ecology & Evolution, 22:42-47.

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Elith J & Leathwick JR, 2009. Species Distribution Models: Ecological Explanation and Prediction Across Space and Time. Annual Review of Ecology, Evolution, and Systematics, 40:677-697. Elith J et al., 2006. Novel methods improve prediction of species’ distributions from occurrence data. Ecography, 29:129-151. Fielding AH & Bell JF, 1997. A review of methods for the assessment of prediction errors in conservation presence/ absence models. Environmental Conservation, 24:38-49. Furnival GM & Wilson RW, 1974. Regressions by leaps and bounds. Technometrics, 16:499-511. Guisan A & Zimmermann NE, 2000. Predictive habitat distribution models in ecology. Ecological Modelling, 135:147-186. Hoeting JA, Madigan D, Raftery AE & Volinsky CT, 1999. Bayesian Model Averaging: a tutorial. Statistical Science, 14:382-417. Hosmer DW & Lemeshow S, 1995. Confidence interval estimates of an index of quality performance based on logistic regression models. Statistics in Medicine, 14:2161-2172. Kass RE & Raftery AE, 1995. Bayes Factors. Journal of the American Statistical Association, 90:773-795. Ludwig D, Hilborn R & Walters C, 1993. Uncertainty, resource exploitation, and conservation: lessons from history. Science, 260:17-36. Madigan D & Raftery AE, 1994. Model Selection and Accounting for Model Uncertainty in Graphical Models Using Occam’s Window. Journal of the American Statistical Association, 89:1535-1546. McKelvey KS & Noon BR, 2001. Incorporating Uncertainties in Animal Location and Map Classification into Habitat Relationships Modeling. In: Hunsaker CT, Goodchild MF, Friedl MA & Case TJ (Ed.). Spatial Uncertainty in Ecology: Implications for Remote Sensing and GIS Applications. New York: Springer-Verlag. p. 72-90. Moilanen A et al., 2006. Planning for robust reserve networks using uncertainty analysis. Ecological Modelling, 199:115-124. Moilanen A, Wilson KA & Possingham HP, 2009. Spatial conservation prioritisation: quantitative methods and computational tools. Oxford: Oxford University Press.

Araújo MB, Williams PH & Turner A, 2002. A sequential approach to minimise threats within selected conservation areas. Biodiversity and Conservation, 11:1011-1024.

Pontius RG & Schneider LC, 2001. Land cover change model validation by an ROC method for the Ipswich watershed Massachusetts, USA. Agriculture, Ecosystems & Environment, 85:239-248.

Burnham KP & Anderson DR, 2002. Model selection and multi-model inference: a practical information-theoretic approach. 2nd ed. New York: Springer.

Raftery AE, Madigan D & Hoeting JA, 1997. Bayesian model averaging for linear regression models. Journal of the American Statistical Association, 92:179-191.

Clyde M, 2000. Model uncertainty and health effect studies for particulate matter. Environmetrics, 11:745-763.

Schwarz G, 1978. Estimating the dimension of a model. Annals of Statistics, 6:461-464.

Davis TJ & Keller CP, 1997. Modelling and visualising multiple spatial uncertainties. Computers and Geosciences, 23:397-408.

Volinsky CT & Raftery AE, 2000. Bayesian information criterion for censored survival models. Biometrics, 56:256-262.

Diniz-Filho JAF et al., 2009. Partitioning and mapping uncertainties in ensembles of forecasts of species turnover under climate change. Ecography, 32:897-906.

Wilson KA et al. 2005b. Measuring and incorporating vulnerability into conservation planning. Environmental Management, 35:527-543.

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Wilson KA, Newton AN, Echeverría C, Weston CJ & Burgman MA, 2005a. A vulnerability analysis of the temperate forests of south central Chile. Biological Conservation, 122:9-21.

to using predicted species distribution data. Biological Conservation, 122:99-112.

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Wrigley N, 1985. Categorical Data Analysis for Geographers and Environmental Scientists. New York: Longman.

Received: August 2010 First Decision: September 2010 Accepted: September 2010

Research Letters

Brazilian Journal of Nature Conservation

Reef Fisheries and Underwater Surveys Indicate Overfishing of a Brazilian Coastal Island Hudson Tercio Pinheiro*, Jean-Christophe Joyeux & Agnaldo Silva Martins Departamento de Oceanografia e Ecologia, Universidade Federal do Espírito Santo, Vitória, ES, Brasil

Abstract The preoccupation about fishing effects on marine ecosystems has increased sharply over the last three decades. However, little is known about the impact of multi-gear artisanal and recreational fisheries on the structure of local reef fish communities in Brazil. Fishing activities around a Brazilian coastal island were monitored while reef fish density was censused during underwater surveys (UVC). The links between frequency of capture, intensity at which species are wished and UVC density were explored. Species were classified according to their frequency of capture as regular, occasional and rare, and classified according to the intensity at which they are wished (based on size and price), as highly targeted, average and non-targeted species. Ninety-seven species were caught by fishing, the majority of them being either rarely caught or non-targeted. Nineteen species were highly targeted but rarely caught. The highly targeted species showed extremely low density in the UVC. These results put in question the sustainability of the local fishing activities. The predominance of non-targeted species in the catches and in the reefs environment studied supports the expectation that these species will be more and more captured, thus collaborating to further change the structure of the reef community. Key words: Rocky Reefs, Fishery, Catch, Discards, Tourism, Sustainable Use.

Introduction The effects of human activities on reef and island environments are a wide-world preoccupation (Roberts et al. 2002). Fishing has been recognized as the principal activity able to alter the structure of fish communities (Jennings & Blanchard 2004) because it is widespread, typically multi-specific and practiced with a variety of methods (Roberts 1995). Recent analysis suggest that the majority of fish stocks are decreasing (Myers & Worm 2003), principally those of predators (Dulvy et al. 2004). This removal often induces a cascade effect of extinctions, thus reducing the stability of the environment and turning it more vulnerable to natural and anthropic disturbances (Friedlander & DeMartini 2002). Although the information about Brazilian reef fish communities is rising in the last years, the conservation status of reef fish communities in Brazil remains largely unknown. Even with the proven effectiveness of some Brazilian MPAs (Floeter et al. 2006, Francini-Filho & Moura 2008), the number of commercially-relevant reef species *Send correspondence to: Hudson Tercio Pinheiro Departamento de Oceanografia e Ecologia, Universidade Federal do Espírito Santo, Av. Fernando Ferrari, nº 514, CEP 29075-910, Goiabeiras, Vitória, ES, Brasil E-mail: htpinheiro@gmail.com

overexploited is still growing (e.g., Klippel et al. 2005, Araujo & Martins 2009). In the other hand, the official Brazilian red list shows only 19 marine species threatened of extinction. This low number of included species probably reflects an important lack of studies (Rosa & Lima 2008). In this context, artisanal and recreational fisheries were monitored at Franceses Island (central coast of Brazil) between 2005 and 2006. In parallel, underwater visual census were realized in the same area to determine the community structure and establish the conservation status of the studied reefs. The relationships among frequency of capture, intensity at which species are wished and UVC density are explored and sustainability of resources and alternatives for coastal reefs management are discussed.

Materials and Methods Study site The study was carried out in the state of Espírito Santo, central coast of Brazil (Figure 1). The state is located in a transition area between tropical and sub-tropical zones, influenced by the warm southward-flowing oligotrophic Brazilian current and strong seasonal upwellings in the south of the state (Schmid et al. 1995). Franceses Island

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Figure 1. Location of the study area, Franceses Island, Espírito Santo, Brazil.

and associated rocky reefs (20° 55’ S and 40° 45’ W) are granitic formations 4 km off the coast. The region, overall, shelters a rich reef fish fauna (Floeter & Gasparini 2000) and the presence of 181 fish species has been recorded at the site (Pinheiro et al. unpublished data). Like the majority of coastal environments in Brazil, the local is not under any special management rules. Activities practiced are diverse and include professional-scale artisanal fishing (using gill nets, seine nets and trolling) and invertebrate exploitation (sea stars, mussels and gorgonians) along recreational activities such as hook-and-line fishing, spear-fishing and mussel extraction.

Fishing activities monitoring Ten expeditions to Franceses Island were done (March, April, June, August, September, October and December 2005, January and 2 in February 2006) totaling 51 field days. Observations were realized in the Island, during both day and night, each time that professional or recreational extractive activities were sighted. Target, by-catch and discard species were censused through direct observation and by questions (about number and species caught; wished species to capture and discard species). Whenever possible (e.g., small specimens, discard species), vouchers were preserved to confirm field identification. The main characteristics of the anthropic activities (extractive or not) monitored at Franceses Island during the ten expeditions are shown in Table 1.

Fish census During the five expeditions between October and February we carried out 208 underwater visual censuses (UVC) to determine the numerical density of fish species in the area directly used by fishers. The fishes were identified and enumerated in replicate belt transects of 20 × 2 m (40 m²), and the density calculated is therefore the number of fish per 40 m². This small-width transect method is much used by Brazilian researchers due to the low water transparency of coastal waters (about 5 m in summer) (see Floeter et al. 2006, 2007).

Data analysis Species were classified according to their frequency of capture as regular (species registered in more that 50% of the expeditions), occasional (30-40% of the expeditions), and rare (up to 20% of the expeditions). The species were also classified according to the intensity at which they are wished, based on size and price. The categories established were highly targeted (larger species [TL > 40 cm] or species with commercial value above R$ 4.00 kg–1 [selling price in 2006; about US$ 2.00 kg–1]), average (medium size [10 cm < TL < 40 cm] or low commercial values [below R$ 4.00]) and non-targeted species (small-sized fishes [TL < 10 cm] and species without commercial value).

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Table 1. Human activities realized at Franceses Island, Brazil, between March 2005 and February 2006. The frequency of occurrence (percentage of expeditions in which the activity was recorded), the number of operations monitored for the present study, the type of interaction the activity has with the local ichthyofauna, the number of fish species directly involved (i.e., caught) and a summary of the characteristics of these activities are given.

Activities

Occurrence Number Interaction Species (%)

Characteristics

Spear-fishing

direct

Practiced by visitors transported to the island by local boats for leisure, subsistence or small commerce. It is carried out in all the island surroundings but beginners are mainly found in the shallow sheltered zone and more advanced fishermen in the deeper exposed zone.

Hook-andline

direct

Practiced by tourists and visitors, rarely by professional fishers. It is done from all rocky shores that surround the island, mainly during the day but also at night by campers. The bait is shrimp and small fish purchased on the continent or fishes and small crabs caught at the island.

Tourism on island

indirect

Done by all the stakeholders of the island. Occurs throughout the year, mainly on weekends but on a daily basis during the summer. It is most practiced by tourists that spend the day in the island. Normally, the same boat carries the tourists both on the island and on a boat tour. On the island, the visitants look for leisure (barbecues, picnic, sun-bathing and sea swimming) and sports (fishing, swimming and walking).

Gill netting

direct

Practiced a few meters off the island shoreline. It is often an alternative extra-income for some fishermen that set their nets when they go out to sea to practice other fishing activities and pick up the net at their return.

Seining

direct

Practiced in the sheltered zone, daily during the summer and sporadically during the rest of the year. A large number of fishermen (at least 6) are needed. Few, however, are professionals. The nets reach over 800m of length. Two boats are used, a smaller one (moved by paddle) to surround the school and a larger motorized boat to transport the crew and the catch. If the lance fails, another haul is carried out at the same locale.

Boat tour

indirect

Practiced by local fishers that carry groups of tourists (up to 20 persons) for a tour around the island.

Invertebrate extraction

indirect

The fishermen and the local community use boats, compressors (for diving) and scrapers to collect sea stars, mussels and gorgonians. Other visitants, such as tourists, also collect mussels and sea stars but in lesser quantities.

Trolling

direct

Trolling is carried out around the entire island, especially in the exposed zone. It is practiced by local fishermen and a few tourists in their own motorized boats. It occurs during the whole year, but is most commonly practiced during summer afternoons.

Trap setting

direct

It is practiced by a few motorized boats that set fishing traps in exposed reefs. It targets small fish that are commercialized as bait and ornamentals. Lobster and octopus are also caught.

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Differences among cells of contingency tables were evaluated through chi-square (X2) tests against the null hypothesis (no difference among cells) while differences in fish density in UVC among categories for intensity of targeting and frequency of capture were tested by Kruskal-Wallis (KW) (Zar 1999) due to the lack of normality of the data.

species had the lowest recorded numeric density (Figure 3). Twenty-nine highly targeted species were rarely caught (Table 2) and were not seen during UVC (Table 3). Most of the abundant species in the community are not targeted by any fishing activities and were only occasionally or rarely caught (Table 3).

Results

Discussion

Species caught by fishing

Preservation status

Ninety-seven fish species were caught by fishing activities (Table 2), which represents 53.6% of the 181 species previously registered at the site. Captured species were classified heterogeneously among targeting categories (X2; p = 0.044) with 43% not been targeted by any activity and 34% being highly target (Figure 2; Table 2). Also, the frequency of capture varied importantly among species (X2; p < 0.001) and, while the majority of species (57%) were rarely caught, a few (13%) were regularly captured (Figure 2; Table 2). Among species that are especially interesting due to price or size, those rarely caught were much more numerous (19) than those regularly caught (5; Table 2). Many species were neither targeted nor frequently captured (26; Table 2) and only three untargeted species were regularly caught. Thirty-two species, corresponding to 17.7% of the whole community (i.e., 181) or to 34% of the species captured (i.e., 97), were present in discards from fishing activities. Most of this (91%) were classified as not targeted species (Table 2).

Coastal islands and their reefs, due to the partial protection offered by distance from shore, show natural environments that are slightly more preserved than those found along the coastline (Floeter et al. 2006). These places attract recreational hook-and-line and spear fishers that come looking for fishes that have been extirpated elsewhere. However, our study shows that the reef fish community of the coastal island and reefs studied is already overexploited, with a low density of highly targeted species and a high predominance of non-target species both in natural environments and in the fishery caught. Fishermen related that many target species were more abundant in the past (Pinheiro HT & Martins AS, unpublished data). However, the majority of highly targeted species are rare nowadays, hardly being found and caught in the area.

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Captured species density Underwater visual censuses registered 90 fish species, among which 51 (57%) were caught by fishing activities done at the site (these are listed in bold in Table 2). Fished species that went unrecorded during UVC (Table 2) were rare species or species that show escape behavior near divers. The highest numeric density was shown by untargeted species (KW; p < 0.001) and occasionally caught species (KW; p < 0.001) while highly targeted and regularly caught

Many species registered in the fisheries catches were not recorded during SCUBA census, being rare in the study area. Many studies have shown that pristine and protected sites have higher relative abundance of target and predators species in UVCs than in overexploited sites (Friedlander & DeMartini 2002, Floeter et al. 2006, Francini-Filho & Moura 2008). Others, however, contest the efficacy of UVC methods to register commercial species in exploited areas due to the behavioral features adopted by target species (Kulbicki 1998). In this context, species that are highly targeted, regularly caught and not observed under water may be underestimated by visual census. On the other hand, species that are highly targeted, rarely caught and have low density, as estimated by UVC, can already be considered

Figure 2. Number of species in function of the intensity at which they are wished (left; as determined from interviews) and their frequency of capture (right; as censused in monitored fishing operations) at Franceses Island, central coast of Brazil.

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Table 2. Reef fish species caught by the various fishing activities monitored at Franceses Island, Brazil, including the intensity at which they are wished and the frequency at which they are captured. Families are ordered alphabetically within each cross-category of decreasing intensity and decreasing frequency (i.e., the same family may appear in several places in the table). Species also recorded in UVC are in bold. Abbreviations for intensity are H: highly targeted; A: average; N: non-targeted. Abbreviations for frequency are Re: regularly; Oc: occasionally; Ra: rarely. Fishing activities are listed in decreasing order of frequency of capture and the frequency at which the species appears in discards is indicated.

Family Clupeidae Carangidae

Species

Intensity Capture

Haemulidae

Opisthonema oglinum Caranx crysos Caranx latus Anisotremus surinamensis

H H H H

Scombridae

Scomberomorus brasiliensis

Pomatomidae

Pomatomus saltator

Carangidae

Fishing activity / discards

Re Re Re Re

seine seine, spear, hook, trolling, gill net seine, discard, spear, hook, trolling, gill net spear, hook

seine, spear, trolling, gill net

spear, seine, trolling

Carangoides bartholomaei Trachinotus goodei Lutjanus jocu Lutjanus synagris Ocyurus chrysurus Diplodus argenteus Chaetodipterus faber

H H H H H H H

Oc Oc Oc Oc Oc Oc Oc

seine, spear, hook, trolling spear, trolling, hook spear, hook seine, hook seine spear, seine, hook spear, seine

H H H H H H H H H H H H H H H H H H

Oc Ra Ra Ra Ra Ra Ra Ra Ra Ra Ra Ra Ra Ra Ra Ra Ra Ra

seine spear spear, hook, seine spear, hook, seine spear, gill net Spear spear, hook Spear spear, seine seine, spear, trolling, gill net spear spear, hook seine, hook spear spear gill net, seine spear seine, spear, trolling, gill net

Tetraodontidae Holocentridae Carangidae Haemulidae Kyphosidae Scaridae

Trichiurus lepturus Ginglymostoma cirratum Rhinobatos horkelii Rhinobatos percellens Centropomus undecimalis Epinephelus itajara Mycteroperca marginata Mycteroperca acutirostris Mycteroperca bonaci Caranx hippos Trachinotus carolinus Trachinotus marginatus Lutjanus analis Lutjanus cyanopterus Archosargus probatocephalus Micropogonias furnieri Scarus trispinosus Euthynnus alletteratus Scomberomorus cavalla Lagocephalus laevigatus Holocentrus adscensionis Oligoplites saliens Anisotremus virginicus Kyphosus sectator Sparisoma axillare

H H A A A A A

Ra Ra Re Re Re Re Re

seine, spear, trolling, gill net seine, hook hook, spear seine, spear, gill net spear, hook, seine, discard spear, hook, seine

Clupeidae

Sardinella brasiliensis

Lutjanidae

Sparidae Ephippididae Trichiuridae Ginglymostomatidae Rhinobatidae Centropomidae Serranidae

Carangidae

Lutjanidae Sparidae Sciaenidae Scaridae Scombridae

Haemulidae

spear, seine, hook seine

Anisotremus moricandi A Oc hook Haemulon plumierii A Oc spear, hook, seine Haemulon parra A Oc hook, spear, seine Orthopristis ruber A Oc hook, seine, discard The species registered in UVC not captured by the fisheries are: Acanthostracyon polygonius, A. quadricornis, Canthigaster figueiredoi, Cantherhines pullus, Chaetodon sedentarius, C. striatus, Coryphopterus spp., Chromis multilineata, Cryptotomus roseus, Ctenogobius saepepallens, Diodon hystrix, Diplectrum radiale, Doratonotus megalepis, Elacatinus figaro, Gramma brasiliensis, Halichoeres penrosei, Hippocampus reidi, Holacanthus ciliaris, H. tricolor, Labrisomus cricota, Malacoctenus delalandii, Malacoctenus aff. triangulatus, Myripristis jacobus, Odontoscion dentex, Pareques acuminatus, Parablennius marmoreus, P. pilicornis, Pempheris schomburgkii, Pseudocaranx dentex, Scorpaena brasiliensis, S. plumieri, Serranus baldwini, S. flaviventris, Sparisoma amplum, S. tuiupiranga, S. radians, Stegastes fuscus, S. variabilis, Synodus intermedius, S. synodus.

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Table 2. Continued ...

Family Mugilidae Sphyraenidae Rhinobatidae Dasyatidae Myliobatidae Carangidae Gerreidae Sparidae Scaridae Balistidae Clupeidae Carangidae Labrisomidae Elopidae Muraenidae Engraulidae Belonidae Dactylopteridae Haemulidae Mullidae Pomacentridae Labridae Diodontidae Narcinidae Muraenidae

Ophichthidae Engraulidae Hemiramphidae Aulostomidae Serranidae Echeneidae Carangidae

Gerreidae Pomacanthidae Labridae Labrisomidae

Species

Intensity Capture

Fishing activity / discards

Mugil incilis Mugil liza Sphyraena tome Zapteryx brevirostris Dasyatis guttata Aetobatus narinari Carangoides ruber Eugerres brasilianus Archosargus rhomboidalis Calamus penna Sparisoma frondosum Balistes capriscus Balistes vetula Harengula clupeola Chloroscombrus chrysurus Labrisomus nuchipinnis Elops saurus

A A A A A A A A A A A A A N N N N

Oc Oc Oc Ra Ra Ra Ra Ra Ra Ra Ra Ra Ra Re Re Re Oc

seine, spear seine, spear seine, hook spear, hook, seine spear, hook seine seine spear seine seine, spear, hook spear hook, hook, spear seine, discard seine, hook, discard hook, discard seine, spear, discard

Gymnothorax moringa Anchoviella lepidentostole Anchoa spinifer Tylosurus acus Dactylopterus volitans Haemulon aurolineatum Haemulon steindachneri Pseudopeneus maculatus Abudefduf saxatilis Halichoeres brasiliensis Halichoeres poeyi Chilomycterus spinosus Narcine brasiliensis Gymnothorax funebris Gymnothorax ocellatus Gymnothorax vicinus Ophichthus ophis Anchoa filifera Cetengraulis edentulus Hemiramphus brasiliensis Aulostomus strigosus Diplectrum formosum Rypticus saponaceus Echeneis naucrates Decapterus punctatus Selene browni Selene vomer Eucinostomus spp. Pomacanthus paru Bodianus rufus Labrisomus kalisherae

N N N N N N N N N N N N N N N N N N N N N N

Oc Oc Oc Oc Oc Oc Oc Oc Oc Oc Oc Oc Ra Ra Ra Ra Ra Ra Ra Ra Ra Ra

hook, spear, discard seine, discard seine, discard Seine seine, hook, discard seine, discard seine, hook, discard seine, discard Hook hook, spear, discard hook, discard seine, discard Spear hook, spear, discard hook, discard hook, discard Spear seine, discard seine, discard seine, discard Spear hook, discard

N N N N N N N N N

Ra Ra Ra Ra Ra Ra Ra Ra Ra

Spear seine, discard seine, discard Seine Seine seine, discard Spear spear, hook hook, discard

The species registered in UVC not captured by the fisheries are: Acanthostracyon polygonius, A. quadricornis, Canthigaster figueiredoi, Cantherhines pullus, Chaetodon sedentarius, C. striatus, Coryphopterus spp., Chromis multilineata, Cryptotomus roseus, Ctenogobius saepepallens, Diodon hystrix, Diplectrum radiale, Doratonotus megalepis, Elacatinus figaro, Gramma brasiliensis, Halichoeres penrosei, Hippocampus reidi, Holacanthus ciliaris, H. tricolor, Labrisomus cricota, Malacoctenus delalandii, Malacoctenus aff. triangulatus, Myripristis jacobus, Odontoscion dentex, Pareques acuminatus, Parablennius marmoreus, P. pilicornis, Pempheris schomburgkii, Pseudocaranx dentex, Scorpaena brasiliensis, S. plumieri, Serranus baldwini, S. flaviventris, Sparisoma amplum, S. tuiupiranga, S. radians, Stegastes fuscus, S. variabilis, Synodus intermedius, S. synodus.

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Table 2. Continued ...

Family Acanthuridae

Bothidae Monacanhidae Tetraodontidae

Species

Intensity Capture

Acanthurus chirurgus Acanthurus bahianus Acanthurus coeruleus Bothus ocellatus Stephanolepis hispidus Sphoeroides testudineus

N N N N N N

Ra Ra Ra Ra Ra Ra

Fishing activity / discards spear, seine, discard spear, seine, discard Spear seine, discard Seine hook, discard

Sphoeroides spengleri N Ra hook, discard The species registered in UVC not captured by the fisheries are: Acanthostracyon polygonius, A. quadricornis, Canthigaster figueiredoi, Cantherhines pullus, Chaetodon sedentarius, C. striatus, Coryphopterus spp., Chromis multilineata, Cryptotomus roseus, Ctenogobius saepepallens, Diodon hystrix, Diplectrum radiale, Doratonotus megalepis, Elacatinus figaro, Gramma brasiliensis, Halichoeres penrosei, Hippocampus reidi, Holacanthus ciliaris, H. tricolor, Labrisomus cricota, Malacoctenus delalandii, Malacoctenus aff. triangulatus, Myripristis jacobus, Odontoscion dentex, Pareques acuminatus, Parablennius marmoreus, P. pilicornis, Pempheris schomburgkii, Pseudocaranx dentex, Scorpaena brasiliensis, S. plumieri, Serranus baldwini, S. flaviventris, Sparisoma amplum, S. tuiupiranga, S. radians, Stegastes fuscus, S. variabilis, Synodus intermedius, S. synodus.

Figure 3. Mean fish density (Number 40 m–2 ± 1 standard error) in UVC in function of the intensity at which they are targeted (left) and their frequency of capture (right) at Franceses Island, central coast of Brazil.

Table 3. Mean density (Number 40 m–2 ± 1 standard error) in UVC by cross-category for intensity at which fish are wished and their frequency of capture at Franceses Island.

