IALE Landscape Research and
Management papers
Understanding Biodiversity Loss
An overview on Forest Fragmentation in South America
Edited by Maria Jose Pacha, Sandra Luque, Leonardo Galetto, Louis Iverson
ISSN: 1570-6532
SANDRA LUQUE, MSc PhD Research Director at Cemagref, France. Landscape ecologist working in the development of rapid assessment methods for biodiversity evaluation and comprehensive landscape monitoring and modelling. At the present she serves as elected vice-president for IALE (International Association for Landscape Ecology). She is serving in the board of the IALE-IUFRO Working Group on Forest Landscape Ecology. For 18 years she has been working on remote sensing/GIS and landscape ecology issues in relation to change detection, forest ecology, biodiversity indicators, and biodiversity habitat quality models. At the present, she is the co-ordinator of the scientific group of experts on forest fragmentation and biodiversity loss in South-America. She is merited coordinator of international and national level research projects.
MARIA JOSE PACHA MSc PhD working for the NGO Fundacion Vida Silvestre, Argentina. Her background is in Vegetation Ecology with links at the landscape level. She has worked in applied conservation projects in Argentina and the United Kingdom in Nature Reserves and National Parks. Also her research interests are about linking social and natural aspects of conservation. She has worked as project coordinator of the Atlantic Forest Programme of FVSA and at the present she coordinates a WWF UNESCO project on financing mechanisms for protected areas in South America.
LEONARDO GALETTO PhD Researcher from CONICET (National Research Council of Argentina) and Professor at the Universidad Nacional de Córdoba (Argentina), working with plant reproductive ecology and forest fragmentation for the last 18 years. He is particularly interested about the changes on animal-plant interactions due to habitat loss, studying different processes as pollination, dispersion and herbivory. He also has worked in ethnobotany and the relationship of rural people perception of local resources and their availability. At present supervises doctoral and postdoctoral students working with different approaches in the Chaco forest. He is Editor of Kurtziana (a regional journal of Botany) and involved as Vice-President with the Asociación Argentina de Ecología
LOUIS R. IVERSON PhD Research Landscape Ecologist for United States Forest Service in Delaware, Ohio. Vice-president for International Association for Landscape Ecology and book review editor for the journal, Landscape Ecology. His current research concerns potential changes in tree species following climate change in the United States, the use of fire and thinning to restore oak-hickory forest communities, and modelling the advance of the emerald ash borer, an insect killing ash trees in the central United States.
IALE electronic publication series
Understanding Biodiversity Loss: An overview on Forest Fragmentation in South America
Edited by Maria Jose Pacha, Sandra Luque, Leonardo Galetto, Louis Iverson
Target audience: The publication is aimed at advanced undergraduate and graduate students, researchers and teachers, professional landscape ecologists, policy makers and practitioners with a special interest in South American forest issues.
Aim: to demonstrate the contribution that Landscape Ecology can make to forest management and the understanding of forest fragmentation and biodiversity loss. This publication is specially targeted to describe and determine causes of forest fragmentation in South America. The papers that contribute to the series present study cases from Argentina, Chile and Southern Brazil. The different forests that are covered in this issue are: Atlantic Forest of Argentina and Brazil, Gran Chaco (Argentina, Paraguay and Bolivia), Valdivian Forest (south of Argentina and Brazil) and Yungas rainforest (Argentina and Bolivia).
Cite as:
Pacha, M.J., Luque, S., Galetto, L. and Iverson, L. (2007) Understanding biodiversity loss: an overview of forest fragmentation in South America. IALE Landscape Research and Management papers. International Association of Landscape Ecology Preface
This publication is the result of a series of papers presented during the workshop organized in Bariloche, Argentina: “Understanding Biodiversity Loss: A Workshop on Forest Fragmentation in South America “ (26 – 30 June 2006). The workshop allowed an assessment of the situation for the region and provided an analysis of the state of the art on the subject and an identification of gaps in research. More importantly the activity, funded by MEDD, France (Ministry of ecology and sustainable development); allowed an assembly of experts working on the evaluation of temperate and subtropical forests within the region (South America). The workshop was focused within the framework of the “Paris Declaration for the biodiversity" (Paris Conference, January 2005). As the Declaration states, we aim at bringing researchers together from developing countries and reinforcing the links between North and South in order to work towards an improved protection of biodiversity. The overall goal of creating a network of experts working more precisely on forest fragmentation, biodiversity loss and conservation issues targeted two main issues: i) improve the knowledge and the relevance of the indicators that can be developed and used in relation to forest biodiversity loss in South America. ii) facilitate building capacity not only in monitoring and evaluating forest fragmentation but also on forest restoration to mitigate the existing trends on biodiversity loss for the region.
Sandra Luque
S Luque Introduction Overview of Biodiversity Loss in South America
S. Luque & M.J. Pacha In response to global concern over the rapid loss of the world’s biodiversity, the 6th Conference of the Parties of the Convention on Biological Diversity (CBD) adopted a global target to reduce the rate of biodiversity loss by 2010 (CBD 2002). This target, which was later endorsed by the World Summit on Sustainable Development (United Nations 2002), has also been adopted by a number of regional scale policies and processes. The European Union Sustainable Development Strategy (2001a) and various other European Union policies (EC 1998, 2001b, c) set similar or even more ambitious biodiversity goals. The Pan-European Ministerial ‘Environment for Europe’ process adopted a resolution on halting the loss of biodiversity by 2010 (UN/ECE 2003). This widespread adoption of targets for reducing the rate of biodiversity loss has highlighted a need for indicators that will allow policy makers to track progress towards these ambitious goals. Recognising this need, the Convention of the parties (CoP) of the CBD identified a series of biodiversity indicators for immediate testing (UNEP 2004). Such indicators are needed at national, regional and global levels. In June 2004, the Environment Council of the EU adopted a set of 15 headline indicators for biodiversity to evaluate progress towards the 2010 target (Council of the European Union 2004). This set of indicators was recommended by the EU Biodiversity Expert Group and its Ad Hoc Working Group on Indicators, Monitoring and Assessment, and the Malahide stakeholder conference (Anonymous 2004). Both the CBD decision and the European documents recommend, among other indicators for immediate testing, indicators of trends in abundance and distribution of selected species. Species trend indicators are considered a sensitive measure of biodiversity change (Balmford et al. 2003; Ten Brink et al. 1991; Ten Brink 2000), and one such approach, composite species trend indicators, has been increasingly applied. In addition to the global-scale Living Planet Index (Loh 2002, and this volume) there are several instances of the successful implementation of such indicators, principally at national scales (Jenkins et al. 2004). The UK Headline indicator of wild bird populations (Gregory 2003a) is one example. The European Bird Census Council (EBCC) has used a similar approach to develop the Pan-European Common Bird Index for farmland and forest birds (Gregory 2003b; Gregory et al. 2004). Another set of indicators is directly related to forest biodiversity and, in particular, to forest cover loss. Valid indicators for this target area are poorly developed. During the International Conference on Biodiversity (Paris, January 2005), the workshop on “Biodiversity Indicators and the 2010 target: scientific challenges in meeting and assessing progress towards the 2010 biodiversity targets and related goals”; identified forest fragmentation as a key indicator to be added to the list. However, consensus and work is needed in the application of the indicator as a tool for monitoring forest status within the Action 6 framework. We need also to reach a consensus on the use of the indicator according to general guidelines. These guidelines, as established in the CBD 2010 targets, need to be set in order to develop suitable indicators for informing the general public on biodiversity trends. The indicators should match the set of requirements as listed in the CBD general guidelines and principles for developing national-level biodiversity monitoring programmes and indicators (UNEP 2003a). These principles require that an indicator be, among other characteristics: policy and biodiversity relevant; scientifically sound; broadly accepted; affordable to produce and update; sensitive; representative; flexible; and amenable to aggregation. Within this context, we intend to build up this network in order to reinforce the local capacity of different actors and to coordinate actions to prepare integrated projects at the international level that can have an impact at the global level. The goal is to work on native forest that has a particular important biodiversity value and that is been neglected until now in International projects. It is only with a proactive co-operation between North and South that we will be able to reach the targets set up by the CBD 2010 and reinforced during the Biodiversity Conference in Paris (January 2005). At present, we have an exchange group but not an actual official network due to the lack of funding (FRAGFORNET (http://sympa.lyon.cemagref.fr/wws/info/fragfornet)). But in the light of our efforts we hope to be able to develop a network based on this core group of experts that work on forest fragmentation, biodiversity loss and conservation issues. The aims are to improve the knowledge and the relevance of the indicators that can be developed, in particular for temperate and subtropical forested habitats in South America. We organize our network of contacts on the basis of decisions taken at the time of the seventh conference of the parts of the CBD (UNEP/CBD/COP/7/21), which relates to the biological diversity of forests (VII/I). The general objective is to build a flexible network in which national and regional actions can be developed according to the priorities and nationals’ interests in order to implement an effective program for the protection of biodiversity in the future.
The critical importance to help reduce the rate of biodiversity loss by 2010 in South America Biological diversity, the variety of all forms of life on Earth, plays a critical role in meeting human needs directly while also maintaining the ecological processes upon which our survival depends (BSP 1996). By any standard of measure, the Latin American and Caribbean (LAC) region is the repository of some of the world’s richest biodiversity, containing 40% of Earth’s plant and animal species (Global Environment Outlook 2000). Nine of the 25 most biodiverse countries are located in the LAC region (Caldecott et al. 1994). Of the 229 terrestrial ecoregions (geographically distinct assemblages of natural communities that share a large majority of species, dynamics and environmental conditions) designated in the region by the World Wildlife Fund (WWF), 57 are considered to be highest priority for conservation at the regional scale. Although South America still maintains vast areas of intact tropical and temperate forest, the region’s biodiversity is facing significant and growing threats, including increased rates of deforestation. One of the problems to monitor, manage and restore biodiversity is the unequal distribution of funding (Castro and Locker 2000). In order to effectively mitigate these threats, practitioners and donors in the conservation community must work together with host countries to improve the conservation of the region’s biodiversity.
The biological importance of the native forest1
Most of the native forests that are represented in this publication are amongst the most threatened ecoregions around the world and are included in the Global 2000 list of the Worldwide Fund for Nature (WWF) which includes representations of all major habitat types in each major biogeographic unit of the world and aims to prioritize conservation actions worldwide. They are located in the southern part of South America. Chile and Argentina together harbour the largest temperate rainforest area of South America, and more than half of the temperate forests in the Southern Hemisphere (Donoso, 1993; Wilcox, 1996). These forests are classified as temperate rainforests because of their geographical location outside the tropics, and because they experience high rainfall and low temperatures in winter. Similar forests are found in Tasmania, New Zealand and the Pacific Northwest in North America. Forests in South America are important as they store vast quantities of carbon that contribute to global climate regulation, flood control, water purification and soil nutrient cycling, as well as providing habitat for a high diversity of species that contribute to the genetic material for valuable new products and a foundation for the resilience of natural systems. First, the Valdivian Rainforest Ecoregion is located from 36º S through 48º S, and extends from southern Chile and adjacent Argentina This ecoregion includes terrestrial, freshwater and coastal marine ecosystems and is a worldwide reserve due to its unique biodiversity and high biogeographic and ecological significance, covering a total area of 10.5 million ha. These forests have a high diversity of trees and shrub species (over 35% of tree genera are endemic), and are the habitat of 60 endemic bird and 38 endemic mammal species. Reports indicate 23% of endemism for reptiles, 30% for birds, 33% for mammals, 50% for fish and 76% for amphibians. Forests in southern Chile include Fitzroya cupressoides, the second longest living tree worldwide that may live over 3600 years. These forests are home to over 900 vascular plant species, including 60 tree species, of which over 90% are endemic (Arroyo et al., 1995). Due to their special biodiversity assemblages, the Valdivian forests provide important ecosystem services that are the basis for several relevant economic activities, including the conservation of biological diversity of aquatic ecosystems, water production (quantity and quality), salmon farming (accounting for over 1 billion US dollars of annual exports, which represent 80% of the exports from the Lake Region in southern Chile), sport fishing and ecotourism.
1 In collaboration with Adriana Rovere, Cristian Echeverria and Daniel Somma Secondly, the Gran Chaco Americano is the largest dry forest in South America and the continent’s most extensive forested region after Amazonia. It occupies territories in Argentina, Paraguay, Bolivia, and Brazil. Extending from tropical latitudes (18°S) to subtropical zones (31°S), the Chaco shows strong climatic gradients, generating (together with geological and topographic characteristics) a range of environments: wide plains, swamps, and dry or seasonally flooded savannas, marshes, salt flats, and a variety of forests and scrublands. This landscape diversity translates into a high diversity of animal and plant species that makes the Chaco a key area for biodiversity conservation. The Chaco also provides important resources for the local population and industries. Because of the fragility of the Gran Chaco’s natural resources, the irreversibility of many of the changes that have already taken place in the region, and the human pressures currently underway, urgent steps are needed to reconcile productive sector interests with the sustainable development and conservation of the region’s natural resources through a long-term vision (TNC et al., 2005).
Thirdly, the Yungas rainforest in the Northwest of Argentina and South of Bolivia. is considered the most extensive biodiversity hotspot in Argentina. There are about 5,000,000 hectares that host 60% of the total bird species of Argentina and some key mammals like the jaguar (Panthera onca) and the Andean taruca (Hippocamelus antisensis). The latter is a mountain deer nominated as a “national natural heritage”. At the same time, this subtropical rainforest acts as an important carbon sink, an important water producer and the watershed shield for an extended irrigated region. The Yungas conditions are changing rapidly, as indicated by an estimated 1,250,000 hectares converted to agriculture between 1975 and 1988 (no later figures are available for the whole region). Synergy will be strengthened in the near future with ongoing local initiatives such as the Yungas biosphere reserve project and the local reforestation program in old degraded forests.
The forth important native forest that is described in this publication is the Atlantic Forest which is one of the most endangered and biologically diverse biomes on the Earth. It is one of the Conservation International's “Hot Spots; it is a WWF Global 200 Ecoregion; it overlaps significantly with Birdlife International's Endemic Bird Areas of the World: and it is one of the IUCN/WWF Centers of Plant Diversity. Unfortunately, with high human population densities and deep soils, only 7% of the original forest cover remains intact. Fifteen different ecoregions can be recognized within the Atlantic Forest biome. The Upper Parana Atlantic Forest constitutes the southwestern portion of the Atlantic Forest ecoregion, and extends from the western slopes of the Serra do Mar in Brazil to eastern Paraguay and the Misiones Province in northeast Argentina. A continuous subtropical semi- deciduous forest originally covered this entire area with a high diversity of animal and plant species. Deforestation in the Upper Parana Atlantic Forest has been most severe in Brazil, where only 3% of the original forest remains intact, most of it in protected areas. Paraguay has approximately 10% of its original forest cover remaining, though with a very high deforestation rate. In contrast, approximately 50% of the original cover remains intact in the Argentine province of Misiones with the bulk located in the almost continuous block in the northcentral portion of the province. This is the largest remaining forest block of the Atlantic Forest, and still contains the original set of large vertebrates including top predators such as harpy eagles, crested eagles, jaguars, pumas, and ocelots, and large herbivores such as tapirs, brocket deer, and peccaries (Di Bitetti et al 2003).
Pressures on the native forests
All the native forests that have been described are suffering from intensive pressures that are threatening the biodiversity and persistence of the ecoregions in the short and long term. In the Valdivian forests, the main threats are degradation due to inadequate logging practices, and destruction due to conversion to agriculture and fast-growing plantations, as well as human-set fires. Current estimates indicate that only 10-20% of the native forests are adequately managed. Most of them are either destroyed or degraded, leading to important negative reductions on the biological diversity of terrestrial ecosystems, as well as on the ecosystem services which depend on this diversity. The Gran Chaco and the Yungas are suffering irrational logging for soybean and cane plantations that bring huge revenues to the international market. Large areas are being cleared and small subsistence farming and communities are losing not only their ecosystem but also their identity and livelihood.
On the other hand, the greatest threat to biodiversity in the Upper Parana Atlantic Forest ecoregion is the extreme degree of forest fragmentation and degradation due to agricultural expansion. Large-scale soybean production, pine plantations, and pasture for cattle ranching, as well as small-scale tobacco, yerba-mate plantations and subsistence agriculture are all contributing to forest fragmentation across the ecoregion. Other causes of forest conversion and degradation include squatting by landless people, the construction of infrastructure (dams, roads, etc.), illegal and unsustainable hunting of wildlife, and unsustainable exploitation of the native forest. However, despite the high degree of forest fragmentation across the ecoregion, there are still excellent opportunities for conserving remaining large forest blocks, as is the case with the Misiones.
As a result of all the pressures on these native forests, they are being degraded, fragmented and large areas of forest are being lost.
Forest fragmentation within the context of this publication
Fragmentation is simply the disruption of continuity (Lord and Norton 1990). When defined in this manner, the concept of fragmentation can be applied to any domain in which continuity is important to the functioning of ecosystems (Lord and Norton 1990). In a restricted way, fragmentation occurs when a large expanse of habitat is transformed into a number of smaller patches of smaller total area, isolated from each other by a matrix of habitats unlike the original (Wilcove et al. ,1986). The fragmentation of natural habitats is usually a result of the expansion of land use that accompanies human population growth. As fragmentation proceeds, average fragment size and total fragment area decreases and insularity of fragments increases (Moore 1962; Webb and Haskins 1980; Burgess and Sharpe 1981). Habitat fragmentation and forest loss have been recognized as a major threat to ecosystems worldwide (Armenteras et al. 2003; Dale and Pearson, 1997; Iida and Nakashizuka, 1995; Noss, 2001). These two processes may have negative effects on biodiversity, by increasing isolation of habitats (Debinski and Holt, 2000), endangering species, and modifying species’ population dynamics (Watson et al., 2004). Fragmentation may also have negative effects on species richness by reducing the probability of successful dispersal and establishment (Gigord et al., 1999; Luque et al., 1994; Luque 2000) as well as by reducing the capacity of a patch of habitat to sustain a resident population (Iida and Nakashizuka, 1995). For example, fragmentation of the Maulino temperate forest in central Chile has affected the abundance of bird richness (Vergara and Simonetti, 2004) and regeneration of shade-tolerant species (Bustamante and Castor, 1998), and has also favoured the invasion of alien species (Bustamante et al., 2003). The ecological consequences of fragmentation can differ depending on the pattern or spatial configuration imposed on a landscape and how this varies both temporally and spatially (Armenteras et al., 2003; Ite and Adams, 1998). Some studies have shown that the spatial configuration of the landscape and community structure may significantly affect species richness at different scales (Steiner and Köhler, 2003). Other authors emphasise the need to incorporate the spatial configuration and connectivity attributes at a landscape level in order to protect the ecological integrity of species assemblages (Herrmann et al., 2005; Piessens et al., 2005). The dynamics of populations inhabiting terrestrial habitat fragments have received considerable research attention, including studies of birds, mammals, invertebrates, and plants (see Herkert 1994 and references therein). Perhaps the most extensively studied system thus far is the breeding birds of eastern North American deciduous forest, where several researchers have shown that habitat fragmentation adversely affects many forest birds species (Herkert 1994; Robbins 1979, Leck et al. 1981, 1988, Askins and Philbrick 1987, Johnston and Hagan 1992). Although there is general agreement on the effects of fragmentation on breeding birds within forest habitats, the mechanisms that account for these trends are not clear (Lynch 1987, Martin 1988). There is a need for studies that provide a quantitative treatment of landscape pattern changes and dynamics to better understand the widespread population decline of several species in fragmented landscapes. In this sense, this publication provides case studies that serve as examples of the type of research that needs to be integrated in collaborative projects. In order to better understand fragmentation, we need to be able to compare different study sites and species information to target the many
unresolved questions that exist within the subject, as has been pointed out by several authors (Farigh 2003, Bissonette & Storch 2002). About this publication
The contributions for this publication are grouped considering three main topics that are considered relevant in these ecoregions such as the need for science-based information on the status of these native forests and to suggest practical tools for their conservation, restoration and management. In this sense all the contributions use different concepts of landscape ecology and they show the growing importance of this discipline in integrating different concepts of remote sensing, geographical information systems, description of the natural world and incorporation of socio economic and cultural aspects. In the first part, Multitemporal changes and forest status, the authors describe changes in forest cover over a period of time using mainly remote sensing techniques. They evaluate the composition and landscape structure of the remaining forest fragments. The first four contributions are from the Gran Chaco region: Torrella et al. describes the changes that have occurred in nearly 50 years in a particular area covering 72.000 ha and have analysed the surface reduction and the forest fragmentation using several landscape indices. Guinzburg et al. used satellite images to analyse the effect of agriculture expansion in the last ten years in an area of approximately 4 million ha. Parmunchi et al. concentrated on the last 5 years and identified landscape pattern changes in the Chaco region along a precipitation gradient beween 1,100 and 600 mm annual rainfall isolines. Menghi and Sueldo focused on the mountainous Chaco region in Central Argentina and described the landscape spatial structure and habitat diversity of Polylepis astralis and Lithraea ternifolia forests in high mountain areas. Echeverria et al. analysed land-use change and spatial patterns in the Valdivian forest over twenty years using landscape indicators and satellite images. Finally, Vibrans et al. presented a forest inventory and characterised the phytosociology of a section of the Atlantic Forest in the Santa Catarina state in Brazil.
In the second part of the publication, Ecological consequences of forest fragmentation, three contributions aim at exploring some aspects on how fragmentation can affect interactions among species and ecological processes. Vidella et al. analysed the effect of fragmentation on herbivory, comparing fragmented and continuous patches of the Chaco forests in Central Argentina. Marchelli et al. concentrated more on reproductive characteristics and evaluated the degree of connectivity between populations of three tree species of the Valdivian Forests of Argentina through pollen flow studies. Finally in this section, Echeverria et al. described the effect of land- use change on the water regime in ñadi soils in the Valdivian Forest of Chile.
The third part, Landscape Ecology for conservation, management and restoration, has contributions of different authors that want to show how the landscape approach can be used to provide recommendations for better management of natural resources based on local needs. Lencinas et al. focused on analysing how alternative forest management methods can improve the species richness of Nothofagus pumilo forests in Patagonia, taking into consideration a landscape level approach. On the other hand, Bachmann et al. introduced different land-use planning strategies in the Yungas region of Argentina, aimed at improving biodiversity conservation. Lara et al. presented the case of experimental plantation of Fitzorya cupressoides, aimed at genetic conservation of the species in southern Chile. Finally, Pacha et al. showed how landscape ecology can serve as a useful tool to identify conservation priorities and help to plan actions for conservation at the species, landscape and ecoregional level in the Upper Parana Atlantic Forest.
We hope that these contributions can foster further research and actions to conserve the temperate forests of South America and their rich ecosystems.
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Part I
Multitemporal changes and forest status
1. Análisis multitemporal de la fragmentación y reducción del Bosque de Tres 19 Quebrachos S. A. Torrella, R. G. Ginzburg y J. M. Adámoli
2. Cuantificación y análisis regional de la expansión agropecuaria en el Chaco Argentino 28 R. G. Ginzburg, S. A. Torrella y J. M. Adámoli
3. Landscape changes due to native forest loss along a precipitation gradient in the Chaco region, Argentina 39 M. G i, J. Bono, M. Stamati, C. Montenegro, M. Brouver, E. Manghi and M. Strada
4. Landscape mosaic, habitat structure and fragmentation of native forests at Córdoba 50 mountain areas (Argentina central). M. Menghi and R. del Sueldo
5. Patterns of land use change and forest fragmentation in the temperate forests in southern Chile 63 C. Echeverría,, D. A. Coomes, A. C. Newton, A. Lara and J. M. Rey-Benayas
6. Forest floristic inventory of Mixed Ombrophilous Forest and Deciduous Forest of Santa Catarina State, Southern Brazil: preliminary results. 74 C. Vibrans; A. Uhlmann; L. Sevegnani; M. Marcolin; N. Nakajima,C. R. Grippa, E. Brogni & M. Braga Godoy
Part II
Ecological consequences of forest fragmentation
1. Habitat fragmentation effects on insect herbivory in Chaco Serrano woodlands 86 M. Videla, L. Cagnolo, G. Valladares, A. Salvo, S. Fenoglio
2. Extensive pollen flow may counteract the effects of landscape fragmentation 94 P. Marchelli, A.C. Moreno & L.A. Gallo
3. Effects of forest loss on soil water regime in the temperate landscape in southern 102 Chile C. Echeverría, O. Thiers, A. Lara
Part III
Landscape Ecology for conservation, management and restoration
1. Mitigation of biodiversity loss in Nothofagus pumilio managed forests of South 112 Patagonia M.V. Lencinas, G. Martínez Pastur, E. Gallo, A. Moretto, C. Busso & P. Peri
2. Estrategias de ordenamiento territorial y conservación de la naturaleza en la Eco- 121 región de las Yungas (noroeste de Argentina) L. I. Bachmann, C. L. Daniele, A. G. Frassetto 133 3. Ecological restoration of Fitzroya cupressoides, a long-lived conifer in southern Chile A. Lara, C. Echeverría, O. Thiers, F. Bustos, E. Huss, B. Escoba
4. Conservation approaches in the Atlantic Forest of Argentina: from eco-region to 142 single-species M. J. Pacha, M. S. Di Bitetti, G. Placci, E. Carabelli1, A. Paviolo, C. D. De Angelo , M. Jaramillo
Part I
Multitemporal changes and forest status
Análisis multitemporal de la fragmentación y reducción del Bosque de Tres Quebrachos
S. A. Torrella1, R. G. Ginzburg1 y J. M. Adámoli 1y2
1 Laboratorio de Ecología Regional, Facultad de Ciencias Exactas y Naturales, Universidad de Buenos Aires. Ciudad Universitaria, Pab. II, 4º Piso (1428), Ciudad de Buenos Aires, Argentina. Tel.: 054-011-4576-3300 int. 214. e-mail: [email protected] 2 Consejo Nacional de Investigaciones Científicas y Técnicas (CONICET).
Abstract
Since the beginning of the 20th century, the Chaco region in Argentina has been the scenario of consecutive agricultural frontier increases. This expansion has been produced disregarding any territorial planning criteria, thus jeopardizing the conservation of many Chaco’s environments, including the object of this study: the Tres Quebrachos Forest. The landscape structure, interpreted as an agricultural matrix of remaining patches of the original Tres Quebrachos Forest, was analyzed in an area of 71.975 hectares in the surroundings of Charata, Chaco Province. Based on aerial photographies and satellite images, the forest patches were digitalized and thematic maps were generated for the years 1957, 1975, 1988 & 2002. Several landscape indices were used to quantify, among others, two key issues: a) the surface reduction (the forest’s surface decreased from 25.898 hectares in 1957 to 14.917 in 2002), and b) the forests fragmentation (163 to 303 patches during the same period). The 1988-2002 period showed the most tangible variations for many of the studied indicators. This analysis provides essential information to design conservation strategies for these forests.
Resumen
Desde principios del siglo XX la región chaqueña en la República Argentina ha sido escenario de sucesivos avances de la frontera agrícola. Esta expansión, realizada al margen de cualquier criterio de planificación territorial, compromete seriamente la conservación de muchos ambientes chaqueños, entre ellos, el Bosque de Tres Quebrachos, objeto de este estudio. En un área de 71.975 hectáreas alrededor de Charata, Provincia de Chaco, se analizó la estructura del paisaje, interpretada como una matriz agrícola con parches remanentes del Bosque de Tres Quebrachos original. Sobre la base de fotografías aéreas e imágenes satelitales, se digitalizaron los parches de bosque y se generaron mapas temáticos para los años 1957, 1975, 1988 y 2002. Se utilizaron diversos índices de paisaje, que permitieron cuantificar, dos aspectos clave: a) la reducción en superficie (los bosques pasaron de 25.898 hectáreas en 1957 a 14.917 en 2002) y b) la fragmentación de los bosques (pasaron de 163 a 303 parches en ese período). El período 1988-2002 presentó las variaciones más marcadas para muchos de los índices estudiados. Este análisis provee información esencial para diseñar estrategias de conservación para estos bosques.
Understanding biodiverity loss: an overview on forest fragmentation in South America 19 Palabras clave
Chaco Subhúmedo Central; Bosque de tres quebrachos; Expansión agrícola; Fragmentación de bosques; Deforestación.
Introducción
El Bosque de Tres Quebrachos es el único ecosistema donde coexisten el quebracho blanco (Aspidosperma quebracho blanco), el quebracho colorado chaqueño (Schinopsis balansae) y el quebracho colorado santiagueño (Schinopsis lorentzii). La distribución espacial de este bosque en la República Argentina coincide con la subregión ecológica del Chaco Subhúmedo Central, que abarca el centro-oeste de la Provincia del Chaco, este de Santiago del Estero y noroeste de Santa Fe (Morello y Adámoli, 1968 y 1974). Dentro de esta subregión, el Bosque de Tres Quebrachos ocupa aproximadamente un 50 % de la superficie, y se manifiesta donde los suelos son profundos, de buena fertilidad natural y sin mayores limitaciones por inundaciones, salinidad ni sequías, es decir que este bosque se corresponde con los mejores suelos agrícolas de la zona, también llamada “el óptimo de los tres quebrachos”. (Morello y Adámoli, 1974). Las áreas ocupadas por el Bosque de Tres Quebrachos han sido transformadas para la agricultura desde las primeras décadas del siglo XX, por colonias de pequeños productores agrícolas.
En un análisis que contempló la expansión agrícola en la región chaqueña (ver “Cuantificación y análisis regional de la expansión agropecuaria en el chaco argentino” en esta publicación), la subregión ecológica del Chaco Subhúmedo Central concentró casi el 50 % de la expansión total (aunque ocupa tan sólo algo más del 10 % del área), y para el año 2002 se hallaba ocupada por agricultura cerca de la mitad de su superficie.
En el presente trabajo se cuantifica la sensible reducción del área y se evalúa la intensa fragmentación de los parches del Bosque de Tres Quebrachos. Asimismo, se alerta acerca de los riesgos de desaparición de este bosque, por la intensidad del avance agropecuario y por la falta de criterios de planificación que compatibilicen los intereses de la producción, con los de la conservación de los bienes y servicios ambientales en ecosistemas únicos.
Materiales y Métodos
El área de estudio abarca 71.975 hectáreas alrededor de la Ciudad de Charata en la Provincia del Chaco, República Argentina (Figura 1), en la zona denominada “el óptimo de los tres quebrachos” (Morello y Adámoli, 1974).
Se analizó la evolución (1957-2002) de la estructura del paisaje, en particular los parches remanentes del Bosque de Tres Quebrachos dispuestos en la matriz agrícola.
Se trabajó con fotografías aéreas del año 1957 (escala 1:40.000) las que fueron escaneadas y luego ensambladas para obtener un mosaico fotográfico del área de estudio. Se utilizaron tres imágenes satelitales, una Landsat 2 de 1975 (path-row 245-79), una Landsat 5 (228-79) del año 1988 y una Landsat 7 (228/79) del año 2002. Se georreferenció la imagen satelital Landsat 7 del 2002, y sobre esta base se registraron las demás imágenes y el mosaico fotográfico.
A partir de la interpretación visual de las imágenes satelitales y el mosaico fotográfico, se procedió al mapeo de los fragmentos de bosque, visualizando las escenas en pantalla a escala 1:150.000, utilizando el programa Arc View Gis 3.2. Understanding biodiverity loss: an overview on forest fragmentation in South America 20 Se aplicaron diversos índices de paisaje que permitieron cuantificar, la fragmentación y reducción de superficie que sufrieron estos bosques a lo largo de 45 años. Para el cálculo de superficies y la estadística de los parches de bosque digitalizados se empleó la extensión Patch Analyst del mismo programa.
Figura 1. Situación geográfica del área de estudio (1.a-República Argentina; 1.b-Provincia del Chaco; 1.c-Área de estudio en torno a la Ciudad de Charata).
1.c 1.a
Ruta Provincial 94
1.b
Resultados
Los resultados muestran la importante reducción de superficie y la fuerte fragmentación que sufrieron estos bosques durante el período analizado (tabla 1 y figura 2). La superficie total pasó de 25.898 ha en 1957 a 14.917 ha en 2002 (figura 3.a). O sea que en 45 años se perdieron 10.981 hectáreas, lo que significó una pérdida del 42,4 % de la superficie de bosques, los que sólo ocupan 20,72 % de la zona estudiada.
Tabla 1. Índices de paisaje calculados.
Año Superficie Número de Tamaño Mediana del Desvío estándar total (ha) parches medio de tamaño de del tamaño de parche (ha) parche (ha) parche (ha) 1957 25.898,48 163 156,96 10,63 830,41 1975 24.143,88 197 122,56 12,15 706,32 1988 22.154,06 227 97,59 10,85 460,79 2002 14.916,74 303 49,23 12,64 126,53
Understanding biodiverity loss: an overview on forest fragmentation in South America 21 Tabla 1. Índices de paisaje calculados (continuación).
Año Borde total Borde Densidad Media de Índice Índice de (m) medio de de borde perímetro/área de forma medio parche (m) (m/ha) (m/ha) forma pesado por medio área 1957 1.195.779 7.247,15 46,17 481,98 1,72 5,91 1975 1.214.357 6.164,25 50,30 399,20 1,67 5,83 1988 1.183.251 5.212,56 53,41 371,95 1,63 4,56 2002 1.091.352 3.601,82 73,16 279,25 1,61 2,58
Figura 2. Evolución de los parches remanentes del Bosque de Tres Quebrachos (gris) dispuestos en la matriz agrícola (blanco), entre los años 1957 y 2002.
2.a 2.b
2.c 2.d
La cantidad de parches aumentó en todos los períodos analizados, pasando de 163 en 1957 a 303 en 2002 (figura 3.b), es decir que la reducción de superficie fue acompañada por un intenso proceso de fragmentación. La intensidad de la reducción fue tal, que más allá de esta fragmentación, entre 1988 y 2002 hubo 35 parches de bosque que fueron eliminados completamente.
La fragmentación afecto al tamaño medio de parche, el cual se redujo 68,64 % durante todas las etapas, cayendo de 157 ha en 1957 a 49 ha en 2002 (figura 3.c). Sin embargo, la mediana del tamaño de parche no se modificó sustancialmente, lo que muestra la conjunción observada del aumento del número de parches y la disminución de sus tamaños, Understanding biodiverity loss: an overview on forest fragmentation in South America 22 o lo que es lo mismo el fraccionamiento de los parches de mayor tamaño (este fraccionamiento de los parches más grandes se ve reflejado también al analizar la disminución conjunta del tamaño medio de parche y del desvío estándar del tamaño de parche, el cual cayó de 830 ha en 1957 a 127 ha en 2002).