Intensity

Frequency of capture Regularly

Occasionally

Rarely

Highly 1.4 ± 0.3 0.8 ± 0.2 0 targeted Average 9.9 ± 0.9 7.9 ± 1.8 0 Not targeted 1.1 ± 0.1* 19.1 ± 1.9 18.1 ± 1.7 *Shows the density of L. nuchipinnis since the other 2 species (H. clupeola and C. chrysurus) aggregate in schools and were, therefore, excluded from the analysis.

threatened in the reefs we studied and nearby shores. The low density or absence of targeted species in the majority of coastal areas, and the high number (and quantity; not shown) of discard species are, most probably, consequences of inappropriate and excessive capture by artisanal and recreational activities. Elasmobranchii, serranids, lutjanids and scarids are particularly susceptible to overexploitation due to their remarkable life history and characteristics, such as slow growth, late maturity, high site fidelity, complex social

structure, sex reversal, group spawning at predictable sites, etc. (Coleman et al. 2000, Hawkins & Roberts 2003). Many works show that these groups are among the first to disappear under overexploitation (Roberts 1995, Oliveira et al. 1997, Coleman et al. 2000). The capture of juvenile specimens by fisheries and the apparent absence of reproductive adults of groupers, snappers and parrotfishes, as is locally observed at the studied site and other coastal areas (Pinheiro, HT personal observation), can reduce the reproductive potential of their population (Friedlander & DeMartini 2002), promoting the total local extirpation of the species. Albeit tropical Brazilian islands and reefs have low diversity compared to many other tropical regions of the world (Rangel et al. 2007, Souza et al. 2007), a high number of fish species was caught at Franceses Island. However, the rareness of species that are regularly captured and are highly targeted puts in question the sustainability of fishing activities. The predominance of species non-targeted in the catches and in situ suggest that, due to the constant increase of anthropic pressure onto explored resources, these species will be more and more heavily captured, thus collaborating to further alter the structure of the reef community.

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Perspectives of sustainability The coastal islands and reefs attract many non-professional fishermen. On the other hand, these areas are not able to economically sustain some professional fishing activities, mainly the most selective ones, due to the low abundance of remaining commercially-sized and -priced fish. Professional fishermen have been compelled to look for other fishing grounds more distant and deeper to sustain their catch rates (Martins et al. 2005), or to get extra alternatives to their income (Pinheiro et al. 2009). However, artisanal fishermen continue to fish in coastal zones, frequently using low-selectivity artisanal gears that are able to catch a great amount of non-target species (Erzini et al. 2002), thus contributing to alter the structure of local fish community. Those fishermen that do not sustain their catches are exploring tourism directed to the islands and reefs, earning direct and indirect additional income (Pinheiro et al. 2009). Much of this tourism is also directly related to recreational fishing because of the fact, both cause and consequence, that visitors transport to the islands and reefs is exclusively provided by local fishermen. Due to the importance of fishing and tourism for coastal communities, educational programs that foment the transformation of uncontrolled and predatory activities into sustainable practices are an alternative way of valorization of the environment and sustainability of the exploited resources. The creation of a protected area with a mosaic of usages would also collaborate to an adequate management of the area and reduce possible conflicts of social, economic and cultural orders (Pinheiro et al. 2009). However, the benefits for the fishing activities and tourism (see in Hall 2001, Friedlander et al. 2003) would be secondary close to the rehabilitation of insular and reef environments. The rocky shores and reefs, which cover small areas and are relatively shallow, would be least-difficult environments to protect and manage (compare to large and deep areas such as trawling grounds for example) and easiest to check for resulting effects of managements practices. Finally, rare species deserve a special attention about their individual size and level of capture. Brazilian laws need to incorporate maximum length to the catch of some reef fishes, once the bigger individuals are disproportionately more fecund than the smaller ones and therefore management strategies should be directed to protect the bigger and not only the smaller specimens against fishing (Palumbi 2004). However, regulation applied to fisheries, such as length limits, limited catches or area closure, has to be applied with caution because, if it offers positives prospects for species recuperation (Beets & Friedlander 1998, Coleman et al. 1999), it also often conflicts with resource users (Kirchner et al. 2001).

Acknowledgements We thank B.P. Ferreira and C.E.L. Ferreira for critically reading an earlier version of the manuscript, F. Frizzera, R. Molina, A. Ferreira, L. Schuler, J.M. Madureira, P. Assis, L.

Natureza & Conservação 8(2):151-159, December 2010

Baião, T. Simon, V. Brilhante for their help on the field, Sr Cazimiro, Carimbo, Josias and Vito for providing transport to the island, R. Sforza (TAMAR/ICMBio Project) and S. Pinheiro for their support in the initial phases of the project, J.L. Gasparini, S.R. Floeter, C.E.L. Ferreira, R. Nogushi and C.G.W. Ferreira for logistics and training in UVC techniques and fish identification, J.B. Teixeira for technical support, the Fundação O Boticário de Proteção à Natureza for project funding (0643-20042) and CNPq and CAPES for financial support to HTP (Pibic 2005-06 and PPGOAm), JCJ (grant 301390 ⁄ 2007-0) and ASM (grant 308867/2006-8). Fundamental partnership and logistical support to diving activities was provided by diving operator Flamar and by NGO Voz da Natureza.

References Araujo JN & Martins AS, 2009. Aspects of the population biology of Cephalopholis fulva from the central coast of Brazil. Journal of Applied Ichthyology, 25:328-334. Beets J & Friedlander A, 1998. Evaluation of a conservation strategy: a spawning aggregation closure for red hind, Epinephelus guttatus, in the U.S. Virgin Islands. Environmental Biology of Fishes, 55:91-98. Coleman FC et al., 1999. Management and conservation of temperate reef fishes in the grouper–snapper complex of the southeastern United States. American Fisheries Society Symposium, 23:233-242. Coleman FC et al., 2000. Long-lived reef fishes: the groupersnapper complex. Fisheries, 25:14-20. Dulvy NK, Freckleton RP & Polunin NVC, 2004. Coral reef cascades and the indirects effects of predator removal by exploitation. Ecology Letters, 7:410-416. Erzini K et al., 2002. A comparative study of the species composition of discards from five fisheries from the Algarve (southern Portugal). Fisheries Management and Ecology, 9:31-40. Floeter SR & Gasparini JL, 2000. The southwestern atlantic reef fish fauna: composition and zoogeographic patterns. Journal of Fish Biology, 56:1099-1114. Floeter SR et al., 2007. Reef fish community structure on coastal islands of the southeastern Brazil: the influence of exposure and benthic cover. Environmental Biology of Fishes, 78:147-160 Floeter SR, Halpern BS & Ferreira CEL, 2006. Effects of fishing and protection on Brazilian reef fishes. Biological Conservation, 128:391-402. Francini-Filho RB & Moura RL, 2008. Dynamics of fish assemblages on coral reefs subjected to different management regimes in the Abrolhos Bank, eastern Brazil. Aquatic Conservation, 18:1166-1179. Friedlander A et al., 2003. Designing effective marine protected areas in Seaflower Biosphere Reserve, Colombia, based on biological and sociological information. Conservation Biology, 17:1769-1784.

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Friedlander AM & DeMartini EE, 2002. Contrast in density, size and biomass of reef fishes between the northwestern and the main Hawaiian islands: the effects of fishing down apex predators. Marine Ecology Progress Series, 230:253-264. Hall CM, 2001. Trends in ocean and coastal tourism: the end of the last frontier? Ocean & Coastal Management, 44:601-618. Hawkins JP & Roberts CM, 2003. Effects of fishing on sex-changing Caribbean parrotfishes. Biological Conservation, 115:213-226. Jennings S & Blanchard JL, 2004. Fish abundance with no fishing: prediction based on macroecological theory. Journal of Animal Ecology, 73:632-642. Kirchner CH, Holtzhausen JA & Voges SF, 2001. Introducing size limits as a management tool for the recreational line fishery of silver kob, Argyrosomus inodorus (Griffiths and Heemstra), in Namibian waters. Fisheries Management and Ecology, 8:227-237. Klippel S et al., 2005. Avaliação dos estoques de lutjanídeos da costa central do Brasil: análise de coortes e modelo preditivo de Thompson e Bell para comprimentos. In: Costa PAS, Martins AS & Olavo G (Ed.). Pesca e potenciais de exploração de recursos vivos na região central da Zona Econômica Exclusiva brasileira. Rio de Janeiro: Museu Nacional. p. 83-98. Kulbicki M, 1998. How the acquired behaviour of commercial reef fishes may influence the results obtained from visual censuses. Journal of Experimental Marine Biology and Ecology, 222:11-30. Martins AS, Olavo G & Costa PAS, 2005. A pesca de linha de alto mar realizada por frotas sediadas no Espírito Santo, Brasil. In: Costa PAS, Martins AS and Olavo G (Ed.). Pesca e potenciais de exploração de recursos vivos na região central da Zona Econômica Exclusiva brasileira. Rio de Janeiro: Museu Nacional. p. 35-55.

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Myers RA & Worms B, 2003. Rapid worldwide depletion of predatory fish communities. Nature, 423:280-283. Oliveira GM, Evangelista JEV & Ferreira BP, 1997. Considerações sobre a biologia e pesca no arquipélago dos penedos de São Pedro e São Paulo. Boletim Técnico Científico do CEPENE, 5: 1-16. Palumbi SR, 2004. Why mothers matter. Nature, 430:621-22. Pinheiro HT et al., 2009. Profile of social actors as a tool for the definition of marine protected areas: the case of the Ilha dos Franceses, southern coast of Espírito Santo, Brazil. Natureza & Conservação, 7:181-194. Rangel CA, Chaves LCT & Monteiro-Neto C, 2007. Baseline assessment of the reef fish assemblages from Cagarras Archipelago, Rio de Janeiro, Southeastern Brazil. Brazilian Journal of Oceanography, 55:7-17. Roberts CM et al., 2002. Marine Biodiversity Hotspots and Conservation Priorities for Tropical Reefs. Science, 295:1280-1284. Roberts CM, 1995. Effects of fishing on the ecosystem structure of coral reefs. Conservation Biology, 9:988-995. Rosa RS & Lima FCT, 2008. Peixes. In: Machado ABM, Drummond GM and Paglia AP (Ed.). Livro vermelho da fauna brasileira ameaçada de extinção. Brasília, DF: Ministério do Meio Ambiente; Belo Horizonte: Fundação Biodiversitas. p. 8-285. Schmid C et al., 1995. The Vitória eddy and its relation to the Brazil current. Journal of Physical Oceanography, 25:2532-2546. Souza AT et al., 2007. Fishes (Elasmobranchii and Actinopterygii) of Picãozinho reef, Northeastern Brazil, with notes on their conservation status. Zootaxa, 1608:11-19. Zar JH, 1999. Biostatistical analysis. 4ªed, New Jersey: PrenticeHall Inc.

Received: August 2010 First Decision: September 2010 Accepted: October 2010

Research Letters

Brazilian Journal of Nature Conservation

Successional and Seasonal Changes in a Community of Dung Beetles (Coleoptera: Scarabaeinae) in a Brazilian Tropical Dry Forest Frederico de Siqueira Neves1,2,*, Victor Hugo Fonseca Oliveira1, Mário Marcos do Espírito-Santo1, Fernando Zagury Vaz-de-Mello3, Júlio Louzada4, Arturo Sanchez-Azofeifa5 & Geraldo Wilson Fernandes2 1

Biologia da Conservação, Departamento de Biologia Geral, Universidade Estadual de Montes Claros, Montes Claros, MG, Brasil

Ecologia Evolutiva e Biodiversidade, Departamento de Biologia Geral, Universidade Federal de Minas Gerais, Belo Horizonte, MG, Brasil

Departamento de Biologia e Zoologia, Universidade Federal de Mato Grosso, Cuiabá, MT, Brasil

Ecologia e Conservação, Departamento de Biologia, Universidade Federal de Lavras, Lavras, MG, Brasil

Erth and Atmospheric Sciences Department, University of Alberta, Edmonton, Alberta, Canada

Abstract We tested the following hypotheses on the dynamics of a dung beetle community in a Brazilian Seasonally Dry Tropical Forest: (1) successional changes of dung beetle community, with species composition, richness and overall abundance increasing with the successional stages; (2) dung beetle community changes between dry and wet seasons, with species composition, richness and abundance decreasing in the dry season. Dung beetles were sampled in 15 plots from three different successional stages in both wet and dry seasons. We sampled a total of 2,752 individuals, representing 38 beetle species and 14 genera. The composition, richness, and abundance of dung beetles changed along the successional gradient and was strongly related to seasonal variation. The highest diversity of dung beetles was found in the intermediate aged forest fragments. These findings highlight the importance of secondary forests to biodiversity conservation and restoration programs in seasonally dry tropical ecosystems. Key words: Biodiversity, Habitat Complexity, Resource Availability, Scarabaeidae, Secondary Succession.

Introduction The loss of tropical habitats due to anthropogenic activities is the major cause of decline in species diversity during recent decades (Balmford et al. 2005; Reid et al. 2005). Seasonally Dry Tropical Forests (SDTF) encompasses approximately 42% of all tropical habitats (Murphy & Lugo 1986) and harbors a unique range of biodiversity (Janzen 1988). SDTFs are among the least protected (Janzen 1988) and most threatened ecosystems in the world (Miles et al. 2006), yet they have generally been neglected by conservation efforts and fall behind other global initiatives aimed to protect tropical rainforests (Sanchez-Azofeifa et al. 2005). For instance, less than 1% of the original 500,000 km2 of *Send correspondence to: Frederico de Siqueira Neves Biologia da Conservação, Departamento de Biologia Geral, Universidade Estadual de Montes Claros, Campus Darcy Ribeiro s/n, CEP 39401-089, Montes Claros, MG, Brasil E-mail: frederico.neves@unimontes.br

STDFs is under protection in Central America (Janzen 1988). In Brazil, only 3.9% of the remaining SDTFs are protected (Sevilha et al. 2004). Human occupation of SDTF areas is usually followed by abandonment and the subsequent natural recover in secondary forest formations (Quesada et al. 2009). As such, SDTF landscapes are often a mosaic of remnants in a range of successional stages depending on the length of recovery time and the degree of the disturbance (Madeira et al. 2009; Quesada et al. 2009). These, in turn, affect the community structure and composition of several groups of plants and animals (Siemann et al. 1999; Neves et al. 2010). It is not yet clear, however, how arthropod communities change across successional gradients in tropical dry systems (see Neves et al. 2010). In addition to forest successional status, seasonal variation in biotic and abiotic factors strongly influence SDTF structure

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and function (Murphy & Lugo 1986). SDTFs are usually characterized by a marked dry season (3-6 months) and concentrated rainfall during the wet season (Murphy & Lugo 1986). The seasonality of these forests affects the phenological patterns of plants and, consequently, the faunal abundance (Wolda 1978). Since insect abundance has often been correlated with rainfall (Wolda 1978), pronounced differences are expected for these organisms between dry and wet seasons in SDTFs (but see Neves et al. 2010). Many groups of arthropods have been used as indicators of habitat quality due to their great sensitivity to environmental changes (Gardner et al. 2008). Particularly, Scarabaeinae dung beetles have been used to evaluate and monitor the relationship between habitat structure and degree of disturbance and community composition (Nichols et al. 2007; Gardner et al. 2008; Nichols et al. 2008). Dung beetles include a very diverse group of detritus-feeding insects in tropical rainforests (Hanski & Cambefort 1991). They are considered to be one of the most important insect components of terrestrial ecosystems, since they contribute to several ecological functions, such as secondary seed dispersal, nutrient cycling, bioturbation and biological control of pests and parasites (Nichols et al. 2008). Casual observations indicate that dung beetles are very common in SDTFs, but there is a lack of information about the responses of this insect group to forest disturbance. A better understanding of the spatial and temporal distributions of the dung beetles across gradients of natural regeneration, and of the effects of habitat structure on this group’s diversity may inform the management and conservation of threatened SDTFs. Therefore, the objectives of this study were to test two hypotheses on a dung beetle community in a Brazilian SDTF. The first hypothesis predicts changes in species composition, richness and overall abundance increasing along the successional gradient. The second predicts seasonal changes in the dung beetle community, with species composition, richness and overall abundance decreasing in the dry season.

precipitation of 818 ± 242 mm. (Madeira et al. 2009). The main land uses in the area before the establishment of the park were cattle ranching and cultivation of beans, tomato and corn in irrigated areas. Approximately 1,525 ha of the PEMS is covered with abandoned pasture where SDTF can be found in different stages of recovery, while the remaining area supports secondary and primary SDTF (Madeira et al. 2009). Samples of dung beetles were conducted inside 15 randomly delimited plots of 50 × 20 m inside forest fragments classified as early, intermediate or late successional stages (five plots per stage). These plots were established in January 2006, and all plants with a diameter at breast height ≥5 cm were tagged and identified (see Madeira et al. 2009 for details). All plots were located along a 7 km transect inside the original area of a single farm, where management practices were similar for all pasturelands over the last 30 years. Plots from the same regeneration stage were located at least 0.2 km apart. The early successional plots were characterized by a forested area composed of sparse patches of woody vegetation, shrubs, herbs and grasses with a single stratum of tree crowns presenting a very open canopy approximately 4 m high. This area had been used as pasture for at least 20 years and abandoned in 2000. The intermediate successional plots were characterized by two vegetation layers: one composed by deciduous trees (10-12 m) and some emergent trees (up to 15 m), and a second layer of dense understory with many young trees and abundant lianas. This area was used as pasture for at least five years and abandoned in 1987. The late successional plots were also characterized by two strata: the first composed of taller deciduous trees forming a closed canopy 18-20 m high, and the second of a sparse understory with reduced light levels and low density of young trees and lianas. There are no records of clear-cutting in this area for the last 50 years. Site history was determined through interviews with the park manager and former farm employees. For a detailed description of plot structure and composition see Madeira et al. (2009).

Dung beetle sampling

Material and Methods Study area This study was carried out in the Parque Estadual da Mata Seca (PEMS), a conservation unit created in the area of four farms in 2000. The PEMS is located on 10,281.44 ha in the São Francisco river valley, Minas Gerais, Brazil (between 14° 48’ 36” – 14° 56’ 59” S and 43° 55’ 12” – 44° 04’ 12” W). The original vegetation is SDTF, growing on plain and nutrient–rich soils (Instituto Estadual de Florestas 2000) and dominated by deciduous trees, with 90-95% leaf loss during the May- September dry season (Madeira et al. 2009). The climate of the region is tropical semi-arid (Köppen’s classification), characterized by a severe dry season during the winter months. The average temperature of the study region is 24 °C (Antunes 1994), with an average annual

Dung beetles were sampled during two periods in 2007: February (wet season) and September (dry season). We placed four pairs (two alternative baits) of pitfall traps in each plot, totaling 120 traps. Traps consisted of a plastic container 14 cm in diameter and 9 cm deep buried with the opening level with the soil. In the inner compartment approximately 50 g of human feces or carrion (rotten chicken liver) was used as bait, and covered with a lid to protect from rain. We filled the surroundings of the inner pitfall compartment with 250 mL of a liquid detergent solution as a killing and preservative agent. After a period of 48 hours, all insects were collected, sorted and identified to the lowest taxonomic level possible. Voucher specimens were deposited at the Entomological Collection of the Laboratório de Biologia da Conservação (Unimontes) and in the FZVM collection at Universidade Federal do Mato Grosso (UFMT).

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Data analyses We used non-metric multidimensional scaling to search for overall differences in species abundance and composition between different successional stages. Ordination was undertaken for abundance and composition using the Bray-Curtis index. We used analysis of similarities (ANOSIM, Clarke 1993) to test the differences in species abundance and composition between successional stages. This is a non-parametric permutation procedure applied to rank similarity matrices underlying sample ordinations (Clarke 1993). The relative differences between R-values from the ANOSIM test were used to determine the patterns of similarity between dung beetle communities in the three successional stages. We used similarity percentage (SIMPER, Clarke 1993) to determine the individual species contribution in each successional stage. The analyses were performed using the software PAST (Hammer et al. 2001).

Natureza & Conservação 8(2):160-164, December 2010

presented a more distinct dung beetle community compared to the advanced stages of regeneration (Figure 1, Table 1). Only four species (C. histrio, Uroxys sp.2, Deltochilum verruciferum and Canthidium manni) contributed more than 69% to the observed differences between stages. The dominant Uroxys sp.1 and C. histrio did not occur in the early successional stage, where the most common species was Canthidium sp.1.The species richness and total abundance of dung beetles were significantly higher in the intermediate stage, followed by the late stage (Figure 2, Table 2).

We used generalized linear models (GLMs) to identify the effects of successional stages (explanatory variable) on dung beetle richness and abundance (response variables). All GLMs were submitted to residual analyses to evaluate adequacy of the error distribution (Crawley 2002). Significant results for factor levels (stages) were compared using contrast analysis by aggregating levels and comparing deviance change (Crawley 2002). If the level of aggregation did not significantly alter the deviance explained by the model, the levels were pooled together (amalgamated), simplifying the model. Minimum adequate models were generated by stepwise omission of non-significant terms. The GLMs were performed with the software R (R Development Core Team 2008).

Results We sampled 2,752 dung beetle individuals, belonging to 38 species from 14 genera. Among the identified species, 56% have a wide geographical distribution, 20% occur in the Caatinga, 12% in the Cerrado, and another 12% are common to both Biomes (Table S1). For all samples combined, the species accumulation curve was close to reaching an asymptote (Figure S1a), but this varied between succecional stages (Figure S1b). Strong changes were observed in the species abundance and composition of dung beetles across the successional gradient. We collected 41 individuals from 13 species in the early successional stage, with one species exclusive to this stage. In the intermediate stage, we found 32 species (2,010 individuals) from which ten were exclusive to this stage. In the late sucessional stage, 27 species were sampled (701 individuals), with five exclusive species. Among the collected species, 32% were found in all stages, whereas 58% were found only in intermediate and late stages (Table S1). The MDS ordination showed clear differences in dung beetle species abundance (Figure 1a) and composition (Figure 1b) between successional stages (Table 1). However, these analyses showed that the early successional stage

Figure 1. Non-metric multidimensional scaling (MDS) ordination of the dung beetle community in three successional stages as sampled by pitfalls traps. a) species abundance and b) species composition. Table 1. Results of Non-parametric analyses of similarity (ANOSIM) testing for differences in the rank similarities for 15 sites in three successional stages (early, intermediate and late) grouped by dung beetles species abundance and composition. ANOSIMs were calculated based on Bray–Curtis similarity.

Parameter

Stage

Species abundance

early × intermediate early × late intermediate × late early × intermediate early × late intermediate × late

1 0.984 0.7 0.722 0.73 0.294

0.008 0.0085 0.0074 0.0081 0.0074 0.0156

Composition

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Figure 2. Dung beetle richness and abundance (Mean ± SE) per plot in the wet season in three successional stages. Distinct letters in the bars represent statistical differences between successional stages detected by contrast analysis (p < 0.05). Table 2. Analyses of deviance of the minimal adequate models showing the effects of successional stage on dung beetle richness and abundance.

Response variable Richness Abundance

Error d. f. Deviance distribution Poisson Quasipoisson

2 2

16.82 1476.49

p <0.001 <0.001

Dung beetle communities also showed a strong seasonal variation, with remarkable differences between the dry and wet seasons. During the wet season 2,748 individuals belonging to 38 species of dung beetles were collected, whereas only four Uroxys sp.1 individuals were sampled in the dry season.

Discussion Dung beetle community changed along a successional gradient in the studied SDTF, with variation in species abundance and composition and increases in species richness and overall abundance from early to advanced stages. The differences between successional stages observed here are consistent with studies conducted on other organisms in the same plots (see Madeira et al. 2009; Quesada et al. 2009; Neves et al. 2010). After seven years of regeneration, only 31.6% of the species present in the early stage was shared with late forests. Additionally, of the five most abundant species in intermediate and late stages, three were not recorded in early stages: C. histrio, Uroxys sp.2 and C. manni. These species are therefore potential indicators of disturbance levels in Brazilian SDTFs, due to their marked responses to habitat conditions, though further studies are needed to validate this suggestion. C. histrio is commonly found in seasonal wooded areas of the Cerrado, Caatinga and Chaco biomes (and is also encountered in southern Amazonia), whereas C. manni is restricted to the Caatinga and northern Brazilian coastal vegetation complexes, apparently always

associated with less open habitats (Fernando Vaz-de-Mello, pers. obs.). Though intermediate and late sucessional stages sustained more similar beetle communities, substantial differences were still observed. Of Of the 32 species recorded in intermediate plots, 22 were also found in late stage plots. All the exclusive species from both intermediate (10) and late (5) stages can be considered rare, and there is no clear explanation for this habitat segregation. It is possible that some beetle species were excluded in late succession plots due to a decreased habitat complexity caused by liana impoverishment (see Madeira et al. 2009; Sanchez-Azofeifa et al. 2009). Thus, it is possible that more structurally complex, liana-rich intermediate forest fragments attract and maintain potential seed dispersers, which in turn provide feeding and reproductive resources for dung beetles, increasing their diversity in these habitats. This explanation is in accordance with the intermediate disturbance hypothesis (Connel 1978), which predicts a higher organism diversity under moderate levels of disturbance. Similar results have been reported in a study of dung beetles by Nichols et al. (2007), who did not find differences in species richness between secondary and primary forests. Dung beetle community changed between dry and wet seasons, with a dramatic decrease in species richness and abundance in the dry season. SDTFs probably have the most pronounced intra-annual differences in dung beetle communities among all tropical forested ecosystems, since these organisms all but cease activity during the dry season. The dry season in the PEMS region is very intense, with up to seven months without rain (typically from mid-April to mid-November). In this period, the superficial soil layer becomes desiccated and compacted, and plants lose up to 95% of their leaves (Pezzini et al. 2008). Though direct observations of dung-producing vertebrates are not currently available for the study site, it is very likely that they respond

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negatively to the lack of water and to the absence of insects as a food resource, and probably migrate temporarily to the riparian forests nearby.

IEF - Instituto Estadual de Florestas, 2000. Parecer técnico para a criação do Parque Estadual da Mata Seca. Relatório Técnico, Belo horizonte, Brasil

Habitat structural complexity and resource availability are important factors in determining the dung beetle community in this SDTF, as indicated by their high susceptibility to both spatial (across successional stages) and temporal (across seasons) environmental changes. These characteristics, combined with the relatively inexpensive sampling techniques necessary for dung beetle survey, strength the role of these insects as good ecological indicators in rapid diversity assessment programs. Usually, secondary forests are often ignored and their importance for conservation purposes is neglected or underestimated. The highest beetle diversity in intermediate forests also highlights the importance of secondary forests maintenance and restoration programs to tropical conservation strategies.

Janzen DH, 1988. Management of habitat fragments in a tropical dry forest: growth. Annals of the Missouri Botanical Garden, 75:105-116.

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Acknowledgements We thank T.A. Gardner, for their useful comments in previous drafts of this manuscript and M. Hesketh for english review. We are very grateful to the staff of the Instituto Estadual de Florestas (IEF) for logistical support at the PEMS, in special to J. L. Vieira (IEF) for his invaluable field assistance. This work was carried out with the aid of a grant from the Inter-American Institute for Global Change Research (IAI) CRN II # 021, which is supported by the U.S. National Science Foundation (Grant GEO 0452325), and from the Fundação de Amparo à Pesquisa de Minas Gerais (FAPEMIG CRA 2288/07). GWF also acknowledges a grant provided by CNPq (309633/2007-9) and FAPEMIG (CRA 697/06, CRA 122/07, APQ-01278-08, 465/07), FZVM is partly funded by FAPEMAT (447441/2009), whereas FSN, VHFO and MMES greatly acknowledge a scholarship from FAPEMIG.

References Antunes FZ, 1994. Caracterização Climática - Caatinga do Estado de Minas Gerais. Informe Agropecuário, 17:15-19. Balmford A et al., 2005. The convention on biological diversity’s 2010 target. Science, 307:212-213 Clarke KR, 1993. Non-parametric multivariate analysis of changes in community structure. Australian Journal of Ecology, 18:117-143. Connel JH, 1978. Diversity in tropical rain forests and coral reefs. Science, 199:1302-1310. Crawley MJ, 2002. Statistical computing: an introduction to data analysis using S-plus. Chinchester: John Wiley and Sons. Gardner TA et al., 2008. The cost-effectiveness of biodiversity surveys in tropical forests. Ecology Letters, 11:139-150. Hammer O, Harper DAT & Ryan PD, 2001. PAST: Palaeonthological Statistics Software Package for education and data analysis. Palaeontologia Electronica, 4:1-9. Hanski I & Cambefort Y, 1991. Dung Beetle Ecology. Princeton: University Press.

Madeira BG et al., 2009. Changes in tree and liana communities along a successional gradient in a tropical dry forest in south-eastern Brazil. Plant Ecology, 291:291-304. Miles L. et al., 2006. A global overview of the conservation status of tropical dry forests. Journal of Biogeography 33:491-505. Murphy PG & Lugo AE, 1986. Ecology of tropical dry forest. Annual Review of Ecology and Systematics, 17:67-88. Neves et al. 2010. Diversity of arboreal ants an a brazilian Tropical Dry Forest: Efects of seasonality and successional Stage. Sociobiology, 56:177-194. Nichols E et al., 2007. Global dung beetle response to tropical forest modification and fragmentation: a quantitative literature review and meta-analysis. Biological Conservation, 137:1-19. Nichols E et al., 2008. Ecological functions and ecosystem services provided by Scarabaeinae dung beetles. Biological Conservation, 141:1461-1474. Pezzini FF et al., 2008. Polinização, dispersão de sementes e fenologia de espécies arbóreas no Parque Estadual da Mata Seca. MG. Biota, 1:37-45. Quesada M et al., 2009. Succession and Management of Tropical Dry Forests in the Americas: Review and new perspectives. Forest Ecology and Management, 258:1014-1024. R Development Core Team, 2008. R: a language and environment for statistical computing. Vienna, Austria: R Foundation for Statistical Computing. Reid WV et al., 2005. Millenium ecosystem assessment synthesis report. Island Press, Washington DC, USA. Sanchez-Azofeifa GA et al., 2005. Need for integrated research for a sustainable future in tropical dry forests. Conservation Biology, 19:1-2. Sanchez-Azofeifa GA et al., 2009. Tropical forest succession and the contribution of lianas to Wood Area Index (WAI). Forest Ecology and Management, 258:941-948. Sevilha AC, Scariot AO & Noronha S, 2004. Estado atual da representatividade de unidades de conservação em Florestas Estacionais Deciduais no Brasil. In: Sociedade Brasileira de Botânica. Biomas ﬂorestais. Viçosa: Editora da Universidade Federal de Viçosa. p. 1-63. Siemann E, Haarstad J & Tilman D, 1999. Dynamics of plant and arthropod diversity during old fiel succession. Ecography, 22:406-414. Wolda H, 1978. Fluctuations in abundance of tropical insects. The American Naturalist, 112:1017-1045.

Received: August 2010 First Decision: September 2010 Accepted: October 2010

Research Letters

Brazilian Journal of Nature Conservation

How Can We Estimate Buffer Zones of Protected Areas? A Proposal Using Biological Data Brenda Alexandre1, Renato Crouzeilles1,2 & Carlos Eduardo Viveiros Grelle1,* 1

Laboratório de Vertebrados, Departamento de Ecologia, Universidade Federal do Rio de Janeiro – UFRJ

Programa de Pós-graduação em Ecologia, Instituto de Biologia, Universidade Federal do Rio de Janeiro – UFRJ

Abstract A strategy to avoid the loss of habitats and preserve large areas is the establishment of protected areas. Brazil’s Conservation Units National System (SNUC) determines that protected areas should be surrounded by buffer zones where human activity is restrict, but the established size of the buffer seems arbitrary. The restrictions provided by SNUC could be based on limits that allow the persistence of species’ ecological function. Here we use the “landscape species” concept as a tool for buffer zone design, using the marsupial Micoureus paraguayanus as a model organism. We used its minimum area requirement for population viability (5,000 ha) to define the size of buffer zones around protected areas with smaller size that the minimum required area. The amount of habitat within protected areas was negatively correlated with the buffer size. Therefore, the largest the protected area, the smaller should be its buffer zone, provided that it does not have an impermeable barrier. Buffer zones are generally in private properties and, therefore, governmental incentives are essential to stimulate land uses compatible with biological flux through a permeable matrix. The method proposed here provides a simple analysis that can be used to establish the limits of buffer zones. Key words: Buffer Zones, Atlantic Forest, Landscape Species, Micoureus paraguayanus, Reserve Design.

Introduction Protected areas are elements of complex landscapes that should be managed, allowing the persistence of viable populations and the use for some economic activities (Crooks & Sanjayan 2006). For protected areas to maintain viable populations at the long-term, a minimum amount of habitat area is needed in a single area, or a network with other habitat patches where potential connectivity is possible. Connectivity can be defined as the degree to which the landscape facilitates or impedes movement among resource patches. It can be divided into “structural connectivity”, which refers to the spatial arrangement of the elements of the landscape, and “functional connectivity”, which refers to the behavior response of organisms to landscape structures (Crooks & Sanjayan 2006).

(SNUC, in the Portuguese acronym), where buffer zones are transition areas that should minimize negative impacts on protected areas. The size of buffer zones is not specified in SNUC, however. In 1990, the National Environment Council Resolution (CONAMA, Portuguese acronym) Nº 13 had already defined a 10km buffer zone around protected areas, where any activity that may affect the biota should be licensed (Gonçalves et al. 2009). The size proposed by CONAMA, however, is arbitrary and may be inadequate to maintain minimum viable population of some species. Studies performed in wetlands already showed the inefficiency of arbitrary buffer zones to population viability of frogs and salamanders (Harper et al. 2008).

On the other hand, to reconcile conservation and land-use one of the alternatives is to establish buffer zones around protected areas, within which human activities are subjected to specific rules and restrictions (Gonçalves et al. 2009). In Brazil, the protected areas system was established in 2000 by law Nº 9.985 of the Conservation Units National System

Conservation strategies need to be based on data, preferably data that allow for the interaction between biological and socio-economic perspectives. The “landscape species” approach has been used to characterize impacts on the structure and functioning of natural systems (Sanderson et al. 2002). It arises as a tool to be used in strategies for conservation and public policies. In this approach, the species are chosen based on the heterogeneity of its habitats, land uses and vulnerability to anthropogenic pressure, ecological functionally and socio-economic significance (Sanderson et al. 2002).

*Send correspondence to: Carlos Eduardo Viveiros Grelle Laboratório de Vertebrados, Departamento de Ecologia, Universidade Federal do Rio de Janeiro – UFRJ, CP 68020, CEP 21941-590, Rio de Janeiro, RJ, Brasil E-mail: grellece@biologia.ufrj.br

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Putting together the landscape species approach and the minimum area needed for species’ long-term persistence, it is possible to draw a proposal to delimit buffer zones. Thus, it is expected that the larger the protected area, the smaller should be its buffer zone to achieve a threshold of habitat amount to long-term persistence of populations. However, in highly fragmented landscapes, besides of the amount of forest cover within the protected area, the buffer zone may be enlarged or reduced depending on the amount and arrangement of the natural remnants surrounding protected areas.

mma.gov.br/mapas/aplic/probio accessed on November 10th, 2008) and the shapes of protected areas (27 strictly areas and 21 sustainable-use) were downloaded from Brazilian Institute of Environment and Renewable Resource (IBAMA in portuguese acronym - www.ibama.gov.br/patrimonio/ accessed on March, 5th, 2008). All geographical information data used were in Albers projection.

In this study, the marsupial Micoureus paraguayanus was used as a landscape species to delimit buffer zones. This small species is endemic to the Brazilian Atlantic forest and needs a minimum area of 5,000 ha to maintain a viable population over the long term (Brito & Grelle 2004). We used the minimum forested area required by the species to define the size of buffer zone around protected areas >5,000 ha. We also evaluated whether the increase of forest cover inside protected areas translates into a decrease in the size of buffer zones.

Micoureus paraguayanus (previously Micoureus travasossi) is a small didelphid whose adult can weigh up to 130 g, males being slightly larger than females (Rossi et al. 2006). It can be found in dense forest and also in secondary vegetation, it is nocturnal and mostly arboreal, although it can descends to the ground (Grelle 2003). The species is widely distributed in the state of Rio de Janeiro, occurring in all vegetation types but campos de altitude (highland fields occurring above 2000 a.s.l) (C.E.V. Grelle unpublished data). The species may cross up to 100 m in matrices composed of grass (Forero-Medina & Vieira 2009), and up to 300 m in matrices composed of grasses, shrubs and pioneer trees (Pires et al. 2002). According to Brito & Grelle (2004), M. paraguayanus needs a minimum of 5,000 ha forested area to support a viable population of 2,000 individuals that will retain demographic and genetic integrity for 100 years. We used M. paraguayanus as landscape species because it has not only the appropriate behavioral ecology, but also sufficient and appropriate data. The species occurs in heterogeneous habitat, being found in different successional stages of vegetation (Forero-Medina & Vieira 2009; Grelle 2003, Pires et al. 2002). It has an important ecological function as seeds disperser (Cáceres et al. 2002), is vulnerable to anthropogenic threats (Umetsu et al. 2008), but even so it may cross different human modified landscapes such as coffee plantation, and has socio-economic significance (Passamani & Ribeiro 2009).

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Material and Methods Region of study Rio de Janeiro state is located in southeastern Brazil. Its original vegetation is Atlantic forest, which has been radically transformed since the 1500’s (Fundação SOS Mata Atlântica & INPE 2010). Rio de Janeiro has more than 16 million people (http://www.ibge.gov.br/estadosat/ perfil.php?sigla=rj accessed on 12 July 2010), 43.6 million hectares of total area, of which about 800 thousand hectares of forest (or 18% of the state’ area, Fundação SOS Mata Atlântica & INPE 2010). Approximately 14% of the forest remnants are under protected areas (Jenkins et al. 2010), which can be divided into two classes: Strictly Protected Areas and Sustainable-Use Areas. The former corresponds to categories I-IV of the IUCN, which is established to maintain biodiversity and increase the protection, whereas the second corresponds to categories V-VI of the IUCN, which has the main focus of conciliate conservation and economical activities. Strictly protected areas have 77% of forest cover (Fidalgo et al. 2009) and protect about 6% of the remaining forest (Jenkins et al. 2010), but sustainable-use areas have only 43% of forest cover (Fidalgo et al. 2009). According the SNUC, protected areas should be surrounded by buffer zones, but this rule has two exceptions in sustainable-use categories: Environmental Protection Area (APA) and Private Natural Heritage Reserve (RPPN). The forest cover map used is from the cartographic base of Fundação SOS Mata Atlântica & INPE (2010), resulting from TM/Landsat bands 5 or 7, obtained at a scale of 1:50,000 in vector format, mapping forest remnants with at least 10 ha. The land-use map used is from Brazilian Environment Ministerial (MMA in portuguese acronym - http://mapas.

Micoureus paraguayanus as a “Landscape species”

Delimiting buffer zones First, we overlie the forest cover and protected areas maps, determining the actual amount of forested habitat for each protected area. Then, we evaluated which protected areas needed buffer zones according SNUC, and which were >5,000 ha (i.e., the estimated minimum area needed for M. paraguayanus). The protected areas that were not large enough, but were structurally connected to other protected areas or forest remnants, were not included in the analysis, even if together with its connecting elements it exceeded 5,000 ha. For the other protected areas, buffers were created according to protected areas boundaries, drawing concentric areas. Buffer size is then expressed as the radius of these circles and expressed in km. The surrounding forest remnants were incorporated until the total achieved the 5,000 ha landmark.

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In each buffer zone, the land use/cover was classified into five categories, according to M. paraguayanus preferences: forest remnants (habitat patches), simple matrix (grass; low permeability), complex matrix (grasses, shrubs and pioneer trees; high permeability), large water bodies and urban areas (impermeable). All analyses were performed using ArcView 9.3 (ESRI 2008). Finally, we tested how the increase of forest cover in protected areas translated into a decrease in buffer zone size, because of the variation in forest cover within protected areas. Spatial autocorrelation in these variables (forest cover and buffer zone size) were tested using Moran’s I correlograms with five distance classes. To take into account the effects of autocorrelation when correlating buffer size and forest cover, we calculated geographically effective degrees of freedom (v*) based on the correlograms (Legendre P. & Legendre L. 1998). All spatial analyses were obtained in SAM (Spatial Analysis in Macroecology), version 4.0, freely available at www.ecoevol.ufg.br/sam (Rangel et al. 2010).

Results Altogether, 31 out of the 48 protected areas in Rio de Janeiro state required buffer zones as proposed in the SNUC. Of these, 16 were smaller than the minimum requirement for M. paraguayanus. For these, the average buffer zones size was 9.3 km (Table 1, Figure 1). The protected area with the largest forest cover was “Ilha Grande State Park” (4,302 ha), and that needed the smallest buffer zone (0.45 km), whereas the “Guaxandiba” Ecological Station (1,241 ha) required the largest buffer zone (27.15 km) (Table 1). Forest Cover and Buffer Zone Size showed a weak and non-significant autocorrelation pattern (Moran’s I = 0.161,

P = 0.358 and Moran’s I = 0.186, P = 0.301, respectively, for the first distance class, around 30 km). Despite the Guaxindiba protected area tends to be an outlier in the relationship, there is a significant correlation between Forest Cover and Buffer Zone Size, even taking into account the small effects of spatial autocorrelation (r = -0.52, v* = 13, P = 0.042; Figure 2). A log-transformation of the buffer size increases the linearity of the relationship, but the relationship in the original data was shown to reveal the larger variance and uncertainty in the buffer size surrounding small protected areas. Five protected areas required buffer zones larger than 10 km and 10 protected areas had urbanization and/or water bodies in its interior, which are impermeable to the movement of M. paraguayanus.

Discussion Protected areas must be effective to conserve nature, and need to be planned in an association of stakeholders and scientists (Albernaz & Souza 2007; Grelle et al. 2010). In reality, most legal actions and guidelines have been defined in an arbitrary way, without considering ecological processes (e.g. Hunter et al. 2009). Ecological processes can be quantified in many ways (e.g. minimum forest cover needed for species persistence in forested biomes), which may allow objective definitions of conservation goals (Ficetola & Denoël 2009). In Brazil, environmental guidelines are determined through legal instruments, as SNUC and CONAMA. However, in some cases, the specifications are insufficient and poorly defined. SNUC determines that the buffer zones should be established by each protected area’s management plan,

Table 1. Forest cover, buffer zone size, and landscape elements within buffer zones for the 16 protected areas in Rio de Janeiro State that were smaller than the minimum requirement for M. paraguayanus. Landscape elements: (a) forest remnants, (b) complex matrix, (c) simple matrix, (d) water bodies, (e) urban area. S.P – State Park, E.R – Ecological Reserve, B.R. – Biological Reserve, S.R. – State Reserve, N.F. – Natural Forest, N.P – National Park, E.S. – Ecological Station.

Protected area and category Chacrinha S.P Jacarepia E.R. Guaratiba B.R Alcobaca S.R. Grajaú S.P. Massambaba E.R. Mário Xavier N.F. Serra da Concordia S.P. Serra da Tiririca S.P. Guaxandiba E.S. Araras B.R. Marinho do Aventureiro S.P. União B.R. Tijuca N.P. Praia do Sul E.R. Ilha Grande S.P.

Forest cover (ha) 12.16 74.88 87.44 98.99 100.65 287.94 499.00 796.65 1,072.69 1,241.30 1,412.33 1,785.85 2,053.53 2,615.37 3,438.82 4,302.10

Buffer size (km) 13.2 16.5 5.45 6.5 12.75 21.4 9.2 6.4 8.0 27.15 3.1 3.8 9.6 4.6 0.9 0.45

Landscape elements a.c.d.e a.b.c.d a.b.c.e a.e a.b.c.d.e a.b.c.d.e a.c.e a.c a.b.c.e a.b.c.d a a a.b.c.d.e a.c a a

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Figure 1. Buffer Zones with forest remnants surrounding protected areas in Rio de Janeiro State

Figure 2. Relationship between forest cover and buffer zone size in protected areas in Rio de Janeiro State that were smaller than the minimum requirement for M. paraguayanus.

without any specification regarding buffer zone’s size (Vitalli et al. 2009), and CONAMA arbitrarily require protected areas to have a 10 km buffer zones surrounding it. The science of determining buffer zones and surrounding areas is very immature, and need more attention. The functional essence of buffer zones is to act as transition areas that should minimize disturbance in protected areas. The size is very important in that sense, as it may be inadequate to deal with the ecological needs of species (e.g. sustain minimum viable populations - Harper et al. 2008).

We found that, for a small landscape species (M. paraguayanus), the average size of buffer zones needed to maintain population viability was lower than 10 km proposed by CONAMA. However, five protected areas would need a larger buffer zone than the proposal. Moreover, not all protected areas needed a buffer zone, because these protected areas were already larger than the minimum required by the species. It is important to note, however, that M. paraguayanus is a small species. The result would be different for other species, such as the primate Brachyteles hypoxantus, which requires a minimum forested area of 11.600 ha (Brito & Grelle 2006) and had been proposed as landscape species (Cunha & Grelle 2008). However, currently this primate has a restrict distribution in the state of Rio de Janeiro (Cunha et al. 2009). Therefore, a single buffer size may be insufficient to maintain ecological processes structured in the landscape. Conservation planning and actions should be based on ecological needs of species, and here we used the landscape species concept as tool for conservation action. Besides landscape species, there are other kinds of approaches related to target species that can be used in conservation planning, such as focal species, umbrella species and flagship species (Caro et al. 2004). The landscape species approach

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differs from these in that it depends on the composition and configuration of the landscapes, focusing not only in the conservation of sympatric species, but also in the conservation of the landscape structure, including the ecological functions that depend on landscape heterogeneity (Sanderson et al. 2002). The restrictions provided by law to buffer zones could be based on limits that maintain the ecological function of species (e.g. connectivity). Therefore, the size of buffers zones should be enlarged up to the point where it allows ecological functionality. For M. paraguayanus, we found impermeable barriers to dispersal (urbanization and/or water bodies) in 10 buffers zones, eight of which included urban areas. Therefore, the buffer zones would have to be larger than proposed here, in order to guarantee sufficient habitat connectivity for the species to maintain minimum viable populations. We found a relationship between buffer zone size and forest cover, besides differences in this cover forest along protected areas, but the relationship is complex, indicating that other factors must be taken into account when determining buffer size. For example, an important issue is matrix quality, which determines the possibility of biological flow (Crooks & Sanjayan 2006). Although in Brazil buffer zones are generally in private properties (Vitalli et al. 2009), it is important that public policies incentive land uses which increase matrix permeability, as in the implementation of agroforestry systems (McNeely & Schroth 2006). This can facilitate the dispersal of organisms establishing a flexible network among protected areas, and their surrounding habitat forests, and also enabling the maintenance of viable populations of the landscape of species. Indeed, higher variance in the estimated buffer size shows that these effects must be more important for smallest protected. Thus, it is necessary to increase the buffer size or small and isolated protected areas to size up its total area to a threshold to allow population viability. If it is impossible, buffer zone (and their fragments) should be viewed as a strategy for improving landscape connectivity. In conclusion, our aim here was to propose a new way to achieve buffer zones with information on minimum habitat necessary to maintain viable population. The method proposed here provides a simple analysis that can be used to establish the limits of buffer zones.

Acknowledgements Mariana Vale, Mirian Plaza Pinto and Valeska B. Oliveira for comments and suggestions in the earlier version, and José Alexandre F. Diniz-Filho provided help with spatial analyses. FAPERJ, CNPq and Conservation International-Brazil for financial support. Capes and Faperj for a scholarship to Renato Crouzeilles, and CNPq for Productivity fellowship to Carlos E. V. Grelle and ESRI that provided a free version of ArcGis 9.3.1.

References Albernaz ALKM & Souza MA, 2007. Planejamento sistemático para a conservação na Amazônia brasileira – uma avaliação preliminar das areas prioritárias de Macapá-99. Megadiversidade, 3:87-101. Brito D & Grelle CEV, 2004. Effectiveness of a reserve network for the conservation of the endemic marsupial Micoureus travassosi in Atlantic Forest remnants in Southeastern Brazil. Biodiversity and Conservation, 13:2519-2536. Brito D & Grelle CEV, 2006. Estimating minimum area of suitable habitat and viable population size for the Northern Muriqui (Brachyteles hypoxanthus). Biodiversity and Conservation, 15:4197-4210. Cáceres NC, Ghizoni-Jr IR & Graipel ME, 2002. Diet of two marsupials, Lutreolina crassicaudata and Micoureus demerarae, in a coastal Atlantic Forest island of Brazil. Mammalia, 66:331-340. Caro T et al., 2004. Preliminary assessment of the flagship species concept at a small scale. Animal Conservation, 7:63-70. Crooks KR & Sanjayan M, 2006. Connectivity conservation. Cambridge: Cambridge University Press. Cunha AA & Grelle CEV, 2008. Landscape species for conservation planning: are muriquis good candidates for the Brazilian Atlantic Forest? Natureza & Conservação, 6:17-24. Cunha, AA, Grelle CEV & Boubli JP. 2009. Distribution, population size and conservation of endemic muriquis (Brachyteles spp.) of the Brazilian Atlantic forest. Oryx, 43:254-257. ESRI, 2008. ArcView 9.3. California, U.S.A.: Redlands. Ficetola GF & Denoël M, 2009. Ecological thresholds: on the correct assessment of abrupt and gradual changes. Ecography, 32:1075-1084. Fidalgo ECC et al., 2009. Distribuição dos remanescentes vegetais no Estado do Rio de Janeiro. In: Bergallo HC et al. Estratégias e ações para a conservação da biodiversidade no Estado do Rio de Janeiro. Rio de Janeiro, Instituto Biomas. p. 91-99. Forero-Medina G & Vieira MV, 2009. Perception of a fragmented landscape by Neotropical marsupials: effects of body mass and environmental variables. Journal of Tropical Ecology, 25:53-63. Fundação SOS Mata Atlântica & Instituto Nacional de Pesquisas Espaciais – INPE, 2010. Atlas dos remanescentes florestais da Mata Atlântica no período de 2008-2010. São Paulo. Gonçalves CN et al., 2009. Buffer zone: creation or delimitation? Natureza & Conservação, 7:38-43. Grelle, CEV, 2003. Forest structure and vertical stratification of small mammal populations in a secondary forest, South-eastern Brazil. Studies on Neotropical Fauna & Environment, 38:81-85. Grelle, CEV, Lorini, ML & Pinto, MP. 2010. Reserve selection based on vegetation in the Brazilian Atlantic forest. Natureza & Conservação, 8:46-53.

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Harper EB et al., 2008. Demographic Consequences of Terrestrial Habitat Loss for Pool-Breeding Amphibians: Predicting Extinction Risks Associated with Inadequate Size of Buffer Zones. Conservation Biology, 22:1205-1215

Rangel TF, Diniz-Filho JAF & Bini LM, 2010. SAM: A comprehensive application for spatial analysis in macroecology. Ecography, 33:46-50.

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Hunter-J ML, Bean MJ & Lindenmayer DB, 2009. Thresholds and the Mismatch between Environmental Laws and Ecosystems. Conservation Biology, 23:1053-1055. Jenkins CN, Alves MAS & Pimm SL, 2010. Avian conservation priorities in a top-ranked biodiversity hotspot. Biological Conservation, 143:992-998. Legendre, P. & Legendre, L. 1998. Numerical Ecology. 2nd ed. Amsterdan: Elsevier. McNeely J & Schroth G, 2006. Agroforestry and biodiversity conservation – traditional practices, present dynamics, and lessons for the future. Biodiversity and Conservation, 15:549-554. Passamani M & Ribeiro D, 2009. Small mammals in a fragment and adjacent matrix in southeastern Brazil. Brazilian Journal of Biology, 69:305-309.