Figura 3. Índices de paisaje calculados para los parches de Bosque de Tres Quebrachos.
Superficie total de bosques (ha) Numero de parches 310 28,000.00 290 26,000.00 270 24,000.00 250 22,000.00 230 20,000.00 210 18,000.00 190 16,000.00 170 14,000.00 150 1957 1975 1988 2002 3.a 3.b 1957 1975 1988 2002
Tamaño medio de parche (ha) Densidad de borde (m/ha) 80.00 160.00 75.00 140.00 70.00 120.00 65.00 60.00 100.00 55.00 80.00 50.00 60.00 45.00 40.00 40.00 3.c 1957 1975 1988 2002 3.d 1957 1975 1988 2002
Media de la relacion perímetro área (m/ha) Índice de forma medio pesado por área
6.00 520.00 5.50 470.00 5.00
420.00 4.50 4.00 370.00 3.50
320.00 3.00
270.00 2.50 3.e 1957 1975 1988 2002 3.f 1957 1975 1988 2002
Tanto el borde total como el borde medio del parche disminuyeron en el transcurso de los 45 años estudiados, lo que es consecuencia por un lado, de la reducción sufrida en su Understanding biodiverity loss: an overview on forest fragmentation in South America 23 superficie (en cada período la superficie total de bosque fue menor y a su vez los parches fueron de menor tamaño) y por el otro de la simplificación de sus formas. Esto mostraría que, si bien, dado el aumento del número de parches se esperaría que aumente el borde total, el proceso de reducción de la superficie predominó sobre el de fragmentación.
El índice de densidad de borde (figura 3.d), que indica la cantidad de metros de borde de bosque que hay por hectárea de bosque, aumentó 58,46 %, trepando de 46 a 73 m/ha entre 1957 y 2002. Esto significa que aunque la cantidad de borde total de bosque se redujo en el área de estudio, su relación con la cantidad neta de bosque remanente, aumentó en todos los períodos.
Para los parches de bosque la relación media de perímetro/área disminuyó en todos los períodos, mostrando que el perímetro de cada parche es cada vez menor respecto de su superficie (figura 3.e). Considerando que tanto el borde medio de parche como el tamaño medio de parche disminuyeron en todos los años analizados, se llega a la conclusión de que en los parches predominó la simplificación de los bordes sobre la pérdida de superficie (la disminución en la cantidad de borde/parche fue superior a la disminución del área de estos parches).
El proceso de simplificación de los bordes se ve claramente al analizar que el índice de forma medio presentó valores decrecientes para todos los años (esta variable mide la complejidad de los parches en comparación con la forma de un círculo; cuanto más irregulares son los parches, mayor es el valor del índice). En tanto que al ponderar el índice y considerar la superficie de cada parche (índice de forma pesado por área) la reducción fue sensiblemente mayor, indicando la tendencia de los bordes de los parches a formas más regulares (figura 3.f).
Es de notar que ambos procesos, de reducción y fragmentación, se produjeron simultáneamente. Durante el período de 45 años considerado, la cantidad de parches de bosque y la superficie que ellos ocupan aumentó en las clases menores a 1000 hectáreas, mientras que se redujo para los parches mayores (tabla 2 y figura 4). Cabe aclarar, por si quedan dudas, que el aumento de superficie ocupada por parches pequeños no se dio por regeneración de bosques, sino debido a la fragmentación de los parches de mayor tamaño.
Tabla 2. Distribución de los parches de bosque en clases de superficies.
Año 1957 1975 1988 2002 Clase Nº Superf. Nº Superf. Nº Superf. Nº Superf. parches (ha) parches (ha) parches (ha) parches (ha) < 10 ha 79 301,48 88 354,61 108 468,17 128 597,63 10-100 ha 61 1.857,99 83 2.612,77 93 2.997,04 140 4.747,00 100-1000 ha 18 5.902,07 23 7.534,23 23 8.396,26 34 8.110,00 > 1000 ha 5 17.836,94 3 13.642,27 3 10.292,59 1 1.462,12
Understanding biodiverity loss: an overview on forest fragmentation in South America 24 Figura 4. Superficie (4.a) y número (4.b) de parches de bosque distribuidos en clases de superficies.
Superficie (ha) por clase de tamaño de Número de parches por clase de tamaño parche 160 18,000 < 10ha 16,000 140 < 10ha 14,000 120
12,000 10 a 100 100 10 a 100 ha ha 10,000 80
8,000 60 100 a 100 a 6,000 1000 ha 40 1000 ha
4,000 20 >1000 ha >1000 ha 2,000 0 0 1957 1975 1988 2002 4.a 1957 1975 1988 2002 4.b
Por último, el período comprendido entre los años 1988 y 2002 presentó las variaciones más marcadas para muchos de los índices estudiados: la superficie total distribuida en parches mayores a 1.000 hectáreas cayó de 10.293 ha a 1.462 ha; la pérdida de superficie de bosque alcanzó a 65,9% (figura 5) y la densidad de borde aumentó 37%.
Figura 5. Pérdida de bosques durante los períodos 1957-1975, 1975-1988 y 1998-2002.
Conclusión
Los resultados obtenidos muestran la significativa reducción de superficie y la intensa fragmentación que ha sufrido el Bosque de Tres Quebrachos, proceso que se ha acelerado en el último período. Dado el alto valor de conservación de este tipo de bosque, es imperioso tomar medidas de protección y manejo para asegurar su preservación.
Understanding biodiverity loss: an overview on forest fragmentation in South America 25 Un criterio básico de conservación indica que debería preservarse al menos una parte de cada tipo de bosque natural que exista (Burkart, 1999), ya que cada tipo de bosque contiene un elenco diferente de especies de plantas y animales, con sus propias interacciones, lo que es fundamental para la conservación. La autoperpetuación del bosque depende de la interacción entre manchones en diversos estados de desarrollo o sucesión. A su vez, las interacciones entre el área protegida y su entorno influyen en el funcionamiento de los ecosistemas protegidos, por lo cual además de proteger un área, es necesario planificar los usos en el entorno, con normas de manejo adecuadas y sustentables (Morello y Matteucci, 1999).
De acuerdo con diversos especialistas, el mínimo a conservar debería estar entre 15 y 30 % (Mackinnon et al., 1986; Reid y Miller, 1989), entre los cuales sugieren 5 a 10% de protección estricta, como mínimo, y 10 a 20 % adicional de áreas no estrictas (un 10 % bajo protección estricta sería insuficiente, sin amplias áreas de amortiguación de bosques manejados).
El Bosque de Tres Quebrachos presenta tal nivel de fragmentación, sobreexplotación y ritmo de deforestación, que si no se adoptan medidas urgentes, en pocos años más es posible que ya no queden masas disponibles con número, tamaño y conectividad mínimas como para asegurar su protección y el cumplimiento de los servicios ecológicos que este ambiente presta. Debido a la falta de tierras fiscales en el área, la única posibilidad real de conservación consistiría en integrar una red de áreas protegidas en propiedades privadas, para poder conservar muestras representativas de la diversidad ecológica de este tipo de bosques.
Entre los especialistas en biología de la conservación, existe consenso acerca de que las curvas especie-área permiten predecir la proporción de especies que se extinguirán en una región con base en la cantidad de hábitat que se pierde (Raven, 1987, 1988a, b; Myers, 1988a; Simberloff, 1986; Lovejoy, 1980; Wilson, 1988; Reid y Miller, 1989). La tasa de extinción de especies en base a diversos escenarios de deforestación, no tiene una relación lineal, ya que para una pérdida de 11 % de la superficie se prevé una pérdida de 2 % de especies, mientras que con 44,8 % de pérdida de superficie, las pérdidas de especies llegarían a 35 % (Reid, 1992).
Los trabajos basados en aplicaciones de la teoría de biogeografía de islas (MacArthur y Wilson, 1967), demuestran que en bosques tropicales, cuando un hábitat pierde el 90% de su extensión, con el tiempo se extingue la mitad de sus especies (Myers, 1988b). Sin embargo, (Wilson, 1988) destaca que estas proyecciones son conservadoras, porque aún cuando una porción de especies sobreviva, probablemente hayan sufrido una significativa reducción en la variación genética entre sus miembros, debido a la pérdida de genes que se dio junto con la disminución del número de individuos. De esta forma, cuando un bosque se reduce, por ejemplo de 10.000 a 1.000 ha, algunas extinciones de especies son inmediatas; otras especies seguirán existiendo pero en poblaciones que se habrán reducido de forma muy peligrosa para su viabilidad futura.
Las pérdidas estimadas en el Bosque de Tres Quebrachos, son del orden del 80 % de la superficie original. Los cambios climáticos podrían exacerbar esta pérdida potencial. Para detener la pérdida de especies, es necesario por un lado disminuir la tasa de deforestación y por el otro, racionalizar el uso sustentable del bosque y proteger los hábitats claves (con alta riqueza de especies y endemismos).
El Bosque de Tres Quebrachos se está perdiendo y fragmentando aceleradamente. En un escenario en el que la frontera agrícola continúa avanzando con intensidad, donde no hay ningún tipo de áreas protegidas y que presenta una notoria escasez de tierras fiscales, Understanding biodiverity loss: an overview on forest fragmentation in South America 26 las perspectivas de conservación de las especies y servicios ambientales de este ecosistema tan particular, está muy comprometida. Por ello surge la necesidad de implementar con urgencia acciones concretas que permitan controlar el patrón espacial y la localización de los fragmentos remanentes, asegurando la existencia de áreas relativamente grandes de hábitats naturales y semi-naturales, para reducir la pérdida de especies.
Agradecimientos: Silvia Matteucci, Patricia Kandus, Martha Gazzano.
Bibliografía
Burkart, R. (1999) Conservación de la biodiversidad en bosques naturales productivos del subtrópico argentino. Matteucci, Solbrig, Morello y Halffter (editores). Biodiversidad y uso de la tierra. Conceptos y ejemplos de Latinoamérica. EUDEBA, Buenos Aires, 589 pp. Lovejoy, T.E. (1980) A projection of species extinctions, in Council on Environmental Quality (CEQ), The Global 2000 Report to the President, Vol. 2. CEQ, Washington, DC, pp. 328-31. MacArthur, R.H. y Wilson, E.O. (1967) The Theory of Island Biogeography. Princeton University Press, Princeton, RI. Mackinnon, J.; Mackinnon, K.; Child, G. y Thorsell, J. (comps). (1986). Managing protected areas in the Tropics. IUCN/UNEP, Gland (Suiza), 290 pp. Morello, J. y Adámoli, J. (1968) Las grandes unidades de vegetación y ambiente del Chaco argentino. Primera parte: Objetivos y metodología. Serie fitogeográfica Nº 10, INTA, Buenos Aires. 125 pp. Morello, J. y Adámoli, J. (1974) Las grandes unidades de vegetación y ambiente del Chaco argentino. Segunda parte: Vegetación y ambiente de la provincia del Chaco. Serie fitogeográfica Nº 13, INTA, Buenos Aires. 130 pp. Morello, J. y Matteucci, S. (1999) Biodiversidad y fragmentación de los bosques en la Argentina. Matteucci, Solbrig, Morello y Halffter (editores). Biodiversidad y uso de la tierra. Conceptos y ejemplos de Latinoamérica. EUDEBA, Buenos Aires, 589 pp. Myers, N. (1988a) Threatened biotas: 'hotspots' in tropical forests. Environmentalist, 8(3), 1- 20. Myers, N. (1988b) Tropical forests and their species - going, going...?. E.O. Wilson and F.M. Peter, eds. Biodiversity. National Academy Press, Washington, D.C., pp. 28-35 Raven, P. H. (1987) The scope of the plant conservation problem world-wide. D. Bramwell, O. Hamann, V. Heywood, and H. Synge, eds. Botanic Gardens and the World Conservation Strategy. Academic Press, London, pp. 19-29. Raven, P. H. (1988a) Biological resources and global stability. S. Kawano, J.H. Connell, and T. Hidaka, eds. Evolution and Coadaptation in Biotic Communities. University of Tokyo Press, Tokyo, pp. 3-27. Raven, P.H. (1988b) Our diminishing tropical forests. E.O. Wilson and F.M. Peter, eds. Biodiversity. National Academy Press, Washington, D.C., pp. 19-22. Reid, W. R. (1992) How many species will there be? T. C. Whitmore and J. A. Sayer, eds. Tropical deforestation and species extinction. New York. Reid, W. y Miller, K. (1989) Keeping options alive. The scientific basis of conserving biodiversity. World Resources Institute, Washington, D.C. Simberloff, D. (1986) Are we on the verge of a mass extinction in tropical rain forests? D.K. Elliott, ed. Dynamics of Extinction. New York, NY, pp. 165-180. Wilson, E.O. (1988) The current state of biological diversity. E.O. Wilson and F.M. Peter, eds. Biodiversity. National Academy Press, Washington, D.C.
Understanding biodiverity loss: an overview on forest fragmentation in South America 27 Cuantificación y análisis regional de la expansión agropecuaria en el Chaco Argentino
R. G. Ginzburg1, S. A. Torrella1 y J. M. Adámoli1,2
1 Laboratorio de Ecología Regional, Departamento de Ecología, Genética y Evolución, Facultad de Ciencias Exactas y Naturales, Universidad de Buenos Aires. Ciudad Universitaria, Pab. II, 4º Piso (1428), Ciudad de Buenos Aires, Argentina. Tel.: 054-011-4576-3300 int. 214. e-mail: [email protected] 2 Consejo Nacional de Investigaciones Científicas y Técnicas (CONICET).
Resumen El aumento en la producción agrícola registrado en el país en los últimos 15 años trajo aparejadas importantes implicancias ambientales. En la región chaqueña se generó una fuerte presión sobre el uso de la tierra, que desencadenó un intenso proceso de deforestación. Mediante el uso de imágenes satelitales se detectó que entre 1992 y 2002 las áreas transformadas en el zona de estudio tuvieron un aumento del 60%, de 3.014.107 a 4.816.502 hectáreas; es decir que se han perdido 1.802.395 ha de ambientes nativos, constituidos principalmente por bosques. Esta expansión tiene a la soja como su principal motor, y todo indica que la demanda de esta oleaginosa se mantendría o incrementaría en los próximos años. A esto hay que agregar la presión generada por el inminente desarrollo del mercado de los biocombustibles. Al desarrollarse este proceso sin una efectiva regulación por parte del estado, se compromete tanto la conservación de la biodiversidad, como también la sustentabilidad de los emprendimientos productivos.
Palabras clave Región chaqueña; Expansión agrícola; Chaco subhúmedo central; Bosque de tres quebrachos.
Abstract The increase of crop production in Argentina during the last 15 years entailed strong environmental implications. In Chaco region a strong pressure on land use appeared and unleashed an intense deforestation process. The use of satellite images detected in the 1992 – 2002 period, an increase of 60% in land conversion: from 3.014.107 to 4.816.502 hectares; this implies that 1.802.395 ha of native environments, mainly forests, have been lost. This expansion is leaded mainly by soy, and all signs indicate that the demand for this oleaginous will stay constant or even be increased during the next years. The pressure generated by the imminent development of the biofuel market must also be taken into account. As this process has been developed in absence of an effective State regulation, both biodiversity conservation and productive endeavor sustainability have been jeopardized.
Key words Chaco region; agricultural expansion; central sub-humid Chaco, Tres quebrachos Forest
Understanding biodiverity loss: an overview on forest fragmentation in South America 28 Introducción
En los últimos años se dio en la Argentina un significativo aumento en la producción agrícola; la superficie dedicada a cultivos anuales se incrementó en el período 1988-2002 en 5.500.000 hectáreas, pasando de 13.800.000 a 19.300.000 ha (INDEC). En este proceso cumple un papel fundamental el cultivo de la soja, que en el mismo período incrementó su superficie de 4.670.000 a 11.639.240 ha sembradas en el país, constituyendo en la actualidad cerca del 50% de la producción total del sector (SAGPyA). Los indicadores internacionales señalan que la demanda de soja no sólo se mantendría en los próximos años sino que además aumentaría sensiblemente, tanto en lo referido a las fuertes demandas de granos, aceites y harinas, como por la inminente expansión del uso de biocombustibles a nivel nacional e internacional.
La producción de soja ha aumentado en las distintas regiones del país, empezando por la región pampeana donde se expandió, en gran medida sobre otros cultivos -principalmente maíz-, así como sobre terrenos antes dedicados a la ganadería (Azcuy Ameghino y León, 2005).
En la región chaqueña el proceso tuvo distintos matices; la soja desplazó a cultivos tradicionales, como el algodón en la Provincia del Chaco, pero motorizó además una importante expansión de la agricultura en general: Entre 1995 y 2005, la superficie sembrada con cultivos anuales en Salta, Chaco y Santiago del Estero aumentó de 1.800.000 a 3.100.000 ha, y la sembrada con soja de 420.000 a 1.760.000 ha (ODSMA - OEA, en preparación).Al mismo tiempo, la región recibió a buena parte de los emprendimientos ganaderos desplazados de la región pampeana por la mayor rentabilidad de la actividad agrícola. Esta combinación constituyó una fuerte presión sobre los bosques nativos, que terminó desencadenando un proceso de desmontes generalizados. Sólo entre 1998 y 2002 se desmontaron 306.000 ha en Santiago del Estero, 194.000 en Salta y 118.000 en Chaco (Montenegro et al., 2004), convirtiendo a la región en la de mayor tasa de deforestación del país.
La expansión agropecuaria en la región chaqueña generó opiniones encontradas: por una parte se celebra la incorporación de nuevas áreas productivas al mapa agrícola del país y los ingresos económicos que ello implica. Por otra parte se alerta sobre los riesgos que conlleva el modelo adoptado, en consonancia con la creciente preocupación por los temas de sostenibilidad agraria (INTA-INDEC, 1994; Viglizzo, 2001). En este sentido se destaca que el proceso se lleva adelante sin que exista un plan de ordenamiento territorial, por lo que se permite que el avance se realice sobre zonas en las que no está garantizada la sustentabilidad de la producción, ya sea por condiciones edáficas o climáticas (Adámoli et al., 2004; Adámoli, 2005; Grau et al., 2005). Tampoco se están contemplando los riesgos ambientales de los procesos de expansión, como la pérdida de biodiversidad (Torrella et al., 2003), la simplificación del paisaje (Forman y Godron, 1985), o la conectividad entre ambientes.
El fenómeno también tuvo implicancias relevantes en el ámbito social, los productores y pobladores tradicionales de la región no se vieron incluidos mayoritariamente en el nuevo modelo, ni percibieron los beneficios económicos que este generó (Reboratti, 2005); también se alerta sobre la preocupante concentración de la renta y la disminución del trabajo rural (Barsky y Gelman, 2001).
El Gran Chaco Americano es una vasta planicie de más de 1.000.000 km2, de los cuales el 60% están en Argentina. Presenta en su extensión una gran variabilidad climática, acompañada por distintas formaciones fisonómicas, entre las que se destacan distintos tipos de bosques, caracterizados por la dominancia de especies del género Schinopsis; sabanas Understanding biodiverity loss: an overview on forest fragmentation in South America 29 y pastizales (Morello y Adámoli, 1974; Prado, 1993). Su alta biodiversidad y grado de intervención antrópica la sitúan en el contexto regional y nacional como una de las áreas prioritarias para la conservación (Dinerstein et al, 1995; Bertonatti y Corcuera, 2000).
En el presente trabajo se describe y analiza la configuración espacial del proceso de expansión agropecuaria en la región chaqueña, en el período 1992-2002, mediante la interpretación de imágenes satelitales y el uso de sistemas de información geográfica, en una aproximación tendiente a desentrañar sus implicancias ambientales en el nivel ecorregional.
Materiales y Métodos
Se definió como área de estudio a parte del Chaco argentino donde se desarrolla con mayor intensidad el proceso descrito de expansión agropecuaria. El área, con una superficie total de 41.748.513 hectáreas, quedó comprendida por el Chaco salteño, el norte de la Provincia de Santa Fe, y las Provincias del Chaco, Formosa y Santiago del Estero (a excepción de su extremo SE) (Figura 1).
Sobre imágenes satelitales se identificaron visualmente y se mapearon, trabajando a escala 1:250.000, todas las parcelas en las que la cobertura vegetal original ha sido sustituida por cultivos, tanto agrícolas como pasturas (en adelante “áreas transformadas”). Esta digitalización se realizó con el programa Arcview 3.2.
Se utilizaron mosaicos “Mr Sid” compuestos a partir de imágenes Landsat de acceso libre en internet (http://glcfapp.umiacs.umd.edu:8080/esdi/index.jsp); para abarcar la totalidad del área de estudio fueron necesarios los mosaicos de las ubicaciones 20-20, 20-25, 21-20 y 21-25. Se utilizó una serie de mosaicos del período 1986-1992 que en el texto y las tablas se indica como 1992 y otra del período 1999-2002, que se indica como 2002.
La ventaja de poder disponer de mosaicos de imágenes tiene la limitación de la amplitud temporal de las series. No obstante, el volumen de información generada, las tendencias detectadas y la localización espacial de los procesos, consideramos que son válidos para la toma de decisiones y la planificación en esta escala de trabajo.
En la enorme mayoría de los casos la diferenciación entre la vegetación nativa y cultivos es inequívoca, pero en algunos potreros de ganadería extensiva sobre campos naturales, pueden generarse confusiones, porque ciertos tipos de manejo pueden presentar un patrón similar al de las parcelas cultivadas. En estos casos la identificación de las parcelas se hizo ampliando sensiblemente la escala de la imagen, para mejorar la definición. Este tipo de errores no son intrínsecos de la metodología, ya que inclusive una clasificación automática sin una exhaustiva verificación a campo también puede presentarlos, incluso en mayor medida.
Para analizar el proceso de expansión en las distintas zonas climáticas, se dividió la región chaqueña a partir de un análisis bibliográfico de datos de precipitación anual (Galmarini y Raffo del Campo, 1964; Bianchi 1981; Bruniard 1987). Así quedaron definidas las siguientes zonas (figura 1):
Chaco Árido: menos de 500 mm, marginal en el área de estudio definida para este trabajo. Chaco Semiárido: 750 a 500 mm, la de mayor extensión territorial. Chaco Subhúmedo: 750 a 900 mm, presenta una faja muy angosta en el borde oeste de la región chaqueña (subhúmedo occidental), y una faja más ancha en la frontera entre Santiago del Ester, Chaco y Santa Fe (subhúmedo central). Understanding biodiverity loss: an overview on forest fragmentation in South America 30 Chaco Húmedo: más de 900 mm, se extiende por el este de las provincias de Formosa, Chaco y Santa Fe.
Figura 1. Área de estudio, división política y zonas climáticas.
Estas zonas no son estables en el tiempo, ya que pueden desplazarse, en función de ciclos plurianuales secos o húmedos. Tomando como referencia a la isohieta de 750 mm en la Provincia del Chaco, durante un ciclo húmedo esa isohieta se desplaza hacia el Oeste, mientras que en un ciclo seco se desplaza hacia el Este (en el borde occidental de la región, los desplazamientos son en sentido inverso).
Así, queda determinada una faja de variabilidad climática. En los últimos 25 años hubo un sensible desplazamiento de las isohietas hacia el Oeste, mientras que en los 5 años más recientes, hay evidencias de una tendencia más seca. Esto indicaría un proceso de reversibilidad climática. Las áreas con riesgo de reversión fueron definidas como aquellas en las que en una situación normal están dentro de la zona del subhúmedo, pero en un ciclo seco quedan incluidas en la zona correspondiente al subhúmedo seco a semiárido. Se identificaron así dos zonas con riesgo de reversión, en las áreas de contacto entre el Chaco Semiárido, y ambas porciones del Chaco Subhúmedo.
Understanding biodiverity loss: an overview on forest fragmentation in South America 31 Resultados
Los resultados obtenidos indican que para la primera serie temporal analizada (1992), las áreas transformadas cubrían 3.014.107 ha, o sea el 7,22 % de la superficie total estudiada. Para el año 2002 este valor trepó hasta el 11,54 % (4.816.502 ha). Las áreas transformadas tuvieron una expansión del 59,8 %, es decir que en este período fue sustituida la cobertura vegetal nativa, constituida principalmente por bosques, en 1.802.395 ha. Como se observa en la figura 2, estas áreas no tienen una distribución homogénea en el área de estudio, sino que se presentan agrupadas en núcleos de diferentes características.
Figura 2. Superficies donde la cobertura vegetal nativa ha sido eliminada.
Para visualizar más claramente los núcleos en los que se concentran las áreas transformadas, se dividió el área de estudio en hexágonos regulares de 10.000 hectáreas. Las áreas transformadas dentro de cada hexágono, fueron expresadas como % de cada polígono. En la figura 3 se muestran de esta manera las áreas transformadas para 1992 y 2002.
Se distinguen claramente los principales núcleos agrícolas de la región:
El grueso de los cultivos se localiza en el núcleo del centro-sur de la región, correspondiendo básicamente a la zona climática del “Chaco Subhúmedo Central”. Este núcleo se presenta dividido por una depresión salobre. La mayor parte del núcleo se localiza en torno al límite entre las provincias de Chaco y Santiago del Estero. La porción sur de este núcleo se encuentra en el límite entre el sudeste de la provincia de Santiago del Estero y el noroeste de Santa Fe.
Understanding biodiverity loss: an overview on forest fragmentation in South America 32
En la porción más húmeda de la región, en el este, predominan los suelos inundables. Por tal motivo, la agricultura sólo se expresa en los terrenos topográficamente más altos: - En el NE del área de estudio, puede observarse que las áreas cultivadas presentan un alineamiento en sentido NO-SE, acompañando el trazado de los albardones de los ríos.
- En el extremo SE del área, la agricultura se localiza en la llamada “dorsal oriental” en el NE de Santa Fe y sur de la provincia de Chaco, de orientación N-S.
El núcleo occidental de la agricultura está formado a su vez por varios centros, que se disponen de norte a sur recostados sobre las primeras estribaciones andinas.
Hay un par de núcleos en pleno Chaco Semiárido (ver figura 5a), que corresponden a zonas de irrigación: el más grande y evidente, ubicado en Santiago del Estero, en el área de riego del río Dulce, y el otro es el de la localidad de Joaquín V. González en la provincia de Salta, sobre el área de riego del río Juramento.
Figura 3. Áreas transformadas para 1992 (a) y 2002 (b), expresadas como porcentaje de hexágonos de 10.000 ha.
a b
En la figura 4 se representa, también como porcentaje de la superficie de los hexágonos, la expansión agropecuaria, medida como la diferencia entre la superficie transformada en 2002 y la transformada en 1992. Se observa que en el período estudiado la expansión no fue homogénea ni se dio en todos los núcleos, sino que también estuvo concentrada en ciertos puntos. Pueden advertirse además diferentes patrones de “expansión”: en algunos casos hubo una “intensificación” o expansión interna, entendiendo por esto que dentro de un núcleo hay más superficie transformada, pero que el núcleo no expandió sus límites; mientras que en otros casos se produjo un expansión propiamente dicha, aparecen nuevas áreas transformadas en las periferias de los núcleos, que expanden de hecho sus límites.
Understanding biodiverity loss: an overview on forest fragmentation in South America 33
Figura 4. Áreas de expansión agropecuaria entre 1992 y 2002, expresada como porcentaje de hexágonos de 10.000 ha.
La principal zona de expansión resulta entonces de una combinación de estos procesos en las dos zonas climáticas del Chaco Subhúmedo (Figuras 4 y 5a). En el límite de Santiago del Estero con Chaco y Santa Fe, se observa por una parte una intensificación de la agricultura que se refleja en los tonos más oscuros de los hexágonos, pero también se ve un avance de la frontera agropecuaria hacia el Oeste.
En el Subhúmedo Occidental, los mayores valores de la expansión se registran en el Norte. La frontera Oeste de Santiago del Estero registra poca expansión, reflejando que ya estaba fuertemente ocupada en 1992, salvo en el límite Sur con Catamarca.
Dentro del Chaco Semiárido, hay algunos puntos importantes en Salta, cerca de la triple frontera con Chaco y Santiago del Estero. La información disponible indica que se trata de grandes campos ganaderos. Puede observarse que el área de irrigación de Santiago del Estero permanece con pocos cambios, con una ligera expansión hacia el Este.
Contrariamente a lo que podría pensarse, las menores tasas de expansión se registraron en el Chaco Húmedo. Las áreas de cultivos sobre albardones en Chaco y Formosa prácticamente no tuvieron crecimiento, mientras que el área de la Dorsal Oriental de Santa Fe, tuvo un pequeño incremento en intensidad, pero no expansión efectiva.
Zonas de riesgo climático
Understanding biodiverity loss: an overview on forest fragmentation in South America 34 Como se explicó, a partir del mapa de zonas climáticas se identificaron dos fajas críticas en cuanto la variabilidad de sus precipitaciones (Figura 5b). Estas fajas históricamente (registros climáticos de largo plazo), formaron parte de la zona semiárida, pero debido a los desvíos de precipitaciones de los últimos 25 años, pasaron a formar parte de la zona subhúmeda. Sin embargo, en los últimos 5 años hay fuertes evidencias de una reversibilidad hacia las condiciones semiáridas antecedentes.
La angosta faja occidental corresponde a la transición entre el Chaco Semiárido y el Chaco Subhúmedo Occidental, que recibe precipitaciones determinadas por la cercanía con las primeras estribaciones andinas. La faja oriental corresponde a la transición entre el Chaco Semiárido y el Chaco Subhúmedo Central, cuyas precipitaciones están originadas por las corrientes húmedas que ingresan al continente desde el Este.
Figura 5. Áreas transformadas según zonas climáticas (a) y zonas de riesgo climático (b).
a b
Los resultados en cuanto a superficies transformadas en cada una de las zonas y períodos que se muestran en la tabla 1, son sumamente preocupantes. Hay casi 1.300.000 ha en zonas de riesgo climático (la cuarta parte del total de las áreas transformadas), de las cuales aproximadamente la mitad corresponden a desmontes realizados a partir del año 1992. Teniendo en cuenta que en la gran mayoría de estos campos no se hacen rotaciones con ganadería, ni siquiera rotaciones con gramíneas como maíz y sorgo, y que los suelos supuestamente cultivados en siembra directa no tienen cobertura superficial y presentan evidencias de procesos erosivos (ODSMA-OEA, en preparación; experiencia personal), existe un riesgo muy fuerte de que se dispare un proceso de desertificación de consecuencias muy graves.
Understanding biodiverity loss: an overview on forest fragmentation in South America 35 Tabla 1. Superficies agropecuarias en zonas de riesgo climático.
Zona de Superficie Sup. Sup. Diferencia Expansión riesgo total (ha) transformada transformada (ha) climático 1992 (ha) 2002 (ha) Occidental 1.471.957 358.271 519.798 161.527 45,09 % Oriental 4.310.011 315.863 769.738 453.876 143,69 % Total 5.781.969 674.134 1.289.537 615.403 91,29 %
Perspectivas de conservación
Finalmente se analizó la ubicación geográfica de las áreas naturales protegidas (considerando las de la administración nacional y provincial, los Sitios Ramsar y las Reservas de Biosfera). Los resultados son preocupantes, puesto que virtualmente no existen áreas protegidas en las zonas con un desarrollo agrícola histórico ni en aquellas en las que se concentra la expansión actual (Figura 6). Resulta evidente entonces que esta red de áreas protegidas no es suficiente para garantizar la conservación de los ambientes nativos frente a la expansión agropecuaria. Peor aún, recientemente el gobierno de la provincia de Salta ha desafectado y vendido parte de la Reserva de Pizarro, única reserva vinculada con el núcleo de expansión de Las Lajitas, los lotes fiscales 32 y 33.
Figura 6. Sistema de áreas naturales protegidas en el área de estudio; se observa que es totalmente periférico a las zonas con desarrollo agropecuario.
Understanding biodiverity loss: an overview on forest fragmentation in South America 36 Discusión y conclusiones
La expansión agropecuaria en la región chaqueña tiene en la soja a su principal motor, ya sea por el cultivo de esta oleaginosa en la región, o por el asentamiento de emprendimientos ganaderos que se ven desplazados por la mayor rentabilidad de la agricultura en las zonas más favorables, como la región pampeana. Los indicadores internacionales señalan en forma consistente que la demanda de soja no es un fenómeno pasajero, ya que no sólo se mantendría en un largo plazo para granos, harinas y aceites, sino que además hay fuertes indicios de que se incrementaría significativamente.
Hay que considerar además el compromiso de muchos países (Estados Unidos, la Unión Europea, China, etc.) de alcanzar en el corto plazo distintas metas de incorporación de biocombustibles (biodiesel y bioetanol) como corte en las fuentes de energía fósil. Argentina aparece como uno de los países con potencial para aumentar significativamente su producción, no sólo por mejoras en los rendimientos, sino también por expansión de su frontera agrícola. Considerando sólo el mercado interno, de acuerdo a la ley 26.093 a partir del año 2010 la totalidad de los combustibles expendidos en el país deberán tener un 5% de biocombustibles en su formulación. Para el biodiesel, la demanda sería de unos 600.000 metros cúbicos, que de producirse a partir de soja implicarían unas 650.960 tn de aceite. Esta cantidad puede ser obtenida por la reducción de las exportaciones, o bien incrementando el área sembrada en una superficie de 1.815.176 ha. (ODSMA - OEA, en preparación) es decir, un área similar a la expansión registrada en este trabajo.
En la región todavía existen vastas superficies disponibles potencialmente para el uso agropecuario, especialmente en las fajas indicadas como zonas de riesgo climático, y su continuidad hacia el semiárido.
Hasta el momento esta expansión se ha dado en forma espontánea (sino anárquica) sin una regulación efectiva por parte del gobierno nacional ni de los gobiernos provinciales. Por lo general, cuando las normas regulatorias existen, los gobiernos no cuentan con la infraestructura necesaria o con la decisión política para garantizar su cumplimiento. Si bien es cierto que existen algunas experiencias y esfuerzos para avanzar en esta tarea, muchas de ellas impulsadas por organizaciones no gubernamentales, y también por organismos del estado como la Administración de Parques Nacionales o el Instituto Nacional de Tecnología Agropecuaria, éstas todavía no se han plasmado en acciones concretas sobre el terreno. De hecho el proceso de expansión estuvo regido por las leyes del mercado y la oferta de tierras, con todos los riesgos que eso implica.