Rossi, RV, Bianconi, GV & Pedro, WA. 2006. Ordem Didelphimorphia. In: Reis N, Peracchi AL & Pedro WA. Mamíferos do Brasil. Londrina: Ed. Universidade Estadual de Londrina. p. 27-66. Sanderson EW et al., 2002. A conceptual model for conservation planning based on landscape species requirements. Landscape and Urban Planning, 58:41-46. Umetsu F, Metzger JP & Pardini R, 2008. The importance of estimating matrix quality for modeling species distribution in complex tropical landscape: a test with Atlantic forest small mammals. Ecography, 31:359-370. Vitalli PL, Zakia MJB & Durigan G, 2009. Considerações sobre a legislação correlata à zona tampão de Unidades de Conservação no Brasil. Ambiente & Sociedade, 12:67-82.

Pires AS et al., 2002. Frequency of movements of small mammals among Atlantic Coastal Forest fragments in Brazil. Biological Conservation, 108:229-237.

Received: September 2010 First Decision: September 2010 Accepted: October 2010

Research Letters

Brazilian Journal of Nature Conservation

Drafting a Blueprint for Functional and Phylogenetic Diversity Conservation in the Brazilian Cerrado Rodrigo Assis de Carvalho*, Marcus Vinicius Cianciaruso, Joaquim Trindade-Filho, Maíra Dalia Sagnori, Rafael Dias Loyola Departamento de Ecologia, Universidade Federal de Goiás – UFG, CP 131, CEP 74001-970, Goiânia, GO, Brasil

Abstract Understanding how different aspects of biodiversity are covered by protected areas and how they could be used to derive efficient conservation actions remains little studied. We mapped mammal functional and phylogenetic diversity in the Cerrado Biodiversity Hotspots to pinpoint sites with high conservation value across the biome. Further, we overlapped sites with higher or lower diversity than expected by chance with the current network of protected areas. Northeast and midwest regions emerged as priority for bats, whereas southern sites were less critical. Midsouth region captured both aspects of non-flying mammals’ diversity more than expected. Current network of protected areas covers 52% of sites with high diversity for non-flying mammals; this value being lower respective to bats (22%). Our approach provides a wall-to-wall picture on the effectiveness of the Cerrado protected areas in capturing different aspects of mammal biodiversity and points to new directions for future establishing conservation actions. Keywords: Bats, Biodiversity Hotspots, Conservation Assessment, Diversity Gradients, Protected Areas, Mammals.

Introduction Human activities are changing climate and landscapes worldwide, leading to a significant increase in extinction rates. Consequences of extinctions are not restricted to the loss of species per se but also to the erosion of phylogenetic diversity (i.e. species evolutionary history, von Euler 2001) and losses of functional diversity (i.e. the diversity of morphological, physiological and ecological traits within biological communities, Ernst et al. 2006, Petchey & Gaston 2006). Several studies have already demonstrated that phylogenetic and functional diversity might be lost faster than we lose species (e.g. Heard & Mooers 2000). Thus, preserving these different aspects of biodiversity poses a new and important challenge for conservation biology. Previous global gap analyses pointed out that biodiversity coverage by networks of existing protected areas is inadequate (Rodrigues et al. 2004). But these studies focused in the effectiveness of protected areas to represent the species and, as far as we know, there is only one study highlighting the importance to evaluate the effectiveness of existing protected areas in capturing phylogenetic and functional diversity (Devictor et al. 2010). An evaluation of protected areas capability to represent different biodiversity aspects is a critical step for conservation because if sites with high *Send correspondence to: Rodrigo Assis de Carvalho Departamento de Ecologia, Universidade Federal de Goiás, CP 131, CEP 74001-970, Goiânia, GO, Brasil E-mail: rodrigoassiscarvalho@gmail.com

phylogenetic and functional diversity are not protected by existing reserves, human impacts on species and extinction risk can be higher than previously forecasted. Decrease in communities’ phylogenetic and functional diversity lead to loss of species evolutionary history and as well as of future options to ensure provision of ecosystem goods and services (Díaz et al. 2007; Forest et al. 2007). Another important step for conservation is to identify places outside protected areas that can integrate future strategies and actions to preserve these different biodiversity aspects. Here we mapped spatial patterns of phylogenetic and functional diversity from bats and non-flying mammals inhabiting the Brazilian Cerrado and evaluated the effectiveness of existing protected areas to represent these different biodiversity facets. Moreover we identified a set of sites that could play an important role in the future to maximize the protection of phylogenetic and functional diversity in the entire biome. We choose the Brazilian Cerrado as a study case because 1) the Cerrado is a Biodiversity Hotspot with high number of endemic and rare species (Mittermeier et al. 2004), 2) it is severely threatened by the expansion of agriculture and cattle-ranching activities (Machado et al. 2004), and 3) establishment of reserves in the biome often follows subjective criteria with political and economical having more weight than biological ones (Diniz-Filho et al. 2008).

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We addressed the following questions: 1) Are protected areas from the Brazilian Cerrado capable to protect bats and non-flying mammal functional and phylogenetic diversity? 2) Which sites outside actual protected areas hold high levels of community functional and/or phylogenetic diversity?

genus/family average values. Traits compiled for bats and non-flying mammals are summarized in Tables S1 and S2 (see on-line Supplementary Material).

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Material and Methods We divided the Brazilian Cerrado into 181 equal-area grid cells of 1° × 1° degree of spatial resolution, excluding isolated and peripheral savanna areas in Amazonian region (see Diniz-Filho et al. 2008, for details). Then we overlaid extent of occurrence maps of 187 mammal species inhabiting the biome (retrieved from natureserve.com). We updated species list from Marinho-Filho et al. (2002) and constructed range maps based on both primary and secondary literature. Extent of occurrence maps entail their own limitations; therefore using a 1° grid cell recognizes the limitations of the data, reflecting a compromise between data quality and spatial resolution. It may also provide guidelines for detailed studies at finer spatial scales (see Hulbert & Jetz 2007). Recording mammals’ presence in each cell we constructed two binary matrices separating mammals into bats and non-flying mammals. We considered this division because bats are functionally very distinct from non-flying mammals, so conservation assessment and actions must be different for these two groups. For each grid cell we calculated values of functional and phylogenetic diversity associated to bat and non-flying mammal species composition. Based on these values we built maps with the spatial patterns of phylogenetic and functional diversity in the Cerrado. Phylogenetic diversity was based on the phylogeny proposed by Bininda-Edmonds et al. (2008) and we used phylogenetic diversity (henceforth PD, Faith 1992) as phylogenetic diversity index. PD is obtained by summing branch lengths of a phylogenetic tree from species that compounds a community. PD is therefore a function of species number and phylogenetic differences among species (Faith 1992). The non-flying mammals, Calomys tocantinsi, Philander frenata, Rhipidomys emiliae, Rhipidomys macrurus, and Thylamys velutinus were not present in the phylogeny used and, thus, we did not consider them in PD calculation. All bat species were present in the phylogeny. We calculated functional diversity (henceforth, FD) using the protocol proposed by Petchey & Gaston (2006): i) construction of a species-trait matrix; ii) conversion of species-trait matrix into a distance matrix; iii) clustering distance matrix into a dendrogram; and iv) calculating functional diversity by summing dendrogram branch lengths of community species. Here we used the method proposed by Pavoine et al. (2009), using Gower distance to create the distance matrix and UPGMA to build up the dendrogram. We collected trait information from the PanTHERIA database (Jones et al. 2009) updated with data collated from Marinho-Filho et al. (2002) and Reis et al. (2006). When trait values were not available for a given species we used

For each grid cell we tested if observed FD and PD were higher, equal or lower than expected by chance, assuming a null-model in which every species could occupy any grid cell in the biome. For each grid cell we fixed the observed species richness, randomized species composition without replacement, and then calculated expected FD and PD values. We repeated this procedure 1000 times for each grid cell producing a distribution of random FD values and another for PD. Finally, we checked whether observed FD and PD values for each cell were within the empirical 95% confidence interval of its simulated distribution. This approach allowed us to identify sites harboring higher or lower values of FD and/or PD than expected by chance and overlap them to existing protected areas; hence, we were able to identify if protected cells are capable to preserve these different aspects of biodiversity. We considered a grid cell as protected if it contained a reserve of at least 10 ha included in IUCN categories I-IV. All analyses were done using the R software (R Development Core Team 2009).

Results The Brazilian Cerrado has several sites with high values of FD and PD for bats (Figure 1a, b) and non-flying mammals (Figure 1c, d). Few sites had low values of mammal FD and/ or PD. However, when compared with expected values of functional and phylogenetic diversity different patterns emerged. For bats, the midwest region had sites harboring more FD than expected by chance (Figure 2a) whereas for PD these areas were located in the northeast region (Figure 2b). On the other hand, the southern region had a great number of sites with lower PD than expected (Figure 2b). For non-flying mammals, both aspects of biodiversity achieved higher values than expected by chance in the midsouth region of the biome (Figure 2c, d). Further, two sites had lower values of FD than expected by chance, while four other sites had PD values lower than expected (Figure 2c, d). Given the criteria used we considered 27 sites to be currently protected in the Brazilian Cerrado. For bats, six of them had higher FD than expected by chance, while 21 did not differ from the null expectation (Figure 3a). Two protected cells had higher values and eight lower values of PD than expected (Figure 3b). When considering the protection of both aspects of biodiversity any protected cell represented them at the same time and just three unprotected cells appears as important (Figure 3c). For non-flying mammals 14 and 12 protected cells had have higher values of PD and FD, respectively, than expected by chance (Figure 3d, e). Finally, 12 protected cells had higher values of both biodiversity aspects than expected by our null model (Figure 3f).

Functional and Phylogenetic Diversity in the Cerrado

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Figure 1. Spatial patterns of functional (a, c) and phylogenetic (b, d) diversity for bats and non-flying mammals occurring in the Brazilian Cerrado.

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Figure 2. Sites with observed functional (a, c) and phylogenetic (b, d) diversity higher (red cells) or lower (blue cells) than expected by chance for bat and non-flying mammal species pool inhabiting the Brazilian Cerrado.

Discussion Mammals compose a diversified group that play key roles in ecosystems and provide important benefits to humans (Schipper et al. 2008). Given the current biodiversity crisis, conservation actions must be taken fast if we want to preserve these species, their evolutionary history and the ecological processes shaping communities and driven diversity at different spatial scales. Here, we presented a first approach that accounts for FD and PD to shed light on how these biodiversity aspects are protected and how to apply them for future conservation actions under a conservation biogeography approach (sensu Whitakker et al. 2005). In the last decades, ecologists, macroecologists and conservation biologists developed analytical tools and methods that made possible a critical evaluation of the differences among species to better understand community structure and composition (Devictor et al. 2010). PD and FD are measures that quantify such differences within and among communities. While the former measure focuses in the historical biogeographical events to depict the evolutionary history of local and regional assemblages (Webb et al. 2002) the later reflects the role of species interactions and trait diversity of communities that are

supposed to be linked to ecosystem functioning (Petchey & Gaston 2006). Thus, preserving phylogenetic and functional diversity may, respectively, guarantee the maintenance of evolutionary processes and features, as well as the continuity in goods and ecosystem services provision. Conservation of functional and phylogenetic diversity in the Brazilian Cerrado is threatened by the expansion of agriculture and cattle ranching, leading to loss of natural habitats. If no action is taken, the biome is likely to disappear until 2030 (Machado et al. 2004), putting in jeopardy species that hold unique evolutionary features, as well as important ecological traits that maintain ecological processes or, in the worst scenario, all of these aspects of biodiversity. We found that the current network of protected areas established in the Cerrado is not entirely capable to represent mammal functional and phylogenetic diversity. For non-flying mammals 52 and 44% of protected sites represent FD and PD, respectively, better than one would expect by chance alone. The situation is worse for bats, as only 22 and 7.4% of protected sites overlap with sites of higher FD and PD, respectively. Moreover, there is an aggravating factor: 30% of the protected sites carry lower PD than expected.

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Figure 3. Spatial overlap of existing protected cells (green cells) and sites with observed functional (a, d) and phylogenetic (b, e) diversity higher (red cells) or lower (blue cells) than expected by chance for bat and non-flying mammal species pool inhabiting the Brazilian Cerrado; c) and f) stands for the spatial overlap of protected areas and sites showing, at the same time, higher functional and phylogenetic diversity, Color codes as above. PC: Protected Cell.

Inefficiency of existing protected areas in the Brazilian Cerrado for these two biodiversity aspects reflects the consequences of government opportunistic old-fashion way to establish nature preserves, choosing sites for conservation under political and economical criteria (Diniz-Filho et al. 2008) or just for their scenic value. Such policy for establishing protected areas is surely not exclusive from the Brazilian Government, being commonly applied worldwide (Margules & Pressey 2000). As a consequence, species evolutionary history and the diversity of ecological traits, which should have a close link to ecosystem processes, may be more threatened than we expect. Under this scenario selecting new areas to maximize the representation of phylogenetic and functional diversity is essential for conservation purposes. In fact, recent attempts to include species evolutionary history (e.g. Forest et al. 2007, Loyola et al. 2008a) and biological traits (e.g. Loyola et al. 2008b, 2009) in conservation planning have been published elsewhere. In the Brazilian Cerrado, in particular, three sites are crucial for conserving higher values of bat FD and PD at the same time (see Figure 3c); non-flying mammal PD and FD could be well conserved in the biome focusing in the midsouth region (see Figure 3d, f). We suggest these areas should

be the focus of future studies aimed at applying a spatial conservation prioritization for the Cerrado. Although we indicate these regions as priority sites for mammal FD and PD conservation we call attention that other sites are also important to the persistence of species. For example, if we consider complementarity, a key concept in spatial conservation prioritization (Moilanen et al. 2009), maybe the northern sites appears as important for non-flying mammals because they have species not represented in current network of protected areas. But if systematic conservation plans built on complementarity indicate priority sites both in northern and southern regions, conservation investments should be first placed in the South because this region captures higher values of FD and PD than expected by a site from the north. In a nut-shell, spatial patterns of different biodiversity facets can add important information to guide decisions of where start to invest aiming to maximizing conservation of all biodiversity. Finally, our study reinforces the idea proposed by Devictor et al. (2010), which is to avoid strategies using a single biodiversity aspect as a cure-all. Despite several areas with high values of FD and PD for non-flying mammals are congruent, bats are not in the same situation. While good areas to preserve FD of bats are in midwest region,

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northeast concentrates better sites for conserving PD. Hence, effective conservation strategies may emerge with biodiversity assessments done under integrative approaches connecting biogeography, evolutionary and functional ecology (Johnson & Stinchcombe 2007).

Loyola RD et al., 2008a. Conservation of Neotropical carnivores under different prioritization scenarios: mapping species traits to minimize conservation conflicts. Diversity and Distributions, 14:949-960.

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Acknowledgements We are grateful to D. Brito, M. Oprea, and F.L. Sobral; for helping us with the functional database. R.A.C. research is supported by a UFG doctoral scholarship; J.T.F. received a CAPES Msc. scholarship; M.D.S. received a CNPq scholarship; R.D.L. research is supported by CNPq (project #475886/2009-7); R.D.L. and M.V.C. work is also funded by CAPES (project #012/09).

References Bininda-Emonds ORP et al., 2008. The delayed rise of present-day mammals. Nature, 446:507-512. Devictor et al., 2010. Spatial mismatch and congruence between taxonomic, phylogenetic and functional diversity: the need for integrative conservation strategies in a changing world. Ecology Letters, 13:1030-1040.

Loyola RD et al., 2008b. Hung out to dry: choice of priority ecoregions for conserving threatened Neotropical anurans depend on their life-history traits. PLoS ONE, 3:e2120. Loyola RD et al., 2009. Integrating Economic Costs and Biological Traits into Global Conservation Priorities for Carnivores. PLoS One, 4:e6807. Machado RB et al., 2004. Estimativas de perda da área do cerrado brasileiro. Brasília: Conservation International. Margules CR & Pressey RL, 2000. Systematic conservation planning. Nature, 405:243-253. Marinho-Filho J, Rodrigues FHG & Juarez KM, 2002. The Cerrado mammals: Diversity, Ecology and Natural History In: Oliveira PS & Marquis RJ, ed. The Cerrados of Brazil: ecology and natural history of a Neotropical savanna. New York: Columbia University Press. 266-284 p. Mittermeier RA et al., 2004. Hotspots Revisited: Earth’s Biologically Richest and Most Endangered Ecoregions. CEMEX, Mexico City, Mexico.

Díaz S et al., 2007. Incorporating plant functional diversity effects in ecosystem service assessments. Proceedings of the National Academy of Science, 104:20684-20689.

Moilanen A, Wilson KA & Possingham HP, 2009. Spatial conservation prioritization: quantitative methods & computational tools. Oxford: Oxford University Press. 304 p.

Diniz-Filho JAF et al., 2008. Spatial patterns of terrestrial vertebrate species richness in the Brazilian Cerrado. Zoological Studies, 47:146-157.

Pavoine S et al. 2009. On the challenge of treating various types of variables: application for improving the measurement of functional diversity. Oikos, 118:391-402.

Ernst R, Linsenmair KE & Rodel MO, 2006. Diversity erosion beyond the specie level: Dramatic loss of functional diversity after selective logging in two tropical amphibian communities. Biological Conservation, 133:143-155.

Petchey OL & Gaston KJ, 2006. Functional diversity: back to basics and looking forward. Ecology Letters, 9:741-758.

Faith DP, 1992. Conservation evaluation and phylogenetic diversity. Biological Conservation, 61:1-10. Forest F et al., 2007. Preserving the evolutionary potential of floras in biodiversity hotspots. Nature, 445:757-760. Heard SB & Mooers AÒ, 2000. Phylogenetically patterned speciation rates and extinction risks change the loss of evolutionary history during extinctions. Proceedings of the Royal Society B, 267:613-620. Hulbert AH & Jetz W, 2007. Species richness, hotspots, and the scale dependence of range maps in ecology and conservation. Proceedings of the National Academy of Science, 104:13384-13389. Johnson MTJ & Stinchcombe JR, 2007. An emerging synthesis between community ecology and evolutionary biology. Trends in Ecology & Evolution, 22:250-257. Jones KE et al., 2009. PanTHERIA: a species-level database of life history, ecology, and geography of extant and recently extinct mammals. Ecology, 90:2648.

R Development Core Team, 2009. R: A language and environment for statistical computing. Vienna: R Foundation for Statistical Computing. Available from: <http://www.R-project.org>. Reis NR et al., 2006. Mamíferos do Brasil. Londrina. 437 p. Rodrigues ASL et al., 2004. Effectiveness of the global protected area network in representing species diversity. Nature, 428:640-643. Schipper J. et al., 2008. The Status of the World’s Land and Marine Mammals: Diversity, Threat, and Knowledge. Science, 322:225-230. von Euler F, 2001. Selective extinction and rapid loss of evolutionary history in the bird fauna. Proceedings of the Royal Society B, 268:127-130. Webb CO et al., 2002. Phylogenies and community ecology. Annual Review of Ecology and Systematics, 33:475-505. Whitakker RJ et al, 2005. Conservation Biogeography: assessment and prospect. Diversity and Distributions, 11:3-23.

Received: September 2010 First Decision: October 2010 Accepted: October 2010

Research Letters

Brazilian Journal of Nature Conservation

The Opportunity Cost of Conserving Amphibians and Mammals in Uganda Federica Chiozza, Luigi Boitani & Carlo Rondinini* Department of Biology and Biotechnology, Sapienza Università di Roma, Viale dell’Università 32, 00185, Roma, Italy

Abstract Despite substantial conservation efforts, biodiversity continues to decline and further conservation action is needed. This imposes significant opportunity cost on local communities, particularly in developing countries where livelihood depends strictly on land use and agricultural activities. Incorporating socio-economic data into methods for the identification of conservation priorities can reduce conflicts between socio-economic development and biodiversity conservation. We present a systematic selection of priority sites for the conservation of 353 Ugandan mammals and amphibians. We used the suitable habitat as an estimate of the area potentially occupied by each species inside its geographic range, and estimated the opportunity cost based on data on agricultural profit. We used the software Marxan to identify the sites that need to be added to the existing protected areas (IUCN categories I-IV) to conserve Ugandan mammals and amphibians at a minimum cost. In addition to the existing protected areas, covering ca. 17,100 km2, ca. 57,500 km2 of land should be protected to achieve the conservation target for amphibians and mammals, bringing the coverage to ca. 38% of the country. The sites that are irreplaceable for the target achievement occupy ca. 32,800 km2, are mostly located in the Western and Eastern regions and overlap with the Eastern Afromontane hotspot and the Albertine Rift. The yearly agricultural profit from these sites amounts to ca. 540,700,000 US$, or 16,524 US$/km2 (2008 value). Key words: Conservation Planning, Conservation Priority Setting, Conservation Conflict, Agriculture, Habitat Suitability Model.

Introduction Despite substantial conservation efforts, biodiversity continues to decline (Hoffman et al. 2010). Further conservation action is thus needed, but this imposes a significant opportunity cost (the income and other benefits from land use, investment and development opportunities precluded or diminished by the need to maintain biodiversity (Emerton & Muramira 1999) on local communities. This is particularly true in developing countries, where livelihood depends strictly on land use and agricultural activities (Norton-Griffiths & Southey 1995). Opportunity cost varies substantially across sites (Adams et al. 2010, Naidoo et al. 2006, Polasky et al. 2001), and this is reflected in variable cost-effectiveness of alternative conservation plans. It has been demonstrated that conservation plans that consider or not consider opportunity cost can protect similar numbers of species, but the former are much cheaper (Naidoo & Adamowicz 2006). The biggest limitation in incorporating cost in conservation plans is the scarcity of spatially explicit data, *Send correspondence to: Carlo Rondinini Department of Biology and Biotechnology, Sapienza Università di Roma, Viale dell’Università 32, 00185, Roma, Italy E-mail: carlo.rondinini@uniroma1.it

especially for developing countries. In these cases proxies of cost are used, such as human population density or the distribution of infrastructures (Williams et al. 2003), but in order to be effective, the surrogate measures chosen should estimate actual cost, or at least have a strong spatial correlations with actual cost (Adams et al. 2010). Uganda is a global biodiversity hotspot (Mittermeier et al. 2004) and includes some the richest spots in Africa in terms of vertebrate species (Rondinini et al. 2005), but its protected area (PA) coverage is relatively low (IUCN & UNEP 2010). Rapid population growth, high population density and heavy reliance on subsistence agriculture for income and park resources for subsistence characterise areas surrounding National Parks (Archabald & Naughton-Treves 2001). Conflicts with local communities in several National Parks are still a major concern for Uganda Wildlife Authority, and a negative attitude towards conservation persists in local people (Chhetri et al. 2003). Consideration of opportunity cost to minimise further socio-economic conflict would provide added value to plans for the conservation of biodiversity in Uganda. In this paper we used the software Marxan to systematically identify sites to be added to the existing PA network to protect the Ugandan amphibians and mammals, while

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minimising the cost of the conservation plan. We estimated the area potentially occupied by each of 353 species of Ugandan mammals and amphibians as the area of suitable habitat inside their geographic ranges (Rondinini et al. 2005; Rondinini & Boitani 2006). We used data on agricultural production and net profit value derived from agricultural activities to estimate agricultural opportunity cost.

Material and Methods Estimation of the area potentially occupied by species We determined the geographic ranges for 377 African vertebrates (65 amphibians and 312 mammals) whose distribution falls within Uganda’s administrative boundaries. To avoid planning reserves for species for which Uganda is at the edge of their geographic distribution, 24 marginal species were excluded from the analysis. These species have less than 1% of their extent of occurrence falling within Uganda’s administrative boundaries and occupy less than 1% of Uganda land area. For 341 of these species we determined the area of suitable habitat inside the geographic range (at 1 km2 resolution) using species’ habitat preferences. We obtained the data from the literature and experts (in collaboration with the IUCN Global Amphibian Assessment and the IUCN Global Mammal Assessment) and built on an existing database on large- and medium-size African mammals (Boitani et al. 1999). We reclassified as suitable or unsuitable the land classes of a land cover map (United States Geological Survey 2000), the elevation values from a digital elevation model (United States Geological Survey 2001), and the distances to water from a map of water bodies and water courses (Environmental Systems Research Institute 1993). For each species we used the intersection of suitable areas from the three environmental layers as the estimated area of suitable habitat. This estimation is more robust than other, more permissive estimates in terms of the prevalence of false positive errors (Rondinini & Boitani 2006), which are dangerous in conservation because they may lead to the protection of sites that do not contain the species of interest (Loiselle et al. 2003, Rondinini et al. 2006). The modeling procedure and validation of results are fully described elsewhere (Rondinini et al. 2005). We performed all the cartographic data processing with ArcInfo GIS 8.3 (Environmental Systems Research Institute, CA, USA). For 12 poorly known species we were unable to collect enough information regarding species-habitat relationships, which prevented us from estimating habitat suitability inside their range. However we included these species in the analysis and replaced their missing suitable area with the estimated geographic range. This was necessary as many of these species have restricted ranges and their removal from the sample would underestimate the high conservation value of the sites where they occur.

Agricultural opportunity cost We used total yearly net profit from agricultural activities (hereafter yearly agricultural profit) as an estimate of yearly agricultural opportunity cost, to estimate the potential benefits foregone as a result of protecting an area rather than using it for agriculture production. We obtained crop statistics from the 2002 Uganda Population and Housing Census (UBOS 2004). The census incorporates an agricultural module inquiring about household-based agricultural activities (MAAIF et al. 2010). Since the census included the enumeration of all households, we aggregated the data to small administrative areas (sub-county level). The crop data is supplied for various crops, including coffee and cotton that represent the two major cash crops cultivated in Uganda, in terms of number of plots for each administrative unit, where a crop plot is defined as a piece of land within the agricultural holding on which a specific crop or crop mixture is cultivated (Table 1). To estimate the average plot size in each administrative unit we divided the total number of plots cultivated by the total cropped area within that unit (UBOS 2004). This value was used to calculate the hectares under a specific crop cultivated in each administrative unit, assuming that all plots within that unit have the same size. In addition, we obtained data from the National Agricultural Advisory Services (NAADS 2003) on yearly net profit per hectare (expressed in Uganda Shillings) for some of the crop types surveyed by the 2002 Uganda Population and Housing Census (UBOS 2004). Net profit from agriculture indicates the revenue of farm activities and is calculated as the difference between the value of outputs and expenditures (Table 1). Yearly agricultural profit for each administrative unit was estimated as: J

∏ =∑ x

Pj + x ui P

(1)

j=1

where xji is the area under crop j in the administrative unit i; Pj is the net profit per hectare for crop j; xui is the area under crop with unknown revenue in the administrative unit i; and P is the average net profit for cultivated crops with unknown revenue. We converted the value of the yearly agricultural profit from Uganda Shillings in 2002 to US dollars in 2008 using the inflation measured by the consumer price index (World Bank 2010), and the exchange rate.