El escenario planteado pone en evidencia que es imprescindible la implementación de un programa de ordenamiento territorial a escala regional. Desde el punto de vista ambiental, es necesario señalar que hay serios riesgos de desertificación, así como riesgos de pérdida de ecosistemas únicos, como es el caso del Quebrachal de tres quebrachos, que ya ha sufrido una reducción muy drástica en su superficie y conectividad (ver “Análisis multitemporal de la fragmentación y reducción del bosque de tres quebrachos” en esta misma publicación). Pero también es necesario que se prevea y regule la expansión de aquí en adelante, de forma de establecer prioridades ambientales, garantizando la conservación y la conectividad de los elementos más relevantes o amenazados, así como también el cumplimiento de los servicios ecológicos que prestan estos ambientes.
Como se señaló al comienzo de este trabajo, el crecimiento agropecuario es una gran oportunidad para el país y para la región chaqueña en particular. Por eso mismo, se impone un mínimo de racionalidad para evitar perjuicios para los productores, para el ambiente y en definitiva para el país.
Understanding biodiverity loss: an overview on forest fragmentation in South America 37 Bibliografía
Adámoli, J.; Ginzburg, R.; Torrella, S.; Herrera, P. (2004) Expansión de la frontera agrícola en la región chaqueña: el ordenamiento territorial como herramienta para la sustentabilidad. Gerencia ambiental 11 (112): 810-823. Adámoli, j. (2005) “Problemas ambientales de la agricultura en la región chaqueña” A. Brown, U. Martinez Ortiz, M. Acerbi y J. Corcuera (Eds). La Situación ambiental Argentina 2005, Fundación Vida Silvestre Argentina, Buenos Aires. Pp 436-442. Azcuy Ameghino, E. y Leon, A. (2005) La sojización, contradicciones, intereses y debates. Revista interdisciplinaria de estudios agrarios 23: 133-157. Bertonatti, C. y Corcuera, J. (2000) Situación Ambiental Argentina 2000, Fundación Vida Silvestre Argentina, Buenos Aires. Bianchi, A. (1981) Las precipitaciones en el Noroeste Argentino. Instituto Nacional de Tecnología Agropecuaria, Salta. Barsky, O. Y Gelman, J. (2001) Historia del agro argentino. Grijalbo-Mondadori, Buenos Aires Bruniard, E. (1987) Atlas geográfico de la Provincia del Chaco I El medio natural. Instituto de Geografía N° 5, Facultad de Humanidades. Universidad Nacional del Noreste. Dinerstein, E., Olson, D.M., Graham, D.J., Webster, A.L., Primm, S.A., Bookbinder, M.P., Ledec, G., (1995) Una Evaluación del Estado de Conservación de las Ecorregiones Terrestres de América Latina y el Caribe. Banco Mundial, Washington DC. Forman, R. Y Godron, M. (1985) Landscape ecology. Wiley and Sons, Editors. Galmarini, A. y Raffo del Campo, J. (1964) Rasgos fundamentales que caracterizan el clima de la región chaqueña. CONADE N° 9. 178 pp. Grau, R.; Gasparri, I.; Aide, M. (2005) Agriculture expansion and deforestation in seasonally dry forests of north-west Argentina. Environmental conservation, 32 (2):140-148. INDEC. Instituto Nacional de Estadísticas y Censos. Censos Nacionales agropecuarios. www.indec.gov.ar/nuevaweb/cuadros/11/d060104.xls INTA-INDEC (1994) Desarrollo Agropecuario sustentable. Eds. L. Verde y E. Viglizzo. Buenos Aires Montenegro, C.; Gasparri, I.; Manghi, E.; Strada, M.; Bono, J. y Parmuchi, G. (2004) Informe sobre deforestación en Argentina. Secretaría de Ambiente y Desarrollo Sustentable, Dirección de Bosques, Ministerio de Salud y Ambiente. 8pp Morello, J. y Adámoli, J. (1974) Las Grandes Unidades de Vegetación y Ambiente del Chaco Argentino. Segunda parte: Vegetación y ambiente de la Provincia del Chaco. Serie Fitogeográfica, 13. INTA, Buenos Aires. ODSMA - OEA (en preparación) “Evaluación regional del impacto de sustentabilidad de la cadena productiva de la soja”. Análisis ambiental. Prado, D.E. (1993) What is the Gran Chaco Vegetation in South América? I A review. Contribution to the study of flora and vegetation of the Chaco. V. Candolle, 48: 145-172. Reboratti, C. (2005) “Efectos sociales de los cambios en la agricultura”. Ciencia Hoy 15 (87), pp 52-61. SAGPyA. Estimaciones agrícolas. http://www.sagpya.mecon.gov.ar Torrella, S., Herrera, P. y Adámoli J. (2003) “Sostenibilidad de la expansión agraria en la región chaqueña: condiciones favorables y factores limitantes” 3ras Jornadas Interdisciplinarias de Estudios Agrarios y Agroindustriales. Buenos Aires. Viglizzo, E. (2001) “La trampa de Malthus: agricultura, competitividad y medio ambiente en el siglo XXI”. Editorial Universiaria de Buenos Aires, Buenos Aires
Understanding biodiverity loss: an overview on forest fragmentation in South America 38 Landscape changes due to native forest loss along a precipitation gradient in the Chaco region, Argentina
M. G. Parmuchi, J. Bono, M. Stamati, C. Montenegro, M. Brouver, E. Manghi and M. Strada
Unidad de Manejo del Sistema de Evaluación Forestal (UMSEF) – Dirección de Bosques – Secretaría de Ambiente y Desarrollo Sustentable de la Nación (Forest Evaluation System Management Unit - Native Forest Division - Secretariat for the Environment and Sustainable Development), San Martín 451 – (1004) Ciudad de Buenos Aires – Argentina [email protected], www.medioambiente.gov.ar/umsef
Abstract
From 1998 to 2002, an important landscape transformation occurred in the Chaco region as a consequence of the expansion of the agricultural frontier. It produced not only a loss of forest ecosystems but also an increase in fragmentation and a decrease in connectivity. Therefore, we evaluated landscape pattern changes in the Chaco region in Argentina along a precipitation gradient. We analysed three transects in the east-west direction, from the 1,100 to the 600 mm annual rain isolines, and established 6 sample units in each transect. We calculated landscape indexes for 1998 and 2002 using land coverage obtained by visual interpretation of satellite images. Land classes were Forest Land, Other Wooded Land and Other Land. Different landscape patterns were evident showing association with natural and anthropic factors. Deforestation is present in some samples situated between 700-900 mm but there is not a clear pattern associated with the rain gradient. In areas with high deforestation, indexes allow us to detect more obvious landscape changes. ______
Introduction
In Argentina, native forests cover approximately 30.300.000 ha and are the most biodiverse systems in the country. During the last years, the agricultural expansion mainly for the culture of soybean produced a significant conversion of native forest into anthropic ecosystems (UMSEF, 2006).
One of the most affected zones is the Chaco region, which is situated in northern Argentina and are dominated by xerophitic forests that cover almost 21.700.000 ha. It is the second largest native forest region in South America after the Amazon rainforest (Eva et al., 2004) and is considered internationally as a key area in terms of biodiversity conservation and a key area for the production of timber and non-timber goods and services (FVSA, 2006). However, the region lost 920.000 ha of forest between 1998 and 2002 with an annual rate of deforestation of –0.9 % (UMSEF, 2006), a process that continues at present. As a consequence, the region has suffered landscape structural changes that have resulted in decreased connectivity among patches, putting species at risk of extinction and modifying population dynamics.
Understanding biodiverity loss: an overview on forest fragmentation in South America 39
However, spatial changes are not homogeneous across the region since they respond to natural and anthropic variables that influence their probability of occurrence. Water availability determines whether or not land can be used for agriculture. For instance, those areas that receive less than 600 mm annual rain are not suitable for soybean crops in the Chaco region. Moreover, there is a rainfall gradient that historically defined areas under agricultural use (primarily cotton and wheat) in the region (Morello et al., 2005). In relation to anthropic variables, proximity to roads is considered an important factor in the land cover change analysis since it determines accessibility to forest and likelihood of replacement (Geist and Lambin, 2002).
It is important to take into account the variation in rain quantity and distribution because climate change, added to economic interests in promoting deforestation to convert land to agriculture, and technological improvements, could modify more intensively forest landscape and consequently their goods and services.
The development of policies and programs for conservation, restoration and sustainable use depends on management and land planning strategies based on information about forest status, changes and responses. In this context, the Native Forest Division (Dirección de Bosques) of the Secretariat for Environment and Sustainable Development of Argentina (Secretaría de Ambiente y Desarrollo Sustentable) is in charge of monitoring status of native forests through analysis of deforestation, fragmentation and their causes. For this, we consider the landscape ecology frame since its theories and methodological tools allow us to build an integrated approach (Forman and Godron, 1986; Gustafson, 1998; Turner et al., 2001; Lawrence et al.; 2000; Vos et al., 2001).
The objective of this work is to analyse landscape pattern status and changes in the Chaco region in Argentina along a precipitation gradient.
Study Area
The Chaco region comprises Humid, Semiarid, Arid and Low Mountain subregions according to climatic and geomorphologic criteria (Figure 1). The Humid subregion is characterized by annual rainfalls that vary from 750 to 1300 mm. The most important tree species are Schinopsis balansae (quebracho colorado chaqueño), Aspidosperma quebracho-blanco (quebracho blanco), Astronium balansae (urunday), Ziziphus mistol (mistol), Phyllostylon rhamnoides (palo amarillo) and several species of the genus Prosopis.
The Semiarid subregion is characterized by annual rainfalls from 750 to 1300 mm and is dominated by xerophyllous open forests where the main tree species are Schinopsis lorentzii (quebracho colorado santiagueño), A. quebracho blanco, Z. mistol, Bulnesia sarmientoi (palo santo) and several species of the genus Prosopis. Forests alternate with wetlands, prairies and palms (Cabrera, 1976; Red Agroforestal Chaco Argentina, 1999).
This study included the provinces of Chaco, Formosa, Santiago del Estero, Salta and Tucumán in Argentine.
Understanding biodiverity loss: an overview on forest fragmentation in South America 40 Figure 1. A) Study area, transects and samples, B) Republic of Argentina in South America and C) Study Area in the Chaco region in Argentina.
Methods
In this study we used the land cover classification carried out by UMSEF for 1998 and 2002 (UMSEF, 2002, 2003a, 2003b, 2004a, 2004b, 2005) through visual interpretation of Landsat 5 TM and 7 ETM images, at a scale of 1:50.000 with a minimum map unit of 10 ha. Land cover classes are Forest Land (FL), Other Wooded Land (OWL) and Other Land (OL) (Table 1). Land cover classes are based on crown cover and physiognomic features, according to Forest Resources Assessment (FRA-FAO, 2000) classification and adapted to Argentinean characteristics.
Table 1. Land Cover Classes Definition.
Land Cover Class Definition Forest Land Land with tree crown cover of more than 20 percent. The trees should be able to reach a minimum height of 7 meters at maturity in situ. It may consist of closed forest formation where trees of various storey and undergrowth cover a high proportion of the ground.
Other Wooded Land Land with tree crown cover of 5-20 percent able to reach a height of 7 m at maturity in situ or crown cover of more than 20 percent of trees not able to reach a height of 7 m at maturity in situ or with shrub cover of more than 20 percent.
Other Land Land not classified as forest land or other wooded land as defined above. It includes agricultural land, meadows and pastures, built-on areas, barren land, among others.
Understanding biodiverity loss: an overview on forest fragmentation in South America 41 In order to study the landscape pattern change, we set up three 500 km-long transects (north: N, center: C and south: S) in a west-east direction considering the rainfall gradient and in each transect established six equidistant circular samples of 100,000 ha (Figure 1). In each sample we calculated the following landscape indexes (Forman and Godron, 1986; Forman, 1995) for 1998 and 2002:
• FL, OWL and OL Area to document the presence and representativity of each class • Number of FL, OWL y OL Patches to characterize landscape configuration, in particular spatial heterogeneity • Mean Size & Largest FL Patches to characterize landscape configuration, in particular continuity • Mean Distance to Nearest FL Neighbor (MDNN) to characterize patch isolation • Deforested Area in the period 1998-2002 • Road Density to characterize accessibility • Annual Deforestation Rate (r) to characterize the rate of native forest loss in relation to forest area at the beginning of the period (Puyravaud, 2003).
In this study we used Arcview GIS 3.2.
Results & Discussion
Figure 2 shows samples for 2002 situated in the three selected transects. Each one presents a particular landscape pattern although in some cases samples share characteristics along the precipitation gradient. Samples in the east (number 6), that receive more than 1000 mm of annual rain, have an OL matrix with several FL small irregular patches while samples in the 700 to 900 mm isohyets shows a similar pattern but their forest patches present regular shapes (C1 and S4). Moreover, we distinguish a set of samples with less than 600 mm of precipitation with a FL matrix and OL or OWL perforations or gap formation (N2 and C2).
Figure 2. 2002 Samples of the North (N), Centre (C) and South (S) transects along a precipitation gradient (PP).
Understanding biodiverity loss: an overview on forest fragmentation in South America 42
In relation to deforestation between 1998 and 2002, we notice that samples C1 and S4 have the highest values (more than 6,000 ha) and also high annual deforestation rate (near -6.7 %) due to the small area covered by FL in 1998 (Fig. 3 and 4). They are located in areas with annual rainfalls above 700 mm. On the other hand, S3 has a large deforested area but lower annual deforestation rate (-1.7 %) because it preserves a large forest area in 1998.
Figure 3. Distribution of deforested areas in the different samples during 1998-2002 period along a precipitation gradient.
Figure 4. Deforested Areas (ha) and Annual deforestation rate.
Understanding biodiverity loss: an overview on forest fragmentation in South America 43 The number of FL patches in samples C1, S4 and C4 increased dramatically as a consequence of a forest fragmentation process in these areas. On the other hand, the number of FL patches in samples in the east and those with precipitation less than 600 mm do not change during the period (Figure 5).
Figure 5. Number of Forest Land Patches in 1998 and 2002.
Figure 6 illustrates different examples of typical and contrasting landscape pattern in the study area (N2, C1 and S6). N2 has low deforestation (conversion to cattle range) since it is placed in the area with the lowest precipitation and poor water access for irrigation. It is characterized by one large FL patch that covers almost the whole area and little variation between years, which causes a low matrix perforation. C1 is located in places where there are not water restraints to agriculture and thus presents high deforestation. This process causes an increase of fragmentation that is evident through new FL patches smaller than 250 ha in 2002 and shrinkage of high FL patches (1,000-10,000 ha). Since C1 is part of an area that was historically designated for agriculture, it presents an OL matrix and several small and irregular FL patches. Finally, S6 is situated in the east where annual rainfalls are above 1,000 mm and presents an OL matrix that corresponds to wetlands. They are designated for cattle range and are not useful for agricultural uses, and so they do not suffer deforestation or spatial changes.
Understanding biodiverity loss: an overview on forest fragmentation in South America 44 Figure 6. Examples of different land patterns.
We distinguish two cases when analysing changes in mean distances between FL patches and their nearest neighbor between 1998 and 2002: a decrease in samples N1 and S4 and an increase in C1, C3, S2 and S3 (Table 2). We detect that there are new patches as a result of dissection of the forest and also a patch size decrease during the period in most samples. The index changes according to the dominance of one of these processes. When fragmentation is the dominant process, the index decreases because there are new FL patches and less distance between them. Contrarily, when the main process is shrinkage, the index increases since there is the same number of patches but they are more isolated. Other samples such as C4 and S1 do not show changes in the index although they have changes between years but they are offsetting.
Table 2. Mean Distance to nearest neighbour (MDNN) Forest Land (FL) Patch in 1998 and 2002 (PP: precipitation ranges).
MDNN (m)
1998 2002
Sample PP (mm) FL FL N1 800-700 547.9 500.2 N2 <600 0.0 0.0 N3 600-800 0.0 0.0 N4 800-900 254.7 254.7 N5 900-1000 225.8 237.5
Understanding biodiverity loss: an overview on forest fragmentation in South America 45 N6 >1000 368.1 368.9 C1 800-700 309.3 370.0 C2 <600 0.0 0.0 C3 600-800 291.9 305.6 C4 800-900 265.6 264.5 C5 900-1000 184.3 184.3 C6 >1000 517.3 517.3 S1 800-700 663.4 660.4 S2 <600 634.4 679.3 S3 600-800 227.7 279.3 S4 800-900 468.5 456.2 S5 900-1000 470.2 470.2 S6 >1000 261.9 260.3
We are not able to calculate the index for samples N2, N3 and C2 because they are composed of only one FL patch that covers almost the whole sample (Table 3). However, N2 and N3 present some deforested areas which results in a matrix perforation and then in the decrease of the largest FL patch size (Table 3). Although C1 and S4 have a similar landscape pattern (a large OL matrix alternated with FL patches) and present a decrease in mean FL size due to high deforestation, both samples behave differently in relation to mean OL size (Table 3). In C1, forest fragmentation and reduction seem to be the dominant processes and they lead to an increase of OL matrix size without an outstanding change in the number of OL patches (from 17 to 13), increasing mean OL size. On the other hand, deforestation in S4 mainly produces a perforation of remnant FL patches causing an increase of the number of OL patches (from 19 to 34) and consequently a decrease in mean OL size. Moreover, these samples show a reduction of approximately 36% in the largest FL patch size (Table 3).
Table 3. Mean and Largest Patch Size in 1998 and 2002 (FL: Forest land, OWL: Other wooded land, OL: Other land, PP: precipitation ranges).
Patch Size (ha) Mean Largest 1998 2002 1998 2002 Sample PP (mm) FL OWL OL FL OWL OL FL FL N1 800-700 4844 731 1219 4197 731 1251 62075 60920 N2 <600 89101 936 168 87773 936 172 89101 87773 N3 600-800 97341 55 204 96873 55 209 97341 96873 N4 800-900 13748 784 462 13748 743 462 62225 62225 N5 900-1000 1743 113 181 1667 106 190 57039 56116 N6 >1000 310 455 1677 289 454 1735 6523 4201 C1 800-700 499 96 3993 279 56 5800 9500 6057 C2 <600 88138 1548 172 88138 1548 172 88138 88138 C3 600-800 888 1655 4060 865 1660 4077 17239 17201 C4 800-900 867 435 1025 730 435 922 20102 19383 C5 900-1000 194 23 1899 194 23 1903 8833 8784 C6 >1000 591 118 2438 591 123 2424 20181 20181 S1 800-700 603 316 992 568 298 992 5742 4324 S2 <600 4629 1382 2551 3889 1076 2715 26892 24118
Understanding biodiverity loss: an overview on forest fragmentation in South America 46 S3 600-800 2516 150 806 2266 150 729 63242 59560 S4 800-900 244 116 3787 175 116 2302 5823 3691 S5 900-1000 176 193 8895 166 191 9005 5706 5706 S6 >1000 69 89 10413 69 89 10412 853 853
The south-centre of the study area shows a high road density that is related to a historical agricultural use (S4) while the west-centre is characterized by a low road density (N3 and C3). Isolated areas have low probability of having important native forest loss. However, we do not find a direct relation between road density and deforestation (Figure 6). In general, roads are non-pavement and samples that have more pavement or consolidated roads do not show more deforestation.
Figure 6. Road Length & Deforestation (a) and Road types (b).
Understanding biodiverity loss: an overview on forest fragmentation in South America 47 Conclusions
Along transects different landscape patterns are evident showing association with natural and anthropic factors that are not at present necessarily associated with the rain gradient.
In the East, where precipitation exceeds 900 mm, there is a natural pattern characterized by an OL matrix (mainly wetlands) and a high density of small FL patches that do not vary in time. In the zones with precipitation lower than 700 mm, in general there is a FL matrix that presents OL perforations due to cattle range or old dry riverbeds. Between 700 and 900 mm, the most common pattern is an OL matrix with small geometric FL patches as a consequence of human pressure mainly for the development of agriculture that is evident during the analyzed period.
Deforestation is an evident process in some samples situated between 700-900 mm but there is not a clear pattern associated with the rain gradient although it is well-known that the 600 mm isohyet is a threshold for the culture of soybean.
Although it is expected that areas with precipitation above 700 mm show high deforestation, some samples present little or no change. In general, this behavior is due to the lack of infrastructure and the presence of wetlands. On the other hand, some deforested areas appear in zones with precipitation below 600 mm. They correspond to areas where shrubs have been eliminated for cattle grazing or those areas that have access to irrigation channels.
In areas with high deforestation, indexes allow us to detect more obvious landscape changes, that are reflected in an increase in FL patches fragmentation and reduction, affecting their connectivity.
Acknowledgements
To Ph. D. Andrew J. R. Gillespie for his useful suggestions.
References
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Understanding biodiverity loss: an overview on forest fragmentation in South America 49 Landscape mosaic, habitat structure and fragmentation of native forests at Córdoba mountain areas (Argentina central).
M. Menghi1 and R. del Sueldo
CONICET1, Centro de Ecología y Recursos Naturales Renovables (CERNAR), Edificio de Investigaciones Biológicas , Facultad de Ciencias Exactas, Físicas y Naturales- Universidad Nacional de Córdoba. Ciudad Universitaria, X5016GCA, Córdoba, Argentina. e- mail:[email protected]
______Abstract The growing of human pressures over mountainous areas that directly affects native ecosystems, in particular forests, is a major issue of world wide concern. In Central Argentina, the mountain region at Córdoba, native biodiversity is under serious threat due to exotics plantations, as well as urbanization that are expanding without an integrated plan that looks after a sustainable use of natural resources, including the visual landscape. The present work, focused mainly on the diagnosis and hypothesis exploration, it presents first results obtained from a current landscape spatial structure analysis, as well as habitat diversity and status study. Native woods of Polylepis australis “tabaquillo” at upper areas, and of Lithraea ternifolia “molle” toward lower ones, have both shown huge retraction and fragmentation into small isolated patches, as well as habitat alteration, which appear to be mainly related to traditional wood cutting, grazing livestock and burnt. Toward middle and lower sectors of the studied basin, the Pinus plantation caused high retraction and fragmentation of the Hetherotalamus alienus “romerillo” shrub-land. Recent intensification of land use increased the areas with complete decline of native woody strata. The interplay of Pinus plantation and/or urbanization, among others pressures, are increasing the rate of wood habitat loss and subsequently transforming the landscape. ______key words: elevation gradient, spatial variation, habitat status, boundary type, diversity
Introduction
Topography and rock substrate are major sources of landscape heterogeneity at Córdoba mountains (central Argentina). Different geological parent materials, their associated landforms and soil conditions, combined with climate are determinants of local and regional plant spatial patterns along an elevation gradient from 650 to 2950 m. a. s. l. (Luti et al.1979; Menghi et al., 1989; Acosta et al., 1992, among others).
Besides natural conditions, present and past human activities have also affected plant spatial patterns, with the increasing trend to substitute the native plant cover. The historical human-nature interaction through livestock rearing and wood-cutting based on native resources, as well as fire, were major sources of vegetation shaping up to 1950s. Related to those disturbances, the presence and abundance of some components were altered but the broad structure of native ecosystems has remained relatively unchanged from a physiognomic point of view, with exception of woods. Toward lower mountainous areas, wood cutting led to overgrowth of the less productive under-story spiny shrub species,. At the opposed elevation end, the same disturbance have mainly promoted the conversion of Understanding biodiverity loss: an overview on forest fragmentation in South America 1 50 woodlands to shrubby growth forms of the same tree species, as well as the wood habitat loss against the growing of secondary grasslands or rock outcrop areas. In agreement with recent global trends related to the advance of agriculture, tree plantation and urbanization over mountainous areas, the human impact over ecosystem and landscape components of Córdoba mountains has been growing in both, intensity and extension. The conversion of wood habitat to open grasslands or its replacement by exotic wood or by bare or paved soil (roads and residences) are pointed out. The expanding of Pinus spp. plantations, incipient 30 years ago (Menghi and Luti, 1982), has increased the loss rate of natives habitat as well as the disruption of extensive patches into small isolated ones, with notable impact on species diversity and on visual landscape.
The process, which has already started, is complex ought to the natural complexity of ecosystems as well as to the presence of actors and social elements interacting at different spatio-temporal scales. In addition, there are very few integrated studies done aiming to support the development of resources use programs. Present work, focused mainly to the diagnosis and hypothesis exploration, intended to stand out landscape ecological elements and interactions that are predominant for large geographical areas (Forman and Godron, 1986; Turner and Ruscher, 1988; Wiens et al., 1989; Turner, 1990; Turner and Gardner, 1991; Gardner and O´Neill, 1991; Matteucci, 1998, among others) as well as to gain insights into those proper of more detailed scales. In the present study, we analyzed current landscape spatial patterns, habitat status, the habitat and boundary variations along the elevation gradient, with emphasis on P. australis “tabaquillo” and of L. ternifolia “molle” native woods.
Study area
The study was carried out at “Los Reartes” water-basin river (31°50'S/64°50'W) (central Argentine) (Fig. 1). It involves an area of 77.622 ha varying in geomorphology, rock substrate and soil along an elevation gradient from 650 to 2400 m. a. s. l. (Ambrosino, 2000).
From a bio-geographical point of view, the basin is included in the Chaco province (Cabrera, 1976). The potential native plant cover is represented by more or less continuous belts with domain of wood of L. ternifolia “molle” (from 900 to1100 m. a. s. l.), of shrub-land of Hetherotalamus alienus “romerillo” and of grasses, from lower to upper sectors respectively (Luti et al., 1979). At the upper sector, the grassland domain could have a coarse-grained spatial pattern including patches of woodland and shrub-land of P. australis “tabaquillo” (over 1500 m. a. s. l.).
Historically, the economic activities were livestock rearing and wood cutting based on native plant resources. Currently, the Pinus plantation is displacing native vegetation over extensive areas (del Sueldo, 2004), and activities like tourism and urbanization are experimenting a sustainable growth.
Methodology
The landscape spatial structure analysis followed two main steps (Fig. 1 A, B). It was based on a subset of a Landsat TM image (pixel of 30 x 30 m) (Path Row 229/ 83) (Comisión Nacional de Actividades Espaciales, CONAE) dated September 2002, when the contrast between native (semi-deciduous) and exotic (evergreen) woods was enhanced. Geo- referencing (to the Gauss/Kruger projection), geometric and radiometric corrections were carried out through a digital elevation model based on topographic maps N° 3166-36-3 and N° 3166-36-4 (scale 1:50.000) (Military Geographic Institute) (Fig. 1 A). Seventeen control points at sites easily identifiable in the satellite image, were obtained with a Magellan 2000 XL GPS. Understanding biodiverity loss: an overview on forest fragmentation in South America 2 51 Land cover classes were separated by a physiognomic criteria based on authors previous experience, as well as on current field observations and image resolution, taking into account a minimum mapping unit of 3 x 3 pixels (0.81 ha). A supervised classification, based on a false color composition (bands 4, 3, 2) and the maximum probability algorithm (ENVI 3.5) (Chuvieco, 1996; Johnston, 1998), was carried out.
The first step included the analysis of all the basin (77.622 ha) (Fig. 1A) at mosaic and patch levels. Broad and clear land cover types at the resolution level of analysis were considered: wood of L. ternifolia “molle”; wood of Pinus spp. “pino”, agriculture, urban, artificial lake, shrub-land of H. alienus “romerillo”; shrub-bland of Acacia spp. “espinillo”, open grasslands. The upland grassland domain was classified like a mosaic ought to the mix of numerous small patches of “tabaquillo” (trees or shrubs), as well as of areas locally dominated by rock outcrop, by tall tussock grasses or by lawns, difficult to discriminate over extensive areas.
The second step (Fig 1 B) was focused on an area of 21.000 ha including extensive Pinus plantations, as well as on analysis at mosaic, patch (Fig. 1 B a) and boundary (Fig. 1 B b, c) levels. Sub-areas of 1.800 ha (N=12), three at each of four sectors (upper, middle upper, middle low, lower) along the elevation gradient (Fig. 1B, a) were selected, excluding the valley and piedmont. This analysis considered the same land cover types of first step, plus the local habitat variation of the grassland belt, previously not discriminated. The variety and number of boundaries (sensu Rescia et al. 1994) were registered within fifteen circles of 240 m of radio, regularly distributed at each area of 1800 ha (N= 180) (Fig. 1 B b, c).
Figure 1. Localization of the study area and simplified scheme of the methodological procedure.
Location of "Los Reartes" river basin
Landscape analysis
Based on the land cover maps obtained, the spatial structure at mosaic, habitat and patch levels (Fragstat 3.3) (Mc Garigal et al., 2002) was analyzed.
Understanding biodiverity loss: an overview on forest fragmentation in South America 3 52 Each habitat was analyzed considering its total area, number, patches density and mean area, area of largest patch. From an existing data base and current field observations (visual inspection and systematic sa mpling along interception lines of 60 m in length), the native habitat was also analyzed at plant community level. The number of strata; the maximum strata height, the plant cover; the floristic composition and the character (native or exotic) of dominant species were considered.
The habitat and boundary diversities of landscape mosaic, as well as their spatial variation along the elevation gradient, were calculated through the Shannon index (H´ = - Σ pi log pi) (Magurran, 1988). The richness (r) was the variety of the considered element (habitat, or boundary type). The relative abundance (pi) of each element was calculated from the percentage of area (habitat) or the number (boundary), represented in their respective totals within the analyzed area.
Results and discussion
The mosaic has shown predominance of rural landscape with a native plant cover matrix, which represents the 77 % of the total area. The matrix is made by grasslands, shrub- lands and woodlands varying in their conservation degree, with irregular edges and gradual transitions between adjoining habitat (Fig. 2, Table 1). It is suggested that this type of contact could function like a biological corridor (Taylor et al., 1993) and, from the high frequency (81%) detected all over the studied area (Table 4), it could provide connectivity to native ecosystems.
Figure 2. Map of land cover types of "Los Reartes" river basin
Understanding biodiverity loss: an overview on forest fragmentation in South America 4 53 Map references: M[G+Pol+R]: Mosaic of grassland (G), wood of Polylepis australis "tabaquillo" (Pol) and rock outcrop (R) (from 1.400 to 2.400 m a. s. l.).
Grassland patch (G) (30%). It includes tall tussock grasses (Deyeuxia hieronymi, Poa stuckertii, Festuca lilloi, F. hieronymi, Stipa pseudohichu), short sized grasses (Stipa neesiana, S. eriostachya, S. flexibarbata, Agrostis montevidensis, Eragrostis lugens, Piptochaetium montevidensis, Poa annua, Carex fuscula) and lawns (Alchemilla pinnata, Muhlembergia peruviana, Gentiana achalensis).
Polylepis patch (Pol) (10%), Woodland/shrubland of Polylepis australis "tabaquillo" (7 - 9 m) , Maytenus boaria "maiten", Berberis spp, Escallonia cordobensis, Heterothalamus alienus, Cassia hookeriana, Colletia spinosissima, Pernettya poeppigii, Baccharis spp., and ferns.
Rock patch (R) (60%). Predominance of rock outcrop with Stevia achalensis, Solanum incisum, Berberis hieronymi, Satureja odora, Blechnum penna-marina, Elaphoglossum gayanum y Polypodiun gilliesii, and grasses.
G: Grassland (from 1.000 to 1.800 m a. s. l.). The floristic composition could vary with altitude and the conservation degree. Deyeuxia and Festuca species are dominant upwards, and Stipa filiculmis, S. trichotoma, S. tenuissima, Paspalum dilatatum, P. notatum downwards. Schizachyrium spicatum, S. imberbe, Sporobolus indicus, Aristida spegazzini, among others, are common on less fertile and/or degraded soils.
Het: Hetherotalamus shrubland (from 1.100 to 1.400 m a. s. l.). It is a low sized shrubland (1-2 m) with predominance of Heterothalamus alienus “romerillo”, Eupatorium buuniifolium, Baccharis spp., over a grassland matrix. Emergent L. ternifolia "molle" trees, and/or spiny shrubs can be frequent downwards.
Aca: Acacia shrub-land (from 900 to 1.100 m a. s. l.). It is tall sized (2-5 m) shrub-land, closed or open, with predominance of spiny shrubs of Acacia spp. “espinillo”, Schinus spp. "moradillo", grasses and forbs. Isolated trees of L. ternifolia "molle" and Fagara coco "coco" can be observed accompanied by Colletia spinosissima, Condalia microphylla, Baccharis articulata, Geoffroea decorticans, Lippia turbinata, Lycium spp., among others.
Lit: Lithraea ternifolia "molle" woodland (from 900 to 1.100 m a. s. l.). The well preserved wood has a tree stratum ( 7 m) made by L. ternifolia, Fagara coco “coco”; Ruprechtia apetala "manzano del campo", accompanied by shrubs (Acacia spp., Schinus spp, Bacharis spp.), grasses and forbs. Towards lower areas the tree stratum could include Aspidosperma quebracho-blanco "quebracho blanco"; Prosopis spp. "algarrobo", Celtis tala "tala", Geoffroea decorticans "chañar", Jodinia rhombifolia "sombra de toro".
Pin: wood of Pinus spp. "pino". Plantation of Pinus elliottii, P. taeda and P. radiata (P. insignis).
Agr: Cultures. Zea maiz "maíz"; Solanum tuberosum "papa"; Sorghum spp., Avena spp., Eragrostis curvula "pasto llorón"; Secale cereale “centeno argentino”.
Urb: Urban. Villages, towns and secondary residences.
Lak: Artificial lake Los Molinos.
With exception of the complex mosaic ( M[G+ Pol + R] ) and the lake (Lak), most of the landscape units have shown numerous patches with averages areas ranging from 0.64 to Understanding biodiverity loss: an overview on forest fragmentation in South America 5 54 6.62 ha, as well as small large patch areas (Table 1). The "romerillo" shrubland (Het) was an exception showing a large patch which represented the 17.84 % of basin surface.
It is pointed out, the current small area of well preserved L. ternifolia "molle" (Lit) wood (3148 ha), which includes only one patch of 415 ha, two of 100 ha each one, and numerous small patches of less than 1 ha. This wood area represents the 19 % of its potential.
From the analysis by elevation sector (Table 2) emerged that the mosaic (M[G+ Pol + R]) (Fig. 2) is made by 30 % of grasslands (G), 10 % of "tabaquillo" woodland/shrub-land (Pol) and 60 % of rock outcrop. The detailed analysis of 5400 ha (Fig. 1B), revealed the presence of 282 ha of "tabaquillo" wood (Pol) fragmented in 1016 patches with an average area of 0.26 ha ± 0.45, or 0.14 ha ± 0.8, depending on the altitude (upper, or middle-upper sectors of the basin). The large part was related to ravines.