Systematic selection of conservation priority sites In addition to the distribution of target species and the cost of land, the analysis for the identification of conservation priority sites required two other pieces of information: 1) the conservation target to be achieved for each target

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Table 1. Total number of plots (UBOS 2004) and yearly net profit (NAADS 2003) of each type of crop in Uganda.

Crop

Total number of plot

Net profit (Ushs/ha)

Beans Cassava Coffee Cotton Cow peas Groundnuts Irish potatoes Millet Onion Pineapple Rice Simsim Sorghum Soya beans Sweet potatoes Tomato Vanilla

2 152 076 2 101 580 334 846 109 961 64 200 754 709 146 810 750 722 31 020 1 800 86 394 192 504 763 779 1 454 1 439 658 2 980 9 786

81 000 414 250 530 000 102 700 225 909 211 000 389 500 346 250 2 028 119 1 677 956 473 000 125 050 174 900 165 508 426 500 649 194 9 075 000

species (Rondinini & Chiozza 2010), and 2) the boundaries of sites (Margules & Pressey 2000). For species whose area of suitable habitat was smaller than 1,000 km2 we set the conservation target to 100% of this area; for more widespread species, with area of suitable habitat greater than 10,000 km2, we set the target to 10% of this area; for species with intermediate area of suitable habitat, we interpolated the conservation target between these two extremes, proportionally to the log of the area of suitable habitat. In an earlier analysis we demonstrated that the use of different targets resulted in only minor differences in the site selection outcome (RondininiÂ etÂ al. 2005). In order to generate a map of sites we selected the Ugandan reserves from the World Database on Protected Areas (IUCN & UNEP 2010). For this analysis only those protected areas that qualified as IUCN categories I to IV were included. We generated a grid of squares of 25 km2 to divide the rest of the country into planning units selectable in the site selection analysis. Cells with an area smaller than 12.5 km2, found along the country administrative boundaries or generated by the intersection of the grid with the protected area map, were removed by merging them with the adjacent ones. Uganda was so subdivided into 8,970 planning units of which 21 are represented by national protected areas with IUCN category I to IV. We performed the site selection analysis with the software Marxan (Ball & Possingham 2000).We selected the analysis input parameters as follows: algorithm, simulated annealing; number of simulations, 10,000; iterations per simulation,

100,000; number of temperature decreases per simulation, 20,000; choice of the initial temperature and cooling factor, adaptive; and boundary length modifier, 200. Marxan simulations produced 10,000 different solutions to the conservation problem. The selection frequency of each site is therefore an estimate of the overall value of the site for the achievement of the conservation target. The sites selected 10,000 times (i.e. with selection frequency equal to 1) were likely to be necessary to the achievement of the target, conditional to the parameters set for the problem. Hereafter these sites will be referred to as irreplaceable. It is important to note that irreplaceability estimated through selection frequency is not absolute but conditional to the problem constraints, including the cost of sites. If for example two sites, one cheap and one highly expensive, contain the same species, the former but not the latter may be always selected (i.e. be considered irreplaceable). Newly selected sites were added to the existing IUCN protected areas I-IV. We assigned a penalty factor of 100 for each species missing in the final reserve system. This way we ensured that the target was met for all species in the selected systems of reserves. To allow conservation targets to be achieved at the minimum cost, we used the opportunity cost from agriculture activities as the monetary value of each planning unit. The software and all procedural details are freely available online from: http://www.ecology.uq.edu. au/marxan.htm (accessed October 2010).

Results The yearly agricultural opportunity cost of land in Uganda varied from 0 to 55,320 US$/km2 with an average value of 13,733 US$/ km2 (2008 value). Western and central regions were the ones with the highest mean values of land, with most of the high-value land concentrated along northern shores of Lake Victoria (Figure 1). The land surface occupied by the existing IUCN protected areas of categories I to IV is approximately 17,100 km2 corresponding to 8.7% of Uganda. In order to achieve our conservation target for 59 amphibians and 294 mammals, another 57,515 km2 (29.1% of the country) was necessary. Of these, 32,723 km2, corresponding to 57% of the total area selected for complementing the current Uganda reserve system, were irreplaceable. The irreplaceable sites were mainly clustered in the western and eastern regions of Uganda (Figure 2). Those in the western region were mainly found along the administrative boundary with Democratic Republic of Congo and Rwanda, while the block of irreplaceable sites in the eastern region surrounded the Mt. Elgon Nature Reserve. Other irreplaceable sites were found along northern shores of Lake Victoria and in the northern region along the administrative boundary with Kenya. Overall, the yearly agricultural profit from the sites that are irreplaceable for the conservation of Ugandan mammals and

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Natureza & Conservação 8(2):177-183, December 2010

Figure 1. Annual revenue from agriculture in Uganda.

amphibians amounted approximately to 540,700,000 US$, equal to 16,524 US$/km2. The yearly agricultural profit of the most efficient Marxan solution (including all the irreplaceable sites plus other replaceable sites needed to achieve the conservation target for amphibians and mammals) was 772,850,000 US$ or 13,437 US$/km2.

Discussion Our approach to estimating opportunity cost relied on a number of assumptions. We assumed that all plots in an administrative unit are the same size. We are aware that if conditions, such as rainfall distributions, soils and topography are suitable, usually the plots cultivated with cash crops are very different in terms of size to the plots cultivated with crops for subsistence farming, thus we may have somewhat underestimated the profit of each administrative unit. We only estimated the agricultural profit of land, and we assumed that productivity and profit for each crop type were even across the country. Agriculture is not the only

reason for land to be valuable. Other types of land use, such as timber and mining extraction, can make land valuable and difficult to set aside for conservation. Yet agriculture plays a key role in Uganda’s economic development. Uganda has put emphasis on the agricultural sector as a strategy for raising rural incomes and reducing rural poverty (NEMA 2005). For the majority of Ugandans, the agricultural sector (including crops, livestock, and fisheries) is the main source for livelihoods, employment, and food security (MAAIF et al. 2010). Despite its slow decline in the past years due to a number of reasons (including drought, instability, pest outbreaks, and productivity and price declines for selected crops and commodities, NPA 2010), combined with the faster growth in the services and industrial sectors, the agriculture’s share of Uganda’s Gross Domestic Product (GDP) was 38% in 2009 (World Bank 2010). The sector provided 73.3 percent of employment in 2005/06, and most industries and services in the country are dependent on it (UBOS 2009).

Opportunity Costs of Conservation in Uganda

181

Figure 2. Irreplaceability of sites in Uganda, i.e. number of times each site was included in a Marxan solution that achieved the conservation target for all amphibians and mammals. Sites with selection frequency = 1 are likely to be not replaceable if the target is to be achieved.

The agricultural profit in Uganda was higher in highlands and in Lake Victoria areas. Lake Victoria shores are characterised by intensive cropping, having soils suitable for agriculture, receiving sufficient rainfall to support perennial crop production and having the most favourable access to infrastructure and markets compared to other regions in Uganda (NEMA 2005). Elsewhere, some 40% of the people in rural areas still live below the poverty line, accounting for 95% of the total number of Uganda’s poor; most of them depending on agriculture as their primary source of livelihood (Fan et al. 2004). More than 15% of Uganda appeared to be irreplaceable for the achievement of the conservation target for amphibians and mammals. These sites were mostly concentrated in western and eastern regions, and many of them overlapped with the Eastern Afromontane hotspot already identified as a region with exceptional levels of endemism and by serious levels of habitat loss (Mittermeier et al. 2004). South-west

Uganda is a key component of the larger Albertine Rift, one of the richest parts of the world in terms of biological diversity (Nantamu 2005). The block of irreplaceable sites in the eastern region correspond to Mount Elgon Nature Reserve. This is a volcanic massif with high conservation and ecological value due to its rare endemic species, valuable forests and water catchment functions (Chhetri et al. 2003). Many sites of the northern region overlapped with areas already identified by IUCN as protected areas with category VI and designated as Controlled Hunting Area. These areas, allowing for a sustainable use of resources, impose a lower opportunity cost than strict reserves. Category VI protected areas are managed to ensure long term protection and maintenance of biological diversity, while providing at the same time a sustainable flow of natural products and services to meet community needs. Categories V and VI protected areas have obvious relevance in the contest of rural poverty as these promote and support traditional

Chiozza et al.

Natureza & Conservação 8(2):177-183, December 2010

livelihoods and cultures as well as protection of biodiversity (Scherl et al. 2004).

(FAO - AGAL). We are grateful to the experts involved in the IUCN Global Amphibian Assessment and IUCN Global Mammal Assessment and in particular to J. Chanson, N. Cox, W. Sechrest, S. Stuart and B. Young.

182

If the conservation of Ugandan amphibians and mammals were to be achieved through strict reservation, profit in excess of 500,000,000 US $ could be lost annually. Although high, this value represents a small fraction, 12.5%, of the Uganda Agriculture GDP for 2008 (4,332 billion US$, World Bank 2010). Other estimates at a national level have shown that states can incur considerable opportunity costs from the loss of agricultural land to protected areas (Howard 1996, Norton-Griffiths & Southey 1995). For developing countries as a whole, James et al. (2001) estimate a total opportunity cost for existing reserves in categories II, III, and IV, occupying 3.62 million km2, of 4.9 billion US$. The high agricultural opportunity cost of conservation in Uganda could be reduced by aiming at coexistence and conflict resolution in place of strict reservation, trading off between biodiversity conservation and poverty alleviation. The problem of poverty is acute in Uganda rural areas (Woelcke 2006). Moreover, even if parks generate an economic return, the distribution of these benefits is so skewed against poor rural people that the role of parks in local development is negligible and they neither justly compensate for lost property nor contribute to poverty alleviation (Brockington 2003). These imbalances act as a constraint to biodiversity conservation. This recognition has inspired adoption of different human-inclusive strategies guided by the philosophy that the success of conservation objectives depends strictly by the interests of the communities (Kideghesho et al. 2007). Our approach to minimising conservation conflict relies on agricultural profit, therefore it does not accounts for conservation conflicts sensu Araujo & Rahbek (2007), i.e. the coincidence of complementarity areas and sites with high human population density. Indeed, it may be important to explore the trade-offs between minimising the cost of a conservation plan and minimising its impact on rural populations (Adams et al. 2010). Even if we accounted for all reasons for land to be valuable, this would represent only a partial estimate of the economic cost of conservation. Real economic cost include acquisition cost, management cost, and transaction cost (Naidoo et al. 2006). Yet, by calculating opportunity cost from a substantial source of profit for a large share of the Ugandan population, our conservation plan can explicitly minimise local socioeconomic impacts, reducing local poverty and displacement of communities (Adams et al. 2010), and providing a plausible basis for a full conservation plan.

Acknowledgments We would first like to acknowledge the collaboration offered by Thomas Emwanu, Senior System Analyst at the Uganda Bureau of Statistics (UBOS), and the support and collaboration of Timothy Robinson and Ugo PicaCiamarra

References Adams VM, Pressey RL & Naidoo R, 2010. Opportunity costs: who really pays for conservation? Biological Conservation, 143:439-448. Araujo MB & Rahbek C, 2007. Conserving biodiversity in a world of conflict. Journal of Biogeography, 34:199-200. Archabald K & Naughton-Treves L, 2001. Tourism revenuesharing around national parks in Western Uganda: early efforts to identify and reward local communities. Environmental Conservation, 28:135-149. Ball IR & Possingham HP, 2000. Marine reserve design using spatially explicit annealing. Available from: <http://www. ecology.uq.edu.au/marxan.htm>. Access in: July 2007. Boitani L et al., 1999. A databank for the conservation and management of African Mammals. Roma: Istituto di Ecologia Applicata. Brockington D, 2003. Injustice and conservation: is local support necessary for sustainable protected areas? Policy Matters, 12:22-30. Chhetri P, Mugisha A & White S, 2003. Community resource use in Kibale and Mt Elgon National Parks, Uganda. Parks, 13:28-38. Emerton L & Muramira E, 1999. Uganda biodiversity: economic assessment. Biodiversity Economics for East Africa. Nairobi, Kenya: The World Conservation Union - IUCN. Environmental Systems Research Institute, 1993. Digital chart of the world for use with ARC/INFO software. Redlands: Environmental Systems Research Institute. Fan S, Zhang X & Rao N, 2004. Public expenditure, growth, and poverty reduction in rural Uganda. Washington, D.C.: International Food Policy Research Institute, Development Strategy and Governance Division. Discussion paper, 4. Hoffman M et al., 2010. The impact of conservation on the status of the world’s vertebrates. Science. Howard P, 1996. The opportunity costs of protected areas in Uganda. In: IUCN Workshop on Economics of Biodiversity Loss. Gland, Switzerland. International Union for Conservation of Nature - IUCN & United Nations Environment Programme - UNEP, 2010. The World Database on Protected Areas (WDPA). Cambridge, UK: UNEP-WCMC. James A, Gaston KJ & Balmford A, 2001. Can we afford to conserve biodiversity? BioScience 51:43-52. Kideghesho JR, Røskaft B & Kaltenborn BP, 2007. Factors influencing conservation attitudes of local people in Western Serengeti, Tanzania. Biodiversity Conservation, 16:2213-2230. Loiselle BA et al., 2003. Avoiding pitfalls of using species distribution models in conservation planning. Conservation Biology, 17:1591-1600.

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Margules CR & Pressey RL, 2000. Systematic conservation planning. Nature, 405:243-253.

estimators of the area of occupancy. Conservation Biology, 20:170-179.

Ministry of Agriculture, Animal Industry and Fisheries - MAAIF, Uganda Bureau of Statistics - UBOS, Food and Agriculture Organization of the United Nations - FAO, International Livestock Research Institute - ILRI & World Resources Institute - WRI, 2010. Mapping a better future: spatial analysis and pro-poor livestock strategies in uganda. Washington, D.C.: World Resources Institute.

Rondinini C & Chiozza F, 2010. Quantitative methods for defining percentage area targets for habitat types in conservation planning. Biological Conservation, 143:1646-1653.

Mittermeier RA et al., 2004. Hotspots revisited: earth’s biologically richest and most endangered terrestrial ecoregions. Mexico City: Conservation International and Agrupación Sierra Madre - CEMEX. Naidoo R & Adamowicz WL, 2006. Modeling opportunity costs of conservation in transitional landscapes. Conservation Biology, 20:490-500. Naidoo R. et al., 2006. Integrating economic costs into conservation planning. Trends in Ecology and Evolution, 21:681-687. Nantamu N, 2005. The Albertine Rift: a globally important endangered space harbors the majority of Africa’s biodiversity. Kampala: USAID/Uganda. Environmental Brief, 3. National Agricultural Advisory Services - NAADS, 2003. Profitability Analysis of Agricultural Enterprises in Uganda. Kampala, Uganda. National Environment Management Authority - NEMA, 2005. State of environment report for Uganda 2004/05. Kampala, Uganda. National Planning Authority - NPA, 2010. National Development Plan 2010/11 – 2014/15. Kampala, Uganda: NPA. Norton-Griffiths M & Southey C, 1995. The opportunity costs of biodiversity conservation in Kenya. Ecological Economics, 12:125-139. Polasky S, Camm JD & Garber-Yonts B, 2001. Selecting biological reserves cost-effectively: an application to terrestrial vertebrate conservation in Oregon. Land Economics, 77:68-78. Rondinini C & Boitani L, 2006. Differences in the umbrella effects of African amphibians and mammals based on two

Rondinini C et al., 2006. Tradeoffs of different types of species occurrence data for use in systematic conservation planning. Ecology Letters, 9:1136-1145. Rondinini C, Stuart S & Boitani L, 2005. Habitat suitability models and the shortfall in conservation planning for African vertebrates. Conservation Biology, 19:1488-1497. Scherl LM et al., 2004. Can protected areas contribute to poverty reduction? Opportunities and limitations. Gland, Switzerland: IUCN. Uganda Bureau of Statistics - UBOS, 2004. Report on agriculture module, piggy-backed onto the population and housing census 2002. Entebbe, Uganda. Uganda Bureau of Statistics - UBOS, 2009. Statistical Abstract. Kampala, Uganda: UBOS. Available from: <http://www.ubos.org/onlinefiles/uploads/ubos/pdf%20 documents/2009Statistical_%20Abstract.pdf>. Access in Oct. 2010. United States Geological Survey, 2000. Global land cover characterisation ver. 1.2. Available from: <http://edcdaac. usgs.gov/glcc/glcc.html>. Access in: Oct. 2010. United States Geological Survey, 2001. GTOPO30. Available from: <http://lpdaac.usgs.gov/gtopo30/gtopo30.html>. Access in: Oct. 2010. Williams PH et al., 2003. Integrating biodiversity priorities with conflicting socio-economic values in the Guinean–Congolian forest region. Biodiversity and Conservation, 12:1297-1320. Woelcke J, 2006. Technological and policy options for sustainable agricultural intensification in eastern Uganda. Agricultural Economics, 34:129-139. World Bank, 2010. World Development Indicators database. Available from <http://data.worldbank.org/>. Access in: Oct. 2010.

Received: October 2010 First Decision: October 2010 Accepted: October 2010

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Brazilian Journal of Nature Conservation

Geoconservação em Áreas Protegidas: o Caso do GeoPark Araripe - CE Geoconservation in protected areas: the case of Geopark Araripe/CE Nájila Rejanne Alencar Julião Cabral1 & Teresa Lenice Nogueira da Gama Mota2,* 1

Departamento de Construção Civil, Instituto Federal de Educação, Ciência e Tecnologia do Ceará – IFCE

Secretaria da Ciência, Tecnologia e Educação Superior do Ceará – SECITECE

Em setembro de 2006, oriunda de iniciativa do Governo do Estado do Ceará, por intermédio da Secretaria da Ciência, Tecnologia e Educação Superior (SECITECE), a UNESCO aprovou a solicitação do GeoPark Araripe, que passou a fazer parte da Rede Global de GeoParks (Global GeoParks Network – GGN). GeoPark é um selo outorgado pela UNESCO para áreas com significativo patrimônio geológico de especial interesse científico, que contenham atributos de valor natural raro, deve integrar sítios naturais e pontos turísticos de interesse cultural, constituindo-se em espaço fundamental para proteção dos recursos naturais, o geoturismo, educação e a popularização da ciência. A Rede Global de GeoParks (Global..., 2010) possui atualmente 64 GeoParks em 19 países do mundo, a saber: Austrália (1), Áustria (1), Brasil (1), China (22), Croácia (1), República Tcheca (1), França (2), Grécia (3), Alemanha (5) Irã (1), Itália (5), Japão (3), Malásia (1), Noruega (1), Portugal (2), Irlanda (1), Romênia (1), Espanha (4) e Reino Unido (8). 2

O GeoPark Araripe, único das Américas, possui 3520,52 km , sendo o território constituído pelos municípios de Barbalha, Crato, Juazeiro do Norte, Missão Velha, Nova Olinda e Santana do Cariri, no Estado do Ceará. Abriga geossítios singulares que refletem a história da evolução da Terra e do Homem, que traduzem a riqueza natural e cultural daquela região. A Unidade Executora do GeoPark Araripe está sob a responsabilidade da Fundação Universidade Regional do Cariri (URCA). Recebe forte apoio institucional do Governo do estado do Ceará, por intermédio da SECITECE, Secretaria das Cidades, da Secretaria de Turismo (SETUR), da Secretaria da Cultura (SECULT), do Conselho de Políticas e Gestão do Meio Ambiente (CONPAM) e da Secretaria da Educação (SEDUC). Possui, ainda, significativa contribuição da sociedade *Send correspondence to: Nájila R. A. Julião Cabral Av. Treze de Maio, 2081, Bairro Benfica, CEP 60040-531, Fortaleza, CE, Brasil. E-mail: najila@ifce.edu.br

civil de todos os municípios envolvidos; com conseqüente respaldo na viabilização do desenvolvimento sustentável. Considerando-se GeoPark como uma província de desenvolvimento sustentável e observando-se seu caráter de possibilitar a proteção dos recursos ambientais ali presentes em consonância com o desenvolvimento socioeconômico das populações residentes, mencionada área adquire importância no contexto do modelo de preservação e conservação dos recursos ambientais de países desenvolvidos e em desenvolvimento, uma vez que desempenha ambos procedimentos citados por Cabral & Souza (2005), concernente ao papel das áreas protegidas: a) interrompe, em alguns casos, a atuação antrópica de modo a permitir a manutenção e a recuperação dos atributos ambientais e b) permite o uso dos recursos ambientais garantindo sua manutenção no longo prazo em condições regulares, minimizando os efeitos adversos da atuação antrópica. GeoParks costumam ocupar vastos espaços territoriais, muitas vezes, extrapolando limites municipais (como por exemplo, o GeoPark Araripe, Ceará, que congrega 6 municípios), aliado ao fato de possibilitarem o desenvolvimento de atividades econômicas em seus limites (em geral, privadas), que costumam ir de encontro aos interesses de conservação de atributos ambientais, notadamente em economias capitalistas, dificultam sobremaneira sua gestão. Nessa situação, a implementação de instrumentos que favoreçam ou permitam práticas econômicas ambientalmente adequadas deve produzir resultados satisfatórios para o maior número possível de agentes socioeconômicos. No modelo brasileiro, a conservação da biodiversidade não é o único objetivo de manejo dos diferentes tipos de áreas protegidas. Há outros objetivos, dentre os quais se citam: conservação de sítios históricos, arqueológicos e culturais; proteção de bacias hidrográficas; fomento à recreação e ao turismo; proteção da paisagem; entre outros. Dentre as áreas protegidas estão as Unidades de Conservação, preconizadas pela Lei 9985, de 18/07/2000 (Brasil, 2000),

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Geoconservação em Áreas Protegidas

as Áreas de Preservação Permanente, as Áreas de Reserva Legal, e outras que possuem diploma legal conferindo a elas delimitação e proteção ambiental. As Unidades de Conservação podem ser instituídas em território nacional, estabelecidas por meio do Poder Público (em suas três esferas: federal, estadual e municipal). GeoPark não se encaixa em nenhuma das categorias de Unidades de Conservação preconizadas no modelo brasileiro de conservação e preservação dos recursos naturais. Legalmente não é uma área protegida. Assim, ao conceituar GeoPark como província de desenvolvimento sustentável, é perfeitamente possível sua instituição e operacionalização no território nacional. Ademais dentro de GeoParks devem ser instituídas categorias mais permissivas (como as Unidades de uso sustentável) e categorias mais restritivas (como as Unidades de Proteção Integral), no intuito de se promover restrições de uso do solo aos agentes socioeconômicos do território. Estabelecer ações em geoturismo, geoconservação e educação na busca de consolidar o território do GeoPark Araripe requer compromissos, não apenas da Unidade Executora, como de toda a sociedade afetada. O planejamento de políticas públicas de turismo, no propósito do manejo integrado dos municípios integrantes do GeoPark Araripe, conduz a compreensão dos processos históricos, econômicos, políticos e culturais dos geossítios. O desafio é possibilitar o desenvolvimento da atividade de geoturismo com as demais atividades realizadas no GeoPark Araripe, de maneira a contribuir para: •

a minimização dos prejuízos/danos ambientais oriundos da não observância da capacidade de suporte dos ecossistemas (ênfase nos geossítios);

•

a conservação do patrimônio geológico;

•

desenvolver produtos regionais, chamados geoprodutos, com incremento de bem-estar da população local;

•

promover a valorização da comunidade local, seus saberes, costumes e tradições; e

•

criar experiências únicas, que só podem ser vividas nas circunstâncias do território.

Quanto à conservação e preservação da biodiversidade presente na área do GeoPark Araripe, mencionado espaço territorial abriga 9 (nove) unidades de conservação (UCs), criadas em âmbito federal, estadual ou municipal, conforme Tabela 1. A questão central da conservação e preservação da biodiversidade é implementar meios de gestão e manejo adequados. Importante, então, vincular as atividades turísticas (que, pela legislação brasileira, são consideradas de uso indireto dos recursos naturais, factíveis em unidades de conservação de ambos os grupos: de Proteção Integral e de Uso Sustentável) às diferentes unidades de conservação existentes no território, bem como aos outros geossítios presentes no território. Levantamentos técnico-científicos elaborados na área do GeoPark Araripe permitiram a identificação de roteiros turísticos diversificados, desde o roteiro religioso (que encontra no Padre Cícero de Juazeiro do Norte interessante respaldo popular e de fé), passando por roteiros culturais (incorporando os saberes e legados culturais locais, a exemplo das festas populares e artesanatos) até roteiros gastronômicos que privilegiam gostos e sabores tradicionais locais (a exemplo da canjica, mugunzá, rapadura e bebidas locais: chamados de geoprodutos na área do GeoPark). Assim sendo, nas estratégias de geoturismo com forte interface em geoconservação, alguns fatores são primordiais e devem ser identificados de maneira a serem incluídos na proposição de estratégias de ação, com vistas ao geoturismo. A Tabela 2 traz alguns fatores e suas descrições. No GeoPark Araripe, o conhecimento dos efeitos das atividades de geoturismo desenvolvidas (e/ou a serem desenvolvidas), tanto individualmente como cumulativamente, nos seus limites deve possibilitar o rearranjo institucional e, dessa maneira, contribuir para novas práticas sustentáveis. Para tanto, a Unidade Executora do GeoPark Araripe em conjunto com Institutos de Pesquisa e Ensino tem feito, sistematicamente, avaliação de gestão e

Tabela 1. Unidades de Conservação integrantes do GeoPark Araripe - CE.