With respect to grassland habitat (G), the mapped area (Fig. 2) involves different community types and processes along the elevation gradient. The 30% of (G) reported for the mosaic at upper sector, includes semi-natural climax and well preserved grasslands and lawns (C), as well as degraded ones (del Sueldo, 2004). The contrasting proportion of (G) and (C) areas detected (Table 2), could be partially explained by the advance of degraded lawns against productive grasslands lost by burnt and overgrazing. In contrast, a large part of grassland habitat (G) mapped towards lower sectors (Fig. 2), is made by secondary communities including many exotic species, which are promoted by “molle” wood or shrub-land alteration.
Table 1. Spatial structure of land-cover types detected on "Los Reartes" river basin
Ta Np Lpi Map Land-cover types (ha) (%) (%) (ha) 1. M[G+ Pol + R] 19.252 (27) 85 24.76 226.5 2. G 4.060 (6) 3.731 0.61 1.09 3. Het 15.547 (22) 2.349 17.84 6.62 4. Pin 7.670 (11) 1.695 2.27 4.52 5. Aca 13.080 (18) 3.201 9.64 4.08 6. Lit 3.148 (4) 2.812 0.53 1.12 7. Agr 7.295 (10) 1.400 7.65 5.21 8. Urb 701 (1) 1.091 0.42 0.64 9. Lak 647 (1) 1 0.83 323.73
References: (M[G + Pol + R]: mosaic of grassland, Polylepis woodland and rock outcrop; G: grassland; Het: Hetherotalamus shrubland; Pin: Pinus plantation; Aca: Acacia shrubland; Lit: Lithraea woodland; Agr: agriculture; Urb: urban; Lak: artificial lake. Ta: total area ; Np: number of patches; Lpi: large patch index; Map: mean patch area.
The rest of basin surface (23 %) is covered by exotic and/or artificial habitat (agriculture, Pinus plantation, lake, urban) (Tables 1), which are more frequent towards lower areas (Fig. 2). This type of habitat has in common sharp and straight edge, as well as a simplified ecosystem structure. The landscape mosaic contrast is enhanced by the boundaries sharpness and the different nature of adjoining habitats (earth/water; native/exotic; natural/build, among others). Concerning organism movements, the described spatial structure, could be resistant to desirable native species, and at the same time permeable for opportunistic native and/or exotic ones.
Understanding biodiverity loss: an overview on forest fragmentation in South America 6 55
Table 2. Spatial structure of land-cover types at each sector of the elevation gradient of "Los Reartes" river basin
Land-cover types / Ta Np Lpi Map Sectors (ha) (%) (%) (ha) (SD) Upper 1. M[G + R] 444 (23.7) 946 2.61 0.47 (2.40) 2. C 385 (21.6) 929 7.98 0.44 (5.03) 3. Pol 235 (13.0) 674 0.43 0.26 (0.45) 4. G 198 (11.0) 733 0.26 0.27 (0.37) 5. Het 446 (24.7) 879 4.08 0.70 (3.42) 6. Pin 107 (6.0) 361 0.19 0.10 (0.20)
Middle Upper 1. M[G + R] 274 (15.3) 746 0.17 0.37 (0.40) 2. C 165 (9.18) 772 0.18 0.21 (0.24) 3. Pol 52 (2.86) 342 0.04 0.14 (0.08) 4. G 354 (19.7) 729 1.45 0.47 (1.54) 5. Het 780 (43.4) 551 19.2 1.53 (17.8) 6. Pin 175 (9.71) 149 0.64 0.88 (1.29)
Middle Low 1. G 587 (32.5) 322 23.5 0.50 (3.30) 2. Het 791 (44.0) 399 21.7 7.43 (65.4) 3. Pin 422 (23.4) 311 5.38 1.47 (6.56)
Lower 1. G 255 (14.3) 777 1.20 0.32 (1.07) 2. Het 552 (31.0) 842 6.27 0.65 (4.88) 3. Pin 296 (16.5) 198 7.44 1.26 (8.88) 4. Aca 555 (31.0) 790 7.40 1.03 (9.05) 5. Lit 109 (6.0) 345 0.48 0.10 (0.87) 6. Agr 33 (1.8) 41 0.06 0.53 (0.51)
References: (M[G + R]: mosaic of grassland and rock outcrop; C: lawn “césped”; Pol: Polylepis woodland G: grassland; Het: Hetherotalamus shrubland; Pin: Pinus plantation; Aca: Acacia shrubland; Lit: Lithraea woodland; Agr: agriculture). Ta: total area ; Np: number of patches; Lpi: large patch index; Map: mean patch area. SD standard deviation.
The native habitat retraction and fragmentation detected (Tables 1, 2), involved decrement of areas with and “interior effect” and increments of the length of boundaries and of the variety of contacts (Table 4) and therefore of "edge effects" . The dynamic and direction of the biotic transition promoted (Peters et al., 2006), as well as the regulating factors, vary along the elevation gradient. These factors would be mainly related to local biotic-abiotic interactions upwards, and to the intensity of human impact toward lower sectors.
With respect to species changes, it was observed that the selective logging and/or complete decline of “molle” trees led to the overgrowth of under-story spiny shrubs with predominance of Acacia spp. (Aca) (Fig. 2, Table 1), which would be promoted by competition process for light. At present, this secondary shrub-land covers the large part (13.300 ha) of the original area of the “molle” wood maintaining gradual transitions with the Understanding biodiverity loss: an overview on forest fragmentation in South America 7 56 remaining well preserved patches. These disturbed areas are being notably invaded by woody exotic species (Ligustrum lucidum “ligustro”, Pyracantha coccinea “crateus”, Gleditsia triacanthos “espina corona”, Melia azedarach "paraiso"; Prunus persica “durazno”, etc.), which have references of high aggressiveness against native plant species related to different disturbed native ecosystems (Vitouseck et al., 1997; Aragón and Morales, 2003; Casavecchia, 2005; Aimar et al., 2006). This process seems to agree, at least in part, with the uni-directional transition hypothesis (Peter et al., 2006), and it is suggested that the spontaneous recovering of the original tree species could be limited.
At the upper sector of the basin, man disturbance has mainly promoted the conversion of “tabaquillo” wood (Pol) to shrubby growth forms of the same species. Also a directional process was observed, causing woody habitat retraction against the growth of secondary grasslands, or eroded and rock outcrop areas. The exotics invasion would be minor at the upland grazed grasslands (Diaz et al., 1994), and, up to present, also at disturbed “tabaquillo” woods. That problem increases at areas higly transformed by man activity, like rural towns (Villa Alpina, La Cumbrecita) and roadsides.
The Pinus plantation (Pin), the most extensive exotic habitat (7670 ha) at the studied basin (Table 1), has replaced the native shrub-land of Hetherotalamus “romerillo” (Het) in the 40 % of its original area, semi-natural grasslands of Festuca and Stipa related to moderated relief and elevation, and a minor proportion of "molle" wood areas already altered (Fig. 2). The Pinus patches have extensive nucleus involving physical conditions adverse for native plant species, as well as 1.200 km of straight edges, sharp and about 20 m height. This type of edge itself, could influence organism movement (native or exotic) and function like a shelterbelt with microclimate effects over adjoining habitats (Ryszkowski, 1992).
Landscape diversity and dynamism
The values obtained of habitat diversity (H´1.85), equity (E 0.80) and dominance (0.34) suggest spatial pattern heterogeneity at the regional scale, with local spatial variation among the sectors analyzed along the elevation gradient (Table 3).
Table 3. Spatial variation of habitat diversity along the elevation gradient of "Los Reartes" river basin.
Sectors Parameters Upper Middle Middle Low Lower Upper Richness (r) 6 6 3 6 Diversity (H’) 2.34 2.18 1.53 2.20 Equity (E) 0.94 0.84 0.97 0.85
The boundary analysis (Table 4) pointed out a similar spatial pattern derived from the diversity index along the elevation gradient, the predominance of native/native boundaries and a clear trend of native/exotic contacts increase in proportion (from 16,7 to 50%) and frequency ( from 3.1 to 21.4) towards lower areas.
Understanding biodiverity loss: an overview on forest fragmentation in South America 8 57 Table 4. Spatial variation of boundary types, abundance and diversity index (H´) along the elevation gradient of "Los Reartes" river basin.
Sectors Upper Middle Middle Lower Boundary types Upper Low M[G + R] / Het 82 33 M[G + R] / C 112 68 M[G + R] / G 77 34 M[G + R] / Pol 42 14 Pol / Het 150 58 Pol / C 18 21 Pol / G 113 32 Pol / Pin 1* C / G 99 40 C / Pin 1* Het / C 74 47 Het / G 221 186 177 144 Het / Pin 10* 14* 46* 26* G / Pin 21* 10* 59* 50* Het / Agr 1* Het / Lit 7 Het / Aca 121 Aca / G 141 Aca / Pin 42* Aca / Agr 1* Aca / Lit 32 G / Agr 5* G / Lit 16
Total number of boundaries (N) 1019 559 282 586 Number of boundaries type (n) 12 14 3 12 Proportion of exotic/native boundaries (%) 16.7 28.6 66.7 50.0 Frequency of exotic/native boundaries (%) 3.1 4.7 37.2 21.4 Diversity (H’) 3.14 2.92 1.26 2.11
References: M[G + R]: mosaic of grassland and rock outcrop; Pol: Polylepis woodland; C: lawn “césped”; G: grassland; Het: Hetherotalamus shrubland; Pin: Pinus plantation; Aca: Acacia shrubland; Lit: Lithraea woodland; Agr: agriculture). The abundance values of contacts between natives and exotics land-cover units are enhanced *.
The structural landscape fragmentation detected would be mainly related to the interaction of natural factors upwards, and to man and natural factors downwards.
Uplands, the natural spatial variability in landforms like extensive areas with rock outcrops, ravines, undulated plains, among others, have shown to be a strong determinant of a cellular plant spatial pattern (Acosta et al., 1992). That pattern is concordant with the richness of habitat types and high frequency of sharp native/native boundaries detected (Table 4). According to the stationary transition hypothesis (Petter et al., 2006), the major coarse- grained spatial structure, typical of the landscape under consideration, could be relatively stable at regional scale, ought to strong control of ecosystem inherent abiotic constraints. Nevertheless, the human impacts, continuous over long periods, has been promoting boundary fluctuations and patch area changes, which have currently landscape evidences of Understanding biodiverity loss: an overview on forest fragmentation in South America 9 58 high concern. The retraction of well preserved wood (Pol) to ravines, and of grassland (G) areas against degraded lawns, as well as the increment of contacts with poor soil and/or vegetation conditions (ex. C/R; G/C, Pol /C, Pol/G, etc.) (Table 4) are mainly promoted by human pressures.
At middle parts of the elevation gradient traditional and new land-uses interactions could enhance landscape dynamics and plant diversity changes at areas relatively stable up to present. The disruption of extensive "romerillo" shrub-land (Het) areas, and of its ecotone with the grassland domain, by Pinus plantations, increased the boundary types and number (Table 4) as well as potential edge effects on the ecosystem (Table 4). The high frequency Het/G contacts, could increase the risk of shrubs invasion on overgrazed and/or burnt grasslands. Furthermore, the Pin/G and Pin/Het contacts could have shelterbelt effects on the adjoining native habitats and influence species composition and productivity.
Towards the valley and piedmont, the landscape transformation is closely related to man activities (roads, villages, agriculture, plantation, selective cutting, open grasslands, etc) involving drastic ecosystem changes (soil loss and pavement, alteration of seed bank, elimination of native species and introduction of exotic ones, retraction and isolation of native patches, etc.). These processes appear irreversible, and the "molle" wood seriously endangered.
Concluding remarks
The descriptions of plant spatial pattern and physiognomy carried out, agree with previous studies at neighbor mountainous areas along the elevation gradient (Menghi and Luti, 1982; Menghi and Intile, 1992) and at upper sector (Cingolani et al., 2004). The present study, also contributed information on the native habitats alteration degree at landscape level, and on boundary types and frequency.
With respect to native woods, the main factors determining current habitat status of “tabaquillo” wood (Pol), are still matter of research and discussion. On one part, there are field observations which support that ravines physical factors (microclimate and relief) would protect the adult and young trees against livestock grazing and fire, and at the same time could provide proper sites for seed germination and seedling growing (Renison et al., 2002). On the other hand, according to habitat models (currently in process), the current wood surface would be quite below its potential area, thus suggesting retraction promoted by human activities. Present results, contributed some evidence for the second hypothesis. Several indicators of burnt trees and/or wood cutting where detected at grassland and at Pol/G boundary areas. This transition, also suggested that the neighbor wood patch is particularly exposed to burnt by fire propagation, ought to frequent use of fire to manage grasslands.
At the opposed elevation end, the “molle” habitat alteration is clearly related to past and present human impacts. References for neighbor areas (Convenio UNC-CONEA 1976-1989; Menghi and Luti, 1982; Menghi and Cabido, 1986) pointed out wood retraction when the Pinus plantation activity was incipient at Córdoba mountain areas. Although present results have also shown that the "pino" plantation was not the major cause of loss of the "molle" wood habitat, this fact could be changing. The deforestation process with urbanization goal is also growing at different Córdoba mountain areas, with alarming references (Gavier and Bucher, 2004).
The native wood habitat loss, dramatic by itself, involves huge impact on wood ecosystem services like the native visual landscape, animal and plant diversity, water-shed and soil Understanding biodiverity loss: an overview on forest fragmentation in South America 10 59 protection, microclimate, among others, which are the bio-geophysical support for many human activities developed at Córdoba mountain and plain surrounding areas. The diversity and intensity of human actions, and the associated landscape transformation, was historically limited by physical factors at Córdoba mountain areas. This situation is currently changing due to recent access to powerful technology and external capital, and to local and regional people migrations looking for job opportunities and new life styles. The development of infrastructure and services (secure roads, hotels, filling stations, etc.) observed at areas considered of low accessibility up to few years ago, also represents a new focus of human impact on natives ecosystem and diversity.
Finally, we conclude that the present study, focused mainly on the diagnosis and hypothesis exploration, therefore help to point out a complex mosaic of native and exotic habitats with multiple edge effects, which are not static. The dynamism and direction of the biotic transitions related to the detected boundary types would affect in different ways the flow of organisms and material (Peters et al., 2006). In the same way, the landscape spatial pattern dynamics described here needs more research.
Acknowledgments This study was partially supported by the Secretary of Science and Technology (SECyT- UNC projects 2003-2004) and by the “Instituto de Altos Estudios Espaciales Mario Gulich (CONAE- UNC). We specially thanks to Dr. Marcelo Scavuzzo and to Lic. Mario Lamfri for their assessing and help with satellite image processing.
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Understanding biodiverity loss: an overview on forest fragmentation in South America 13 62 Patterns of land use change and forest fragmentation in the temperate forests in southern Chile
C. Echeverría1, D. A. Coomes2, A. C. Newton3, A. Lara1 and J. M. Rey-Benayas4
1Millennium Scientific Nucleus FORECOS – Instituto de Silvicultura, Universidad Austral de Chile. Casilla 567, Valdivia, Chile. e-mail: [email protected] 2Department of Plant Sciences, University of Cambridge, UK. 3School of Conservation Sciences, Bournemouth University, Poole, Dorset, UK. 4Departamento de Ecología, Edificio de Ciencias, Universidad de Alcalá, Alcalá de Henares, Spain.
Abstract The spatial patterns of land use change and forest fragmentation were assessed using multitemporal land-cover maps derived from satellite imagery in southern Chile. There was a reduction in forest area of 24% between 1976 and 1999, equivalent to an annual forest loss rate of 1.1%. However, the highest rate is recorded for the period 1976-1985 with 1.6%yr-1. Most of the native forest was converted to shrublands which, in turn, were replaced by crops pasture and agricultural lands. The native forests became increasingly surrounded by arboreus shrublands, which reveal the forest logging around fragment edges. A decline in size and core area and an increase in isolation of forest fragments were observed through time. Results revealed that land-cover change is leading to a transitional landscape, where the matrix of native forests is being converted to a matrix of non-forest elements. Fragmentation was associated with forest logging and expansion of agricultural land.
Key words: satellite imagery, landscape indices, forest matrix, forest loss, spatial configuration
Introduction
Few studies of the consequences of fragmentation on landscape configuration have been made in temperate forests (Staus et al., 2002). Although previous studies have examined the consequences of fragmentation on biological processes within the remnant habitats (Willson et al., 1994; Bustamante and Castor, 1998), it is also important to quantify the extent of fragmentation at the landscape level in order to establish the population-level effects of fragmentation.
Analysis of previous fragmentation studies revealed the need to analyze the matrix habitat to understand the dynamics of remnant fragments (Debinski and Holt, 2000). The interaction of patches with the surrounding matrix appears to be a determinant for species to be retained in isolated patches. If a native forest matrix shifts to a predominance of other land uses, then species dependent on a forest habitat are likely to be adversely affected (Formand and Godron, 1986). For instance, fragmentation of temperate forests in central Chile due to the expansion of a matrix of industrial pine plantations (Echeverría et al., 2006) has reduced the flow of biotic pollinators and dispersers among individuals of the endangered tree Nothofagus alessandrii (Bustamante and Castor, 1998).
The purpose of this study is to assess the temporal and spatial changes of landscape elements in a forested matrix influenced by forest fragmentation. In particular, I examined the
Understanding biodiverity loss: an overview on forest fragmentation in South America 63 spatial changes of land-cover types over time intervals using remotely sensed imagery. Forest fragmentation was assessed over time utilizing selected landscape indices.
Method The study area corresponds to 503,287 ha located between 41° 30’ S, 73° W and 42° 20’S S, 74° W in southern Chile (Fig.1). The zone is characterised by a rainy temperate climate with an oceanic influence and without dry-periods (Di Castri and Hajek, 1976), with a mean annual precipitation of 2090 mm. The landscape is dominated by Valdivian temperate rain forests, surrounded by crops and pasture lands. Approximately 40% of the remaining forest occurs on acidic, shallow, poorly-drained soil referred to as ñadis, which is classified as Gleysol. A similar percentage of native forest is found in volcanic ash soil, and the remaining portion corresponds mainly to marine sediments. The forests are characterised by the presence of several broad-leaved evergreen tree species. In some sites, the long-lived conifers such as Fitzroya cupressoides and Pilgerodendron uvifera (both Cupressaceae) can also be found. Anthropogenic disturbance has led to widespread early successional stages of the forest, which are characterized by a high abundance of young trees. In the middle of the 20th century a significant area of native forests was cut down and burnt as a result of European settlements.
Satellite data
To analyze the spatial and temporal changes of the land use types, a set of three Landsat satellite scenes were acquired at different time intervals: 1976 (Multispectral Scanner), 1985 (Thematic mapper), and 1999 (Enhanced Thematic Mapper). The smallest patches (less than 5 pixels) were removed from all the images. It was necessary to correct the images geometrically, atmospherically and topographically before they could be used to assess changes in forest cover and fragmentation (Chuvieco, 1996; Rey-Benayas and Pope, 1995). Supervised, maximum likelihood classifications were performed on each of the three images to classify the land cover types using training locations, obtained from field surveys. Each land cover map was validated using ground-based data. Overall agreement of classification was 88.8% for the 1976 MSS, 89.6% for 1985 TM image, and 91.9% for the 1999 ETM+ image.
Forest loss and landscape spatial pattern analysis
The resulting categories of land cover were grouped into forest or non-forest categories to create a binary forest/non-forest map. This map was analyzed using ARC VIEW 3.2 software1 and its extension Arc View Spatial Analyst 2.0 for Windows to quantify land cover change and forest loss. The formula used to determine the annual rate of deforestation was (FAO, 1995): − 1/(t2 t1) P = A2 −1 *100 (Equation 1) A1 where P is the percentage loss per year, A1 and A2 are the forest area at time t1 and t2 respectively.
Next, landscape spatial indices were computed using FRAGSTATS (version 3) (McGarigal et al., 2002). The following indices were calculated: a) mean patch size (ha), b) patch density (number of patches per 100 hectares), c) mean proximity index (ratio between
1 ESRI 1996-2000. Environmental Systems Research Institute, Inc. 380 New York St., Redlands, CA92373-8100, USA. Understanding biodiverity loss: an overview on forest fragmentation in South America 64 the size and proximity of all patches whose edges are within 1 km-search radius of the focal patch), and d) total interior forest area (total patch size remaining after removing a specific buffer edge of 100m) (ha).
Results
Forest loss
Approximately 23% of the native forests in 1976 had disappeared by 1999 (Fig. 1). During the whole study period, the annual forest loss was of 2,614 ha year-1, equivalent to 1.1%yr-1. Most of the forest loss was concentrated in the first nine years of the study period, at a deforestation rate of 1.6% yr-1, corresponding to 4,049 ha yr-1. In the second time interval, the rate decreased considerably to approximately 0.62% yr-1, equivalent to 1,341 ha yr-1.
The proportion of the landscape represented by native forests decreased gradually across the time periods. In 1976, this land cover type covered 266.8 thousand hectares, equivalent to 53% of the total study area (Fig. 1). In 1985, native forests represented 45% of the total area (Fig. 3.2), and the large fragments had been divided into smaller patches (Fig. 1). Fourteen years later, the area of native forests declined to 206.7 thousand hectares, which corresponded to 41% of the total land area (Fig. 1).
Approximately 16% of native forests (equivalent to 4,700 ha yr-1) in 1976 was transformed into shrublands and 8% (1,230 ha yr-1) were converted into crops and pasture and other land cover types in 1985. During the same time interval, 18% of shrublands was replaced by crops and pasture. Between 1985 and 1999, 18% (3,027 ha yr-1) of the native forests was converted to shrublands and 4% (equivalent to 660 ha yr-1) was cleared for crops and pasture. Of the total area of shrublands in 1985, 24% was converted to crops and pasture. Across the two study periods, 62% of the original forests remained as native forest and 29% was replaced by shrublands and 6% by crops and pasture. In both study periods, most of the area covered by crops and pasture was derived from shrublands. The rates of transition derived from the analysis of land cover change show the transition of a landscape dominated by native forests in 1976 (53%) to a landscape where non-forest land uses are becoming the dominant land cover type (60%).
Trends in forest fragmentation
Results showed an increasing proportion of the total area occupied by small patches over time. In 1976, 77% of the forest area was concentrated in patches of more than 10,000 ha. In 1985, patches larger than 10,000 ha had decreased to 47% of the total area, while the patches of less than 100 ha had increased to 20% of the total forest area (Fig. 2). In 1999, there was a considerable division of forest fragments which led to a 27% increase in the area occupied by patches less than 100 ha (Fig. 2).
The mean size of forest patches decreased significantly from 47 ha in 1976 to 24 ha in 1985 (Table 1). This decline in the patch size during the first time interval was associated with an increase in the patch density for the same period (Table 1). Patch density showed an increasing trend overall through time, reaching its maximum value of 0.65 fragments per 100 ha in 1999 (Table 1). The main change in the mean proximity was observed from 1976 to 1985, when the value decreased to almost one-fifth of its initial value (Table 1). Between 1985 and 1999, the mean proximity also demonstrated a significant decline. During these periods, the neighborhood of forest patches rapidly became occupied by areas of a different land cover type, as native forest patches became further apart and less contiguous in distribution. Interior areas in 1976 significantly declined in 1985. However, interior area of
Understanding biodiverity loss: an overview on forest fragmentation in South America 65 patches in 1985 did not present significant difference with those areas that were interior in 1999.
Understanding biodiverity loss: an overview on forest fragmentation in South America 66
1976 1985 1999
CHILE
ANTARTICA 90 º 53º
Figure 1.Temporal variation of the major land cover types over 23 years in southern Chile.
Understanding biodiverity loss: an overview on forest fragmentation in South America 67 Table 1. Temporal variation in landscape indices for the native forests in southern Chile. Kruskal-Wallis tests were applied to assess significant differences over time between indices estimated at the patch level Minimum and maximum values are given for mean patch size and mean proximity.
Landscape indices 1976 1985 1999 χ 2 χ 2 1 1 1976-1985 1985-1999 Mean patch size (ha) 47 (0.45 - 132,971.2) 24 (0.45 - 49,767.4) 19 (0.45 - 42,785.3) 201.2 *** 0.12 n.s. Patch density (n/100 ha) 0.36 0.60 0.65 - - Mean proximity 19,350 (0.0 - 4,380 (0.0 - 152,583.1) 2,552 (0.0 - 120,135.5) 168.9 *** 18.5 *** 369,603.5) Total interior forest area (ha) 143,428 89,007 69,900 3.85 * 0.45 n.s.
Understanding biodiverity loss: an overview on forest fragmentation in South America 68
150,000
100,000
50,000
- 123456789 150,000
100,000
50,000
Total forestTotal area (ha) - 0-100 100-500 500-1,000 1,000-2,000 2,000-5,000 5,000-10,000 10,000-20,000 20,000-100,000 >100,000 150,000
100,000
50,000
- 0-100 100-500 500-1,000 1,000-2,000 2,000-5,000 5,000-10,000 10,000-20,000 20,000-100,000 >100,000
Fragment size (ha)
Figure 2.Temporal variation of the forest fragment size for the years 1976, 1985 and 1999.
Understanding biodiverity loss: an overview on forest fragmentation in South America 69 DISCUSSION AND CONCLUSIONS
Deforestation and land use change
In the last three decades, the study landscape has experienced forest loss at an annual rate higher than that estimated between 1995 and 1998 further north to the present study area, in which the annual forest loss reached 0.3% (CONAF et al., 1999). Deforestation in other temperate forests has proceeded at a similar rate. In western Oregon, almost 20% of the forests were cleared between 1972 and 1995, giving an annual deforestation rate of 0.5%-1.2% (Cohen et al., 2002). In the same region, between 1972 and 1988, deforestation by clearcutting reached 1.2% of the entire study area including wilderness areas (Spies et al., 1994).
Changes in the native forest matrix
The pattern of change reveals the transition of the landscape dominated by a large contiguous extent of native forest to a landscape characterized by smaller and more isolated forest fragments surrounded by arboreus shrublands, shrublands, crops and pasture lands. This pattern was associated with an increasing adjacency between arboreus shrublands and native forest as a result of logging in the borders of forest patches. In contrast to this, in central Chile, native forests were primarily surrounded by arboreus shrublands at the earliest stage of deforestation (Echeverría et al., 2006). With progressive deforestation, pine industrial plantations largely dominated the neighboring areas of native forest patches.
The total area of crops and pasture land seemed to be relatively constant over the latter period, increasing by 9,000 ha in 14 years, while shrubland and other land cover types displayed much greater temporal variation. The results contrast with changes occurring in an agriculture landscape in southern Wisconsin, where the amount of natural vegetation (excluding forest) remained stable, and patches of forest appeared and disappeared at the average rate of about 4% per year (Forman and Godron, 1986). This stationary distribution (described as a shifting mosaic by Forman and Godron, 1986) was not found in the current study area.
The landscape element that has the greatest area constitutes the matrix (Forman and Godron, 1986). Up to 1976, native forests predominated both in terms of relative area and percent of aggregation. However, the land-cover change analysis revealed a substantial modification of the landscape composition associated with a lower dominance of native forests in the landscape across time. This pattern has led to a transitional landscape characterised by a matrix of shrublands and crops and pasture in 1999. This change in the matrix may lead to changes in the functioning of the landscape (Forman and Godron, 1986; Bennet, 2003 ), as some ecological processes associated with native forests can be altered by the dominance of non-forest land cover types. This new less desirable matrix may reduce the movement of many species from fragment to fragment, isolating gene pools and reducing local genetics (Barnes, 2000). A matrix of open lands surrounding forest patches may alter species composition (Echeverria et al., in press), increase the susceptibility of windthrow of existing trees, and allow for the invasion of exotic species (Barnes, 2000). In Central Chile, the rapid expansion of exotic-species plantations in the last decades has led to a dominance of this land-cover type in the landscape. This matrix of plantations has caused an isolation of the remnant forest patches which has affected the regeneration of shade-tolerant plants depending on biotic pollinators and dispersers (Bustamante and Castor, 1998; Grez et al., 1998).
Understanding biodiverity loss: an overview on forest fragmentation in South America 70 Spatial patterns of forest fragmentation
The reduction and division of the forest habitat is one of the recognizable results of the process of fragmentation (Ingegnoli, 2002). Owing to this, the size-frequency distribution of remnants in fragmented landscapes is strongly skewed towards small blocks (Bennett, 2003). With progressive forest loss and fragmentation, large forest areas were divided into smaller patches, which led to a decline in the mean patch size and a change in the patch size distribution by 1999. A similar pattern was observed in the highlands of Chiapas, Mexico between 1976 and 1996, in which the mean patch size and the number of fragments greater than 10,000 ha of dense forest decreased over time while small fragments increased (Ochoa-Gaona, 2001). In Bolivia, the fragmentation of the tropical deciduous forest showed a similar trend in which the smallest classes of patch size increased from 1977 to 1998 (Steininger et al., 2001). The progressive increase in these types of patches may have some negative consequences on the occurrence of certain species. For instance, in the Atlantic forest of Brazil, the smallest fragments (less than 50 ha) were associated with a diminishing mammal community as these fragments were below the carrying capacity of these species (Pardini et al., 2005).
The increase in the number of forest fragments also demonstrates that native forests of the study area were continuously affected by fragmentation over time. A similar situation was observed between 1975 and 1990 in central Chile, where the number of patches increased at higher densities from 0.93 to 1.65 (Echeverría et al., 2006) compared to 0.36 to 0.6 from 1976 to 1985 in southern Chile. In Limón, Costa Rica, the patch density reached to 0.4 in 1997 (Van Laake and Sanchez-Azofeifa, 2004), similar to the density recorded in 1976 by the present study.
The decline in the amount of large forest fragments with interior areas reveals the loss of large fragments with high quality habitat due to fragmentation. A very similar trend was described in Costa Rica in 1997, using the same distance from the edge, where large-scale deforestation for agricultural expansion has reduced the core area representing 34.8% of the total forest (Van Laake and Sanchez-Azofeifa, 2004). This decrease in area of large forest patches in the study area may jeopardise the survival of some species dependent on this type of habitat. For instance, some studies conducted in Chile have demonstrated that the reduction of large forest fragments with core area may have some influence the population of rodents (Donoso et al., 2003) and birds of temperate forest (Wilson et al., 1994; Cornelius et al., 2000, Vergara and Simonetti, 2004).
This study demonstrates that the progressive fragmentation and forest loss are associated with dramatic changes in the spatial structure of the temperate forest landscape in southern Chile.
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Understanding biodiverity loss: an overview on forest fragmentation in South America 73 Forest floristic inventory of Mixed Ombrophilous Forest and Deciduous Forest of Santa Catarina State, Southern Brazil: preliminary results.
A. C. Vibrans1; A. Uhlmann1; L. Sevegnani1; M. Marcolin1; N. Nakajima1, C. R. Grippa1 , E. Brogni2 & M. Braga Godoy1
1 Universidade Regional de Blumenau, Blumenau-SC, Brazil. Email: [email protected] 2 Fundação Municipal do Meio Ambiente de Itajaí-SC, Brazil.
Abstract
Forest floristic inventory of Santa Catarina State in Soutjern Brazil aimes quantification of forest resources and increase of knowledge on vascular plant diversity. The study includes sampling of trees, shrubs, herbs and vascular epiphytes. Santa Catarina State’s surface is about 95,443 km², covered by four major vegetation types: Dense Tropical Humid Forests, Mixed Ombrophilous Forests with Araucaria angustifolia, highland grasslands and Seasonal Deciduous Forests. Forest land cover is about 30%, although extremely fragmentized, specially in the western part. This pilot study has been realized in Mixed Subtropical Forest and Seasonal Deciduous Forest during 2005, measuring 74 sample plots (2000 m² each) randomly distributed in the forest type’s area (41,380 km²). Myrtaceae, Fabaceae, Lauraceae are the most numerous families, followed by Asteraceae and Solanaceae. There are a restricted number of very frequent tree species and a large number (120) of species occurring in less than 10 of the total of 74 sample plots; 46 species are present in only one plot, showing vulnerability of species diversity. Multivariate analysis starting from a species density matrix was executed in order to aggregate similar sample plots from the floristic point of view and identify species groups which preferentially occur together. The analysis shows two well distinguished plot clusters, one of them dominated by typical mixed forest species; the other cluster clearly dominated by supposed to be typical seasonal deciduous forest represents, although located in geographical region until now attributed to mixed forests. Other studies are needed to confirm this floristic tendency which could refine existent vegetation maps.
Introduction
The drastic reduction of the Forest cover of the state of Santa Catarina during the second half of the 20th century and the constant pressure exerted by the economic activities on the forest remains lead to the Secretary of Agriculture and Rural policies to formulate in 2003, the Project of the Forest floristic inventory of Santa Catarina State, according to the Resolutions of the Brazilian Environmental National Council, Brazil- CONAMA 278/2001 and 309/2002. The forest floristic inventory has as its main objective to catalogue and analyze the
Understanding biodiverity loss: an overview on forest fragmentation in South America 74 floristic composition and the horizontal and vertical structure of the forest remains of Santa Catarina state. The main purpose of the project is the need to record arboreal diversity of the forests and to collect data on the plant endangered species. Mixed ombrophylous forest (IBGE, 1990) or araucária forest (KLEIN, 1978) includes the forest types on the plateaus in the western regions of Rio Grande do Sul, Santa Catarina, Paraná, São Paulo, Rio de Janeiro and Minas Gerais states and also the northeast of Argentina and southeast of Paraguay, originally covering 400 thousand km2, today remaining less than 5%. The original covering of Santa Catarina was about 2/3 of the whole area of the state, mainly over high altitudes (more than 500m a.s.l.), on the Serra Geral plateau up to the border of Argentina. Physiognomically this type of forest had the emergent stratum composed by Araucaria angustifolia (Bertol.) Kuntze (reaching up to 45m tall). The canopy is located about 10 m below and it is mainly composed by large broadleafed species which composition varies according to the region but includes the Lauraceae (Ocotea porosa (Nees) Barroso, O. odorifera (Vell.) Rohwer, O. puberula (Rich.) Nees, Persea major Kopp, Cryptocarya aschersoniana Mez, Nectandra lanceolata Nees), Myrtaceae (Eugenia involucrata DC, E. pyriformis Cambess., Campomanesia xanthocarpa O. Berg), Meliaceae (Cedrela fissilis Vell.), Dicksoniaceae (Dicksonia sellowiana Hook), Winteraceae (Drymis brasiliensis Miers), and others (GAPLAN, 1986). Lumber exploitation, agriculture, cattle-raising, reforestation with exotic species (Pinus spp.) as well as the expansion of urban areas cities, are factors, which had caused in the past and in the present drastic reduction and fragmentation of this peculiar forest type. Seasonal deciduous forest – in the western part of Santa Catarina State - Brazil covered the side hills and affluents of the Uruguay River penetrating towards the north (GAPLAN, 1986). Typical species of the seasonal forest co-occurred with Araucaria angustifolia and all the other species of mixed forest, leading to Klein’s conclusion that the major portion of the western region of Santa Catarina constitutes a transition area between two phytogeographical regions (Klein 1979). This kind of forest shows remarkable characteristic leaves-fall above 50%, among the species, which compose the canopy. The low temperatures mark the climatic seasonality with an annual average below 15º C in the winter and above 20º C in the summer. The leaves-fall of the species of this forest is determined by a canopy dominated by deciduous leguminous depicting Apuleia leiocarpa McBride and Parapiptadenia rigida (Benth.) Brenan. also recording Patagonula americana Kuntze, Holocalyx balansae Micheli, Syagrus romanzoffiana (Cham.) Glassman, Cedrela fissilis Vell., Tabebuia spp., Cordia trichotoma (Vell.) Steud., Diatenopterix sorbifolia Radlk, Luehea divaricata Mart., Ocotea puberula Ness, Nectandra megapotamica Mez, among dozen others (KLEIN, 1978). The presence of many lumber species of great economic value and the clearing of the areas for the agricultural and cattle-raising purposes has stimulated the raising of cycles of economic exploitation which has lead to the deforestation accounting for only 3% of the remaining forests that nowadays are extremely fragmented.
Objective
The main task of this work phase is the inventory of forest remnants of mixed ombrophilous forest, as well as the forests located in the area of natural grasslands and the Understanding biodiverity loss: an overview on forest fragmentation in South America 75 seasonal deciduous forests, according to the phytogeographical map of Santa Catarina, Brazil (Klein, 1978).
Methods
Sixty temporary sample units were implemented and measured in order to survey the actual state of the vegetation. Fourteen permanent sample units were installed aiming to follow the forest dynamics. The sample units were randomly distributed in the mixed forest remnants identified through visual interpretation of Landsat 2003 images. The sample units distribution took place by a raffling system with 15.023 fragments bigger than 10 ha identified in the image mentioned above (Figure 1). Within each raffled fragment the sample units were allocated randomly. Each sample unit includes the following component: adult trees, thin trees, natural regeneration, herbs, epiphytes and shrubs. Each unit was subdivided in order to make a more appropriate sample as described in Figure 2: K – Herbs, shrubs, and natural regeneration with DBH < 1 cm; surveyed by using a 2 x 2 sample unit (4 m2); L – Natural regeneration with 1 cm ≤ DBH < 10 cm; surveyed by using a 5 x 5m sample unit (25 m2); 2 M – Thin trees with DBH ≥ 10 cm – subunits S1a of 10 x 25 m (250 m ); in this subunit epiphytes were inventoried by means of raffling 02 hostage trees; 2 N – Adult trees DBH ≥ 20 cm – subunits S1b of 10 x 25 m (250 m ); 2 O – Adult trees DBH ≥ 30 cm – subunits S2 of 10 m x 50 m (500 m ); 2 P – Adult trees DBH ≥ 40 cm - subunits S3 of 10 m x 50 m (500 m ); 2 Q – Adult trees DBH ≥ 50 cm - subunits S4 of 10 m x 50 m (500 m ).
In the permanent sample units, were measured all arboreal species above 10 cm DBH in all subunits (S1, S2, S3, S4).
Understanding biodiverity loss: an overview on forest fragmentation in South America 76 Mixed ombrophilous forest
Seasonal deciduous forest
Permanent sample units (14)
Temporary sample units (60) Grassland
Figure 1: Phytogeographical map of Santa Catarina state (Klein, 1978) with 60 temporary sample units and 14 permanent sample units within the mixed ombrophilous forest and (originally) grassland areas.
The collecting data of each individual measured includes: perimeter at breast high (PBH), commercial high, total high and stem quality. All the individuals that were imprecisely identified in the field were collected and taken to the laboratory to be corrected identified. The identified specimens were included in the collection of herbarium Roberto Miguel Klein of Universidade Regional de Blumenau and Herbário Barbosa Rodrigues (HBR).
S1 S2 S3 S4
S1a
S1b 1 RN<1cm Figure 2: Schematic representation of a sample unit and its subunits used in the Santa Catarina State Forest Inventory. For each forest fragment a physiognomic description was made aiming at characterizing the canopy and understorey dominant species, besides describing the relief forms, soil types, conservation state and apparent human disturbances. Understanding biodiverity loss: an overview on forest fragmentation in South America 77 A Correspondence Analysis (CA) was applied in a matrix built with the surveyed data. The matrix was composed by density data (columns) by species (lines). After these procedures which allow the segregation of the total sample units, also a phytosociological analysis of the segregated sets of sample units was conducted at that time. Results Floristic In 74 sample units allocated in the Mixed Forest 1 Pterophyta, 2 Coniferophyta and 179 Angiospermae shrubby-arboreal were collected, all of them belonging to 58 botanical families. Among those arboreal sampled species with DBH≥ 10 cm were found 133 species, being 25 exclusive for this forest type. The natural regeneration (individuals between 1 cm and 10 cm DBH) presents 101 sampled species (8 exclusive) and the herbs (DBH < 1 cm) included 105 species among them 10 exclusive. Among the 58 families surveyed Myrtaceae and Fabaceae were the richest showing 30 and 24 species, followed by Lauraceae (14 spp.), Asteraceae (10 spp.) and Solanaceae (7 spp.). The plant diversity found is lesser than those stated by Reis (1993), who cited 244 arboreal species in the mixed forest in Santa Catarina State as well as by RadamBrasil Project (IBGE, 1986) which listed 90 arboreal species in the states of Santa Catarina and Rio Grande do Sul. Regarding to the species frequency it has been observed that a restrict number of species has occurred in a great number of sample units, meanwhile a great number of species (120) has occurred in less than 10 of the 74 sample units and 46 species in only one sample unit (Figure 3). Araucaria angustifolia, Matayba elaeagnoides, Sapium glandulosum and Clethra scabra are constant species (Figure 4). 45 40 35 30 25 20 15 10 units sample of number 5 0 0 1020304050 arboreous species number Figura 3: Species frequency (species presence in sample units). Understanding biodiverity loss: an overview on forest fragmentation in South America 78 Araucaria angustifolia (Bertol.) Kuntze 41 Nectandra megapotamica (Spreng.) Mez 29 Matayba elaeaginoides Radlk. 23 Cinnamomum amoenum (Nees) Kosterm. 23 Sapium glandulosum (L.) Morong 22 Casearia decandra Jacq. 21 Clethra scabra Pers. 20 Ocotea pulchella (Nees) Mez 19 Capsicodendron dinisii (Schwake) Occhioni 19 Cryptocarya moschata Nees & Mart. ex Nees 19 Lithraea brasiliensis Marchand 19 Campomanesia xanthocarpa O. Berg 18 Zanthoxylum rhoifolium Lam. 18 Prunus myrtifolia (L.) Urb. 18 Figure 4: Major species according to their frequencies in the sample units. The results of correspondence analysis clearly showed two sets of groups of sample units: the first one, formed by typical species of mixed forest, and the latest with a strong influence of seasonal forest species (Figure 5). It is important to notice that the sample units included in the mixed forest group have been distributed all over the original area of this forest type in Santa Catarina state, as well as in the natural grasslands regions (Klein 1978) (Figure 6). This conclusion can be extended to the seasonal forest group. These results show the need of more survey efforts towards new data collection and assessment in order to verify the actual existence of a floristic differentiation throughout the borderline indicated in Figure 6. The next steps of this project can be useful to support this hypothesis. Understanding biodiverity loss: an overview on forest fragmentation in South America 79 3 27 909 905 2 31 901 transition SDF MOF 38 908 1 28 320 16 Axis 2 44 413 289 14 265 32 33 912 158 49 913 914 9 235 57 46 0 6 19 322 418 41 171 24 907 902 20 312 1 23 910 13 370 15 90347 256911 50 7 59 4208 55 26 162 17 131 90430 40 42 25392 43 53 132 10 906 -1 48 29 21 39 11 51 -1 0 1 2 3 4 Axis 1 Figure 5: Diagram depicting three sample units sets (mixed ombrophilous forest - MOF, seasonal deciduous forest – SDF and transition group) based on correspondence analysis (CA). Mixed ombrophilous forest Seasonal deciduous forest Transition Figure 6: Localization of forest remnants groups ordinated according to Correspondence Analysis (CA). Understanding biodiverity loss: an overview on forest fragmentation in South America 80 Table 1: Species aggroupment by correspondence analysis applied on the species by sample unit’s data matrix of the total survey of the Forest Floristic Inventory of Santa Catarina State. Segregated species in the ombrophilous mixed forest Lithraea brasiliensis Marchand Araucaria angustifolia (Bertol.) Kuntze Vernonia discolor (Spreng.) Less. Piptocarpha angustifolia DusTn ex Malme Ilex theezans Mart. Jacaranda puberula Cham. Capsicodendron dinisii (Schwacke) Occhioni Clethra scabra Pers. Lamanonia speciosa Weinmannia paulliniifolia Pohl ex Ser. Dicksonia sellowiana Hook. Sloanea monosperma Vell. Cinnamomum amoenum (Nees) Kosterm. Ocotea pulchella (Nees) Mez Myrceugenia euosma (O. Berg) D. Legrand Myrsine umbellata Podocarpus lambertii Klotzsch ex Endl. Styrax leprosus Hook. & Arn. Drymis brasiliensis Miers Segregated species in the seasonal forest group Dalbergia frutescens (Vell.) Britton Machaerium paraguariense Hassl. Myrocarpus frondosus M. Allemão Parapiptadenia rígida (Benth.) Brenan Persea venosa Nees Diatenopteryx sorbifolia Radlk. Undefined species Syagrus romanzoffiana (Cham.) Glassman Sapium glandulosum (L.) Moroni Ocotea porosa (Nees) Mez Cedrela fissilis Vell. Campomanesia xanthocarpa O.Berg Prunus myrtifolia (L.) Urb. Cupania vernalis Cambess. Casearia decandra Jacq. Understanding biodiverity loss: an overview on forest fragmentation in South America 81 Phytosociology Once the segregation process was taken place by the use of the CA procedures, three forest groups were formed: the mixed ombrophilous forest group (which included 48 sample units), the seasonal deciduous forest group (which included 21 sample units), and the transition group (which included 5 sample units). In the sequence below were presented the phytosociological analysis results of the two formerly cited groups. Mixed Ombrophilous Forest Araucaria angustifolia (Bertol.) Kuntze, Myrceugenia euosma (O. Berg) D. Legrand, Dicksonia sellowiana Hook., Cinnamomum amoenum (Nees) Kosterm., Lithraea brasiliensis Marchand, Capsicodendron dinisii (Schwake) Occhioni, Sapium glandulosum (L.) Morong, Matayba elaeaginoides Radlk., Ocotea pulchella (Nees) Mez and Ocotea puberula (Rich.) Nees, are the ten major species concerning to amplified importance value (in decreasing order). These species set represent 50% of the individuals with DBH bigger than 10 cm and also representing 56% of the total of the average basal area of the fragments. The other 92 species present low values of density and dominance totalizing the other half of the community. Forty species occur with less than two individuals per hectare, of which 26 with less than one individual. As it was expected Araucaria angustifolia represents 10% of individuals as well as 20% of the basal area therefore dominating the physiognomy of the forest. Seasonal Deciduous Forest In this group the analysis of the fragments structure concluded that the ten major importance value species are: Cryptocarya moschata Nees & Mart. ex Nees, Nectandra megapotamica (Spreng.) Mez, Machaerium stipitatum (DC.) Vog., Luehea divaricata Mart. & Zucc., Persea cf. venosa Nees, Machaerium paraguariense Hassl., Parapiptadenia rigida (Benth.) Brenan, Cupania vernalis Cambess., Ilex brevicuspis Reissek e Matayba elaeagnoides Radlk. These species totalize in average 40% of the total individuals with more than 10 cm of DBH per hectare and 46% of the total basal area that reached 25,82 m² ha-1. The low similarity between the two groups of fragments is also evident when comparing the 20 major species of both groups. Only Matayba elaeagnoides, Casearia decandra Jacq. and Prunus myrtifolia (L.) Urb. co-occured in both forest types. The great number of species with low importance values is a common characteristic in the two forest types mentioned above: 55 species (54%) of the 102 total surveyed species found in araucária forest and 34 species (36%) of the 94 total surveyed species found in the seasonal forest showed an importance value lesser than 2. This is a consequence of very low values of density and dominance, as well as the very low frequency of these species in the surveyed forest fragments. Major conclusions • The very low frequency of a large number of species may depict a high level of vulnerability of these, or to the populations of these species according to the continuous Understanding biodiverity loss: an overview on forest fragmentation in South America 82 of fragmentation and degradation of the forest remnants of the araucária forest biogeographical region. • As the average index of forest remnant coverage in Santa Catarina state is only about 5% (SOS MATA ATLÂNTICA, 2004) it is possible that the impact of suppression and fragmentation upon the plant species is very high. The fragmentation possibly allowed the occurrence of species that previously did not exist or presented very low densities (LAURANCE, 1997, KAPOS et al., 1997). • The detected factors that lead to reduction, simplification and alteration of the forest remnants are mainly: the selective species extrativism, burning, the cutting of lower forest strata, the cattle-raising and the agriculture, as well as the exotic species reforestation. This structural and floristic simplification of the forest remnants jeopardizes their maintenance also allowing the invasion by other species, which become dominant. • In order to support the structural and floristic simplification of the forest, it is convenient to observe that according to REITZ et al. (1979), Araucaria angistifolia was a canopy dominant species. According to the data of this work the species shows low frequency which is a strong argument supporting the conclusion that the forest remnants are very degradated • One can be observe that the set of species include a great number of pioneer species. The alteration and degradation of the forest structure allow the invasion of other species from different regions (LAURANCE, 1997, KEANE & CRAWLEY, 2002; WOLFE, 2002, ESPÍNDOLA et al., 2005). The environmental disturbance opens for the possibility of the species overcoming the biogeographic barriers and the colonization of new regions, which may provoke a mixture of flora that has been originally segregated (VERMEJ, 1991). • The quality and the quantity of forest remnants of Santa Catarina state in the central- western region and in the interior plateaus demand concern, mainly due to the isolation and the continuous pressure provoked by anthropic activities (agriculture, cattle-raising and reforestation with exotic species) in the surrounding areas, which can cause strong influence upon the genetic flow and the microclimate. Acknowledgments: The authors are grateful to Marta Helena Cúrio de Caetano for translation support. References ESPÍNDOLA, M., B., BECHARA, F. C., BAZZO, M. S., REIS, A. (2005). Recuperação ambiental e contaminação biológica: aspectos ecológicos e legais. Biotemas 18 (1): 27- 38. GAPLAN. (1986) Atlas de Santa Catarina. Aerofoto Cruzeiro, Rio de Janeiro. IBGE. (1990) Geografia do Brasil: Região Sul. Rio de Janeiro, IBGE. KAPOS, V. WANDELLI, J. L. C., GANADE, G. (1997) Edge-related changes in environment and plant responses due to forest fragmentation in Central Amazonia. In: W.F. LAURANCE, R. Jr. BIERREGAARD (Eds). Tropical Forest Remnants. University Chicago, Chicago, pp. 33-44. Understanding biodiverity loss: an overview on forest fragmentation in South America 83 KEANE, R. M., CRAWLEY, M. J. (2002) Exotic plant invasions and the enemy release hypothesis. Trends in Ecology and Evolution, 17: 164-170. KLEIN, R. M. (1978) Mapa fitogeográfico de Santa Catarina. Flora Ilustrada Catarinense (FIC), Itajaí. LAURANCE, W. F. (1997). Hyper-disturbed parks: edge effects and the Ecology of isolated rainforest reserves in tropical Australia. W.F. LAURANCE, R. Jr. BIERREGAARD (Eds). Tropical Forest Remnants, University Chicago, Chicago, pp. 71- 83. SOSMATATLANTICA. (2004) Atlas da Mata Atlântica. www.sosmatatlantica.org.br. Accessed on: 15th March 2005. VERMEJ, G. J. (1991) When biotas meet: understanding biotic interchange. Science, 253: 1099-1104. WOLFE, L.M. (2002) Why alien invaders succeed: support for the escape-from-enemy hypothesis. American Naturalist , 160: 705-711. Understanding biodiverity loss: an overview on forest fragmentation in South America 84 Part II Ecological consequences of forest fragmentation Understanding biodiverity loss: an overview on forest fragmentation in South America 85 Habitat fragmentation effects on insect herbivory in Chaco Serrano woodlands M. Videla, L. Cagnolo, G. Valladares, A. Salvo and S. Fenoglio Centro de Investigaciones Entomológicas de Córdoba (C.I.E.C.), Facultad de Ciencias Exactas, Físicas y Naturales, Universidad Nacional de Córdoba, Av. Vélez Sársfield 299, Córdoba (X5000 JJC), Argentina. e-mail: [email protected] Abstract Loss of species and changes in community composition derived from habitat fragmentation could affect interactions among species and ecological processes. Herbivory is a key process for energy and nutrient transfer to higher trophic levels, nutrient cycling and plant competitive balance. Despite its importance and widespread study in other contexts, herbivory have rarely been investigated in relation to habitat fragmentation. We compared insect herbivory rates in continuous and fragmented woodlands, separately considering damage by different feeding guilds. Herbivory, mostly by leaf chewers, was lower in remnants than in the original woodlands. This result was not attributable to changes in plant species composition or in relative damage by the different guilds. We have provided the first evidence of habitat fragmentation effects on overall herbivory rates at the plant community level. The consistency of these effects and the underlying mechanisms deserve further study. Keywords: Chaco Serrano woodland; Habitat fragmentation; Herbivory; Insect feeding guilds. Introduction Habitat fragmentation has large, consistently negative effects on biodiversity, mostly due to the loss of habitat (Fahrig, 2003). In plant - insect herbivore systems, loss of species (Steffan- Dewenter and Tscharntke, 2000; Cagnolo et al., 2006) and changes in community composition (Sumerville and Crist, 2003) have been linked to habitat loss. Such changes could affect interactions among species and ecological processes (Saunders et al., 1991). Herbivory is a key process for energy and nutrient transfer to higher trophic levels (Cebrian, 2004), nutrient cycling (Hunter, 2002; Chapman et al., 2003) and plant competitive balance (Carson and Root, 2000; Dyer et al., 2004). Despite their importance and their widespread study in other contexts, herbivory rates have rarely been investigated in relation to habitat fragmentation (Didham et al., 1996; Tscharntke and Brandl, 2004). The few available studies show inconsistent results, with herbivory decreasing (Arnold and Asquit, 2002; Kruess, 2003; Valladares et al., 2006), increasing (Rao et al., 2001) or being unaffected by habitat reduction (Benitez-Malvido et al., 1999). These studies consider herbivory rates on one or a few plant species or by particular guilds of insects. To our knowledge, habitat fragmentation effects on overall insect herbivory of a diverse plant assemblage have not yet been assessed. Here we compare leaf damage on the vegetation of continuous vs. fragmented Chaco Serrano woodlands. Besides analysing aggregate herbivory patterns, we have separately considered damage by different insect feeding guilds, since these can show differential responses to Understanding biodiverity loss: an overview on forest fragmentation in South America 861 environmental conditions (Mazía et al., 2004) and their relative representation can be affected by habitat size (Sumerville, 2004). Methods We studied an area in Central Argentina (31o 10’S to 31o 30’S and 64 o 00’W to 64 o 30’W), with an altitude of about 700m, originally covered by Chaco Serrano Woodland. Details of vegetation structure and composition can be found in Cabido et al. (1991). The Chaco Serrano has suffered a ninety-four percent decrease in cover over the last 30 years, evolving in a fragmented landscape with patches of native vegetation isolated within an intensely managed matrix resulting from human activities (Zak and Cabido, 2004). This system offers a “ready- made experiment” for insularization effects (Saunders et al., 1991), in which the frequent application of herbicide and pesticides enhances the unsuitability of the matrix for forest species. Four remnants (0.7ha - 12 ha) and four sites, distant at least 1000 m from each other, within continuous woodlands (over 1000 ha) were selected after analysis of digital satellite images (Landsat TM 1997). Three transects (10 m long) were run at each site, distant 10-20 m from each other and at least 10 m from the woodland edge. Along each transect, all leaves in contact with the transect line were collected on 12-16 April 2004, towards the end of the growing season, in order to assess cumulative herbivore damage. From a random subsample of 100 leaves for each transect (300 leaves per site), herbivory rates were estimated as frequency of damaged leaves and as amount of damage. The latter variable was calculated by assigning leaves to damage categories 0 = undamaged; 1= 1-5%; 2 = 6-12%; 3 = 13-25%; 4 = 26-50% and 5 = 51-100% (Domínguez and Dirzo, 1995). A herbivory index (HI) was calculated for each transect as: HI = sum xi.ni/N where xi = central value in damage category, ni = number of leaves within the category and N = total number of leaves (Benitez-Malvido et al., 1999). Leaf damage from different feeding guilds, including chewers, miners and sap-suckers (Gullan and Cranston, 1994), was separately measured. This index is commonly employed in herbivory studies and closely reflects levels of damage estimated from quantitative scanning methods (Rodríguez-Auad and Simonetti 2001). In order to consider possible effects of floristic composition on herbivory patterns, a vegetation census was carried out by recording all plant species in a 1000 m2 quadrat at each site and assigning each plant species to one of the following cover abundance categories: <1%, 1-5%, 6-25%, 26-50%, 51-75% and >75% (Braun–Blanquet, 1950). Mean (from 3 transects) frequency and amount of leaf damage, total and for each feeding guild, as well as the proportion of total damage attributed to each feeding guild, were compared between remnants and continuous woodlands by means of t test. We also performed Pearson´s correlation analysis to explore possible relationships between total herbivory levels (frequency and amount of damage) and the damage inflicted by each feeding guild. Percentage data were arcsine square root transformed prior to statistical analyses. Changes in floristic composition between remnants and continuous woodlands were investigated with MANOVA. Results On average, slightly over fifty percent of 2400 leaves from 48 plant species in the Chaco Serrano woodlands suffered some kind of insect herbivory, with approximately eight percent Understanding biodiverity loss: an overview on forest fragmentation in South America 872 leaf area being affected (Appendix 1). The frequency of damaged leaves (Fig. 1a) in fragmented and continuous habitats was similar (t = 1.38, df = 3, P = 0.21). However, habitat fragmentation resulted in lower overall herbivory levels (t = 3.23, df = 3, P = 0.018) as leaf area lost due to herbivores in remnants decreased by nearly 30% in comparison to the original woodland (Fig. 1b). Figure 1. Mean (n=4) herbivory rates, as measured by (a) number of leaves and (b) percentage leaf area damaged by different insect guilds, on vegetation of continuous and fragmented Chaco Serrano Woodlands. The bars represent ± 1 SE. Different letters indicate significant differences at P < 0.05. (a) (b) 70 10 a s a 9 e 60 ) v a a 8 % e ( b l 50 d d 7 e e g g sap-sucker a a 40 6 m m a a miner 5 d d 30 f a 4 o chewer e r r a e 3 20 f b a m e 2 u 10 L N 1 0 0 continuous fragmented continuous fragmented Woodland type Woodland type Damage by chewers accounted for most of the recorded damage (Fig. 1) and was highly correlated with overall herbivory levels, regarding both number of damaged leaves (P = 0.001, r = 0.93) and amount of damage (P = 0,00024, r = 0,95). This was also the only type of damage showing a significant relationship with habitat area, with fewer leaves (t = 2.50, df = 3, P = 0.046) and lower amounts of leaf area (t = 3.04, df = 3, P = 0.022) being consumed in fragmented than in continuous woodlands. The relative contribution of chewers to the total damage did not vary with fragmentation, whether frequency (t = 1.66, df = 3, P = 0.14) or amount of leaf area removed (t = -0.001, df = 3, P = 0.99) were considered. Community plant composition was similar in fragmented and continuous woodland ( = 0.28, F = 0.64, df = 1, 4, P = 0.72). Discussion Our results show that habitat loss due to fragmentation leads to a reduction in overall insect herbivory rates across the vegetation of Chaco Serrano woodlands. This trend seems to be driven mainly by a decrease in the intensity of damage inflicted by chewers, which explained most of the observed damage. It must be noticed that actual damage inflicted by sap-suckers can be much greater than suggested by the perception of small marks revealing their feeding, and could be underestimated by a direct comparison with amount of leaf area removed by Understanding biodiverity loss: an overview on forest fragmentation in South America 883 chewers (Landesberg and Ohmart 1989). For this and other types of damage such as mining, a different sampling approach or larger sampling effort may be necessary to obtain a more accurate assessment of habitat fragmentation effects, as suggested by the fact that a separate study in the same region, in which thousands of mined leaves were analysed, found a significant reduction in the number of mined leaves along a gradient of woodland area (Valladares et al 2006). Changes in herbivory rates could result either from changes in folivore species abundance and/or community composition. Just as relative species richness and abundance of different herbivore guilds can be altered by habitat fragmentation (Sumerville and Crist 2003; Sumerville 2004), the decrease of leaf chewing damage in remnants could result in changes in the relative contribution of this feeding guild to total herbivory. However, no such changes were observed, and chewing damage kept a steady representation (about 80% of total damage) in both, fragmented and continuous woodlands. Different patterns of herbivory could also arise from actual changes in feeding behaviour of herbivore insects, maybe as a consequence of changes in vegetation composition, palatability, etc. Fragmentation has been shown to favour establishment of fast- growing plants (Janzen 1983; Meiners and Pickett 1999), which can be more palatable than slow-growing species (Landesberg and Ohmart 1989; Perez-Harguindeguy et al 2003). In the present case, the observed changes in herbivory rates did not seem related to changes in plant composition, since vegetation species in woodland remnants did not differ from their continuous counterparts. The higher trophic levels of food webs might be the most strongly affected by habitat fragmentation, leading to herbivore release from their natural enemies and favouring herbivore outbreaks (Denys and Tscharntke 2002). Such scenario could result in increased herbivory in small habitats (Kondoh 2003), a possibility that was not supported in this study. Our results have shown a reduction in overall insect herbivory rates across the vegetation of Chaco Serrano woodlands, as a consequence of habitat loss. Given the importance of herbivory in ecosystem diversity, productivity and long-term sustainability (Chapin et al 1996; Crawley 1997; Mulder et al 1999), the impact of habitat fragmentation on this ecological process should be further assessed on a variety of systems. The underlying mechanisms need also be understood. Acknowledgements We are grateful to M. Cabido and M. Zak for helping in site selection and plant census. This research was supported by CONICET, FONCYT and National Geographic Society, and CONICET grants to M. Videla, L. Cagnolo and S. Fenoglio. References Arnold, A.E. & Asquit, N.M. (2002) Herbivory in a fragmented tropical forest: patterns from islands at Lago Gatún, Panama. Biodiversity and Conservation 11: 1663-1680. Benitez-Malvido, J; García-Guzmán, G. & Kossmann-Ferraz, I.D. (1999) Leaf-fungal incidence and herbivory on tree seedlings in tropical raiforest fragments: an experimental study. Biological Conservation 91: 143-150. Braun-Blanquet, J. (1950) Sociología Vegetal. ACME, Buenos Aires. Cabido, M; Carranza, M.L; Acosta, A. & Páez, S. (1991) Contribución al conocimiento fitosociológico del Bosque Chaqueño Serrano en la provincia de Córdoba, Argentina. Phytocoenología 19: 547-566. Understanding biodiverity loss: an overview on forest fragmentation in South America 894 Cagnolo, L; Cabido, M. & Valladares, G. (2006) Plant species richness in the Chaco Serrano Woodland from central Argentina: Ecological traits and habitat fragmentation effects. Biological Conservation 132: 510-519. Carson, W.P. & Root, R. (2000) Herbivory and plant species coexistence: Community regulation by an outbreaking insect. Ecological Monographs 70:73-99. Cebrian, J. (2004) Role of first-order consumers in ecosystem carbon flow. Ecology Letters 7: 232-240. Chapin, F.S; Torn, M.S. & Tateno, M. (1996) Principles of ecosystem sustainability. American Naturalist 148: 1016-1037. Chapman, S.K; Hart, S.C; Cobb, N.S; Whitham, T.G. & Koch G.W. (2003) Insect herbivory increases litter quality and decomposition: an extension of the acceleration hypothesis. Ecology 84: 2867-2976. Crawley, M.J. (1997) Plant-herbivore dynamics. Crawley M.J. (Ed). Plant Ecology, Blackwell Scientific, Oxford, pp. 401-474. Denys, C. & Tscharntke, T. 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(1999) Effects of area, age and diversity of forest patches in Belgium on plant species richness, and implications for conservation and reforestation. Biological Conservation 87: 73-84. Hunter, M.D. (2002) Landscape structure, habitat fragmentation, and the ecology of insects. Agricultural and Forest Entomology 4: 159-166. Janzen, D.H. (1983) No park is an island: increase in interference from outside as size decreases. Oikos 41: 402-410. Kondoh, M. (2003) Habitat fragmentation resulting in overgrazing by herbivores. Journal of Theoretical Biology 225: 453-460. Kruess, A. (2003) Effects of landscape structure and habitat type on a plant-herbivore- parasitoid community. Ecography 26: 238-290. Landesberg, J. & Ohmart, C. (1989) Levels of insect defoliation in forests: patterns and concepts. Trends in Ecology and Evolution 4: 96-100. Mazía, C.N; Kitzberger, T. & Chaneton, E.J. (2004) Interanual changes in folivory and bird insectivory along a natural productivity gradient in northern Patagonian forests. Ecography 27: 29-40. Meiners, S.J. & Pickett, S.T.A. (1999) Changes in community and population responses across a forest-field gradient. Ecography 22: 261-267. Mulder, C.P.H; Koricheva, J: Huss-Danell, K; Högberg, P. & Joshi, J. (1999) Insects affect relationships between plant species richness and ecosystem processes. Ecology Letters 2: 237-246. Understanding biodiverity loss: an overview on forest fragmentation in South America 905 Perez-Harguindeguy, N; Díaz, S; Vendramini, F; Cornelissen, J.H.C; Gurvich, D.E. & Cabido, M. (2003) Leaf traits and herbivore selection in the field and in cafeteria experiments. Austral Ecology 28: 642-650. Rao, M; Terborgh, J. & Núñez, P. (2001) Increased herbivory in forest isolates: Implications for plant community structure and composition. Conservation Biology 15: 624-633. Rodríguez-Auad, K. & Simonetti, J. (2001) Evaluación de la folivoría: una comparación de dos métodos. Ecología en Bolivia 36: 65-69. Saunders, D; Hobbs, R.J. & Margules, C.R. (1991) Biological consequences of ecosystem fragmentation: a review. Conservation Biology 5: 18-32. Steffan-Dewenter, I. & Tscharntke, T. (2000) Butterfly community structure in fragmented habitats. Ecology Letters 3: 449-456. Sumerville, K.S. & Crist, T.O. (2003) Determinants of lepidopteran community composition and species diversity in eastern deciduous forests: roles of season, eco-region and patch size. Oikos 100: 134-148. Sumerville, K.S. (2004) Do smaller forest fragments contain a greater abundance of lepidopteran crop and forage consumers? Environmental Entomology 33: 234-241. Tscharntke, T. & Brandl, R. (2004) Plant-insect interactions in fragmented ladscapes. Annual Review of Entomology 49: 405-430. Valladares, G; Salvo, A. & Cagnolo, L. (2006) Habitat fragmentation effects on trophic proceses of insect-plant food webs. Conservation Biology 20: 212-217. Zak, M. & Cabido, M. (2004) Do subtropical seasonal forests in the Gran Chaco, Argentina, have a future? Biological Conservation 120: 589-598. Understanding biodiverity loss: an overview on forest fragmentation in South America 916 Appendix 1: Life form and percent cover (mean ± SE) of plant species damaged by invertebrate herbivores in continuous and fragmented Chaco Serrano Woodlands. Plant Life form Cover Continuous Fragmented AMARANTHACEAE Iresine diffusa Herb 0.1 ± 0.0 0.06 ± 0.02 ANACARDIACEAE Schinus fasciculatus Tree 10.15 ± 10.1 0.05 ± 0.02 Schinus longifolius Tree 10.18 ± 10.08 3.79 ± 3.74 APOCYNACEAE Aspidosperma quebracho- Tree 1.53 ± 0.03 - blanco Mandevilla pentlandiana Vine 0.08 ± 0.03 2.31 ± 1.78 ASCLEPIADACEAE Morrenia odorata Vine 0.05 ± 0.05 0.04 ± 0.01 ASTERACEAE Baccharis sp. 1 Shrub 0.03 ± 0.03 - Eupatorium sp. 1 Herb 0.08 ± 0.03 0.8 ± 0.73 Eupatorium argentinum Herb 0.03 ± 0.03 0.43 ± 0.01 Eupatorium hookerianum Herb 10.0 ± 10.0 0.43 ± 0.38 Eupatorium inulaefolium Herb - 0.0 ± 0.0 Ophryosporus axilliflorus Herb - 0.05 ± 0.03 BASELLACEAE Anredera cordifolia Vine 0.1 ± 0.0 0.45 ± 0.37 BERBERIDACEAE Berberis ruscifolia Shrub 0.05 ± 0.05 - BRASSICACEAE Brassica sp. 1 Herb 0.0 ± 0.0 - CELASTRACEAE Maytenus spinosa Shrub 0.03 ± 0.03 - CHENOPODIACEAE Chenopodium album Herb - 1.91 ± 1.88 COMMELINACEAE Commelina erecta Herb - 0.01 ± 0.01 CONVOLVULACEAE Ipomoea purpurea Vine - 10.78 ± 5.33 EUPHORBIACEAE Croton sarcopetalus Shrub 1.53 ± 1.48 4.16 ± 3.63 Euphorbiaceae sp. 1 Herb 0.1 ± 0.0 0.04 ± 0.01 FABACEAE Geoffroea decorticans Tree 3.83 ± 3.73 - LAMIACEAE Hyptis mutabilis Herb 1.55 ± 0.0 0.05 ± 0.03 Lepechinia floribunda Shrub 0.05 ± 0.0 - LYTHRACEAE Heimia salicifolia Shrub - 0.04 ± 0.01 MALPIGHIACEAE Understanding biodiverity loss: an overview on forest fragmentation in South America 927 Janusia guaranitica Vine 0.1 ± 0.0 - MALVACEAE Abutilon grandifolium Shrub 0.78 ± 0.78 0.03 ± 0.01 Malvaceae sp. 1 Herb 0.05 ± 0.03 Malvastrum coromandelianum Herb 0.08 ± 0.03 1.95 ± 1.87 Pseudoabutilon sp. 1 Herb 0.1 ± 0.0 Sida rhombifolia Herb 1.55 ± 1.45 0.46 ± 0.36 Sida sp. 1 Herb 0.0 ± 0.0 0.41 ± 0.38 Sphaeralcea cordobensis Herb 0.05 ± 0.05 - PASSIFLORACEAE Passiflora sp. 1 Vine 0.08 ± 0.03 0.06 ± 0.01 POACEAE Oplismenus sp. 1 Grass 13.13 ± 7.58 Setaria hunzikeri Grass - 0.41 ± 0.36 RANUNCULACEAE Clematis montevidensis Vine 1.55 ± 0.0 0.1 ± 0.0 RUTACEAE Fagara coco Tree 7.53 ± 7.48 9.45 ± 9.35 SANTALACEAE Jodina rhombifolia Tree - 0.04 ± 0.02 SOLANACEAE Cestrum parqui Shrub 1.55 ± 0.0 4.19 ± 1.96 Capsicum sp. 1 Herb 0.05 ± 0.05 - Lycium cestroides Tree 0.05 ± 0.05 0.06 ± 0.02 Salpichroa origanifolia Herb 0.1 ± 0.0 0.08 ± 0.01 Solanum argentinum Shrub 0.8 ± 0.7 0.43 ± 0.38 ULMACEAE Celtis tala Tree 14.4 ± 4.4 13.01 ± 8.37 VERBENACEAE Lantana balansae Shrub 0.08 ± 0.03 - Lantana camara Shrub 0.75 ± 0.75 0.01 ± 0.01 Lippia turbinata Shrub 0.8 ± 0.75 0.01 ± 0.01 Understanding biodiverity loss: an overview on forest fragmentation in South America 938 Extensive pollen flow may counteract the effects of landscape fragmentation P. Marchelli1,2, A.C. Moreno1,2 & L.A. Gallo1 1 Unidad de Genética Forestal, INTA EEA Bariloche, CC 277, 8400 Bariloche, Argentina. [email protected], cmoreno@ bariloche.inta.gov.ar, lgallo@ bariloche.inta.gov.ar 2CONICET Abstract Pollen flow might constitute a fundamental factor in balancing some of the negative effects of forest fragmentation. Wind pollinated species in Patagonia can overcome the isolation of populations by mechanisms of extensive pollen flow favored by the strong directional winds of the region. Several species of Patagonia have passed through a fragmentation process. In order to evaluate the degree of connectivity between populations and therefore the effect of fragmentation, pollen flow studies are being conducted in Nothofagus nervosa, Austrocedrus chilensis and Araucaria araucana. In N. nervosa, we began with an intrapopulation pollen flow study in order to establish the pollen dispersal curve (and the effective pollination distance). We collected buds and seeds from 25 mother trees and we are analysing between 10 to 20 seedlings per mother. For A. chilensis and A. araucana we have recently begun a project in order to study pollen flow in fragmented populations in the steppe. For that purpose we have sampled mothers and seeds from different locations with a spatially explicit design (total of 140 mothers and 30 seeds/ mother). We are analyzing the material with microsatellites markers and applying the TWOGENER method. Preliminary results, problems and perspectives of both studies are presented and discussed. Keywords: Gene flow, molecular markers, microsatellites, Nothofagus, Araucaria. Introduction Habitat fragmentation is considered as one of the greatest threats for the conservation of the biodiversity (Noss & Csuti, 1994). Most forest tree species in the world have suffered in the last centuries a continuous process of fragmentation. Such process acts on the dynamics of the ecosystem mainly through habitat loss, reduction in size of the forests fragments and increase of the distance among them (Turner, 1996). Several species have a fragmented distribution that goes into the Patagonian steppe like Austrocedrus chilensis and Araucaria araucana. This fragmentation is mainly due to a natural process but also influenced by anthropogenic activities. On the other hand, some Nothofagus species have a naturally fragmented distribution restricted to the lake watersheds but also suffered overexploitation and drastically reduction