Unidade de conservação

Categoria

Âmbito administrativo

Diploma legal

Floresta Nacional do Araripe-Apodi Parque Ecológico das Timbaúbas

Uso sustentável Proteção integral

Federal Municipal

Parque Estadual do Sítio Fundão

Proteção integral

Estadual

APA da Chapada do Araripe RPPN Arajara Park Monumento Natural Sítio Canabrava Monumento Natural Pontal da Santa Cruz Monumento Natural Sítio Riacho do Meio Monumento Natural Cachoeira do Rio Batateira

Uso sustentável Uso sustentável Proteção integral Proteção integral Proteção integral Proteção integral

Federal Domínio privado Estadual Estadual Estadual Estadual

Decreto-Lei nº 9.226, de 02.05.46 Dec. nº 1.083, de 23.03.95 Dec. nº 29.307, de 05.06.08 (Criação) e Dec. 29.179, de 08.02.08 (Desapropriação) Decreto-Lei nº 9.226, de 02.05.46 Port. IBAMA nº 024-N, de 23.02.99 Dec. nº 28.506, de 01.12.06 Dec. nº 28.506, de 01.12.06 Dec. nº 28.506, de 01.12.06 Dec. nº 28.506, de 01.12.06

Cabral & Mota

Natureza & Conservação, 8(2):184-186, December 2010

Tabela 2. Fatores usados na priorização de estratégias de Geoturismo no GeoPark Araripe.

viabilidade econômica, por permitir o incremento de bem estar às comunidades locais diretamente afetadas, sobretudo com a valorização dos produtos locais (geoprodutos); e de justiça social, por permitir equidade de participação de todos os atores sociais envolvidos no processo.

186

Fatores

Descrição

Biodiversidade

Conservação e preservação da biodiversidade (flora e fauna)

Comunidades

Setores da sociedade civil: organizações não-governamentais, associações, grupos tradicionais, entre outros

Solo

Uso e ocupação do solo: agricultura, pecuária, paisagem, usos potenciais, dinâmica territorial

Água

Recursos hídricos superficiais e subterrâneos, seus usos múltiplos e preponderantes

Reconhecimento dos valores sociais Aspectos históricos, geológicos e arqueológicos

Eventos culturais locais, produção de bem-estar, incorporação de saberes regionais, tradições, gastronomia Sítios únicos de valor que demandam proteção, interpretação e transmissão para futuras gerações

Fonte: Adaptado de Hajkowicz (2008).

manejo do território do GeoPark, no intuito de readequar as ações de geoturismo, educação e geoconservação. Entende-se que o GeoPark Araripe representa importante instrumento para viabilizar o desenvolvimento sustentável na porção sul do estado do Ceará, incorporando-se as premissas de prudência ecológica, por internalizar a variável ambiental nos procedimentos e na tomada de decisão; de

Por fim, infere-se que GeoParks são, então, um instrumento que vem se somar aos demais existentes no arcabouço jurídico disponibilizado por meio do Sistema Nacional de Unidades de Conservação da Natureza (SNUC), contribuindo, significativamente, para conservação da natureza.

Referências Brasil. 2000. Lei 9985, de 18 de julho de 2000. Sistema Nacional de Unidades de Conservação da Natureza. Diário Oficial da União, Brasília. Cabral NRAJ & Souza MP, 2005. Área de proteção ambiental – planejamento e gestão de paisagens protegidas. 2. ed. rev. atual. São Carlos: RiMa. Global GeoParks Network – GGN, 2010. Global Network of National Geoparks (assisted by UNESCO). Available from: http://www.globalgeopark.org Access in: Mar. 2010. Hajkowicz SA, 2008. Supporting multi-stakeholder environmental decisions. Journal of Environmental Management, 88:607-614.

Recebido: Abril 2010 Primeira Decisão: Maio 2010 Aceito: Julho 2010

Forum

Brazilian Journal of Nature Conservation

Conhecimento Científico Scientific Knowledge Rogério Parentoni Martins1,* & Francisco Ângelo Coutinho2 1

Departamento de Biologia, Centro de Ciências, Universidade Federal do Ceará – UFC

Departamento de Métodos e Técnicas de Ensino, Faculdade de Educação, Universidade Federal de Minas Gerais – UFMG

Quando admitimos que algo seja conhecido ou, em nosso caso, cientificamente válido, deixamos implícito que sabemos precisamente o significado de conhecer. Ou achamos que sabemos. Dúvidas sobre a natureza do conhecimento as há desde as primeiras reflexões ocidentais sobre o tema, iniciadas na Grécia Antiga. A história da filosofia do conhecimento é, portanto, antiga e marcada por diversos momentos importantes, mas que não cabe aqui serem discutidos (ver Coutinho 2002, para tal discussão). Por serem contemporâneas e abrangentes, talvez nos interessem mais as reflexões de alguns filósofos da ciência do século XX: Karl Popper, Irme Lakatos, Thomas Kuhn e Paul Feyrabend. Mas, à exceção de Popper, as reflexões dos demais não tiveram impacto significativo entre os biólogos. Por isso e por também ter de certa forma influenciado o modo de desenvolvimento dos trabalhos científicos em Ecologia no Brasil, discutiremos brevemente o contexto do conhecimento ecológico em nosso país sob a perspectiva da epistemologia popperiana. Popper começou a ser lido tardiamente no Brasil, por isso o impacto de suas idéias sobre a biologia brasileira repercutiu muito no meio acadêmico brasileiro, especialmente nos anos 1980. Um dos objetivos de Popper foi o de estabelecer critérios para demarcar o que é conhecimento científico, distinguindo-o de outros tipos de conhecimento tal qual o adquirido por meio da fé ou da arte. E o fez por intermédio da proposição de que o conhecimento científico ou uma teoria científica só poderia ser válida se fossem elaboradas hipóteses (afirmativas sobre a natureza ou causalidade de fenômenos ecológicos) que pudessem ser refutadas ou falsificáveis. Se uma hipótese resiste à refutação, após suas previsões terem sido exaustivamente testadas, o conhecimento que a afirmação propõe está corroborado e é o mais provável de ser próximo da verdade (verossimilhança), pois a verdade absoluta é inatingível tanto davido às limitações de nossos *Send correspondence to: Rogério Parentoni Martins Departamento de Biologia, Centro de Ciências, Universidade Federal do Ceará – UFC, Fortaleza, CE, Brasil E-mail: wasp@icb.ufmg.br

sentidos perceptivos, quanto pelas limitações na estrutura abstrata de nossas teorias. Em contraste, o conhecimento adquirido por intermédio da fé não é passível de refutação empírica e, por isso, fica fora do âmbito da ciência. Por outro lado, muitos ecólogos adotaram um tanto entusiasticamente o método “strong inference” (Platt 1964) pelo qual se testa as previsões de uma hipótese considerada pelo pesquisador como a mais plausível, mas também se testa previsões de outras hipóteses teoricamente também plausíveis (ver também Martins & Coutinho 2004). Em primeiro lugar, testam-se as previsões menos plausíveis e se uma a uma forem descartadas, a elegida pelo pesquisador como a mais plausível, torna-se extremamente “forte” como uma explicação causal de determinado fenômeno. No entanto como o próprio Platt adverte, sua proposta metodológica seria adequada apenas em condições experimentais controladas, a exemplo das que são realizadas pelos físicos. No entanto, muitos ecólogos a ignoraram. Contudo, retornando a Popper, como é comum quando surgem novidades, os ânimos de vários pesquisadores brasileiros, e não apenas brasileiros, se exaltaram até um ponto tão extremo de se tornarem cegamente hiperpopperianos, ou seja, mais intensamente popperianos que o próprio. Este foi o caso de um pesquisador norteamericano, com que um dos autores teve um curto, mas significativo diálogo. Ele achava que estaria resolvido o problema do conhecimento em ecologia, pois seria possível tornar a ecologia parecida com a física, o xodó dos filósofos do conhecimento. E isso se daria pela aplicação rigorosa do método hipotético-dedutivo, que atribuía a Popper. Enfim, tudo em ecologia que não seguisse o método seria uma ciência de segunda categoria e, portanto, a nosso ver, preconceituosamente, relegada a segundo plano. Ora, com isso ele estava desdenhando excelentes estudos de naturalistas presentes e passados, inclusive aqueles que pavimentaram o caminho para a construção da ciência ecológica (Kingsland 1991). Outro episódio significativo foi o de um zoólogo brasileiro entusiasmado pela sistemática filogenética, que na época se difundia pelo Brasil. Dizia o pesquisador, com

Martins & Coutinho

Natureza & Conservação 8(2):187-189, December 2010

ironia, que Ecologia, segundo Popper, não era ciência e que ciência de fato era a sistemática filogenética, pois a ecologia não tinha hipóteses falsificáveis. Tais atitudes extremistas ensinaram-nos que seria contraproducente ignorar outras opções válidas para construção do conhecimento ecológico, pois uma atitude pluralista enriqueceria as possibilidades de sua ampliação. Aprendemos também que tais posturas extremistas não contribuem para o aprimoramento das reflexões e elaboração de teorias abrangentes e consistentes.

possam ser bons instrumentos para compreendermos respectivamente diversos fenômenos físicos e biológicos.

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Mais recentemente, o filósofo Lipton (2005) resolveu examinar quais seriam as vantagens de se iniciar um trabalho científico formulando em primeiro lugar hipóteses ou, em contrapósição, iniciá-lo a partir de um conjunto de dados e só então formular uma hipótese explicativa, induzindo uma generalização (em suas palavras ‘acomodação’ dos dados, sem significado pejorativo). A conclusão é a de que devido à possibilidade de ocorrer “maquiagem” dos dados, contida na acomodação, há uma tendência de os cientistas aceitarem mais o método de previsão de hipóteses previamente formuladas. Porém, ambos os tipos de conhecimentos gerados pela aplicação de cada um dos dois métodos seriam confiáveis e igualmente válidos. Por outro lado, um argumento não considerado por Lipton (2005) é o de que formular hipóteses previamente poderá economizar tempo e dinheiro empregados para realização da pesquisa. Mais uma vez, isso não significa que formular hipóteses a posteriori seja um “crime” praticado contra a “imaculada” Ciência. Finalmente, essa “popperização” excessiva provocou também reações contrárias no Brasil. Um físico, claramente irritado com a aplicação suprema do critério de falseamento, publicou um artigo em jornal contestando Popper quando este rotulou evolução, marxismo e freudianismo como metafísica. Ou seja, a garantia da autenticidade do conhecimento seria a possibilidade de ser refutado e, na análise de Popper, esses conhecimentos não a apresentavam, daí não serem científicos. O físico concluiu o trabalho dizendo que ciência é o que o cientista faz: uma opinião peculiar e simpática e, de certa forma, alinhada à epistemologia de Thomas Kuhn, mas também sujeita a discussões, pois é necessário explicitar como os cientistas fazem ciência. Apesar disso, ele não deixa de ter certa razão na prática científica cotidiana. O conhecimento científico em ecologia, previamente à sua publicação, sofre críticas de revisores que certificam a qualidade e veracidade do mesmo. Portanto, é esta crítica feita por pares que legitima o conhecimento proposto. Por outro lado, considerando-se que nossos sentidos são insuficientes para processar e entender a natureza de certos fenômenos “invisíveis”, formular teorias é o meio de que dispomos para adquirir conhecimento sobre o que supomos possa acontecer, mas temos dificuldade de visualizar. Não há relatos de quem tenha visto seleção natural, empurrando um macaco de uma árvore ou a gravitação universal como se fosse um monte de fios estabelecendo conexões entre dois corpos. Mas, teoricamente, supomos que entidades teóricas tais como gravitação universal e seleção natural

Tendo em vista a utilidade da teoria para o aprimoramento do conhecimento ecológico, cabe-nos discutir sobre a estrutura de teorias a fim de que possamos entender sua utilidade, aprimorá-la e até mesmo construir teorias mais abrangentes (Pickett et al. 2007). A história da ciência em geral e também em ecologia (Kingsland 1991) está repleta de exemplos. Basta citar o caso da deriva dos continentes e a concepção moderna baseada na tectônica de placas. Aliás, nesse caso, comparações entre faunas de mamíferos, filogeneticamente distintas, em florestas equatoriais africana e brasileira, mas morfológica e ecologicamente convergentes, foram importantes para reforçar a hipótese de que os continentes se derivaram, um fenômeno absolutamente “invisível”. Nesse caso, como também em vários outros, a partir de evidências que podem ser interpretadas que ocorreram por causa de determinado fenômeno, infere-se a sua veracidade. Richard Levins (1968), disse que há três ingredientes que deveriam fazer parte de uma teoria ideal: generalização, precisão e realismo. Generalização é o que almeja qualquer cientista que formule uma teoria. Quanto mais geral for, maior número de casos particulares em seu domínio seria por ela explicado. A teoria de evolução por meio da seleção natural, que é geral, diz respeito a qualquer organismo, explica modificações que organismos sofrem em cada geração, cujas modificações selecionadas podem se tornar adaptações em integrantes de gerações posteriores. Se há variação entre atributos individuais em uma população, por exemplo, tamanho, e se indivíduos de tamanhos relativamente menores sobreviverem e reproduzirem mais, relativamente aos de tamanho maiores, haverá uma tendência em predominar indivíduos menores nas gerações posteriores, desde que tamanho seja geneticamente determinado. Todavia, embora possa ser considerada uma teoria madura, ainda não são conhecidos detalhes de como a seleção natural atua em condições naturais, faltando-lhe, portanto, realismo. Em contraste, há teorias configuradas para situações mais aplicadas como, por exemplo, em agroecologia, que lidam com situações tão específicas, que lhes falta generalização, mas têm precisão e realismo, pois incluem em sua estrutura detalhes que lhe garantem essas características. No entanto, tais teorias só terão validade nas condições específicas para as quais foram configuradas. Nesse caso, por exemplo, um modelo computacional que inclua numerosas variáveis poderia ser uma teoria para explicar um fenômeno em um domínio bem restrito e, portanto inválida para outras situações em outros locais e em outros tempos. Enfim, há teorias e teorias e, por isso, devemos reconhecer uma pluralidade de estados em uma teoria. Uma teoria em estado nascente, por exemplo, pode ter em sua estrutura, apenas uma hipótese. Mas não deixa de ser uma teoria. Com o seu desenvolvimento por meio da incorporação de fatos, definições, conceitos, generalizações, modelos e leis ela poderá tornar-se uma teoria consistente, madura

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Conhecimento Ecológico

e assim reconhecida e legitimada pela grande maioria dos cientistas da área (Pickett et al. 2007). Scheine & Willing (2008) propõem uma teoria geral para ecologia que inclua todas as teorias existentes as quais chamam teorias constituintes. Embora não tenham formulado rigorosamente tal teoria, defendem que ela consista na descrição do domínio da ecologia e um conjunto de princípios fundamentais. O domínio da ecologia seria os padrões temporais e espaciais de distribuição e abundância de organismos, incluindo as causas e conseqüências. Os princípios fundamentais são afirmações amplas a respeito dos padrões e dos processos que operam no âmbito do domínio. Os sete princípios fundamentais (são fundamentais, pois as proposições de quaisquer teorias em ecologia podem ser verificadas como uma conseqüência desses princípios e de outros provenientes do domínio de outras ciências, o que seria um modo de translação; ver Pickett et al. 2007 ) são: (1) os organismos se distribuem de forma heterogênea no tempo e espaço; (2) Organismos interagem com seu ambiente biótico e abiótico; (3) A distribuição dos organismos e de suas interações é contingencial; (4) Condições ambientais são heterogêneas no tempo e espaço; (5) Recursos são finitos e heterogêneos no tempo e espaço; (6) Todos os organismos são mortais e (7) as propriedades ecológicas de espécies resultam do processo evolutivo. Se refletirmos e verificarmos, de fato todas as teorias da ecologia tem a ver com esses sete princípios, e a tarefa de integrar todas as teorias constituintes em um arcabouço teórico único e coeso é tarefa formidável. No entanto, esse trabalho indica que a ecologia está a caminho de uma fase de maturidade como disciplina científica.

Finalmente, conhecer a história e um pouco da filosofia da ciência de sua área de pesquisa é importante para os que estão dispostos a refinar suas pesquisas e alcançar resultados coerentes, significativos e de qualidade. Nesse processo de aquisição de conhecimento, discutir as dúvidas e questionar os conhecimentos estabelecidos são combustíveis para o crescimento intelectual e científico.

Refêrencias Coutinho FA, 2002. Conhecimento. In: Martins RP & Mari H (Ed.). Universos do Conhecimento. Belo Horizonte: Faculdade de Letras da UFMG. p. 91-115. Kingsland S, 1991. Defining ecology as a science. In: Real LA & Brown JA. Foundations of Ecology. London: The University of Chicago Press. p. 1-13. Levins R, 1968. Evolution in Changing Environments. Princeton: Princeton University Press. Lipton P, 2005. Testing hypothesis: prediction and prejudice. Science, 317:219-221. Martins RP & Coutinho FA, 2004. O Fantasma Teoria. In: Coelho AS, Loyola RD & Souza MBG (Ed.). Ecologia Teórica: desafios para o aperfeiçoamento da Ecologia no Brasil. Belo Horizonte: O Lutador. p. 15-26. Pickett STA, Kolasa J & Jones CG, 2007. Ecological Understanding. Amsterdam: Elsevier. Platt JR, 1964. Strong inference. Science, 146:346-353. Scheine SM & Willing MR, 2008. A general theory of ecology. Theoretical Ecology, 1:21-28.

Recebido: Agosto 2010 Primeira Decisão: Setembro 2010 Aceito: Setembro 2010

Forum

Brazilian Journal of Nature Conservation

O Desafio da Normatização de Informações de Biodiversidade para Gestão de Águas: Aproximando Cientistas e Gestores The Challenge of Biodiversity Information Normatization for Water Management: Putting Scientists and Managers Together Tadeu Siqueira1 & Fabio de Oliveira Roque2,* 1

Laboratório de Ecologia Teórica e Síntese, Departamento de Ecologia, Universidade Federal de Goiás – UFG, Goiânia, GO, Brasil

Faculdade de Ciências Biológicas e Ambientais, Universidade Federal da Grande Dourados – UFGD, Dourados, MS, Brasil

Os rios, lagos e áreas alagáveis contém uma porção minúscula – aproximadamente 0,01% – do total de água do planeta. Nesses ecossistemas estão presentes cerca de um terço de todas as espécies de vertebrados. Plantas aquáticas e invertebrados também apresentam alta diversidade e endemismo (veja o volume especial sobre biodiversidade aquática em Hydrobiologia, 595, 2008). Nos últimos anos a preocupação com esses ecossistemas tem aumentado, principalmente por que agora sabemos que: 1) para todos esses grupos citados existem registros de espécies extintas recentemente, criticamente ameaçadas, ameaçadas e vulneráveis; 2) a taxa de perda de biodiversidade em ecossistemas aquáticos continentais é muito maior que em ecossistemas terrestres; e 3) sistemas aquáticos têm sido fortemente alterados por mudanças climáticas atuais, introdução de espécies exóticas, modificação de habitats, superexploração e fragmentação (Millennium Ecosystem Assessment 2005). Isso coloca esses ecossistemas como os mais ameaçados no mundo. Além da biodiversidade per se, os ecossistemas aquáticos continentais provém serviços essenciais para a humanidade, com destaque para aqueles relacionados ao funcionamento (ciclagem e produção), à regulação (regulação climática, regulação de erosão), aos valores culturais (espirituais e religiosos, recreação, educação) e bens (alimento, água para consumo, recursos genéticos). A gestão desses ecossistemas, incluindo sua conservação e monitoramento, realizados de maneira eficaz e com amparo legal, é, portanto, de extremo interesse para a sociedade em geral.

*Send correspondence to: Fabio de Oliveira Roque Faculdade de Ciências Biológicas e Ambientais, Universidade Federal da Grande Dourados – UFGD, Rua João Rosa Goes, nº 1761, Vila Progresso, CEP 79825-070, Dourados, MS, Brasil E-mail: roque.eco@gmail.com

A gestão de águas embasada em informações científicas é relativamente bem desenvolvida em alguns países da Europa, na Austrália e nos Estados Unidos, mas tem crescido mundialmente nas últimas décadas. Tradicionalmente, os programas de gestão fazem uso de indicadores baseados em variáveis físicas e químicas da água, e em métricas da biota – peixes e macroinvertebrados são os mais usados. Indicadores são utilizados pelas agências ambientais competentes para análise do nível de qualidade ambiental, que então, baseadas em leis ambientais específicas, determinam ações. O uso de dados de biodiversidade em monitoramento de águas não é uma ideia nova. O final do século XIX e início do XX marcou o surgimento do primeiro índice biótico para a análise de qualidade de água – o índice Saprobiótico – e de muitas outras métricas baseadas na estrutura taxonômica da biota (veja Dolédec & Statzner 2010 para uma revisão sobre o surgimento dos índices mais usados). A partir de então muitos pesquisadores desenvolveram seus próprios índices bióticos ou adaptaram os já existentes para uso em sua região. Nos últimos anos, devido à demanda por avaliações em escala regional e com o aumento da disponibilidade e qualidade de dados, novas abordagens têm sido usadas (e.g., Biological Condition Gradient – BCG; Davies et al. 2006) e esforços têm sido direcionados para padronização e integração entre sistemas de avaliação e biomonitoramento. Por exemplo, a Convenção Internacional de Diversidade Biológica intensificou a construção de um sistema global de monitoramento de biodiversidade, incluindo indicadores de sistemas aquáticos. Esse sistema visa avaliar suas metas de conservação da biodiversidade, do uso sustentável de seus componentes, e da repartição justa e equitativa dos benefícios advindos da utilização de recursos genéticos. Os países que avançaram em políticas e ações específicas para gestão e monitoramento de suas águas interiores há algum tempo, hoje já são capazes não apenas de avaliar impacto, mas também de monitorar alterações na qualidade ambiental e ações de restauração.

Normatização de Dados de Biodiversidade para Gestão de Águas

A tendência de inclusão de dados de biodiversidade para gestão das águas soa muito animadora, mas será que as informações de comunidades biológicas estão realmente sendo usadas como instrumentos de gestão, disparando mecanismos legais de tomadas de decisão? Em tempos de preocupação devido à recente proposta de alteração do Código Florestal Brasileiro, essa pergunta se coloca ainda mais oportuna. Para ilustrá-la, imagine uma demanda de avaliação de impactos ambientais de uma determinada atividade sobre ecossistemas aquáticos em uma dada região do Brasil. Os agentes públicos do órgão de licenciamento ambiental pertinente, ao elaborar o termo de referência para o estudo se deparariam com a inexistência de padrões legais de referência para orientar a avaliação de componentes bióticos (algumas exceções existem para variáveis microbiológicas; Resolução CONAMA 357: Brasil 2005). A equipe de consultores, por sua vez, necessitará executar o estudo cumprindo o termo de referência com embasamento científico e legal. Considere que, no aspecto técnico-científico, a equipe realizou um ótimo trabalho: fez um desenho amostral adequado para o problema em questão, coletou informações sobre variáveis físicas e químicas da água e de paisagem; avaliou diversas métricas de comunidade em diferentes grupos (e.g., riqueza, abundância relativa, organização funcional, padrões de composição de comunidade); realizou análises estatísticas coerentes com o desenho amostral e concluiu que a atividade ou empreendimento proposto gerará respostas negativas nas medidas de comunidade consideradas. Se o relatório evidenciar alguma variável com valores fora dos padrões estabelecidos legalmente (variáveis físicas, químicas ou microbiológicas da água), provavelmente será disparado algum dispositivo legal de resposta. Entretanto, se apenas as métricas de comunidade forem potencialmente alteradas pelos efeitos deletérios do empreendimento/atividade proposto, é bastante provável que o relatório não disparará qualquer mecanismo de tomada de decisão. Mesmo se alguma espécie listada como ameaçada de extinção for encontrada, a equipe de consultores e os agentes do órgão ambiental não terão referência para tomadas de decisão. Mas por que isto ocorre? A resposta mais simples e objetiva para esta questão é que no Brasil não existe normatização de parâmetros de comunidades biológicas na legislação ambiental. Parâmetros normativos são aqueles que dependem de normas definidas segundo o ideal de valor humano. Embora a legislação brasileira referente à gestão de águas tenha avançado nos últimos anos (Buss et al. 2008), os parâmetros normativos ainda são predominantemente físicos, químicos e microbiológicos (Política Nacional de Recursos Hídricos, CONAMA: Brasil 2005). É preciso ressaltar que a resolução CONAMA (Brasil 2005) contempla o uso de informações de biodiversidade, mas não estabelece qualquer parametrização. Portanto, hoje, ao avaliar um sistema aquático, uma equipe de consultores poderá enquadrá-lo numa faixa de qualidade com base nos valores de algumas variáveis ambientais – informação que também será considerada

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pelos agentes públicos – mas não terá qualquer condição legal de fazer isto com base em métricas de comunidade. Consequentemente, a mitigação de impactos ambientais inevitáveis fica restrita a variáveis físicas e químicas da água. Por exemplo, normalmente o agente responsável pelo impacto assume, como condição para obtenção de licença ou autorização ambiental, medidas de manejo para garantir a recuperação da condição ambiental afetada (e.g., concentração de Fósforo na água) em conformidade com o padrão legal. Mas, e a comunidade biológica que teve toda sua estrutura alterada pelo impacto? Qual o padrão de qualidade ambiental deverá ser mantido ou recuperado para a biota? Muito tem sido feito e publicado no Brasil (veja o volume especial sobre monitoramento biológico em Oecologia Brasiliensis, 12(3), 2008), mas ainda há pouca discussão sobre a normatização de indicadores de biodiversidade. Mesmo nos países onde a prática do uso de bioindicadores é mais avançada, na maioria dos casos, eles não são utilizados como parâmetros normativos legais, mas apenas indicadores de condições ambientais. Este problema não está relacionado apenas à falta de vontade política e administrativa, mas também a problemas conceituais, técnicos e operacionais. Por exemplo, quando nos referimos à qualidade de água no Brasil, o conceito normativo subjacente está explicito na resolução CONAMA 357 (Brasil 2005). Porém, quando os trabalhos científicos se referem a conceitos como integridade ambiental e saúde ambiental, estes não possuem respaldo legal e tão pouco são operacionalizados cientificamente. A falta de clareza conceitual e operacional tem resultado em muita confusão e isolamento entre ciência e gestão (Rogers & Biggs 1999). Isso tem gerado fortes críticas sobre o baixo rigor científico em trabalhos de avaliação e monitoramento ambiental. A ciência segue, principalmente, uma abordagem hipotético-dedutiva contextualizada teoricamente, já a gestão ambiental segue uma abordagem normativa/legal com objetivos operacionais. A integração entre conceitos normativos e científicos operacionais é um desafio para aproximar cientistas e gestores. Isto ajudaria a evitar problemas de comunicação e tomadas de decisão. Por exemplo, o uso de conceitos como saúde/integridade/ qualidade de ambientes aquáticos não deve ser confundida com a forma de operacionalizar e medir estes conceitos cientificamente. Em outras palavras, no contexto normativo eles representam valores humanos importantes para tomadas de decisão em meio ambiente. Cientificamente eles são ambíguos e frágeis, não permitindo rigor em seu uso e construção teórica, devendo, portanto, ser traduzidos hierarquicamente em conceitos operacionais mensuráveis. Em termos técnicos, uma provocação pode exemplificar o nosso desafio. Podemos encontrar vários laboratórios de análises químicas ou de ensaios ecotoxicológicos certificados pelo INMETRO, mas um laboratório de ecologia de comunidades certificado? Com todos os seus procedimentos padronizados, incertezas de medidas, erros previstos, controle de informações, credenciamento

Siqueira & Roque

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de pessoal, parece até algo contra intuitivo. Em relação aos procedimentos (tipo de amostrador, processamento das amostras, sistema de banco de dados, capacitação de profissionais), basta esforço, financiamento e organização. No Brasil, algumas iniciativas já estão surgindo, como a da CETESB (órgão ambiental do Estado de São Paulo) que vivencia um processo de acreditação junto ao INMETRO de seus laboratórios para análises de dados de comunidade.

mudar o foco para funções ecológicas (Poff et al. 2006). Os traços funcionais representam características biológicas gerais que estão conectadas a funções ecossistêmicas, fornecendo, portanto, uma medida unificadora que pode ser aplicada para avaliar qualidade ambiental em regiões com diferentes composições taxonômicas (Dolédec & Statzner 2010). A especificidade aqui é em relação ao tipo de impacto. Isto é, a relação traço funcional–resposta deve ser definida para tipos específicos de impacto. Por exemplo, diminuição no fluxo d’água (efeito comum em rios com barragens) deve favorecer organismos maiores e com boa capacidade natatória; poluição orgânica e consequente diminuição de O2 dissolvido deve favorecer organismos com respiração aérea (Dolédec & Statzner 2010). Nesse aspecto ainda estamos bem atrasados aqui no Brasil. A quantidade de informação sobre traços funcionais de grupos aquáticos ainda é incipiente – para peixes a situação é um pouco melhor. Ou seja, atualmente não temos condição alguma de fazer uma proposta de parametrização de traços funcionais para gestão de águas no Brasil.