of their population size. Understanding biodiverity loss: an overview on forest fragmentation in South America 94 Nothofagus nervosa (Phil) Dim. et Mil. (Raulí) is one of the species of the Nothofagus genus that was highly and selectively exploited in the past. This species endemic to Chile and Argentina occurs in the Andes and Pacific Coastal Mountains between 35º and 41º S. The main distribution is along the Andes where it forms continuous forests to the western slopes from 37º to 40º 30´ S, but a narrow and fragmented distribution to the eastern slopes (in Argentina) following the west-east oriented glacial valleys between 39º 25´and 40º 30´S. Due to its high wood quality most of the populations were exploited during the first part of the Twentieth century leading to a severe reduction of the forests (Ing. Lebedeff, Internal Report of the Lanín National Parks Administration). Nowadays the whole distribution of the species in Argentina is included within two National Parks (Lanín and Nahuel Huapi) and conservation and breeding programs have begun (Gallo, 1993). The first step within these programs was the study of the genetic diversity in morphometric characters (Marchelli & Gallo, 1999) as well as with different genetic markers (Marchelli & Gallo, 2001; 2004; 2006) in order to establish the centres of highest diversity. It was detected that the highest variation occurs within a managed area, in Hua Hum, westward of the Lácar lake watershed. Given the importance of these results, the Lanín National Park Administration changed the status of this region to Tourist Use exclusively and promoted further studies in the area (Gallo et al., 2003). Araucaria araucana (Molina) K. Koch (Pehuén) and Austrocedrus chilensis (D.Don) Pic. Ser et Bizzarri (Ciprés de la Cordillera), are conifers endemic to the north region of the Sub Antarctic temperate forests of Argentina and Chile. To the limits of the natural distribution of both species (especially to the eastern extreme, towards the Argentinean steppe) the distribution pattern can be described as fragmentary, and mainly determined by the topography and the climate of the region. In the ecotone between the steppe and the continuous forests, specially in the case of Araucaria araucana, fragmented populations are the consequence of overexploitation, replacement for exotic conifers, large forest fires of antrophogenic origin and introduced livestock that impedes the natural regeneration and leads to a physical erosion of the soil. Within this region, small and dispersed groups of forests patches would be able to connect the larger and continuous forest to the west. The populations of Austrocedrus chilensis that occur in the oriental extreme are extremely isolated, exposed to an arid precipitation regime (330 mm per year) and subjected to strong western winds. The phenotype of the Patagonian Cypress highly varies according to site and genotypes. Although in marginal populations it is a small tree, it can reach a height of more than 35 m at the best sites (Tortorelli, 1956). Its natural range covers an extensive range of precipitation, from 3000 mm in the west to 300 mm in the east (Gallo et al, 2004). Following this precipitation gradient in West-East direction, A. chilensis forms mixed forests (with Coihue), pure forests (in mesic places), and marginal forests (in xeric sites) (Dezotti & Sancholuz, 1991). The reduction in the forests of Patagonian Cypress in Chile and Argentina after the European colonization reaches 37,16 % (Lara et al, 2001). Recent studies showed that the greatest genetic diversity of the species is found in those marginal populations from the steppe (Pastorino & Gallo 2002; Pastorino et al., 2004). Some of these populations could have been originated in glacial refuges situated in the east, where the species would have survived the glaciations, and from which the recolonization would have happened. These fragmented populations of Patagonian Cypress have shown a good relative growth in plant nursery, greater genetic diversity and are probably better adapted to xeric conditions. They are excluded from the protected area of the National Parks system, in private properties with extensive cattle raising, and sometimes competing with a monoculture of Ponderosa Pine. In many cases, plantations with exotic pines between the Patagonian Cypress Understanding biodiverity loss: an overview on forest fragmentation in South America 95 patches have been carried out. These plantations could be acting as physical barriers to gene flow. It is important to initiate strategies for protection of these valuable remnants of potential importance for the conservation and the genetic improvement of the species. The natural distribution of Pehuén (A. araucana) extends from 37º 20' S to 40º 20' S. However, this area represents the remaining forests of a more extensive previous distribution, which has been severely diminished since the European colonization. Estimations made with historical data and SIG techniques indicate that the reduction of the original forests of Araucaria reach a value of 34, 47 %, in Chile and in Argentina (Lara et al, 2001). The eastern most populations are found in the Patagonian steppe, with small groups of isolated trees of relatively low height (10 to 15 m) (Veblen et al, 1995; Donoso, 1993). In addition to the above mentioned factors provoking serious fragmentation, the Pehuén forests have a particular situation. These forests are used by the Mapuche communities that live within them since ancestral times. The current situation of extreme poverty lead to an exceeded increment of livestock which is feed with the edible seeds of Araucaria and which also provokes a soil erosion that avoids natural regeneration (Sanguinetti et al, 2002; Bekessy et al., 2004). Besides, the seeds are collected for human consumption and sale. The Pehuén is currently on risk of extinction and included in the Red Apendix of CITES (Farjon & Page, 1999), threat that is increased due to its restricted present distribution, slow growth and limited dispersion. Significant genetic variation within and among populations was detected in this species (Bekessy et al, 2002), being greater the variation in the eastern populations (Izquierdo & Gallo, 2003). Several genetic processes are more or less affected when the populations are drastically reduced and the landscape is fragmented (Hartl & Clark, 1988; Hanski & Simberloff, 1997). The fragmentation reduces the effective size of the population leading to a higher rate of inbreeding, the possible fixation of loci (as a result of genetic drift) and the reduction of gene flow among populations. These facts determine the loss of genetic diversity of the population and the increase of the genetic differentiation (Ellstrand & Elam, 1993). Nevertheless, these general expected effects cannot always be verified, since they depend on the scale of fragmentation and the biology of the species affected (Young & Boyle, 2000). Moreover, it was suggested that sometimes the fragmentation can lead to an increment in the pollen flow (Robledo-Arnuncio et al, 2004). This increment could be due to the low number of male individuals inside the fragment (so that great part of the pollen able to fertilize efficiently should come from outside of the fragment), or to the fact that the pollen moves further through the open landscape (result of canopy opening). The amount of gene flow among the patches will determine whether the fragments behave as a metapopulation where possible exchange of genetic information exists (Hanski & Simberloff, 1997) or each patch remains isolated from the rest and develops particular evolutionary strategies (Giles & Goudet, 1997). In order to study the contemporary gene flow in plants, it is necessary to determine the mean distance of the effective genetic dispersion (via pollen or seeds) also estimating its variability, namely the effective dispersion range. The distance of dispersion has a high impact in the evolutionary processes, conditioning, for example, the rate of dispersion of a favourable gene or the effective population size. The migration of seeds actually makes a small contribution to the gene flow in real time, being more important the pollen movement. The three species objects of this study are wind pollinated (anemophilous). The anemophily is characterized by the random distribution of pollen, and this result in an ineffective mechanism for the transport of large quantities of pollen (Finkeldey, 1998). Understanding biodiverity loss: an overview on forest fragmentation in South America 96 In order to determine the effects of forest fragmentation in the levels of pollen gene flow in the three mentioned species, we are conducting studies within the fragmented areas. We present here preliminary results and on going research lines in this topic. In particular we present first results to obtain the pollen dispersal curve for Nothofagus nervosa, and preliminary strategies to estimate the pollen flow between patches of Araucaria araucana and of Austrocedrus chilensis. This knowledge is essential to decide concrete measures of conservation of the genetic diversity in populations of the three species. It will permit to define the amplitude of the Evolutionary Significant Units (ESU) that should be conserved and thus to mitigate the effects of the fragmentation process. Materials and Methods Measurement of pollen flow There are several methods for the estimation of gene flow, but there is one that combines the estimation of contemporary pollen gene flow with landscape studies of large sample sizes. This is the method called TWOGENER (Smouse et al., 2001), an indirect method for estimation of actual pollen movement, and it can be seen as an intermediate method between the traditional laborious paternity analysis and the indirect analysis through genetic structure. The TWOGENER approach is based on the idea that the mother plants spaced throughout the landscape sample different sets of pollen clouds, partial or totally different. Mothers dispersed along the landscape with increasing distance are analysed to examine the differentiation of pollen pools and the effect of the distance among them. This method requires to know the genotype of mother and descendant, but the genotype of the potential parents is ignored. The paternal contribution must be determined for a sample of embryos collected from known mother plants. From the paternal genotypes, the parameter φft is estimated. This parameter is the proportion of the total variation among male gametes that is attributable to the fact that different seed-parents (females), spaced across the landscape, are sampling from different sets of pollen donors. The parameter φft is a measure of pollen structure. From this, we can also obtain an estimate of the effective number of pollen donors (Nep), (Austerlitz and Smouse, 2001). Given Nep and the adult stem density, we can compute the effective pollination neighbourhood area, Aep. Knowing these data permits to identify the pattern of gene movement, to estimate the average distance of pollination and the curve of pollen dispersion. Sampling strategies Nothofagus nervosa In order to determine the pollen curve and pollen effective dispersion distance we selected a pure and rather isolated population within Tromen Lake watershed (39º37’02’’ S y 71º20’23’’ W). Twenty five mother trees were sampled and mapped with a GPS. From each mother tree progeny tests were already installed in field trials of the Forest Genetic Unit of INTA EEA Bariloche. Therefore we collected buds of between 10 to 20 progenies of each mother (a total of 410 seedlings). Bud tissue of mothers and progenies was kept at –20ºC until used in the laboratory analyses. Austrocedrus chilensis Understanding biodiverity loss: an overview on forest fragmentation in South America 97 To determine the adequate number of mothers and seeds per mother in the two conifers an analysis with previous genetic information (Izquierdo & Gallo, 2003; Pastorino et al., 2004) was performed. This analysis implies to estimate the exclusion probability of a set of genetic markers in the species in order to determine with the required precision the φFT value. The analysis showed that for the worst case a total of 28 mothers with 28 seeds/mother would be necessary to reach the desired precision (Moreno et al., in prep). In addition the TWOGENER method increases the precision when sampling more mothers at expenses of less seeds. Therefore, we decided to collect more than 28 mothers to have a better representation and to analyse 10 seeds and have enough extra material in case it is necessary to increase the sample size up to 28. Sampling was carefully designed in order to have a large forest fragment and smaller patches nearby. Such a scenario was found in San Ramón Ranch, some 25 km westward of San Carlos de Bariloche. Isolated trees that could function as “bridges” were also sampled. Leaves and seeds from a total 62 trees were collected, cleaned and keep at -20ºC until laboratory analyses. Trees were mapped. Araucaria araucana The sampling was conducted in the region of Tromen lake watershed, to the east. A large population, fragments and isolated trees in the steppe were sampled. A total of 40 mothers were sampled and mapped, collecting at least 30 seeds per mother. Material was kept at –20º C until laboratory analyses. Mixed population In order to study the possible differences of pollen movement in the two conifers a mixed population was sampled. The population was located in Pulmarí (39º07’12’’ S y 71º05’49’’ W). Leaves of thirty mothers of each species and 30 seeds per mother were collected and mapped. Material was treated as previously mentioned. Genetic markers and data analysis To achieve the highest precision it is better to count with highly polymorphic genetic markers as microsatellites (SSRs). For Nothofagus nervosa we have analysed 5 SSRs developed for the genus (Azpilicueta et al., 2004; Marchelli et al., submitted). For A. araucana we have established the protocols to transfer six primers developed mainly for A. angustifolia by Salgueiro et al (2005). Finally for A. chilensis we are developing the markers at the moment. For the TwoGener analysis we employed the program GeneAlEx (Peakall & Smouse, 2005). Preliminary results and discussion The study of the two conifers is still on going and we have done the sampling, DNA extraction and establishment of the protocols. Therefore no results will be presented. In the case of N. nervosa, we have preliminary results of 13 mothers and a total of 204 seedlings that were already analysed with the five SSRs. The TwoGener analysis of these data shows a high genetic structure since the among mothers variation is almost 29 % (Table 1). This high value gives an effective neighborhood pollination (Nep) of two males and an average effective pollination distance of 5 m. Understanding biodiverity loss: an overview on forest fragmentation in South America 98 These results are preliminary since not all the mothers have been analysed yet, so caution should be taken in the discussion of the results. However, the tendency is towards a high differentiation between the mothers and the pollen clouds they capture. If this tendency is right then the pollen flow in Nothofagus nervosa is very limited. Table I: Components of the AMOVA analysis and φFT value. Source df SS MS Est. Var. % Among Mothers 12 64.161 5.347 0.313 29% Within Mothers 178 139.081 0.781 0.781 71% Total 190 203.243 1.094 100% Stat Value Prob PhiFT 0.286 0.001 Future perspectives The on going research lines will be complemented with additional genetic studies in order to have a comprehensive and detailed study of the current and past gene flow. In this sense, we are also conducting fine-scale genetic structure studies in N. nervosa within Lácar lake watershed. These studies will allow the determination (among others) of the limits of the area to be protected in Hua Hum. Concerning the conifers, a genetic characterization with nuclear microsatellites of fragmented populations of the Patagonian Cypress is also in course and the impact of genetic drift in small size populations (due to fragmentation). Finally, seed dispersal in Pehuén is also the subject of a future study in order to complement the data on pollen gene flow. The final aim in all these research lines is to contribute with genetic arguments for conservation policies. Acknowledgements We would like to thanks A. Martínez Meier, A. Martínez and M. Millerón for field assistance. For laboratory help we are grateful to M. Millerón and C. Soliani. We thanks M. Pastorino for helpful comments in the manuscript . These studies are being financed by BES (British Ecological Society, grant N 2254 to PM); PIP CONICET (5451); PICT Joven (25266) and BBVA Fundation (Cooperation INIA-Madrid – INTA EEA Bariloche). References • Austerlitz, F; Smouse, P.E. (2001): Two-Generation Analysis of Pollen Flow Across a Landscape. II. Relation Between φft, Pollen Dispersal and Interfemale Distance. Genetics 157: 851-857. Understanding biodiverity loss: an overview on forest fragmentation in South America 99 • Azpilicueta M. M., H Caron, C Bodénès, L. A Gallo. (2004) SSR markers for analysing South American Nothofagus species. 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(1996) Species loss in fragmented tropical rain forest: a review of evidence. Journal of Applied Ecology 33: 200-209. • Veblen, T.T., Donoso, C, Kitzberger, T., Lara. A. & Villalba, R. (1995) The ecology of the conifers of southern South America. En Ecology of the Southern Conifers, Ed. N. J. Enricht & R. S. Hill: 120-155. • Young, A.G. & Boyle, T.J. (2000) Forest fragmentation. En: Forest Conservation Genetics. Principles and Practice. CSIRO Publishing-CABI Publishing. 123-134. Understanding biodiverity loss: an overview on forest fragmentation in South America 101 Effects of forest loss on soil water regime in the temperate landscape in southern Chile C. Echeverría1,2, O. Thiers2, A. Lara1,2 1Millennium Scientific Nucleus FORECOS –Universidad Austral de Chile, Instituto de Silvicultura, Casilla 567, Valdivia, Chile. e-mail: [email protected] 2 Universidad Austral de Chile, Instituto de Silvicultura, Casilla 567, Valdivia, Chile Abstract The effect of land use change on water regime in ñadi soils (Duric Histic Placaquands) was assessed using multitemporal land-cover maps and monitoring the water table oscillations in southern Chile. Over the last three decades there has been an increase in 3,800 ha of wetlands as a result of the forest loss in ñadi soils at a rate of 1.9 % yr-1. Results also revealed that the deforestation in ñadi soils has caused an increase in the level of the water table leading to an overland flow for longer periods from winter to the beginning of spring. This increase in the water table is related to a decline in evapotranspiration in terms of interception and water consumption due to the loss of the forest cover. The progressive deforestation in southern Chile is facilitating substantial changes in the spatial configuration of the temperate forest and also the creation of a mosaic of sites with different infiltration characteristics. Key words: water table, ñadi soils, deforestation, forest fragmentation, GIS Introduction Temperate rain forests in Chile, which are classified as a biological “hotspot” because of their high endemism (Dinerstein et al., 1995; Myers et al., 2000), are rapidly disappearing as a result of land-cover change (Echeverria et al., 2006). Some of the key causes of habitat destruction are large-scale agricultural activities, subsistence farming, selective logging and infrastructure development (Lara et al., 2000). Some researchers emphasise that the clearance of forest cover may cause serious ecological problems in the ecosystem (Forman and Godron, 1996; Hill and Curran, 2001; Pardini et al., 2005; Uezu et al., 2005). Deforestation and forest fragmentation in temperate zones are causing diverse negative impacts on the bird species (Cornelius et al., 2000; Vergara and Simonetti, 2003), mammals (Donoso et al., 2003) and forest composition and structure (Echeverría et al., in press). However, there are other types of impacts that may severely alter the functioning of the ecosystem. For instance, it is widely known that forest store water both above and below ground. When a forest is removed, evapotranspiration is minimised so that runoff is more favoured in the hydrological cycle (Ward and Trimble, 2004). Some tropical studies have demonstrated that the soil storage water capacity is exceeded in shallow soils when the forest cover is converted to pasture (Zimmermann et al., 2006). This results in saturation at the soil surface, also known as overland flow which is increased with the subsequent addition of rain water flow across the surface (Dunne and Black, 1970). Depending on local climate and geomorphology, this could lead to more or less water stored in the soil. In temperate zones with low slope the loss of a forest could cause the water table to rise due to the reduction in evapotranspiration (Sahagian, 2000). In southern Chile (38º 30´ S and 43º 00´S), an important area of temperate forest occurs on acidic, shallow, poorly-drained soil known as ñadi soils (Besoaín 1985, INIA 2001). These soils are derived from modern volcanic ashes settled on glacio-fluvial deposits, located in flat Understanding biodiverity loss: an overview on forest fragmentation in South America 102 areas in the Central Depression (Schlatter et al., 1995). The ñadi soils are characterised by a slow drainage regime due to the flat topography and the presence of an impermeable solid iron-silicate-layer at a depth of 20-80 cm (named “fierrillo”). The high levels of rainfall (1,000- 2,000 mm/year) contribute to overland flows in deforested areas. The native forests situated on ñadi soils have been affected by progressive forest fragmentation and forest loss over the last three decades as a result of forest logging and agricultural expansion (Echeverría, 2005). At present, there is no documented information about the impacts of such forest loss on water table changes at the landscape level. The purpose of this study was: i) to assess spatiotemporal changes in the land cover types associated with water table oscillations and ii) to relate these changes to oscillations in the water table in deforested areas that occur in ñadi soils. Methods Data for land cover change analysis To analyze the spatial and temporal changes of the land use types, a set of three Landsat satellite scenes were acquired at different time intervals: 1976 (Multispectral Scanner), 1985 (Thematic Mapper), and 1999 (Enhanced Thematic Mapper). It was necessary to correct the images geometrically, atmospherically and topographically before they could be used to assess changes in forest cover (Chuvieco, 1996; Rey-Benayas and Pope, 1995). Supervised, maximum likelihood classifications were performed on each of the three images to classify the land cover types using training locations, obtained from field surveys. Each land cover map was validated using ground-based data. Overall agreement of classification was 88.8% for the 1976 MSS, 89.6% for 1985 TM image, and 91.9% for the 1999 ETM+ image. The following categories of land cover were identified from each image: native forests (including secondary and old-growth forests), seasonal wetlands, agricultural lands, shrublands, and urban grounds among other. In the present study, we assessed the conversion from native forests to seasonal wetlands in ñadi soils from 1976 to 1999. A digital cover of ñadi soil polygons was generated from INIA (2001). This cover was used to mask the landscape in order to select those areas in which the deforestation occurred in ñadi soils. Conversion from forest cover to wetlands was analyzed using ARC VIEW 3.2 software1 and its extension Arc View Spatial Analyst 2.0 for Windows. The formula used to determine the annual rate of deforestation was (FAO, 1995): − 1/(t2 t1) P = A2 −1 *100 [Equation 1] A1 where P is the percentage loss per year, A1 and A2 are the forest area at time t1 and t2 respectively. Measurement of water table 1 ESRI 1996-2000. Environmental Systems Research Institute, Inc. 380 New York St., Redlands, CA92373-8100, USA. Understanding biodiverity loss: an overview on forest fragmentation in South America 103 The experiment to measure water table oscillations was conducted on private land located within the study landscape (white circle in the Fig. 1). The land is flat and is situated at 65 m a.s.l. The external drainage is frequently slow to very slow (Schlatter et al., 2003), and usually does not vary across the year. The experiment was conducted in two sites: i) a 2 ha seasonal wetland (peat bog) originated by the clearance of the forest and currently covered by shrub, sedges, grasses and Sphagnum sp. mosses, ii) a secondary forest of Fiztroya cupressoides and diverse evergreen broad-leaved species. A total of 60 piezometers were distributed across the two sites using a 15 x 15 m grid. The piezometer is a plastic tube of 150 cm length and 1” diameter. The part of the tube that goes in the soil (ca. 60 cm) was perforated to favour the flow of water through the piezometer. Data was collected twice a week during the winter in 1999 and 2000. The oscillation of the water table was determined by estimating the percentage of water- saturated soil, calculated from the difference between the water table level and the total soil depth. The spatial variation of the water table in the two study sites was analysed by creating a contour map by kriging (Golden Software Surfer 8) the percentage of saturated soil (100% means a completely saturated soil) of the winter months during 1999 and 2000. Results Forest cover change Approximately 36% of the native forests that occurred in ñadi soils in 1976 had disappeared by 1999 (Fig. 1). During the whole study period, the annual forest loss averaged 1.9%yr-1. Most of the forest loss was concentrated in the first nine years of the study period, at a deforestation rate of 2.9% yr-1. In the second time interval, the rate decreased considerably to approximately 1.3% yr-1. The proportion of the ñadi soil covered by native forests decreased gradually across the time periods. In 1976, this soil type was covered by 114.1 thousand hectares, equivalent to 64% of the total area (Fig. 1). In 1985, native forests represented 49% of the total area of ñadis soil while fourteen years later, the area of native forests declined to 73 thousand hectares, equivalent to 41% (Fig. 1). This decline in forest area was explained mainly by a conversion of 29,220 ha of native forests to shrublands between 1976 and 1985 and by another 22,671 ha between 1985 and 1999. Contrary to the decline in native forest cover, the area of wetlands in ñadi soil increased consistently over time. Wetlands exhibited a considerable increase from 250 ha in 1976 to 668 ha in 1985 and then 4,045 ha in 1999. The conversion of the land use to wetlands was mainly associated with the clearance of shrublands and native forests. Approximately 486 ha of shrublands and 181 ha of native forests in 1976 were transformed into wetlands in 1985. During the next time interval, the conversion of shrublands and native forests to wetlands occurred at a higher rate of 3,071 and 982 ha respectively. The rates of transition derived from the analysis of forest cover change in ñadi soil show a transition to an area where non-forest land uses such as crops and pasture, shrublands and wetlands are becoming the dominant land cover types in the study landscape (56%). Understanding biodiverity loss: an overview on forest fragmentation in South America 104 1976 1985 1999 CHILE ANTARTICA 90 º 53º n a e c O ic if c a P Figure 1. Loss of native forests on ñadi soils between 1976 and 1999 in southern Chile. The white circle indicates the location of the water table experiment. Deforested area in ñadi soils (including shrublands, crops and pasture land) Native forest Other types of soil parent material Understanding biodiverity loss: an overview on forest fragmentation in South America 105 Substantial changes were observed in the distribution of forest patch size in ñadi soil between time intervals (Fig. 2). In 1976, 73% of the forest area was concentrated in large forest patches greater than 5,000 ha. This trend varied in 1985 when most of the forest area (60%) occurred in smaller patches below 5,000 ha. In 1999, patches of less than 500 ha represented 42% of the forest area whereas patches greater than 5,000 ha contained only 16%. 90,000 80,000 1976 70,000 1985 60,000 1999 50,000 40,000 Forest area (ha) area Forest 30,000 20,000 10,000 - 0-250 250-500 500-1,000 1,000-5,000 >5,000 patch size (ha) Figure 2. Temporal variation of forest patch size in ñadi soils. Water table variations The total soil depth did not vary significantly (P < 0.05) between sites. In the seasonal wetland the soil depth ranged from 47 to 113 cm with a median of 76 cm while in the native forest the depth varied from 29 to 96 cm with a median of 66 cm. During the winter months (June to September), the seasonal wetland showed an average soil water-saturation of 90% of the total depth (Figure 3a). In some places, overland flow was observed, indicating 100% saturation. In the native forest, 77% of the soil profile was water- saturared during the same months (Figure 3a). In some specific places, the soil presented a water-saturation of only 60%. During the summer season (December to February), both sites exhibited lower levels of water saturation (Figure 3b); in the seasonal wetland 30-70% of the soil profile was water- saturated while in the native forest this value varied from 20 to 50% of the soil profile. In the summer, the native forest showed a faster reduction in the water saturation of the soil than in the seasonal wetland. Similarly, the native forest exhibited the greatest variation in percentage of water saturation between summer and winter (Figure 3a,b). The soil in the seasonal wetland presented a low to very low level of aired soil (0 - 5 cm) from the winter to the beginning of the spring. However, in the native forest the aired soil reached up to 10-20 cm for the same period. The depth of aired soil is highly related to the air regime and is negatively affected when the water table saturates the soil profile. This may be a constraint for the establishment and growth of plants that require a balanced distribution in water and air in the soil. Understanding biodiverity loss: an overview on forest fragmentation in South America 106 100 A B 90 Native Native forest forest 80 N 70 60 50 40 Seasonal Seasonal 30 wetland wetland 20 90 m 10 Figure 3. Spatial variation of the water table measured as the percentage of water-saturated soil in a native forest and a seasonal wetland originated by deforestation. Values correspond to mean percentage (1999 and 2000) of water-saturated soil for the winter (a) and summer (b) months. Discussion and conclusions The native forest in ñadi soil has undergone high degrees of forest clearance during the three decades analysed compared to many other forested landscapes. In the study landscape, the native forests have disappeared at a rate of 1.9% yr-1 over the whole period, with a maximum of 2.9% yr-1 in the first nine years. Forest losses recorded for other temperate forests have showed different magnitudes. For instance, the Maulino forests in the coast range of central Chile have disappeared at a higher rate of 4.5% yr-1 between 1975 and 2000 (Echeverria et al., 2006). In the Klamath-Siskiyou ecoregion, USA, there has been a reduction of the native forests at a lower rate of 0.53% yr-1 over the period 1972-1992 (Staus et al., 2002). In western Oregon, deforestation rates by clearcutting between 1975 and 1995 varied from 0.5% to 1.2% with almost 20% of the total forest impacted (Cohen et al., 2002). In the present study reductions in the forest cover resulted in a considerable decline in patch size, as was also observed in other fragmentation studies (Zheng et al., 1997; Fitzsimmons, 2003). In our study landscape, the size of the forest fragments declined consistently over time. Some researchers state that the progressive reduction in the size of forest habitats is a key component of ecosystem fragmentation (Armenteras et al., 2003). In central Chile, the Maulino forest has exhibited a severe decline in the patch size during the whole time period (Echeverría et al., 2006). In 1975, 22% of the forest area was concentrated in fragments of less than 100 ha, however, this percentage increased to 69% in 2000. The progressive loss and fragmentation of the native forests in ñadi soils is leading to a substantial modification of the water regime in the study landscape. The seasonal wetland is characterised by overland flows during the winter and the beginning of spring. This reduces Understanding biodiverity loss: an overview on forest fragmentation in South America 107 the probability of establishment of plant species due to water saturation of the soil. Additionally, the fragmentation of the forest cover is leading to a mosaic of surfaces having distinct infiltration characteristics and differing in the propensity to generate overland flows. Similar to our results, the fragmentation associated with human activity in northern Vietnam has caused a reduction in the infiltration and altered near-surface flow paths which have resulted disproportionately in overland flow response during typical rainfall events (Ziegler et al., 2004). Our results demonstrate saturation at the surface soil in the deforested area due to the storage capacity of the soil being exceeded via a rising water table. This excess in the soil capacity to store water is caused by several concurrent factors that cause overland flow in the seasonal wetland (deforested area) during the winter season (Janssen et al., 2004). One of the main factors is the presence of the fierrillo layer in ñadi soils that prevents the vertical movement of the water in the soil profile (Schlatter et al., 2003). On the other hand, the high quantity and intensity of precipitation in winter contributes to a rapid saturation of the soil (Janssen et al., 2004). Also, the presence of short vegetation (shrub and herb species) implies a low evapotranspiration rate, which is contrary to what occurs in the native forest where the presence of tree cover increases substantially the rates of evapotranspiration by water consumption and interception (Echeverria, 1994). Some researchers have observed that the water saturation in ñadi soils causes an oxygen deficiency for plants (Janssen et al., 2004). This may seriously affect the natural regeneration of native plants in deforested areas growing in ñadi soils. Similarly, the presence of an overland flow limits afforestation which requires plants that do not grow in water-saturated sites. In conclusion, due to the presence of a mosaic of surfaces having distinct infiltration characteristics, it is highly recommended to conduct a stratification of the landscape in order to identify homogeneous areas in terms of soil water and air regimes. This study would enable the identification of problematic habitats for the growth of vegetation, which is necessary to carry out an adequate restoration programme of the landscape. The deforestation and fragmentation of the study area has caused changes not only in the habitat conditions but also in the spatial configuration of the forest cover. References Armenteras, D; Gast, F. & Villareal, H. 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Understanding biodiverity loss: an overview on forest fragmentation in South America 110 Part III Landscape Ecology for conservation, management and restoration Understanding biodiverity loss: an overview on forest fragmentation in South America 111 Mitigation of biodiversity loss in Nothofagus pumilio managed forests of South Patagonia M.V. Lencinas1(*), G. Martínez Pastur1, E. Gallo1, A. Moretto1, C. Busso2 & P. Peri3. 