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Normatização de parâmetros de comunidades também enfrentam desafios no âmbito ecológico e biogeográfico. Por exemplo, como estabelecer parâmetros normativos para comunidades biológicas dada a variabilidade natural nesse tipo de informação (os riachos da Mata Atlântica possuem comunidades com características diferentes dos riachos do Cerrado)? Não é tão simples, mas os cientistas defendem duas alternativas principais. É possível regionalizar o estabelecimento de parâmetros normativos considerando as particularidades da biota regional num contexto biogeográfico. Existe um ótimo exemplo para essa alternativa. A Comissão Europeia criou um diretório de trabalho específico para tratar a gestão das águas (Water Framework Directive – WFD, para detalhes consulte o sítio http://ec.europa.eu/environment/water/water-framework/). Essa diretiva tem como objetivo principal “estabelecer um enquadramento jurídico que visa proteger e recuperar a qualidade da água na Europa, bem como assegurar a sua utilização sustentável em longo prazo”. A abordagem de gestão da água é feita com base nas bacias hidrográficas e estabelece prazos para os Estados-Membros protegerem os ecossistemas aquáticos. Os Estados-Membros identificam as bacias hidrográficas que se encontram no seu território, incluem cada uma delas numa região hidrográfica – que geralmente abrange mais de um Estado-Membro – e realiza uma caracterização geral dos corpos d’água, das pressões de impacto e uma análise econômica dos custos relacionados a sua proteção. Cabe a WFD estabelecer padrões de referência para regiões hidrográficas considerando elementos físicos, químicos, geológicos, hidromorfológicos e biológicos (veja Nõges et al. 2009 para uma análise das estratégias da WFD). Por exemplo, para o elemento Invertebrados Bentônicos, são considerados de excelente qualidade aqueles rios cuja “a composição taxonômica e a abundância correspondem totalmente ou quase às que se verificam em condições não perturbadas” (Anexo V–WFD: Directive 2000). Para algumas regiões do Brasil há informação básica e organizada sobre distribuição de biodiversidade aquática que permitem estabelecer parametrização regionalizada. Por exemplo, os dados sobre distribuição de macroinvertebrados em córregos do Sudeste do Brasil, produzidas no âmbito do Programa Biota–FAPESP, permitem inferências sobre padrões de distribuição em áreas de referência considerando escala regional (Roque et al. 2010). Outra alternativa é utilizar traços funcionais das espécies – como tamanho máximo e formato do corpo, estratégia alimentar e duração do ciclo de vida – e então estabelecer parâmetros aplicáveis em escala espacial mais ampla, e

Em ambas alternativas, precisamos entender melhor a distribuição das espécies e organização funcional considerando largas escalas e fatores que influenciam estes padrões em potenciais áreas de referência, para o estabelecimento de limiares que possam ser usados no âmbito normativo. Isto é imperativo no contexto de mudanças climáticas, uma vez que projeções têm indicado uma diminuição das potencias áreas de referência no mundo e os modelos de análise de qualidade de água baseados em bioindicadores ainda não são sensíveis aos efeitos de mudanças climáticas nos organismos e comunidades (Hamilton et al. 2010). Em síntese, a mensagem principal é: se quisermos realmente inserir informações de biodiversidade em processos decisórios que envolvam mecanismos legais de gestão de águas no Brasil, precisamos aproximar a ciência da tomada de decisão, identificando seus pontos de interação e particularidades, elaborar conexões entre os conceitos normativos e as medidas de biodiversidade a serem usadas, testar a racionalidade, implementação e desempenho de indicadores de comunidade para subsidiarem a construção de instrumentos legais. Além disso, temos que gerenciar laboratórios certificados para avaliar estas medidas e capacitar profissionais para os implementarem. Por fim, imaginem um mundo onde os dados de biodiversidade, particularmente de comunidade, sejam efetivamente usados em tomadas de decisão de forma organizada, testada, normatizada, envolvendo pessoas credenciadas e treinadas para este fim. É neste local que esperamos viver.

Agradecimentos Agradecemos as sugestões feitas por Marcus Vinicius Cianciaruso, Paulino Medina Júnior e Sidinei Magela Thomaz durante a elaboração desse manuscrito.

Normatização de Dados de Biodiversidade para Gestão de Águas

Referências Brasil. Ministério do Meio Ambiente. Conselho Nacional do Meio Ambiente - CONAMA, 2005. Resolução CONAMA nº 357, de 17 de março de 2005. Diário Oficial [da] República Federativa do Brasil, Brasília, DF. Buss DF, Oliveira RB & Baptista DF, 2008. Monitoramento biológico de ecossistemas aquáticos continentais. Oecologia Brasiliensis, 12:339-345. Davies SP et al., 2006. The Biological Condition Gradient: a descriptive model for interpreting change in aquatic ecosystems. Ecological Applications, 16:1251-1266. Directive, 2000. Directive 2000/60/EC of the European Parliament and of the council of 23 October 2000 establishing a framework for community action in the field of water policy. Official Journal of the European Communities, L327:1-72. Dolédec S & Statzner B, 2010. Responses of freshwater biota to human disturbances: contribution of J-NABS to developments in ecological integrity assessments. Journal of the North American Benthological Society, 29:286-311.

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Hamilton AT et al., 2010. Implications of global change for the maintenance of water quality and ecological integrity in the context of current water laws and environmental policie. Hydrobiologia, in press. Millennium Ecosystem Assessment, 2005. Ecosystems and Human Well-being: Synthesis. Washington, D.C.: Island Press. Nõges P et al., 2009. Assessment of the ecological status of European surface waters: a work in progress. Hydrobiologia, 633:197-211. Poff NLR et al., 2006. Functional trait niches of North American lotic insects: traits-based ecological applications in light of phylogenetic relationships. Journal of the North American Benthological Society, 25:730-755. Roger K & Biggs H, 1999. Integrating indicators, endpoints and value systems in strategic management of the rivers of the Kruger National Park. Freshwater Biology, 41:439-451. Roque FO et al., 2010. Untangling associations between chironomid taxa in Neotropical streams using local and landscape filters. Freshwater Biology, 55:847-865.

Recebido: Setembro 2010 Primeira Decisão: Setembro 2010 Aceito: Outubro 2010

Forum

Brazilian Journal of Nature Conservation

Mudanças Climáticas e a Biodiversidade dos Biomas Brasileiros: Passado, Presente e Futuro Climate Change and Biodiversity of Brazilian Biomes: Past, Present, and Future Alexandre Aleixo1*, Ana Luisa Albernaz2, Carlos Eduardo Viveiros Grelle3, Mariana Moncassim Vale3 & Thiago Fernando Rangel4 1

Coordenação de Zoologia, Museu Paraense Emílio Goeldi, CP 399, CEP 66040-170, Belém, PA, Brasil

Coordenação de Ciências da Terra e Ecologia, Museu Paraense Emílio Goeldi, CP 399, CEP 66040-170, Belém, PA, Brasil

Departamento de Ecologia, Universidade Federal do Rio de Janeiro – UFRJ, CP 68020, CEP 21941-902, Rio de Janeiro, RJ, Brasil

Departamento de Ecologia, Instituto de Ciências Biológicas, Universidade Federal de Goiás – UFG, CP 131, CEP 74001-970, Goiânia, GO, Brasil

Os biomas brasileiros abrigam uma porção significativa da biodiversidade mundial, constituindo importantes centros de biodiversidade pela combinação de altos níveis de riqueza e endemismo. No entanto, essa rica biodiversidade vem sendo crescentemente ameaçada por atividades antrópicas, principalmente aquelas ligadas à conversão das paisagens naturais em áreas de produção agro-pecuária e ocupação imobiliária. Altíssimos níveis de devastação ambiental já colocaram dois biomas brasileiros – a Mata Atlântica e o Cerrado – na lista dos “Hotspots” de biodiversidade, que são conjuntos de ecorregiões prioritárias para conservação em nível mundial (Myers et al. 2000). Para piorar a situação, o conhecimento sobre a real diversidade dos grupos biológicos que compõe a biodiversidade brasileira ainda pode ser considerado bastante incipiente, mesmo para aqueles grupos que tradicionalmente sempre foram considerados bem conhecidos, como é o caso das aves (Vale et al. 2008), o que pode inviabilizar seu uso como fonte confiável de informações para planejamentos sistemáticos e desenvolvimento de políticas de conservação (Aleixo 2010). Além das alterações recentes nas paisagens naturais, mudanças climáticas em curso e previstas constituem um segundo fator de ameaça à biodiversidade dos biomas brasileiros, com especial ênfase para aqueles predominantemente florestais e com maior riqueza de espécies e endemismo: a Amazônia e a Mata Atlântica. Apesar da sua importância, a produção científica nesta área do conhecimento é ainda pequena se comparada *Send correspondence to: Alexandre Aleixo Coordenação de Zoologia, Museu Paraense Emílio Goeldi, CP 399, CEP 66040-170, Belém-PA, Brasil E-mail: aleixo@museu-goeldi.br

à outras relacionadas com a Biologia da Conservação (Grelle et al. 2009). Na América do Sul, temperaturas mais altas e uma maior duração da estação seca poderão aumentar a frequência de estiagens sazonais rigorosas iniciadas pelo episódio El Niño / Oscilação Sul (ENSO) e de anomalias da temperatura da superfície do mar (SST) no Atlântico, contribuindo para incêndios cada vez mais frequentes e intensos, os quais ameaçarão a distribuição e integridade ambiental dos biomas brasileiros, em particular os predominantemente florestais (Marengo et al. 2009). Um conhecimento essencial para o desenvolvimento de políticas de minimização dos efeitos das mudanças climáticas sobre a biodiversidade brasileira é a reconstrução do grau de suscetibilidade e resposta das biotas dos diferentes biomas às alterações climáticas intensas de passado recente, como aquelas ocorridas ao longo dos últimos 2 milhões de anos (i.e. período Quaternário). Dados recentes sobre a filogeografia e modelagem de nicho ecológico obtidos para várias linhagens de organismos da Amazônia e Mata Atlântica, dão uma ideia da suscetibilidade dos diferentes biomas Enquanto para alguns grupos os últimos ciclos de diversificação correlacionaram-se aparentemente com essas alterações climáticas recentes, em outros, estas pouco ou nada contribuíram para seu padrão contemporâneo de riqueza de espécies e distribuição geográfica (Carnaval et al. 2009, Antonelli et al. 2010). O entendimento das variáveis determinantes deste padrão dicotômico de resposta frente às alterações climáticas intensas, em curtos períodos de tempo, adquire uma importância fundamental no contexto do aquecimento global já verificado e previsto para várias regiões do planeta, inclusive para os biomas brasileiros (Marengo et al. 2009).

Mudanças Climáticas no Brasil

Mudanças Climáticas e Biodiversidade: o Passado como Chave para Entender o Futuro Uma das hipóteses até hoje mais difundidas e intensamente debatidas sobre a origem da grande biodiversidade Neotropical e brasileira contemporâneas (i.e. hipótese dos refúgios pleistocênicos), coloca as mudanças climáticas no centro do processo de geração de barreiras que promovem a vicariância e geração de novas linhagens e espécies, num processo conhecido como cladogênese (Haffer 2008). Segundo a hipótese, as máximas glaciais, como aquela ocorrida há cerca de 20 mil anos atrás, teriam levado a temperaturas mais baixas e a uma drástica redução na pluviosidade média de alguns biomas brasileiros, com inferida redução e fragmentação dos biomas florestais e expansão de formações abertas; já os períodos interglaciais, com características semelhantes ao clima atual, teriam fornecido, de modo inverso, condições ideais para a expansão dos biomas florestais e a retração de formações abertas (Vonhof & Kaandorp 2010). O efeito destes ciclos nas populações dos diferentes organismos teria variado, dependendo do bioma. Durante períodos glaciais, populações de espécies florestais teriam sua distribuição fragmentada, ficando isoladas em “refúgios” florestais e se diferenciando uma das outras, ao passo que espécies de ambientes abertos expandiriam suas distribuições acompanhando a expansão destes ambientes. Em períodos inter-glaciais, o oposto teria ocorrido e espécies florestais recém diferenciadas expandiriam suas distribuições, entrando em contato, enquanto que espécies de área abertas teriam suas distribuições fragmentadas e passariam a se diferenciar pelo isolamento. Embora a hipótese dos refúgios tenha sido aquela mais discutida para a biota Neotropical até hoje e inclusive corroborada por diferentes estudos (ver revisões em Haffer (2008) e Antonelli et al. (2010)), diferentes estudos paleoclimáticos e filogenéticos falsificaram algumas das suas mais importantes previsões, evidenciando claramente as limitações desta hipótese. Por exemplo, alguns estudos palinológicos não encontraram evidências de vegetações abertas durante a Última Máxima Glacial (UMG; 20 mil anos atrás) em grandes extensões da Amazônia ocidental, conforme previsto pela hipótese dos refúgios, mas sim indícios de uma cobertura florestal similar à atual (Colinvaux 1993). De modo análogo, a maior parte das datações obtidas para a separação de espécies irmãs de vertebrados no bioma Amazônia, aponta para uma origem bem anterior à UMG, demonstrando que períodos glaciais, por si só, não seriam responsáveis pelo mais recente ciclo de cladogênese nestas linhagens (Antonelli et al. 2010). Independentemente de ter contribuído decisivamente ou não para formação de novas linhagens e espécies em diferentes contextos, é indiscutível que as mudanças climáticas ocorridas durante a UMG afetaram de modo bastante intenso o nicho ecológico potencial ocupado por espécies associadas aos

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biomas brasileiros. Isso pode ter ocasionado, no mínimo, grandes mudanças na distribuição geográfica destas espécies em relação ao contexto atual. Estima-se que, particularmente a Amazônia e Mata Atlântica, ficaram bastante reduzidas em relação à sua distribuição atual, conforme revelado pela distribuição do nicho ecológico potencial de várias linhagens associadas a esses biomas durante períodos glaciais, incluindo a UMG (e.g. Carnaval et al. 2009). Talvez um dos pontos mais obscuros sobre os efeitos dos ciclos glaciais de mudanças climáticas na biodiversidade dos biomas brasileiros seja o seu potencial de promover extirpações locais e até extinções completas de linhagens e espécies. Infelizmente, as assinaturas desses eventos não podem ser recuperadas de modo direto através de reconstruções filogenéticas (Rabosky 2010). De todo modo, mesmo com a drástica redução da cobertura dos biomas Amazônia e Mata Atlântica na UMG, isso parece não ter sido suficiente para a extinção de várias linhagens modernas, conforme evidenciado por diversos estudos filogenéticos e de modelagem do nicho ecológico pretérito e contemporâneo. Desse modo, três grandes questões inter-relacionadas se impõem hoje no estudo da biodiversidade brasileira num contexto de mudanças climáticas globais intensas em curso e previstas para o futuro próximo: (1) quais as características ecológicas das linhagens que apresentam uma resistência natural ou vulnerabilidade à mudanças climáticas, conforme revelado por seu passado evolutivo?, (2) qual o limiar de alteração climática necessário para gerar não apenas alterações em padrões de distribuição, mas também alterações nas fitofisionomias e extinções de linhagens inteiras nos biomas brasileiros? e (3) mudanças climáticas podem continuar atuando, como no passado, como fatores que contribuem, em diferentes contextos, para a geração de novas linhagens e espécies em biomas cada vez mais descaracterizados por paisagens antropizadas?

O Componente “Biodiversidade” da Rede Brasileira de Pesquisas sobre Mudanças Climáticas Globais: Modelando o Presente para Diagnosticar o Futuro A Rede Brasileira de Pesquisas sobre Mudanças Climáticas Globais (conhecida como Rede-Clima) foi criada em dezembro de 2007 e tem sua sede no Instituto Nacional de Pesquisas Espaciais (INPE) em São José dos Campos (SP), contando com diversas atribuições na formulação e acompanhamento de políticas públicas sobre Mudanças Climáticas Globais no Brasil, inclusive “estudar alternativas de adaptação dos sistemas sociais, econômicos e naturais do Brasil às mudanças climáticas”. Para esse fim, foi constituída uma sub-rede dentro da Rede-Clima, com foco na biodiversidade dos biomas brasileiros e cuja finalidade principal é entender melhor os efeitos das mudanças climáticas sobre a biota brasileira,

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Aleixo et al.

possibilitando a elaboração de estratégias para minimizar seus efeitos deletérios. Atualmente, a sub-rede Biodiversidade da Rede Clima executa projetos de pesquisa na sua temática de especialidade nos biomas Amazônia, Cerrado e Mata Atlântica, devendo em breve expandir sua rede de colaboradores para os demais biomas brasileiros. Neste primeiro momento, os projetos em execução na sub-rede têm três focos principais: (1) usando modelagem de nicho ecológico, mapear possíveis redistribuições geográficas para determinadas linhagens de organismos, sob cenários diversos de emissão de gases de efeito estufa, identificando espécies e áreas prioritárias para a conservação num futuro marcado por mudanças climáticas em andamento; (2) reconstruir os contextos temporal, espacial e ecológico do histórico de diversificação de determinadas linhagens de organismos, com vistas a identificar um conjunto de variáveis bio-geo-climáticas responsáveis pela diversificação e não extinção destes grupos no passado, que podem ser determinantes para a sua sobrevivência num futuro marcado pelas mudanças climáticas em andamento e (3) modelar o comportamento de diferentes tipos de fitofisionomias dos biomas brasileiros frente à mudanças climáticas, com o objetivo principal de prever alterações na sua distribuição futura. Para atingir estes objetivos, é necessária uma abordagem que integre diversas disciplinas, como, por exemplo, a sistemática filogenética molecular e a modelagem de nicho ecológico Grinneliano. Estudos das relações filogenéticas entre táxons próximos, muitos deles considerados atualmente como subespécies ou complexos de espécies, são importantes para se delimitar apropriadamente espécies e evidenciar padrões temporais e espaciais de diversificação. A partir desses resultados, é possível verificar se há padrões comuns que possam ser relacionados a eventos históricos. Para compreender os processos históricos mais recentes, em especial os relacionados a alterações climáticas Pleistocênicas, são necessários estudos populacionais, intra-específicos, que podem indicar padrões demográficos comuns a diferentes grupos que ocorrem nas mesmas regiões geográficas, além de técnicas de modelagem de nicho ecológico com base na distribuição geográfica das linhagens investigadas. A sub-rede biodiversidade da Rede Clima espera gerar um conjunto de dados acadêmicos/científicos com resolução e robustez inéditos sobre a origem, evolução, ecologia histórica e redistribuição em cenários de mudanças climáticas para linhagens de diversos organismos dos biomas brasileiros, que subsidiem diretamente a elaboração de estratégias de conservação futuras eficientes para estes biomas num cenário crescente de mudanças climáticas. Neste sentido, a sub-rede biodiversidade pretende aglutinar um número cada vez maior de pesquisadores e instituições focados em seus objetivos, além de promover encontros entre especialistas com o objetivo de promover a troca de experiências, o compartilhamento de dados e a formulação de sínteses de aproveitamento imediato para formulação de políticas públicas na área da conservação do meio-ambiente e biodiversidade.

Natureza & Conservação 8(2):194-196, December 2010

Agradecimentos Os autores são gratos ao Ministério da Ciência e Tecnologia (MCT) e Instituto Nacional de Pesquisas Espaciais (INPE) pelo apoio através da Rede-Clima. A. Aleixo e A. L. Albernaz agradecem ao CNPq pelo apoio através do “Instituto Nacional de Ciência e Tecnologia (INCT) em Biodiversidade e Uso da Terra da Amazônia” (auxílio nº. 574008/2008-0). A. Aleixo recebe uma bolsa de produtividade em pesquisa do CNPq (auxílio nº. 310593/2009-3). C. E. V Grelle agradece ao CNPq pelos auxílios e bolsa de produtividade e a FAPERJ pelo auxílio Jovem Cientista do Estado do Rio de Janeiro. M. M. Vale agradece ao CNPq pela bolsa de Pós-doutorado Júnior.

Referências Aleixo A, 2010. “Incerteza taxonômica” na biodiversidade amazônica: por que resolvê-la é imprescindível para a conservação do bioma? In: Themoteo R (Ed.). Cadernos Adenauer - Amazônia e desenvolvimento sustentável. 4 ed. Rio de Janeiro: Fundação Konrad Adenauer. v. 10, p. 35-57. Antonelli A et al., 2010. Molecular studies and phylogeography of Amazonian tetrapods and their relation to geological and climatic models. In Hoorn C & Wesselingh F (Ed.). Amazonia, Landscape and Species Evolution: a look into the past. Blackwell Publishing. p. 386-404. Carnaval AC et al., 2009. Stability predicts genetic diversity in Brazilian Atlantic forest hotspot. Science, 323:785-789. Colinvaux PA, 1993. Pleistocene biogeography and diversity in tropical forests of South America. In: GoldBlatt, P (Ed.). Biological relationships between Africa and South America. New Haven: Yale University Press. p. 473-499. Grelle CEV et al., 2009. Uma década de Biologia da Conservação no Brasil. Oecologia Brasiliensis, 13:420-433. Haffer J, 2008. Hypotheses to explain the origin of species in Amazonia. Brazilian Journal of Biology, 68:917-947. Marengo JA et al., 2009. Global warming and climate change in Amazonia. In: Keller M et al., (Ed.) Amazonia and Global Change. Washington DC: American Geophysical Union. v. 186, p. 262-273. Myers N et al., 2000. Biodiversity hotspots for conservation priorities. Nature, 403:853-858. Rabosky DL, 2010. Extinction rates should not be estimated from molecular phylogenies. Evolution, 64:1816-1824. Vale MM et al., 2008. Effects of Future Infrastructure development on threat status and occurrence of amazonian birds. Conservation Biology, 22:1006-1015. Vonhof HB & Kaandorp RJG, 2010. Climate variation in Amazonia during the Neogene and the Quaternary. In: Hoorn C and Wesselingh F (Ed.). Amazonia, Landscape and Species Evolution: a look into the past. Blackwell Publishing. p. 201-210.

Recebido: Novembro 2010 Primeira Decisão: Novembro 2010 Aceito: Novembro 2010

Forum

Brazilian Journal of Nature Conservation

O Protagonismo do Brasil no Histórico Acordo Global de Proteção à Biodiversidade Brazil’s Leading Role in the Historical Global Agreement for the Protection of Biodiversity Russell Mittermeier1, Patrícia Carvalho Baião2,3, Lina Barrera1, Theresa Buppert1, Jennifer McCullough1, Olivier Langrand1, Frank Wugt Larsen1 & Fabio Rubio Scarano2,4,* 1

Conservation International, 2011 Crystal Drive, Suite 500, 22202, Arlington, VA, Estados Unidos

Conservation International, Rua Buenos Aires 68, 26 andar, CEP 20070-022, Rio de Janeiro, RJ, Brasil

Programa de Pós-graduação em Biodiversidade, Universidade Federal do Amapá – UNIFAP, CEP 68908-130, Macapá, AP, Brasil

Laboratório de Ecologia Vegetal, Departamento de Ecologia, Centro de Ciências da Saúde, Instituto de Biologia, Universidade Federal do Rio de Janeiro – UFRJ, CP 68020, CEP 21941-970, Rio de Janeiro, RJ, Brasil

Sob a sombra do fracasso de Copenhagen e após duas semanas de difíceis negociações, poucos acreditavam que a décima Conferência das Partes (COP10) da Convenção da Diversidade Biológica (CDB) das Nações Unidas (veja histórico na Tabela 1) teria um final feliz. Entretanto, nas primeiras horas do dia 30 de outubro de 2010, na cidade japonesa de Nagoya, 193 países do mundo encontraram o tão esperado consenso. O Brasil foi um importante protagonista das negociações e, com freqüência, auxiliou na intermediação e busca de acordo entre as nações cujas posições eram mais extremas. O novo acordo significará um passo decisivo na redução da atual taxa de extinção de espécies e garantirá que países em desenvolvimento e seus povos tradicionais possam se beneficiar das riquezas geradas por seus ecossistemas terrestres e aquáticos. Nessa coluna, avaliamos os principais resultados da COP-10 e discutimos o papel do Brasil no sucesso das negociações.

O Acordo Áreas protegidas Uma das 20 metas acordadas no plano estratégico para o período 2011- 2020 foi a de aumento das áreas protegidas no âmbito global. Até 2020, 17% da superfície continental do planeta e 10% da área oceânica deverão estar sob proteção formal. Considerando que o mundo conta hoje com 12,9% da sua superfície continental e cerca de 0,7% da área oceânica *Send correspondence to: Fabio Rubio Scarano Conservation International, Rua Buenos Aires 68, 26 andar, CEP 20070-022, Rio de Janeiro, RJ, Brasil E-mail: f.scarano@conservacao.org

total nessa categoria, a meta é particularmente ambiciosa no que diz respeito aos ambientes marinhos, e moderadamente ambiciosa para ambientes terrestres. Ao longo das duas semanas de negociações em Nagóia, a discussão dessa meta esteve polarizada, tendo de um lado países como a Colômbia, a Costa Rica e o Equador que defendiam 25/15 (% continental e marinha, respectivamente), e de outro, países como a China que defendiam uma meta menos ambiciosa de apenas 10/6 (note que 10% seria menos que o planeta já tem e que a meta vigente era de 10% terrestre em 2010 e 10% para oceanos em 2012). A posição defendida pelo Brasil foi de 20/10 e o resultado acordado ao final, 17/10, indicou a flexibilidade de todos e a habilidade dos negociadores brasileiros em trazer a decisão para bem perto da proposta inicial do país. A Conservação Internacional organizou, no segundo dia da COP10 em Nagóia, um evento que discutiu resultados recém obtidos por seus cientistas que indicavam que seriam necessários pelo menos 25% de áreas protegidas continentais e 15% de áreas protegidas marinhas para proteger a biodiversidade e garantir a manutenção dos serviços ambientais. Os 25% da superfície continental, indicado pela análise da Conservação Internacional, é a soma de um mínimo de 17% de território global necessário para preencher as lacunas de áreas não protegidas classificadas como de alta prioridade, tais quais as KBA (Key Biodiversity Areas ou Áreas-Chave para a Biodiversidade), mais uma faixa adicional de 6 a 11% para garantir estocagem adequada de biomassa de carbono em ecossistemas naturais. Portanto, trata-se de uma estimativa bastante conservativa, já que não inclui outros aspectos ligados à biodiversidade ou a proteção de outros serviços ambientais como, por exemplo, a proteção de mananciais de água. Quanto à área

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Natureza & Conservação 8(2):197-200, December 2010

marinha, o quinto Congresso Mundial de Parques (Durban. África do Sul, 2003) da UICN (União Internacional para a Conservação da Natureza) recomendara a meta de 20-30% para adequadamente preservar a biodiversidade oceânica e assegurar ambientes marinhos produtivos e saudáveis (http:// cmsdata.iucn.org/downloads/recommendationen.pdf). A estimativa conservativa da Conservação Internacional de 15% para 2020 decorre do julgamento que esse já seria um primeiro largo e significativo passo no sentido de alcançar a meta projetada pelo congresso de parques da UICN. Esses estudos estão sendo ampliados, contemplando a inclusão de serviços ambientais adicionais tais como segurança de qualidade e acesso a água e segurança alimentar. Dados preliminares indicam que para garantir efetivamente a manutenção dos serviços ambientais globais, talvez seja preciso proteger aproximadamente 35-50% da superfície do planeta. Consideramos, portanto, que o acordo 17/10 para 2020 é ainda abaixo do necessário, mas que já representa um passo significativo em direção ao ideal.

espécies. A discussão acerca dessas metas também esteve polarizada na COP10, com alguns países propondo eliminar completamente a extinção de espécies até 2020, e outros defendendo texto sem meta numérica, com indicação vaga de diminuição das taxas atuais dentro das capacidades nacionais. Com relação à extinção de espécies, o consenso foi de evitar a extinção de todas as espécies sabidamente ameaçadas até 2020, posição essa defendida pelo governo brasileiro. Em relação à perda de habitats naturais, as partes concordaram com a meta de reduzir a pelo menos à metade até 2020 e aonde possível chegar à zero.

198

Redução das taxas de extinção e de perda de habitats Duas metas do Plano Estratégico para 2011-2020 se relacionam a esses tópicos. A meta 5 lida com perda de habitats naturais, enquanto a meta 12 com extinção de

Baseado em estimativas recentes lançadas no relatório “A Economia dos Ecossistemas e da biodiversidade (TEEB, do Inglês, The Economics of Ecosystem and Biodiversity), a Conservação Internacional defendeu a meta de extinção zero até 2020 e uma meta ambiciosa para a redução da perda de habitats naturais. De acordo com o TEEB, as taxas atuais de extinção e de destruição de habitats implicam em grandes perdas financeiras anuais e que áreas protegidas proporcionam 100 vezes mais benefícios que custos à economia global. Assim, ainda que o acordo represente um grande passo, a redução da taxa de perda de habitats pela metade até 2020 ainda implicará prejuízo significativo para o mundo, especialmente para países em desenvolvimento.

Tabela 1. Breve cronologia da Convenção da Diversidade Biológica (CBD) das Nações Unidas.

Ano

Evento

Principais fatos

1992

Eco-92, Rio de Janeiro, Brasil

São criadas três convenções, no âmbito da ONU, para conter e reverter a crise ambiental: a CBD, a Convenção de Mudanças Climáticas (UNFCCC) e a Convenção de Combate à Desertificação (UNCCD).

1993

CBD entra em ação com três objetivos globais: 1) conservar a diversidade biológica; 2) garantir o uso sustentável dos componentes da biodiversidade; e 3) garantir a repartição justa e equitativa dos benefícios que derivam dos recursos genéticos.

2002

COP-6, CBD, Johanesburgo, África do Sul

Lançamento do primeiro plano estratégico para reduzir a perda de habitats e extinção de espécies.

2006

COP-8, CBD, Curitiba, Brasil

Demandou ao grupo de trabalho ad hoc sobre acesso e repartição de benefícios (ABS) que completasse seu trabalho antes da COP-10.

2010 (Maio)

Subsidiary Body on Scientific Technical and Technological Advice (SBSTTA), CBD, Nairobi, Quênia

Constatação pelas partes que o planeta fracassou em alcançar as metas lançadas em 2002.

2010 (Outubro)

COP-10, CBD, Nagóia, Japão

Lançamento de novo plano estratégico com 20 metas para o período de 2011-2020; acordo em torno do acesso e repartição de benefícios (ABS) associados à biodiversidade.

2012

COP-11, local por definir, Índia

Modalidades dos compromissos financeiros para ABS serão discutidas.

2012

Rio+20, Rio de Janeiro, Brasil

Re-encontro das três convenções criadas em 1992, para avaliação de seu sucesso.

199

Acordo Global de Proteção à Biodiversidade

Acesso e repartição de benefícios Acesso e repartição de benefícios (ABS, do inglês Access and Benefit Sharing) é relacionado ao terceiro objetivo da CBD (ver Tabela 1) e é um dos mais controvertidos aspectos da convenção desde seu mandato em 1998, tendo sido objeto de intensas negociações desde 2002. Sem dúvida, foi também o ponto que mais suscitou divergências entre os países durante a COP10. Essencialmente, o protocolo de ABS negociado em Nagóia trata de como ter acesso aos recursos genéticos da natureza e de como compartilhar os benefícios derivados do uso dessa biodiversidade, baseado em termos mutuamente acordados entre as partes. O protocolo toca em assuntos delicados como direitos sobre patentes, propriedade intelectual, direitos de comunidades indígenas e locais ao conhecimento tradicional associado a recursos genéticos, bem como acerca de que forma benefícios devem ser compartilhados quando produtos medicinais, cosméticos e outros bens são criados a partir da biodiversidade. No mercado atual, muitos desses produtos têm sua origem biológica no mundo em desenvolvimento, mas o benefício por vezes fica no mundo desenvolvido que, historicamente, é onde se encontram os laboratórios que desenvolvem os produtos comercializáveis. O Brasil desde o início da COP10 condicionou a aprovação dos demais acordos ao acordo em ABS, no que foi seguido por vários outros países do mundo em desenvolvimento. Esse posicionamento foi essencial para impulsionar os demais países na direção de um acordo quanto ao ABS. Agora acordado, o “Protocolo de Nagóia sobre Acesso aos Recursos Genéticos e o Compartilhamento Justo e Equitativo dos Benefícios Derivados de Seu Uso” (http:// www.cbd.int/nagoya/outcomes/) marca uma conquista histórica da CBD em direção ao cumprimento do seu terceiro objetivo. Graças ao Japão ter submetido um texto, um dia antes do término da COP10, que constituía um bom compromisso entre países de norte e sul, consenso foi alcançado em tópicos que dividiam as partes signatárias há anos - incluindo o escopo do protocolo e a definição de utilização e derivados – em geral através do uso de linguagem aberta e da eliminação de trechos particularmente polêmicos. O protocolo também inclui considerações acerca de um mecanismo multilateral global de repartição de recursos que pode tratar de situações específicas onde ainda não haja consenso estabelecido. Embora essa abordagem possa criar algumas ambigüidades e questões acerca de sua efetiva implementação, o Protocolo de Nagóia emerge como um importante avanço.

Financiamento Outro ponto de contenção para o governo brasileiro, no que também foi seguido por vários outros países em desenvolvimento, dizia respeito ao financiamento das várias ações previstas no plano estratégico para o período 2011-2020. Diante do fracasso no alcance das metas previstas

para 2010, esses países condicionaram o acordo no novo plano estratégico a um compromisso de financiamento por parte de países do mundo desenvolvido, reconhecendo também a necessidade de investimentos nacionais por parte dos países em desenvolvimento. Novos compromissos financeiros, além do já anunciado pela Alemanha ao término da COP9 (sediado em Bonn, 2008), demoraram a aparecer, mas a três dias do fim das negociações, o Japão anunciou um compromisso de 2 bilhões de dólares, e foi seguido pela Grã-Bretanha, ainda que com uma menor soma bem inferior, e pela França, que se comprometeu com 200 milhões de dólares por ano ao longo dos próximos quatro anos e, a partir de 2014, 500 milhões de dólares por ano. Adicionalmente, os doadores reabasteceram o Global Environmental Facility (GEF), um fundo fiduciário multilateral estabelecido para financiar as convenções do clima e da biodiversidade, na ordem de 4,25 bilhões de dólares para o período de 2010-2014, que representa um aumento de 37% sobre os quatro anos anteriores. A posição apresentada conjuntamente pela Conservação Internacional e pela Birdlife International era de investimentos na ordem de 0,3% do PIB dos países do OECD (Organisation for Economic Co-operation and Development), que representa um montante de 125 bilhões de dólares. Ainda que os compromissos já anunciados estejam bastante aquém dessa proposta, esses foram bem recebidos e permitiram o acordo.

Outros pontos importantes Merecem destaque ainda os seguintes aspectos: •

Ficou definida a criação, independente da CBD, do IPBES (Painel Intergovernamental de Biodiversidade e Serviços Ambientais), que seria o análogo ao IPCC (Painel Intergovernamental de Mudanças Climáticas), até Janeiro de 2011. O Brasil é um dos candidatos para sediar o Painel;

•

Alinhamento com a convenção do clima: houve um reconhecimento da crescente interrelação entre conservação da biodiversidade, mudanças climáticas e desenvolvimento. Isso ficou evidente no enquadramento das metas para 2020 e também na série de decisões acerca das contribuições da natureza para adaptação e mitigação às mudanças climáticas, assim como na erradicação da pobreza.

Brasil: Protagonismo versus Lacunas Conforme já mencionado, o Brasil teve um papel decisivo nas negociações em Nagóia, assim como já o fizera em Copenhagen na 15a Conferência das Partes (COP15) da Convenção do Clima. O reconhecimento do sucesso da delegação brasileira na COP10 se deveu a diversos fatores tais como: boa estratégia de negociação, negociadores habilidosos e bem treinados; envolvimento e coordenação

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Natureza & Conservação 8(2):197-200, December 2010

entre vários ministérios; a eloqüência e o compromisso da Ministra do Meio Ambiente, Izabella Teixeira; boa representação do setor empresarial brasileiro; boa e atuante representação de ONGs atuantes no país; assim como vários acadêmicos de destaque. Agora é o momento de passarmos do discurso para a ação e do papel para a prática. Se o Brasil não quiser ficar devendo para o planeta em termos de áreas protegidas, precisará pelo menos aumentar em 70% a atual cobertura de áreas protegidas no continente e aumentar 10 vezes a cobertura protegida marinha, conciliando isso com geração de energia, produção de alimentos e combustíveis e construção de infra-estrutura. Se por um lado a Amazônia brasileira está protegida em quase 50% (incluindo unidades de conservação e terras indígenas), biomas como o cerrado, pantanal, caatinga, mata atlântica e pampa estão bem abaixo dos 17% agora lançados como meta para o planeta. O quadro marinho então, com seus 1,5% atuais, é particularmente sério, ainda mais em tempos de pré-sal.

aqui na “Natureza & Conservação” (e.g., Lewinsohn 2010, Metzger 2010, Scarano & Martinelli 2010).

200

O país precisará o quanto antes desenhar uma estratégia robusta para alcançar essas metas em 10 anos e prontamente pô-la em prática, incluindo criação e consolidação de áreas protegidas; criação de mecanismos de geração de renda a partir da biodiversidade e de arcabouço legal em nível nacional para pagamentos por serviços ambientais; revisão de listas de espécies ameaçadas de extinção e monitoramento e recuperação das mesmas; dentre vários outros pontos. Tudo isso irá requerer desembolso e financiamento em escala muito superior ao que se tem atualmente, mas também irá demandar ciência e pessoal qualificado. Se por um lado, a expansão da rede de pós-graduação em áreas ligadas à biodiversidade está em franca expansão, assim como o impacto internacional da ciência produzida nesses temas pelo país, por outro lado, a ciência e a tomada de decisão ainda se comunicam muito mal, como vem sendo discutido

O balanço de vinte anos da ECO-92 se aproxima. O evento que está sendo chamado de Rio+20, em 2012, será hora do Brasil definitivamente despontar como líder ambiental global. Para isso será preciso mais que habilidade de negociação, atributo que o país já possui. Será preciso liderar pelo exemplo, pondo em prática em território nacional as idéias que tão habilmente os negociadores brasileiros têm usado para convencer outros países da importância e valor da biodiversidade.

Agradecimentos À delegação brasileira em Nagóia pelo trabalho comprometido e dedicado sob a liderança da Ministra Izabella Teixeira e do Ministro Paulino Carvalho e, particularmente, ao grande conservacionista e habilidoso negociador Secretário Braulio Dias. Somos gratos também à delegação da Conservação Internacional pelo apoio e trabalho em equipe.

Referências Lewinsohn TM, 2010. A ABECO e o Código Florestal Brasileiro. Natureza & Conservação, 8:100-101. Metzger JP, 2010. O Código Florestal tem base científica? Natureza & Conservação, 8:92-99. Scarano FR & Martinelli G, 2010. Brazilian list of threatened plant species: reconciling scientific uncertainty and political decision-making. Natureza & Conservação, 8:13-18.

Recebido: Novembro 2010 Primeira Decisão: Novembro 2010 Aceito: Novembro 2010

Acknowledgement to the Reviewers

This journal could not exist without the support of scientists to act as peer reviews. Natureza & Conservação gratefully acknowledges the time and effort contributed by the following experts who served as referees for papers received in 2010. Their contribution was indispensable for improving the quality of published papers and was much appreciated. Adriano S. Melo

Marc Dourojeanni

Ana Albernaz

Marcos Figueiredo

Bianca W. Bertoni

Maria Alice S Alves

Carlos Eduardo V. Grelle

Mariana Pires C Telles

Claudio Padua

Mário Luís Orsi

Fernando Fernandez

Miguel Araújo

Flávia Souza Rocha Gustavo Henrique Gonzaga da Silva Irineu Bianchini Jr Jean Vitule João Nabout John Lamoreux

Renata Alves da Mata Ricardo Dobrovolski Rogério Parentoni Martins Ronaldo Angelini Rosane Collevatti

Katia Torres Ribeiro

Samraat Pawar

Lázaro Jose Chaves

Sidinei M. Thomaz

Leandro Duarte

Thiago Fernando Rangel

Leonardo de Carvalho Oliveira

Thiago Santos

Levi Carina Terribile

Viviane G. Ferro

Luis Mauricio Bini

Walfrido Tomas

Natureza & Conservação, 8(2) December 2010

Errata

In the paper by Collevatti et al. (Natureza & Conservação, 8(1): 54-59, 2010), there is an error in one of the authors’ name. Jacqueline Silva Lima should read Jacqueline Santos Lima.

Author Guidelines

Aims and Scope Natureza & Conservação (Brazilian Journal of Nature Conservation, N&C) is a scientific journal devoted to improving theoretical and conceptual aspects of conservation science. It has the main purpose of communicating new research and advances to different actors of society, including the academy, conservationists and practitioners, government and decision-makers. N&C publishes original papers on biodiversity conservation and its implications, including population viability analysis, biological invasion, species distribution modeling, diversity patterns, phylogeography and conservation genetics, reserve design and selection, ecosystem management, conservation policy, among others. There is no bias towards particular biogeographic regions, organisms or ecosystems, although we expect more submissions from Tropical and Neotropical regions. Scientific papers must focus on new conceptual or methodological developments with practical implications. Case studies will be considered only if inserted in these more general contexts. Authors are encouraged to submit reviews and essays on arising issues, in agreement with the editorial board. Purely descriptive papers will not be considered.

Submission Original contributions should be sent as a Word file (*.doc or *.rtf) or PDF attached to the editor, addressed to the Editorin-Chief (other formats may be considered after discussions with the editor). In the initial e-mail, corresponding author must explicitly state that the manuscript was not submitted to other journal and that all co-authors are aware of the submission. Also in the submission message, authors must indicate three preferred reviewers (along with their e-mails and affiliations), as well as stating if someone is to be avoided as a reviewer (non-preferred reviewer). Files with large sizes must be zipped, but it is preferable that authors paste their figures within the word file in JPG format to reduce overall file size (up to 2 Mb is preferable). Supplementary material or other appendices must be submitted as separate files, but with the same submission caption. Submissions will be sequentially numbered and the reference-number will be sent to the authors in the initial acknowledgement e-mail. All further communications between N&C and the corresponding author will refer to

this number. Manuscripts will be screened by the editorial board before being sent out for external review and may be rejected editorially. Editorial reject decisions are based on how well a manuscript fits the scope of the journal as well as on the quality of the manuscript. Authors are expected to receive a first decision within approximately 45 days. When submitting a revised version, authors must also submit a marked-up copy of the manuscript using the “track changes” panel in Word or in any other open-source application.

Sections N&C publishes original papers in English, following basically two formats: Essays & Perspectives will deal with longer essays and reviews updating recent topics of interest in conservation science. Essays & Perspectives will be usually invited, but submissions can be discussed with the editors in advance and suggestions are welcome. Original scientific research will be published in the format of Research Letters, which are short and concise manuscripts with up to 3000 words in length, up to 4 figures and/or tables and 25 references. The other sections of N&C will be invited foruns dealing with specific topics in conservation (written in Portuguese), including book reviews, highlights from the literature, and comments on previous papers published in N&C.

Manuscripts Format Manuscripts should be double-spaced throughout (including tables, figure legends, literature cited) with all lines and pages numbered.

Text The title page must contain the section for which the manuscript is intended, the title of the manuscript and the authors’ name, associated by superscript numbers indicating their affiliations, the total word count (including references, tables and figure legends) and a short title. The second page must contain 5 keywords for indexation purposes and an abstract with up to 300 words for Essays

Natureza & Conservação, 8(2) December 2010

Author Guidelines

& Perspectives, and up to 150 words for Research Letters. The next pages will contain the main text, with up to 3000 words for Research Letters and 7000 words for Essays & Perspectives. For Research Letters, main captions will follow the standard format (introduction, material and methods, results, and discussion), but there is no specific format for Essays & Perspectives. In both cases, main captions must be typed in uppercase bold font, and subtitles within each main caption must be italicized. A limit of 25 references is established for Research Letters, and about 50 references are suggested for Essays & Perspectives. For Research Letters, a limit of 4 figures and/or tables is recommended and authors must write as concise as possible, especially the methods section. It is strongly advised that the details of methods or original data are assigned to a supplementary material (with all figures and tables referred in the main text as Table S1, Figure S1, S2 and so on), which will be published online only. Avoid right margin justification and hyphenation. Double-check the contents of your manuscript before submitting. Only printer’ mistakes in proofs will be changed free of charge.

Figures For initial submission, figures must be inserted at the end of the word file. After acceptance of the manuscript, high-resolution figures in TIF, WMF or EMF formats will be required, with a minimum resolution of 300 dpi. Figure legends must be, as much as possible, stand-alone and must be typed separately, appearing in the end of the main text. Tables must be inserted at the end of the main text with the title, and built using the “Table” option of word processor or any open-source application (and not typed “manually” or pasted from spreadsheet applications). Within figures, authors must be aware that symbols must be large-enough to be readable after reduction in size in the final publication.

Units Use SI units as far as possible.

Nomenclature Binomial Latin names should be used in accordance with International Rules of Nomenclature.

References Citations in the main text will follow the author-year standard format [i.e., Rabelo 2007; Bini & Diniz-Filho 2005; Loyola et al. 2008; or Loyola et al. (2008)]. In the reference list, papers with more than 3 authors must be referred as “et al.” as well, and references to articles, books and book chapters are as follows: Silva JMC, 1995. Birds of the cerrado region, South America. Steenstrupia, 21:69-92. Balmford A et al., 2001. Conservation conflicts across Africa. Science, 291: 2616-2619. Marinho-Filho J, Rodrigues FHG & Juarez KM, 2002. The Cerrado mammals: diversity, ecology, and natural history. In Olivera PS & Marques RJ (eds.), The Cerrados of Brazil. New York: Columbia University Press, p. 266-284. Legendre P & Legendre L, 1998. Numerical ecology, Amsterdam: Elsevier.

Titles of journals should be abbreviated following Biological Abstracts. If in doubt, give the title in full. Do not refer to unpublished material. The reference list should be arranged alphabetically on authors’ names and chronologically per author. If the author’s name is also mentioned with co-authors, the following order should be used: publications of the single author, arranged chronologically - publications of the same author with one co-author, arranged chronologically - publications of the author with more than one co-author, arranged chronologically. Publications by the same author(s) in the same year should be listed as 2009a, 2009b, etc. Reference lists not conforming to this format will be returned for revision.

Language and style Manuscripts will be checked for style and language and authors are invited to ask native speakers or use available online services to improve correctness of language and style. For standardization purposes, authors must check for spelling using the US-English option in their word processor or any open-source application. Editors and reviewers are invited to help in the process of improving as much as possible language and style of the manuscript. The impact of the paper and, consequently, of the journal, will largely depend upon the quality of the English. After acceptance of the manuscripts, editor(s) will deserve the right to do minor changes to improve language and style.

Pesquisa científica na

Reserva Natural Salto Morato A Fundação Grupo Boticário de Proteção à Natureza lhe convida a estudar temas da conservação da natureza e do manejo de áreas naturais na Reserva Natural Salto Morato, Guaraqueçaba - Paraná. Além de um centro de pesquisa equipado, alojamento, e da possibilidade de apoio financeiro, o pesquisador terá à disposição mais de 2 mil hectares de Mata Atlântica em uma região reconhecida pela Unesco como Sítio do Patrimônio Natural da Humanidade. Inscreva seu trabalho e contribua com seu talento para a conservação da natureza brasileira.

Crédito das fotos: José Paiva

Saiba mais. Acesse: www.fundacaoboticario.org.br ou escreva para: morato@fundacaoboticario.org.br

Research Letters Predicting Patterns of Beta Diversity in Terrestrial Vertebrates Using Physiographic Classiﬁcations in the Brazilian Cerrado André Andrian Padial, Luis Mauricio Bini, José Alexandre Felizola Diniz-Filho, Nayara Pereira Resende de Souza & Ludgero Cardoso Galli Vieira ............................................................................... 127

Conhecimento Cientíﬁco Rogério Parentoni Martins & Francisco Ângelo Coutinho ............................................................................................... 187