1CADIC-CONICET; 2Departamento de Agronomía-UNSur; 3INTA-UNPA-CONICET (*) CADIC - Cc 92 (9410) Ushuaia (Tierra del Fuego), Argentina, +54-2091-422310 int 111, [email protected] Abstract Forest management is mainly based on economical and forest structure variables, but ecology must be considered to develop new silvicultural systems. The aim was to analyze the impact of traditional methods compared to aggregated and dispersed retention proposals, over species richness in Nothofagus pumilio forests. Understory plants, birds and insects were sampled after three years of harvesting. Total richness inside the aggregates was more similar to primary forests than harvested sectors (dispersed retention or clear- cuts). The major impact was the richness increment due to species incorporation from associated environments. Harvesting allowed the introduction of exotic plant species and birds from open places. The combined retention treatment conserved most of the original insect richness inside the aggregates (74%). The regeneration methods with aggregated and dispersed retention appear as an effective alternative for biodiversity conservation, because aggregates conserve some of the original heterogeneity and the environmental conditions of the primary forests. Keywords: conservation, understory, insects, birds, aggregated retention. Introduction Since their beginnings, forest management in South Patagonia has been mainly based on economical and forest structure variables, striving to enhance the natural regeneration of the harvested stands. Several cuttings have been applied in these forests to assist regeneration already present or to make natural regeneration possible. Cutting treatments include light selective cuts to clear-cuts (Gea et al. 2004). The most widely used method has been the shelterwood cut (Martínez Pastur et al. 2000; Rosenfeld et al. 2006), which opens the overstory canopy and increases light availability and the effective rainfall at the understory level (Caldentey et al. 2005a). This method also increases global radiation, temperature and wind speed, causing greater evapotranspiration (Caldentey et al. 2005b). This management strategy affects insect communities, plants, birds, and mammals, and also modifies their inter-specific relationships (Pulido et al. 2000; Deferrari et al. 2001; Spagarino et al. 2001; Martínez Pastur et al. 2002). To develop new silvicultural systems, it is necessary to understand the impact of the traditional methods on the biotic and abiotic components, as well as understand the biodiversity of the original ecosystem. These ecological criteria with wide social acceptance have been included in the planning for forest management. The aim of this work was to analyze the impact of traditional and new silvicultural techniques (with aggregated and Understanding biodiverity loss: an overview on forest fragmentation in South America 1 112 dispersed retention) on species richness (understory plants, birds and insects) after three years of harvesting in Nothofagus pumilio forests of South Patagonia. Materials and methods Study site description Nothofagus Blume is the main component of the Magellanic forests, with a wide range of natural distribution from 36° 50' to 55° 02' SL. The southern forests are predominantly pure and deciduous, mainly of N. pumilio (Poepp. et Endl.) Krasser (lenga) and N. antarctica (Forster f.) Oersted (ñire). In Tierra del Fuego, Argentina, forest uses are mainly for harvesting (only in N. pumilio forests), cattle grazing, and tourism, with the objectives of natural ecosystem management primarily economic (Martínez Pastur et al. 2000), and secondarily, for conservation purposes (Martínez Pastur et al. 2002). Homogeneous N. pumilio forests in the San Justo Ranch (54°05' SL, 68°37' WL) were selected, where the “Los Castores” sawmill administers the logging. Forest use was mainly for timber, but natural browsing pressure of Lama guanicoe (“guanaco”) on Nothofagus seedlings and saplings existed (Pulido et al. 2000). Climate is characterized by short, cold summers and long, snowy and frozen winters. Mean monthly temperatures vary from about 3°C below zero to 9°C. Only three months/year are free of mean temperatures under 0°C, and the growing season extends for about five months. Precipitation is between 400-500 mm.year-1 (Cuevas 2000). Sampling was carried out in a permanent plot established in the year 2001. New techniques of regeneration systems were compared to a shelterwood cut (DR - 23.6 ha) and an old-growth primary forest as a control (C - 8.6 ha). The new tested methods were 1) a treatment with aggregated retention (one circular island per hectare of original forest with 30 m radius) combined with dispersed retention (15 m²/ha basal area of the most dominant trees distributed between aggregates) (CR - 10.7 ha) (Figure 1); and 2) a treatment with aggregated retention and clear cut between aggregates (AR - 18.5 ha). The forest before the harvesting had a basal area of 65-75 m²/ha, 400-600 trees/ha, 40-45 cm quadratic mean diameter and 680-780 m³/ha of total over bark volume. Figure 1. Combined retention with aggregated and disperssed retention in Nothofagus pumilio forests in Tierra del Fuego. Understanding biodiverity loss: an overview on forest fragmentation in South America 2 113 Understory plant, bird, and insect sampling Each community of study organisms was characterized in each treatment three years after the logging (February 2004). For understory plant species, ten plots were taken, each with four 0.25 m2 subplots orthogonally placed 5 m apart from the centre (Martínez Pastur et al. 2002). Vascular plants (Dicotyledonae, Monocotyledonae and Pteridophytae) were taxonomically classified by species, following Moore (1983) and Correa (1969-1998). Birds were sampled using a point sampling method during a four-hour period following sunrise (Lencinas et al. 2005). Sampling was carried out under similar climatic conditions, and days with fog, strong wind, or rain were discarded. Six sampling plots were established in each treatment, and each plot was visited 5 times. A 10-minute observation period was used, consisting of a 2-minute accustoming period and an 8-minute counting period. Sampling used a direct recognition method (sight, by binoculars), without limit in the observation range. Taxonomy was recorded following Narosky and Yzurieta (1987). For the Insecta class, adult mobile epigean individuals of Diptera, Lepidoptera, Hymenoptera and Coleoptera were collected, during two 24 h periods (day and night) in each treatment. Collections were carried out under equivalent climatic conditions, discarding days of strong winds or heavy rain, and for three stratum levels along the vertical forest structure (litter, understory and ¾ total height of overstory). A wide spectrum trap system was utilized (Spagarino et al. 2001), with smell ethanol attractive traps (20 cm diameter), black and white cold fluorescent light traps (20 cm diameter with 4 watt lamps), pit-fall traps (100 x 15 x 8 cm), and color attractive trays (10 x 10 x 5 cm), using yellow, white and sky-blue colors. Water was used as a retention agent and formaldehyde as a preservative. Commercial detergent was employed to diminish surface tension. Trap systems were located with enough distance between them (50 m) to avoid interferences. After trapping, individuals were classified under a dissecting scope to order, following the classification proposed by Richards and Davies (1984) and Romoser and Stoffolano (1998). Because Patagonian insect systematic is still incomplete, either the recognizable taxonomic unit (RTU) or the morphospecies concept was utilized (Oliver and Beattie 1993). Data analysis Matrixes of understory plant species, bird, and insect richness in each treatment were utilized for multivariate analysis. Relationships among site types were studied by cluster analysis, using a complete linkage amalgamation rule and Euclidean distance measurement. Then, detrended correspondence analysis was done, without down-weighting rare species and rescaling of the axis. Results and discussion Understory plant richness A total of 23 understory plant species was identified across all treatments (Table 1), which did not included any of the three Pteridophytae species of the Nothofagus pumilio primary forests (Moore 1983; Martínez Pastur et al. 2006). In the primary forests, we sampled 12 understory species, while in the harvested stands the richness increased, reaching 17 species in the shelterwood cut, 16 in the dispersed retention of CR, and 20 in the clear-cut of AR. Inside the aggregates, the richness was comparable to that observed in the primary forest (9 species in CR and 13 species in AR). These values are lower than those described in other N. pumilio forests studies. Martínez Pastur et al. (2002) found 19 species of superior plants and Pteridophytae in primary forests and 35 in harvested stands, while Lencinas Understanding biodiverity loss: an overview on forest fragmentation in South America 3 114 (2005) found 18-28 species in timber quality forests without logging. In this study, only one species, Viola magellanica, was not found in the harvested stands but was sampled in the primary forest (C). Table 1. Undestory plant and bird species. Understory plant species Status Code Birds species Status Code Acaena magellanica Native ACMA Aphrastura spinicauda Native APSP Acaena ovalifolia Native ACOV Campephilus magellanicus Native CAMA Agrostis flavidula Native AGFL Carduelis barbata Native CABA Cardamine glacialis Native CAGL Elaenia albiceps Native ELAL Cerastium arvense Native CEAR Enicognathus ferrugineus Native ENFE Cerastium fontanum Exotic CEFO Falco peregrinus Native FAPE Deschampsia flexuosa Native DEFL Falco sparverius Native FASP Dysopsis glechomoides Native DYGL Milvago chimango Native MICH Festuca magellanica Native FEMA Phrygilus patagonicus Native PHPA Galium aparine Native GAAP Phygarrihchas albogularis Native PHAL Geum magellanicum Native GEMA Scytalopus magellanicus Native SCMA Nothofagus pumilio Native NOPU Tachycineta leucopyga Native TALE Osmorhiza depauperata Native OSDE Theristicus caudatus Native THCA Phleum alpinum Native PHAL Troglodytes aedon Native TRAE Poa pratensis Exotic POPR Turdus falcklandii Native TUFA Ranunculus biternatus Native RABI Xolmis pyrope Native XOPY Ribes magellanicum Native RIMA Zonotrichia capensis Native ZOCA Rumex acetosella Exotic RUAC Taraxacum officinale Exotic TAOFF Trisetum spicatum Native TRSP Uncinia lechleriana Native UNLE Veronica serpyllifolia Exotic VESE Viola magellanica Native VIMA Cluster analysis highlighted the similarity between understory plant richness inside the aggregates and the primary forests, compared to the harvested stands. The most dissimilar treatment was the shelterwood cut, followed by the harvesting methods applied between aggregates (Figure 2). In the DCA analysis, the aggregates plots (ARI and CRI) were located closer to the primary forests than the harvested ones. It was possible to observe that most of the species are located in the center of the graph because they are generalist species (e.g. Gallium aparine, Osmorhiza depauperata). Other species (e.g. Agrostis flavidula) get into the harvested stands from open natural areas or Nothofagus antarctica forests. On the other hand, many of the sampled exotic species get into the system after harvesting (e.g. Rumex acetosella, Veronica serpyllifolia or Taraxacum officinale), and their presence is related to the degree of the interventions. The introduction of exotic species in the harvested stands was also cited by Martínez Pastur et al. (2002), who found six invading species after shelterwood logging and none in the primary forests. Understanding biodiverity loss: an overview on forest fragmentation in South America 4 115 Figure 2. Cluster analysis and DCA for understory plant species richness. ARI: Agregated retention inside; ARO: Agregated retention outside; CRI: Combined retention inside; CRO: Combined retention outside; DR: dispersed retention; C: control. Species code in Table 1. Bird richness Only 17 bird species were found during the sampling (Table 1), all of them commonly observed in Nothofagus forests of Tierra del Fuego (Deferrari et al. 2001; Lencinas et al. 2005). In the primary forests, nine species were found, while in the harvested stands richness was slightly greater due to additional food sources and the presence of open places. We sampled 11 species in shelterwood cuts, 13 in clear-cuts of AR, and 8 in the dispersed retention of CR, while inside the aggregates 11 species were found in CR and 12 species in AR. Richness of the aggregates were higher than the primary forests due to edge effects and refuge characteristics, which were highly visited by many bird species. For example, Falco sparverius uses the borderline trees as a vantage point for hunting. Food availability of the understory in the dispersed retention benefited some passeriform species (Scytalopus magellanicus and Phrygilus patagonicus), which used the aggregates as refuge. Other species were not found in open treatments (Elaenia albiceps, Milvago chimango and Enicognathus ferrugineus) due to specific breeding or tree hollow requirements. Bird richness was higher than that observed by Deferrari et al. (2001) (7 species in primary forests and 15 in harvested stands), but fewer than Lencinas et al. (2005) found in timber quality forests without logging (12 species). These differences could be explained by differences in the sampling effort (e.g. Falco peregrinus was found in very small densities) or due to the season of sampling (e.g. Lencinas et al. 2005 founded 11 species in spring and 6 in summer on timber quality forests without logging). Cluster analysis highlighted the similarity between bird richness inside the aggregates of CR and the primary forests. This group was also similar to AR treatments. The third group was composed of species in the dispersed retention (CRO and DR) and was the most dissimilar (Figure 3). In the DCA analysis, the aggregates plots (ARI and CRI) were located closer to the primary forests than the harvested ones, and the clear-cut (ARO) treatment was the most divergent. Most of the species were again located near the center of the graph because they are generalist species (e.g. Aphrastura spinicauda or Zonotricha capensis). Some species got into the forest due to the connection of open areas (e.g. Falco sparverius) used for hunting, while others benefited from the major food availability of insects and Understanding biodiverity loss: an overview on forest fragmentation in South America 5 116 understory plants in the harvested stands, and the refuge of the aggregates (e.g. Xolmis pyrope). Figure 3. Cluster analysis and DCA for birds species richness. ARI: Agregated retention inside; ARO: Agregated retention outside; CRI: Combined retention inside; CRO: Combined retention outside; DR: dispersed retention; C: control. Species code in Table 1. Insect richness During the sampling, 184 insect morphospecies were determined, where 45% were Diptera, 25% Hymenoptera, 17% Lepidoptera and 13% Coleoptera. Diptera dominance was in agreement with Lanfranco (1977), Solervicens (1995), Spagarino et al. (2001) and Lencinas (2005), but with different percentages (34% to 97%). Total richness was higher than observed by Spagarino et al. (2001) (96 morphospecies) in primary and harvested forests. During sampling in the primary forests, we recorded 88 insect morphospecies. These values are greater than those sampled by Spagarino et al. (2001) (51 morphospecies), and lesser than Lencinas (2005) (between 126 and 144 morphospecies). In the harvested stands, insect richness was significantly higher, reaching 139 morphospecies in DR, 120 in CR and 91 in AR. In CRI 62% of the morphospecies corresponded to primary forests species (65 original morphospecies and 39 introduced morphospecies), while in ARI the 81% of the morphospecies corresponded to primary forests species (50 original morphospecies and 12 introduced morphospecies). The combined treatment conserved most of the original insect richness inside the aggregates (74%). However, the harvested areas allowed the introduction of many species from associated non-timber quality environments. Cluster analysis highlighted the similarity between insect richness inside the aggregates of CR and the primary forests. A secondary group included the harvested areas with dispersed retention (CRO and DR), and this with the treatment inside the aggregates of AR. The last treatment, the most dissimilar one, was the clear-cut in the AR treatment (Figure 4). In the DCA analysis, the aggregates plots (ARI and CRI) were located closer to the primary forests than the harvested ones, and the treatment with shelterwood cuts (DR) was the most divergent compared to CRO and ARO. Diptera was the most generalist group and was located in the center of the graph. All the Lepidoptera morphospecies belonged to the Heterocera subclass, and were mainly captured under a closed canopy, as in the primary forests and inside the aggregates (CRI and ARI). Many Hymenoptera were also generalist Understanding biodiverity loss: an overview on forest fragmentation in South America 6 117 morphospecies, while others were more related to some of the treatments (e.g. 16 morphospecies were only found in DR). Figure 4. Cluster analysis and DCA for insect morphospecies richness. ARI: Agregated retention inside; ARO: Agregated retention outside; CRI: Combined retention inside; CRO: Combined retention outside; DR: dispersed retention; C: control. Forest management implications Harvesting of Nothofagus forests produces a large impact, mainly on insect diversity. Spagarino et al. (2001) determined a loss rate of one morphospecies every eleven years during the cycle of shelterwood cutting, and that it also allows the introduction of species from other environments which quickly colonize the impacted stands. Harvesting also modifies the community structures of understory plants and birds. These communities will recover to some extent when the forest structure closes its canopy. . Deferrari et al. (2001) found the same bird richness in the original and mature secondary forest, while Martínez Pastur et al. (2002) described the main impact of harvesting as the introduction of exotic plant species into the managed forest understory. One alternative for biodiversity conservation in managed landscapes is to maintain untouched, unproductive associated forest environments (Lencinas et al. 2005), where species could survive until the forest structure of the timber forest has recovered. These could be a solution for birds and understory plant species, but this alternative does not offer a solution for insect conservation, as many insect species only live in timber quality forests. Many alternative silvicultural methods have been proposed (Franklin et al. 1997), which conserve some of the original heterogeneity of the old-growth forest. One of these strategies proposes to select cut groups (Bava and López Bernal 2005), affecting a small percentage of the forest area, while another proposes to leave several degrees and retention types of the original old-growth forests (Martínez Pastur and Lencinas 2005; Vergara and Schlatter 2006). After three years of harvesting, the regeneration methods with aggregated retention appear as an effective alternative for biodiversity conservation. The aggregates conserve some of the original heterogeneity and the environmental conditions of the primary forests. Many Understanding biodiverity loss: an overview on forest fragmentation in South America 7 118 species (some plants, such as Viola magellanica, and many insect morphospecies) only survive under these conditions. The size, shape, and distribution of the retention in the timber forest, as well as the distribution patterns of the different types of unproductive associated environments along the forest landscape, must be considered to develop efficient long-term conservation strategies that allow planning for sustainable forest management. Finally, the BACI approach (Before-After-Control-Impact) is highly recommended because it allows the control of environmental variation between sites and year-to-year fluctuation in several conditions (Niemelä 2001). References Bava, J. & López Bernal, P.M. 2005. Cortas de selección en grupo en bosques de lenga. IDIA-XXI 5, 39-42. Caldentey, J., Ibarra, M., Promis, A. & Hernández, P. 2005a. Effects of shelterwood system on photosynthetically active radiation (PAR) and plan regeneration in Nothofagus pumilio stands in Chile. International Forestry Review 7, 46. Caldentey, J., Ibarra, M. & Promis, A. 2005b. Microclimatic variations in a Nothofagus pumilio forest caused by shelterwood systems: Results of seven years of observations. International Forestry Review 7, 46. Correa, M.N. 1969-1998. Flora Patagónica. Colección Científica INTA Tomo 8. Parts II, III, IVb, V, VI y VII. Buenos Aires, Argentina. Cuevas, J. 2000. Tree recruitment at the Nothofagus pumilio alpine timberline in Tierra del Fuego, Chile. Journal of Ecology 88, 840-855. Deferrari, G.; Camilion, C.; Martínez Pastur, G. & Peri, P. 2001. Changes in Nothofagus pumilio forest biodiversity during the forest management cycle: Birds. Biodiversity and Conservation 10, 2093-2108. Franklin, J., Berg, D., Thornburgh, D. & Tappeiner, J. 1997. Alternative silvicultural approaches to timber harvesting: Variable retention harvest systems. In: K. Kohm & J. Franklin, J. (Eds.). Creating a Forestry for the 21st Century. Island press, Washington. 111- 139. Gea, G., Martínez Pastur, G., Cellini, J.M. & Lencinas, M.V. 2004. Forty years of silvicultural management in southern Nothofagus pumilio (Poepp. et Endl.) Krasser primary forests. Forest Ecology and Management 201, 335-347. Lanfranco, D. 1977. Entomofauna asociada a los bosques de Nothofagus pumilio en la región de Magallanes. 1º parte: Monte Alto (Río Rubens, Última Esperanza). Anales del Instituto de la Patagonia 8, 319-346. Lencinas, M.V. 2005. Biodiversidad en el bosque productivo de Nothofagus pumilio y sus ambientes asociados en Tierra del Fuego. Tesis doctoral en Agronomía de la Universidad Nacional de Sur. Bahía Blanca. Lencinas, M.V.; Martínez Pastur, G.; Medina, M. & Busso, C. 2005. Diversity and density of birds in timber Nothofagus pumilio forests and its unproductive associate environments. Biodiversity and Conservation 14, 2299-2320. Martínez Pastur, G., Cellini, J.M., Peri, P., Vukasovic, R. & Fernández, M.C. 2000. Timber production of Nothofagus pumilio forests by a shelterwood system in Tierra del Fuego (Argentina). Forest Ecology and Management 134, 153-162. Martínez Pastur, G.; Peri, P.; Fernández, C.; Staffieri, G. & Lencinas, M.V. 2002. Changes in understory species diversity during the Nothofagus pumilio forest management cycle. Journal of Forest Research 7, 165-174. Understanding biodiverity loss: an overview on forest fragmentation in South America 8 119 Martínez Pastur, G. & Lencinas, M.V. 2005. El manejo forestal en los bosques de Nothofagus pumilio en Tierra del Fuego. IDIA-XXI 5, 107-110 Martínez Pastur, G., Lencinas, M.V., Escobar, J., Quiroga, P., Malmierca, L. & Lizarralde, M. 2006. Understory succession in areas of Nothofagus forests affected by Castor canadensis in Tierra del Fuego (Argentina). Applied Vegetation Science 9, 143-154. Moore, D.M. 1983. Flora of Tierra del Fuego. Anthony Nelson, England – Missouri Botanical Garden, USA. Narosky, T. & Yzurieta, D. 1987. Guía para la identificación de aves de Argentina y Uruguay. Asociación Ornitológica del Plata. Buenos Aires. Niemelä, J. 2001. Carabid beetles (Coleoptera: Carabidae) and habitat fragmentation: a review. European Journal of Entomology 98, 127-132. Oliver, I. & Beattie, A.J. 1993. A possible method for the rapid assessment of biodiversity. Conservation Biology 7, 562-568. Pulido, F.; Díaz, B. & Martínez Pastur, G. 2000. Incidencia del ramoneo del guanaco (Lama guanicoe) sobre la regeneración de lenga (Nothofagus pumilio) en bosques de Tierra del Fuego, Argentina. Investigación Agraria: Sistemas y Recursos Forestales 9, 381-394. Richards, O.W. & Davies, R.G. 1984. Tratado de entomología Imms. Volumen 2: Clasificación y Biología. Ed. Omega, Barcelona. Romoser, W.S. & Stoffolano, J.G. 1998. The science of entomology. WCB/McGraw-Hill, Boston. Rosenfeld, J.M., Navarro Cerrillo, R.M. & Guzmán Álvarez, J.R. 2006. Regeneration of Nothofagus pumilio (Poepp. et Endl.) Krasser forests after five years of seed tree cutting. Journal of Environmental Management 78, 44–51. Solervicens, J. 1995. Enthomology. Informe del Subproyecto 94-14. Estudios de línea base: Proyecto Río Cóndor, Chile. Spagarino, C.; Martínez Pastur, G. & Peri, P. 2001. Changes in Nothofagus pumilio forest biodiversity during the forest management cycle: Insects. Biodiversity and Conservation 10, 2077-2092. Vergara, P. & Schlatter, R. 2006. Aggregate retention in two Tierra del Fuego Nothofagus forests: Short-term effect on bird abundance. Forest Ecology and Management 225, 213- 224. ------ Understanding biodiverity loss: an overview on forest fragmentation in South America 9 120 Estrategias de ordenamiento territorial y conservación de la naturaleza en la Eco-región de las Yungas (noroeste de Argentina) L. I. Bachmann1, A. G. Frassetto1 , C. L. Daniele1 1 Programa de Investigación y Desarrollo en Reservas de Biosfera (ProMAB) Instituto de Geografía, Facultad de Filosofía y Letras, Universidad de Buenos Aires, Puán 480, (1406) Buenos Aires, Argentina. E-mail: [email protected]/ [email protected] / [email protected] Abstract The Eco-region of the Yungas (subtropical mountain rainforest) is one of the main centers of biological diversity in Argentina. Increasing threats, specially caused by the non-arranged distribution of human activities, affect this area and its patrimony in a rising way. From here emerges the need of national and regional organizations of planning and implementing united actions of conservation and territorial arrangement, together with international cooperation, to consider the great diversity of involved actors, and the multiple dimensions that overcome the problematic. Three cases focused on different ways and strategies of territorial arrangement are presented on three different scales: in a local-regional level, the BioAndes Project of identification of the impacts produced over the Calilegua National Park caused by the changes in the public politics about the conservation of biodiversity in the Andes; in a regional scale, the revision of the Zonation of the Yungas’ Biosphere Reserve (MAB- UNESCO); and in a binational scale, the planning of the Binational Ecological Corridor Tariquía - Baritú – Calilegua (Bermejo River Binational Commission-OAS-UNEP-GEF) about the Eco-region of the Yungas in Bolivia and in Argentina. The different instances of public participation carried out and the main achievements and challenges found are put forward. Palabras claves: ordenamiento territorial, biodiversidad, conservación, Yungas, Reserva de Biosfera, corredor ecológico Introducción El ordenamiento territorial constituye la expresión espacial de las políticas económicas, sociales, culturales y ecológicas de una sociedad. Su objetivo es promover el desarrollo equilibrado de las regiones y la organización física del espacio según un concepto rector (Consejo de Europa, 1993). Su implementación se realiza en un escenario dinámico por los procesos que se manifiestan en el territorio, considerando tanto las dinámicas naturales como los procesos y los actores sociales, apuntando a una equidad socioespacial y a un mejoramiento de la calidad de vida que trascienda el mero crecimiento económico. En Argentina, la necesidad de un ordenamiento territorial es reconocida en ámbitos académicos y de ONGs así como por sectores gubernamentales. Sin embargo, son limitadas las iniciativas que lo incluyen como instrumento concreto de planificación y de resolución de Understanding biodiverity loss: an overview on forest fragmentation in South America 1211 conflictos territoriales. En consecuencia, se evidencia un mosaico complejo y heterogéneo e incluso frecuentemente conflictivo de usos del suelo, como resultado de la aplicación de lógicas sectoriales e inconexas sobre el territorio. Esta condición es característica del área de estudio, la Eco-región de las Yungas (en las provincias de Salta y Jujuy, Argentina), en donde la conservación de la naturaleza se ve amenazada por diversas actividades humanas, principalmente el desmonte. Como respuesta, se han desarrollado estudios y propuesto diversas estrategias de ordenamiento territorial, a fin de orientar los procesos de transformación espacial hacia usos sustentables del territorio. Al respecto se presentan tres ejemplos con escalas diferentes: el Proyecto BioAndes a escala local-regional; la Reserva de la Biosfera de las Yungas (RBYun) a escala regional y el Corredor Ecológico Binacional Tariquía - Baritú - Calilegua a escala binacional. En este trabajo se analizan los principales productos, logros y desafíos de los estudios y estrategias analizados. La Eco-región de las Yungas en la actualidad En Argentina, la Eco-región de las Yungas se extiende como una franja longitudinal ubicada sobre las laderas orientales de las Sierras Subandinas (Figura 1), en las provincias de Jujuy, Salta, Tucumán y Catamarca, entre los 400 y 3.000 msnm. Pueden distinguirse distintos pisos de vegetación: la selva pedemontana, hasta los 800 m; la selva montana, entre los 800 y 1.700 m; el bosque montano, entre los 1.700 y 2.500 m; y los pastizales de altura, desde los 2.500 m en adelante, que se alternan con manchones de bosque montano y arbustales. Esta Eco-región constituye uno de los Figura 1. Localización de la Eco- principales centros de biodiversidad de Argentina región de las Yungas (Daniele y otros, 2002). Contiene el 60% de las especies de aves del país, y especies de mamíferos de importancia para la conservación, como el yaguareté (Panthera onca), la taruca (Hippocamelus antisensis), y el tapir (Tapirus terrestris). Además presenta un alto grado de endemismos, y posee especies de valor económico. Brinda importantes servicios ambientales ya que constituye la mayor fuente de agua de la región (altas cuencas de los ríos Bermejo y Pilcomayo), cumple funciones estabilizadoras de laderas y constituye un enorme sumidero de carbono. Por otro lado, posee valiosos recursos paisajísticos y una amplia diversidad cultural. Con respecto a las amenazas, se cuentan principalmente la agricultura (cultivo intensivo de caña de azúcar, tabaco, soja y hortalizas), la deforestación (para aprovechamiento maderero y apertura de tierras para cultivo), la construcción de infraestructura, como vías de comunicación y gasoductos, y la extracción de petróleo. Aunque con menor impacto, también constituyen amenazas la caza de fauna silvestre, especialmente de Fuente: Sistema de Información de la Biodiversidad, Administración de Parques Nacionales. Understanding biodiverity loss: an overview on forest fragmentation in South America 1222 grandes mamíferos, y el pastoreo de ganado bovino. Como consecuencia se percibe un avanzado proceso de fragmentación del paisaje natural, que genera un continuo empobrecimiento del bosque nativo. La selva pedemontana es el piso de vegetación más afectado, y prácticamente ha desaparecido por desmonte. Tales transformaciones agravan los procesos de erosión, sedimentación y desbordes fluviales, aumentan el proceso de reducción y pérdida de poblaciones silvestres e incrementan los procesos y riesgos de contaminación. Con respecto a las características socioeconómicas, en el oeste y norte del área de estudio se encuentran sectores caracterizados por una estructura netamente rural marginal, que practican actividades agropecuarias de subsistencia, y presentan índices medios a altos de necesidades básicas insatisfechas. Además, se han detectado procesos migratorios de pequeños productores desde estas áreas hacia las ciudades capitales provinciales, a causa de los efectos de la crisis económica. A modo de actividad productiva sustituta, algunos de ellos Figura 2. Localización de las estrategias realizan horticultura de primicia en las periferias de ordenamiento en la Eco-región de las urbanas, cerca del mercado de consumo y Yungas en Argentina comercialización. Otros se emplearon en el sector público. En el este y sur del área de estudio se identifican sectores con un alto dinamismo económico que, aunque siempre ligado mayoritariamente al sector agropecuario, presenta un continuo proceso de modernización y crecimiento, protagonizado por sectores de alto nivel económico (capitales transnacionales diversificados y/o integrados). Estos sectores tienden, en el actual proceso de globalización, a concentrar aún más la riqueza a partir de procesos fuertemente expansivos, logrados especialmente a través de la incorporación de nuevas tierras y la implementación de nuevos cultivos, competitivos en las condiciones actuales del mercado. En los próximos parágrafos se presentan tres casos en la Eco-región de las Yungas (Figura 2), donde se tratan las problemáticas generadas por las amenazas a la conservación y alternativas de resolución a través del Fuente: Elaboración propia ordenamiento territorial. Un diagnóstico de procesos territoriales a escala local-regional: el proyecto Bioandes El proyecto Bioandes de “Monitoreo y Modelización de los Impactos producidos por los cambios en las Políticas Públicas sobre la Conservación de la Biodiversidad en los Andes” Understanding biodiverity loss: an overview on forest fragmentation in South America 1233 realizado entre 1998 y 2003 (Unión Europea ERBIC18CT980299)1 tuvo como objetivo investigar, desarrollar y aplicar técnicas innovadoras para el monitoreo y la conservación de la biodiversidad en los alrededores de tres áreas protegidas en la selva subtropical de montaña de los Andes: el Parque Nacional Calilegua (PNC) en Argentina, el Parque Nacional Carrasco en Bolivia y el Parque Nacional Río Abiseo en Perú. Sus principales resultados contribuyeron a un diagnóstico de las amenazas a la conservación de la biodiversidad, en Argentina especialmente a causa de la evolución de los usos del suelo en el entorno del PNC (Provincia de Jujuy), y un modelo conceptual de los principales procesos que causaron los conflictos identificados, a escalas global, regional y local. Estos resultados constituyeron luego un insumo para el diseño de políticas de ordenamiento territorial en la zona, tales como la creación y consolidación de la RBYun y el Corredor Ecológico Binacional Tariquía - Baritú - Calilegua. Los principales productos aportados por el Proyecto Bioandes fueron: • un mapa de amenazas y potencialidades (Figura 3) para la conservación de la biodiversidad en el PNC, elaborado en base al desarrollo e integración de estudios sectoriales a diferentes escalas: aspectos biológicos, sociales y usos del suelo en el PNC y su entorno; análisis de la distribución y abundancia relativa de especies de la flora y de anfibios, reptiles, aves y mamíferos, la comparación de sitios con diferentes tipos e intensidad de uso; del relevamiento a escala local y análisis de los usos de los recursos y de la economía doméstica en un conjunto de comunidades en el entorno del PNC (Departamento Valle Grande); del relevamiento a escala regional de información sociodemográfica y económica (Departamentos de Ledesma y Valle Grande); y del relevamiento regional del uso del suelo2. El mapa de la Figura 3 muestra la localización y la diversidad de las amenazas según su origen. Son más intensas en el sector este y sur, debido especialmente a la presencia de sectores agroindustriales históricos (cuya superficie agrícola se encuentra en avance sobre los pisos de selva pedemontana y montana), de los centros urbanos de mayor importancia en el área y de una vasta red de transporte vial y de energía que comunica sus ciudades. En el oeste, por el contrario, se identificaron amenazas de menor intensidad, vinculadas con el pastoreo estacional de ganado vacuno, la extracción de leña, la caza menor y el turismo de bajo impacto. • Un modelo conceptual gráfico que presenta las principales relaciones entre los actores y procesos a niveles global, nacional y regional, que afectan la biodiversidad en el área de estudio. A nivel global y nacional, los principales factores involucrados son: globalización, modernización tecnológica, políticas neoliberales, integración económica, achicamiento del Estado, polarización social, transnacionalización de la economía, concentración del capital 1 Las instituciones participantes fueron: University of Leicester – Departament of Gepgraphy (Gran Bretaña), Universidad de Buenos Aires – Instituto de Geografía (Argentina), Universidad Nacional de Salta (Argentina), Administración de Parques Nacionales (Argentina), University of Gotemburg (Suecia), Zoological Museum of Copenhagen (Dinamarca), University of Helsinki (Finlandia), Universidad Mayor de San Simón (Bolivia), Asociación Peruana para la Conservación de la Naturaleza (Perú). 2 Dicho relevamiento se realizó sobre la base del procesamiento de imágenes satelitales Landsat TM a escala 1:100.000, y se desarrolló un análisis multi-escalar espacio-temporal integrando la visión local y regional, el uso histórico (1973), y su evolución (1986, 1997 y 2000). Understanding biodiverity loss: an overview on forest fragmentation in South America 1244 • Figura 3. Mapa síntesis de amenazas y potencialidades para la conservación de la biodiversidad en el Parque Nacional Calilegua (Provincia de Jujuy, Argentina) Fuente: Elaboración propia. Proyecto BioAndes. privado, diversificación y reconversión productiva, limitaciones en las gestiones de conservación e internacionalización de la gestión ambiental3. A nivel regional y local, los principales factores, vinculados en múltiples formas a los de nivel global y nacional, son: intensificación de la producción agrícola y agroindustrial, concentración territorial de capital e infraestructura, avance de la frontera agropecuaria, explotación de hidrocarburos, aumento de la actividad maderera, aumento del turismo, degradación de bosques y suelos, fragmentación del paisaje, pérdida de hábitat, disminución de la capacidad de carga y pérdida de diversidad. Ambos productos fueron un importante aporte para identificar el estado de situación de la conservación y la configuración espacial de las actividades en el entorno del PNC, y de los principales procesos políticos, económicos, sociales y ecológicos que se desarrollan en la zona. 3 Cabe aclarar que muchos de los conceptos de nivel global y nacional son compartidos por los diagnósticos y esquemas conceptuales de Bolivia y Perú. Understanding biodiverity loss: an overview on forest fragmentation in South America 1255 Una estrategia de ordenamiento territorial a escala regional: la Reserva de Biosfera de las Yungas (RBYun) Las Reservas de Biosfera (RBs) son áreas de ecosistemas reconocidas internacionalmente por el Programa sobre el Hombre y la Biosfera (MAB) de la UNESCO para contribuir a la conservación de paisajes, ecosistemas, especies y poblaciones, fomentar un desarrollo económico, cultural, social y ecológicamente sustentable a nivel local, y proporcionar el apoyo logístico para la investigación científica, la educación ambiental y el intercambio de información. Territorialmente, las RBs están organizadas en diferentes zonas: núcleo, de amortiguamiento y de transición, delimitadas según los objetivos de manejo y protección que se implementan en las áreas legalmente protegidas y según los usos del suelo desarrollados en el resto de la Reserva (UNESCO, 2000). Como resultado de una iniciativa local de la provincia de Salta y la Administración de Parques Nacionales (APN) la RBYun, fue aprobada por UNESCO en el año 2002. Su objetivo es conservar la mayor biodiversidad, mantener las funciones y procesos ecológicos, integrar a las comunidades a la gestión y promover el desarrollo sustentable, en sus dimensiones social, ecológica y económica para ese sector de la Eco-región de las Yungas. Posee una superficie aproximada de 1.550.000 ha, de las cuales 700.000 se encuentran cubiertas por bosques, 500.000 ha por pastizales naturales y casi 2.000 ha por parcelas agrícolas y barbechos de agricultura migratoria. Sólo 170.000 ha han sido destinadas a la conservación a través de áreas naturales protegidas. La mayoría de la superficie de la RBYun se encuentra bajo el dominio de propietarios privados, y unas 140.000 ha se encuentran bajo manejo tradicional de comunidades indígenas. La población, mayoritariamente rural y con una alta incidencia de menores de 14 años de edad, presenta grados medios a altos de necesidades básicas insatisfechas. En un proceso promovido por la Secretaría de Ambiente y Desarrollo Sustentable de Salta (SEMADES) y la APN, con la participación de la Secretaria de Producción y Medio Ambiente de Jujuy, ONGs y el apoyo de Universidades de Buenos Aires, Salta y Jujuy, se definió una primera configuración de la RBYun (Figura 4), que cumple los requerimientos de la UNESCO: • Aplica el modelo de Zonificación propuesto por el Programa MAB de la UNESCO, ya que responde a los requerimientos referentes a las funciones y usos de las zonas núcleo, de amortiguamiento y de transición. • Abarca una superficie adecuada en relación con los objetivos de conservación y desarrollo. • Comprende un espacio diverso, compuesto, por un lado, por un mosaico de sistemas ecológicos (pisos ecológicos) representativos de las Yungas; y por otro, de un abanico de usos del suelo, en el cual los más intensivos se ubican hacia el este (dominancia de actividades agropecuarias con alta tecnificación), descendiendo el grado de intervención humana hacia el oeste, con zonas de mínima intervención antrópica en las zonas núcleo. • Otorga protección a una Eco-región que presenta numerosas especies amenazadas. • Protege a una parte importante de la alta cuenca del río Bermejo. • Brinda protección a sectores con un valioso patrimonio cultural, ya que incluye numerosos e importantes sitios arqueológicos. Algunas debilidades identificadas son la fragmentación de unidades catastrales, y un limitado grado de participación pública durante el proceso de creación de la Reserva y de diseño de su zonificación. Understanding biodiverity loss: an overview on forest fragmentation in South America 1266 Con el fin de mejorar el diseño mencionado y resolver las debilidades indicadas y para fortalecer el proceso de consolidación de la RBYun y formular su Plan de Gestión, en el año 2003 se realizó, en el marco del Programa Estratégico de Acción para la Cuenca del Río Bermejo (COBINABE4-PNUMA-OEA-FMAM5), una revisión de la zonificación. Con esta iniciativa se buscaba optimizar el modelo de ordenamiento territorial propuesto en la zonificación inicial, teniendo en cuenta y profundizando el conocimiento de los principales conflictos y escenarios territoriales identificados a partir de diversos estudios, en especial del proyecto Bioandes. Para ello se desarrolló y se montó un Sistema de Información Geográfico (utilizando el programa ArcView), y se organizaron talleres y reuniones del Comité de Gestión de la RBYun. A través de la consideración de ocho criterios de revisión6 se elaboró una nueva versión de zonificación (Figura 5), que propone: • Una mayor representatividad del principal patrimonio natural, la Eco-región de las Yungas, ya que incluye una mayor superficie de bosque montano en el oeste, y de parches remanentes de selva pedemontana en el este. • Un aumento de la superficie de la zona de amortiguamiento, reforzando los objetivos de conservación y desarrollo y otorgando mayor protección a las zonas núcleo. • El refuerzo de la conservación de especies amenazadas a través de la preservación de corredores biológicos en el interior de la Reserva. • El refuerzo de la protección de las altas cuencas a través de la ampliación del límite externo oeste (nacientes de ríos) y la consideración de las divisorias de aguas para la delimitación externa y la zonificación, facilitando la gestión integrada de las cuencas. • La inclusión de mayor cantidad de sitios de conservación del patrimonio cultural. • La corrección de límites que implican un marcado efecto de borde, como ocurre en la zona adyacente al parque Provincial Potrero de Yala, en el límite suroeste de la Reserva. • Sumar a los límites catastrales como criterio de delimitación. • Poner en consideración criterios adicionales, como el de líneas de pendientes, en zonas conflictivas desde el punto de vista de la intensidad de los efectos de usos del suelo, en la zona agroindustrial y urbana del este. El principal propósito de esta nueva propuesta de zonificación se vincula con aumentar la conectividad entre los ecosistemas más amenazados y reducir las debilidades del diseño original. Los principales desafíos de la implementación de la nueva propuesta de zonificación se basan en las dificultades encontradas referidas a la participación y el compromiso de la población de la zona, especialmente los grandes propietarios de tierras, los empresarios vinculados al sector forestal y los agroindustriales. La nueva propuesta fue presentada a la SEMADES y puesta en consideración del Comité de Gestión de la RBYun. 4 Comisión Binacional para el Desarrollo de la Alta Cuenca de la Cuenca del Río Bermejo y Río Grande de Tarija 5 Fondo para el Medio Ambiente Mundial (Global Environment Facility-GEF) 6 Los criterios seleccionados fueron: representatividad del principal patrimonio natural del área, usos del suelo, existencia de sitios de importancia para la conservación del patrimonio natural y cultural, minimización del efecto borde, situación socioeconómica, líneas divisorias de cuencas y microcuencas, otros fenómenos geográficos lineales (infraestructuras), y límites catastrales Understanding biodiverity loss: an overview on forest fragmentation in South America 1277 Figura 4. Zonificación de la Reserva de Biosfera de las Yungas Figura 5. Nueva propuesta de zonificación de la RBYun Fuente: Elaboración propia. Fuente: Elaboración propia basándose en: Bachmann, L; Volante, J; Cabral, M; Frassetto, A. & Negri, M. (2003). Understanding biodiverity loss: an overview on forest fragmentation in South America 8 128 Una estrategia de ordenamiento territorial a escala binacional: el Corredor Ecológico Tariquía (Bolivia) - Baritú - Calilegua (Argentina) Un Corredor Ecológico es una porción del territorio con características singulares que cumple una función de conectividad de hábitats para una gran variedad de especies7. En este caso, se trata de un corredor a escala regional que involucra dos países, Argentina y Bolivia y tres áreas protegidas de importancia como son la Reserva Nacional de Flora y Fauna de Tariquía (Bolivia) y los Parques Nacionales Baritú y Calilegua (Argentina). Su desarrollo y consolidación fueron incluidos en el conjunto de proyectos del Programa Estratégico de Acción para el Desarrollo Sustentable de la cuenca del Río Bermejo (PEA - Bermejo) ejecutado por la COBINABE, como una estrategia innovadora de escala regional adoptada para el desarrollo sustentable. La superficie del Corredor en Argentina es del orden de las 1.120.000 ha, con alrededor del 2,5% de tierras agrícolas y un 97,5% de superficie con un bajo grado de antropización, incluyendo mas de 165.000 ha de áreas protegidas. Casi un 20% del Corredor se encuentra en fincas actuales del Pueblo Kolla, habitantes desde épocas prehispánicas. La densidad poblacional es relativamente baja, debido al bajo grado de ocupación general y a la exclusión de los principales centro urbanos en su diseño preliminar. Los principales asentamientos son Los Toldos, Yuto y Colonia Santa Rosa, que suman aproximadamente 21.000 habitantes. Este Corredor, conformado por un mosaico de tierras con los mas variados usos, desde la protección hasta el uso intensivo, deberá ser manejado de manera integrada para garantizar la supervivencia del mayor número posible de especies, a través de la continuidad de sus hábitats y procesos ecológicos y del mantenimiento o restauración de la conectividad ecológica de sus ecosistemas. Hasta el momento, no existe ningún marco legal que lo avale y en el plano gubernamental, las principales acciones se han emprendido en el marco de la COBINABE, desde donde se ha promovido la utilización de la actual estructura institucional de la RBYun como marco para la futura gestión del Corredor. Su implementación se deberá realizar a través de aplicación diferencial de un conjunto de herramientas y acciones estratégicas de diferente escala y características, tendientes a la modificación de los actuales usos del suelo y de las tendencias dominantes hasta el momento. Su consolidación está prevista con la participación de las jurisdicciones intervinientes (nacionales, provinciales y departamentales), los organismos públicos con incumbencias en el tema, las universidades, las ONGs y otros actores de la sociedad civil, en ambos países. El principal objetivo del corredor es asegurar la conectividad biológica de las áreas protegidas de la Eco-región de las Yungas en Argentina y Bolivia, mediante el manejo integral de los recursos naturales y la gestión de dichas áreas y su entorno. De esta forma constituye un proyecto complementario de los ya existentes esfuerzos de conservación y desarrollo, como las áreas naturales protegidas y la RBYun. 7 Según Noss (1991) pueden distinguirse tres tipos de corredores para la vida silvestre a distintas escalas espacio-temporales. El corredor local, que conecta parches de hábitat próximos entre si (por. ej. parches de bosque usando franjas estrechas de hábitat adecuado como cortinas de árboles o arbustos, para permitir desplazamientos de pequeños mamíferos), estos corredores son exclusivamente hábitat de borde, son útiles para especies de hábitat interior. El Corredor zonal, que funciona a escala del mosaico de paisaje, para permitir desplazamientos diarios, estacionales y/o permanentes de especies de borde y de hábitat interior e integran un mosaico de reservas a nivel de paisaje; este tipo de corredor incluye franjas anchas de bosque que conectan reservas separadas, bosques fluviales o hábitats que siguen gradientes topográficos. Finalmente, el corredor regional es la mayor escala para un corredor y conecta reservas en redes regionales. Understanding biodiverity loss: an overview on forest fragmentation in South America 1299 Para ello, las provincias de Salta y Jujuy y la APN en Argentina, junto a la Prefectura de Tarija en Bolivia deberán promover, desde un ordenamiento territorial escala regional, hasta pequeñas actividades de escala local en los correspondientes municipios tendientes a favorecer la conectividad de sus ecosistemas. Puede mencionarse que el Municipio de Los Toldos (provincia de Salta) ya ha elaborado y aplicado su propio ordenamiento territorial. A modo de síntesis del proyecto, en el marco del Plan de Gestión del Corredor Binacional se identificaron y diseñaron distintos proyectos para consolidarlo, agrupados en tres componentes: ambiental, institucional y económica. Entre los proyectos ambientales se destacan: la elaboración e implementación de un Plan de Ordenamiento Territorial para el Corredor en su conjunto, la identificación de sitios críticos de conectividad, la evaluación y monitoreo de actividades agrícolas con indicadores de sustentabilidad, la reducción del conflicto entre producción ganadera y los grandes felinos predadores. La institucionalización del Corredor se promoverá a través de: el fortalecimiento del Comité y Subcomités locales de la RBYun, la creación de una Reserva de la Biosfera Transfronteriza de las Yungas (binacional) y la elaboración y aplicación de un Plan de Gestión para el largo plazo. La búsqueda de la sustentabilidad económica del ordenamiento propuesto será promovido y evaluado a través de un conjunto de proyectos piloto. Finalmente debe mencionarse que el Fondo Francés para el Medio Ambiente y la Fundación Proyungas, se encuentran actualmente en el desarrollo del Proyecto Alto Bermejo como una “propuesta de ordenamiento territorial y planificación estratégica basada en la generación de información ecológica y en la participación de distintos sectores de la sociedad”. Integración de las estrategias de ordenamiento territorial La Eco-región de las Yungas, ha sido desde la última década escenario de un conjunto de iniciativas de diagnostico y ordenamiento territorial que se han ido complementando y articulando en una compleja estructura. Las tres principales aquí analizadas, se desarrollaron desde diferentes marcos institucionales y a diferentes escalas de trabajo, pero con objetivos y productos que se integran en la construcción del desarrollo sustentable de esta Eco-región, incluso fuera de las fronteras de Argentina hacia Bolivia. El Proyecto Bioandes, en un estudio de caso a escala local en el PNC (provincia de Jujuy) y su entorno, logró diferentes resultados en el conocimiento de la biodiversidad y los procesos ecológicos asociados, en la identificación de los principales conflictos entre los diferentes actores sociales y los usos del suelo) y en la construcción de un modelo conceptual que vincula a diferentes escalas (local, nacional y global) los procesos transformadores del ambiente y sus consecuencias sobre la biodiversidad, centrado en la década 1990-2000. La APN y los actores locales fueron los principales destinatarios de estos resultados, utilizados en parte para la gestión dentro del parque nacional y en relación con las actividades en su entorno directo. Los resultados de Bioandes y la experiencia de los equipos técnicos gubernamentales de la APN y de la Universidad de Buenos Aires, se sumaron a la iniciativa de la SEMADES para la elaboración de la propuesta de RBYun. El conocimiento de los procesos de transformación del territorio, fue un insumo de importancia para definir la zonificación interna de la reserva de biosfera, que incluye en PNC. El proceso de elaboración de la propuesta a la UNESCO requirió inicialmente de un proceso participativo en la definición de las estructuras de gestión (un comité central y subcomités locales) y en su zonificación. A partir del 2002, el posterior proceso de Understanding biodiverity loss: an overview on forest fragmentation in South America 13010 implementación de la RBYun y la inclusión de los diferentes actores sociales en las zonas núcleo, de amortiguamiento o de transición evidenciaron las restricciones emergentes a este ordenamiento territorial. Como consecuencia de ello y en el marco del Programa Estratégico de Acción para la Cuenca del Río Bermejo (PEA-Bermejo), se inició en el 2003 un proceso de revisión de la zonificación, nuevamente con un proceso de consulta y participación, esta vez centralizado a través de los Subcomités locales, representantes de los diferentes actores locales y canal de comunicación y toma de decisiones. Esta revisión permitió identificar nuevos componentes y procesos naturales, sociales y económicos que debieron ser tenidos en cuenta en el ordenamiento territorial que significa la zonificación de la RBYun. La RBYun constituye entonces la principal estructura de gestión y ordenamiento territorial, que con criterios hidrológicos, ecológicos, sociales y administrativos, definió una unidad espacial funcional y crea su propia arquitectura institucional y de gestión, de conformación intergubernamental e intersectorial. Esta iniciativa, apoyada por el Programa Estratégico de Acción para la Cuenca del Río Bermejo y la COBINABE, se constituye como un caso demostrativo para el desarrollo sustentable de la Eco-región y un punto de partida para el desarrollo de un proyecto de ordenamiento territorial más amplio, el Corredor Ecológico. En efecto, esta propuesta, cuyos primeros estudios de base fueron financiados por el PEA- Bermejo en el período 1997-2000 y profundizados hasta el 2005, identificó las acciones prioritarias que deben ser implementados a fin de consolidar el Corredor entre sectores de la Eco-región de las Yungas en Bolivia y Argentina. La discusión sobre los objetivos, alcances y modalidad de gestión del Corredor, se dio fundamentalmente con un importante grado de participación en diferentes talleres, entre los mismos actores sociales de la RBYun y sobre su misma estructura de gestión, es decir el Comité y los Subcomités locales. El desarrollo del proyecto incluyó la sensibilización y comunicación de los objetivos a las autoridades gubernamentales en sus diferentes jurisdicciones, logrando diferente grado de compromiso según los casos. Como resultado de esta etapa, se ha reconocido la necesidad de no generar nuevos mecanismos de gestión para el Corredor, sino de aprovechar la existencia y experiencia del Comité de Gestión de la RBYun ya conformado y en el marco de la COBINABE, la propuesta de formación de una reserva transfronteriza, ampliando la RBYun hacia Bolivia. Conclusiones Sobre un mismo territorio se han desarrollado tres iniciativas de conservación de la naturaleza y usos del suelo, con leves diferencias en la delimitación geográfica de dos de ellas, la RBYun y el Corredor. La experiencia del PNC fue la mas temprana y con mayor detalle a escala comparativamente local. Las otras dos, con objetivos específicos de conservación de la naturaleza y de continuidad de los hábitats y procesos ecológicos dirigidos al mantenimiento o restauración de la conectividad ecológica, necesitan de un ordenamiento territorial que defina los usos del suelo prioritarios y limite las actividades no deseadas. Una primera configuración de este ordenamiento ya esta vigente en la RBYun y el principal desafió es su compatibilización y expansión a nivel binacional en función del Corredor Ecológico. Tales objetivos deben ser promovidos desde el Estado a través de una gestión participativa que incluya a todos los actores sociales involucrados, de la integración interdisciplinaria de los conocimientos científicos en el diseño técnico del modelo territorial y en su gestión, y de la articulación del conocimiento técnico con la visión política y con el acuerdo social. Understanding biodiverity loss: an overview on forest fragmentation in South America 13111 Bibliografía Bachmann, L.; Volante, J.; Cabral, C.; Frassetto, A. & Negri, M. (2003) Revisión de la Zonificación de la Reserva de Biosfera de las Yungas. Instituto de Geografía (UBA) - Secretaría de Medio Ambiente y Desarrollo Sustentable de Salta – PEA Bermejo (GEF-OEA PNUMA). Bachmann, L. (2006) Enfoque Ecosistémico: algunos aportes desde la Geografía para la conservación y el desarrollo sostenible en Reservas de Biosfera (RB). XXII Reunión Argentina de Ecología “Fronteras en Ecología: hechos y perspectivas”, Córdoba, Argentina. Brown, A.; Grau, H.; Malizia, L. & Grau, A. (2001) Los bosques nublados de Argentina. M. Kapelle & A. D. Brown (Eds.), Bosques nublados del Neotrópico. Editorial INBio, Costa Rica. 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(2003) Procesos territoriales en la Reserva de Biosfera de las Yungas, Argentina. Bertoncello, R.; A. Fani & C. Alessandri (comp.) Procesos territoriales en Argentina y Brasil. Universidad de San Pablo, Brasil, pp. 337-355. Daniele, C.; Somma, D.; Bachmann, L.; Aued, M. & Frassetto, A. (2002) International Symposium on Remote Sensing of Environment, CONAE, Buenos Aires, "Land Use Planning and Nature Conservation in the Northwest of Argentina: evolution of the landscape fragmentation and its consequences". Fundación Proyungas Fondo Francés para el Medio Ambiente (2006) Proyecto Alto Bermejo. Balance 2005-2007. Levy, G. & otros (1999) La reordenación socio-económica y espacial de las familias de Alto Calilegua: ¿desestructuración o modalidad estratégica de acceso y control de los recursos? II Taller Regional del Proyecto Bioandes, Calilegua, Jujuy, Argentina. Noss, R.F. (1991) Landscape connectivity: Different functions at different scales. Hudson, W. E. Landscape Linkages and Biodiversity. Washington, DC: Island Press; 1991; pp. 27-39. PEA-OEA-PNUMA-FMAM (2000) Diagnóstico Ambiental Transfronterizo de la Cuenca del Río Bermejo. Programa Estratégico de Acción para la Cuenca del Río Bermejo. Comisión Binacional para el Desarrollo de la Alta Cuenca del Río Bermejo y Grande de Tarija. Proyecto BioAndes (2002) Monitoring and Modelling the Impacts of Changing Government Policies on Biodiversity Conservation in the Andes. Instituto de Geografía, UBA; Univ. of Leicester; Univ. of Helsinsky, Zoological Museum of Copenhagen, Univ. San Simón de Bolivia, APECO de Perú; European Comision ERBIC18CT980299. Reboratti, C. (1998) El Alto Bermejo. Realidades y conflictos. La Colmena. Buenos Aires. Somma, D. (2006) Implementación del Corredor Biológico Baritú – Tariquía. Programa Estratégico de Acción para la Cuenca del Río Bermejo. COBINABE. Somma, D.; Aued, B. & Bachmann, L. (2003) The ecological network development in the Yungas, Argentina: planning economic y social aspects. R. Jongman & G. Pungetti (Eds.) Ecological Networks and Greenways. Concept, design, implementation. Cambridge Universidad Press, Cambridge, pp. 207 – 220. Understanding biodiverity loss: an overview on forest fragmentation in South America 13212 Ecological restoration of Fitzroya cupressoides, a long-lived conifer in southern Chile. A. Lara1, C. Echeverría1, O. Thiers2, F. Bustos1, E. Huss3, B. Escobar2 1 FORECOS Millennium Scientific Nucleus – Instituto de Silvicultura, Universidad Austral de Chile. Casilla 567, Valdivia, Chile. e-mail: [email protected] 2Instituto de Silvicultura, Universidad Austral de Chile. Casilla 567, Valdivia, Chile 3 Corporación Nacional Forestal, Oficina Provincial Valdivia, Chile Abstract Alerce (Fitzroya cupressoides) is a long-lived conifer, endemic to Southern Chile and adjacent Argentina. The destruction of these forests by human-set fires and logging in the Central Depression in Chile was almost total. The objective of this work is to systematize the experience of an experimental F. cupressoides plantation aimed at genetic conservation. Survival after five years in most sites, ranged from 77% to 84%. The lowest survival occurred in fertilized plants in a Sphagnum bog (< 13%). The annual periodic increment in height was highest in the moist micro site, excluding fertilized plants (mean value 12.8 -15.8 cm/year), with a maximum of 30 cm/year. The lowest height increments occurred in the Sphagnum bog (3.4- 5.0 cm/year). We propose an integral approach for the long- term restoration of F. cupressoides, combining natural processes and human interventions through plantations and exclosures, based on the understanding of ecological succession .H\ZRUGV $OHUFHJURZWKELRGLYHUVLW\WKUHDWHQHGWUHH6SKDJQXPERJ Introduction Alerce (Fitzroya cupressoides (Molina) Johnston) is an endemic evergreen conifer that may reach 5 m in diameter and 50 m in height. Fitzroya can live longer than 3,600 year (Donoso 1993a, Lara and Villalba 1993). Fitzroya forests have a unique composition and structure and a high economic value. Fitzroya grows in the temperate rain forests of southern Chile and adjacent parts of Argentina. Specifically, it occurs in Chile in three discontinues areas: Coastal Range (39º 50’ S – 43º 30’ S), Central Depression (41º 20’ S-41º 50’ S) and Andes (40º 30’ S – 43º 30’ S, Fig. 1) (CONAF et al., 1999). Although it was placed on Appendix I of the Convention on International Trade in Endangered Species (Trade) in 1975, and was declared a national monument in Chile in 1976, outlawing the cutting of live trees, enforcement of these laws has been problematic and illegal harvesting is still an important threat (Wolodarsky and Lara, 2005). According to the IUCN, Fitzroya is an endangered species (Oldfield et al., 1998). Fitzroya forests in the Central Depression were rapidly destroyed due to the expansion of the European settlement during the second half of the XIX century. It was thought that the species had completely disappeared in that area (Veblen et al., 1976; Veblen et al., 1995). However, further studies revealed the occurrence of the species in a dozen of small forest fragments (1-2 ha). The isolation of these populations has led to a genetic differentiation with respect to those populations occurring in the Andes and Coastal ranges (Alnutt et al., 1999; Premoli et al., 2003). Understanding biodiverity loss: an overview on forest fragmentation in South America 1 133 Within this context, we decided to start an ecological restoration experiment of Fitzroya in the Central Depression in order to determine its survival and growth rates. The objective of the present article is to describe the main achievements of this trial in terms of survival and growth of Fitzroya plants and water table changes as a consequence of vegetation recovery. From these results we propose recommendations for the restoration of the species. The present research is the first ecological restoration experiment conducted in Chile. Figure 1. Location map of the study area and sub-populations of F. cupressoides in the Central Depression in southern Chile. Methods Study Area The research was carried out in a small property named Chaqueihua (41º 26’ S y 73º 07’ W), owned by Mr. Alfredo Nuñez, where it is possible to find one of the well-conserved stands of Fitzroya in the Central Depression. The property is at 65 m elevation, near Puerto Montt city (Fig. 1). The trial was conducted in a 2.74-ha area that includes a secondary forest of Fitzroya. The prevailing climate is wet-temperate with strong oceanic influences (Di Castri and Hajek, 1976). Rainfall occurs throughout the year, with a mean annual precipitation of 1,783 m, concentrated in winter (June to September), and mean annual temperature of 9.9º C (Pezoa, 2003). Soil corresponds to an acidic, shallow, poorly drained soil referred to as ñadi (< 50 cm deep), derived from glacio-fluvial deposits which are classified as Gleysols (FAO- Understanding biodiverity loss: an overview on forest fragmentation in South America 2 134 UNESCO, 1971). Soil is rich in organic matter with a pH ranging from 4.3 to 5.4 (Tosso, 1985). Water-table fluctuation We monitored the water table oscillations through 16 plastic tubes 150 cm in length and 1” in diameter, named piezometers. These were distributed across the study area using a 15 x 15 m grid, which included the different conditions of drainage of the trials. The measurement of the water table depth respect to the ground surface was carried out twice a week from January 1999 to April 2005, with some interruptions. Establishment of the experimental plantation In April 1996 we collected seeds of Fitzroya from the trees that grow at Chaqueihua as well as from other subpopulations situated at less than 20 km in the Central Depression. In July –August 1997 we collected cuttings of Fitzroya in the same sub-populations. The maximum number of cuttings collected from each individual tree was 25. This in order to prevent an excessive reduction in the genetic variability of the plants in the restoration trial. In a first stage the plants were maintained in a nursery and then were put in a bed in an open site. In September 1998 we planted 360 Fitzroya plants. A percentage of the plants were fertilized (Phosphate, Nitrate and Calcium). Assessment of the plantation Success of the plantation was assessed estimating the survival and growth in height and diameter of the Fitzroya plants at year five after the plantation. A non parametric analysis of variance was conducted using one or two of the following factors: type of plant (cutting or seed), micro site (Sphagnum bog, very moist, moist) and fertilization (fertilised and not fertilized) (Table 1). Height growth was assessed in terms of the annual periodic increment (API) at the fifth year after the establishment of the plantation in 1999. Table 1. Number of live Fitzroya plants per treatment at the fifth year and at the beginning of the trial (figures in parenthesis), in the three tested micro sites for plants produced from seed and from cuttings. Sphagnum bog Very moist Moist Fertilized Not Fertilized Not Fertilized Not Plant fertilized fertilized Fertilized origin Seed 2 (16) 16 45 (84) 76 10 (18) 21 (26) (19) (99) Cutting 1 (9) 4 (7) 0 (0) 0 (0) 9 (37) 39 (45) Results Water-table fluctuations The piezometers located in the Sphagnum bog, very moist and moist micro sites exhibited seasonal oscillations in the depth of the water table. The minimum values were recorded in winter and spring (June to November) in which the water table reached a depth close to the soil surface (Fig. 2). Conversely, the maximum values were observed in summer and autumn Understanding biodiverity loss: an overview on forest fragmentation in South America 3 135 (December to April). These oscillations were associated with drainage limitations of the site and the seasonality in precipitations that are concentrated in winter, recharging the soil and causing the raise in the water table. 0 15 30 No data No data Water45 table depth [cm] 60 Feb Apr Jun Aug Oct Dec Feb Apr Apr Jun Aug Oct Dec Feb Apr Dec Feb Apr Jun Aug Oct Dec Feb Apr Jun Aug Oct Dec Feb Apr Dec 99 00 01 02 03 04 05 Hydrological years M icrosites Sphagnum bog Very m oist Moist Figure 2. Oscillations of the water table in different micro site conditions recorded between April 1999 and April 2005. The oscillations in the water table correspond to the mean values estimated from the reading of the piezometers within each given micro site. Total number of piezometers: 16. Survival and growth Survival at year five ranged between 77% and 84% for the moist and very moist micro sites without fertilization. In the Sphagnum peat bog the unfertilised plants reached survival rates between 57% and 84%. When plants were fertilized, survival at year five decreased to 22% - 55%, in the moist and very moist sites, whereas in the Sphagnum bog fertilization decreased survival rate to 11% - 13%. Annual periodic increment (API) in height reached the greatest values in the moist micro site, similar to the trend observed for survival of the plants (F= 1.4, P < 0,0001; Fig. 3). Results show an inverse relation between the mean value of API in height and its variation coefficient for the three micro sites studied. Understanding biodiverity loss: an overview on forest fragmentation in South America 4 136 30 Seed (F) Seed (NF) 25 Cutting (F) Cutting (NF) 20 15 10 Increase height in (cm) 5 0 Sphagnum bog Very moist Moist Micro sites Figure 3 Annual periodic increment (API) in height at year 5 in the three micro sites. F: fertilised, NF: not fertilised. Vertical bars indicate the standard deviation. Discussion Changes in the water-table level The fluctuations in the water table show a decreasing trend between 2002 and 2005 for the three micro sites that were analyzed (Fig 2). This decrease cannot be explained due to a reduction in the precipitations in this period, since rainfall has increased in this period. This change in the water table may be a response to the increase in both the interception of precipitation and evapotranspiration derived from the important increase in plant cover and biomass of grasses, forbs, shrubs and small trees in the area. This increase in plant cover and biomass is a result of the successional process determined by the exclosure of the area and elimination of grazing (Fig. 4). Nevertheless, more years of observation of the water table, and starting the monitoring of plant cover, biomass, as well as interception and evapotranspiration rates are needed to validate the proposed explanation. Understanding biodiverity loss: an overview on forest fragmentation in South America 5 137 Figure 4 Picture taken at the boundary of the experimental plantation, showing the increase in plant cover and biomass after seven years of exclosure compared to the area outside the fence that remains under a high grazing pressure. Photo taken by C. Echeverría in 2005. Survival and growth The survival rates after five years reported here for unfertilized conditions in moist and very moist micro sites are within the range reported for Fitzroya experimental plantations done in the Coastal Range (200 – 800 m of elevation above mean sea level) with survival rates between 65% and 75% ( Donoso et al., 2000; Donoso, 2003b). Nevertheless, survival rates when plants were fertilized decreased to 11% - 55% in the various micro sites. These results indicate that it would be better not to fertilize in the future plantations. The survival rates for not-fertilized plants in moist and very moist sites indicate the feasibility of carrying out Fitzroya plantations with restoration purposes in the Central Depression, using the available afforestation subsidies from the government. Survival rates > 75% at the first year make these plantations eligible for subsidies that range between 600 and 790 euros per hectare (903 USD) for plantations densities of 700 and 1,700 plants per hectare, respectively. In the very moist and moist micro-sites which are less restrictive for vegetation growth, we obtained annual periodic increments in height for the first five years ranging from mean values of 12.8 to 15.8 cm/year. Donoso et al. (2000) determined a mean annual increment of 15.9 cm/year for the Coastal Range after 5 years. The highest increment that has been recorded in the Chaqueihua experimental plantation for plants established in the moist site reach a maximum of 30 cm/year. This increment determines that there are Fitzroya plants that after 7 years have reached 2.0 meters of height (Fig 5). Understanding biodiverity loss: an overview on forest fragmentation in South America 6 138 Figure 5 Left: Mr. Alfredo Núñez planting a Fitzroya seedling (September 1998). Right: Mr. Núñez in front of one of the trees with the highest rates of height growth (30 cm/year, November 2005). Photos taken by A. Lara. Conclusions The initial objective of this research which was to determine the survival and growth rates in Fitzroya experimental plantations in the Central Depresion have been reached. In addition, we have achieved other interesting results, which were not originally planned. This first experience in ecological restoration in Chile has inspired other similar initiatives with threatened long-lived conifers in Chile, including Araucaria araucana and Pilgerodendron uviferum. The Chaqueihua experimental plantation has been visited by hundreds of people, including, professionals, government officials, forest owners, as well as students and researchers from Chile and abroad. Therefore, it has played and important role in the education and dissemination of the concepts, possibilities and challenges for ecological restoration. One limitation of the experiment has been the small number of plants used in some of the treatments. Another limitation has been that the genetic structure and variability of the restoration plantation compared to those of the natural sub-populations of the Central Depression that provided the seed and cuttings has not been studied yet, and therefore remain unknown. From this experiment we propose an integral approach for the long-term restoration of F. cupressoides, combining natural processes and human interventions through plantations and exclosures, based on the understanding of ecological succession. This approach requires also the reduction of the degradation and fragmentation processes, due to grazing, human- set fires and logging, which reverse the successional trends (Fig. 6). Plantations would establish new stands dominated by Fitzroya with similar structure and composition than the remaining fragments, and at the same time increase the viability, size and connectivity of the remnant sub-populations. These remnants are characterized for being young second-growth even-aged stands (mean age ranging from 60 to 120 years), with different degrees of human disturbance due to grazing, fire, and logging. Their structure substantially differs from the old- growth pre-European settlement forests that existed before 1850 (Fraver et al., 1999; Silla et al., 2002; Lara et al., 2003). Therefore, we propose that it is possible to develop a landscape with a larger number and area of better connected Fitzroya sub-populations with a similar composition and structure of the present remnants in the Central Depression within a horizon of 50 to 100 years. Understanding biodiverity loss: an overview on forest fragmentation in South America 7 139 Condition sucesional on ci advanced ec ss Su on ti ra to es R n io at ad gr De Degraded Structural or functional atribute condition 100 Time (years) 1500 Figure 6. Proposed approach for ecological restoration of Fitzroya cupressoides in the Central depression, considering the integration between human intervention and natural ecologic successional processes. The circles represent the various stages considering both succesional and degradation processes. Developed from Stanturf and Madsen (2004). The increased connectivity between fragments includes both functional and structural aspects. The first referred to the flux and dispersal of genetic material and the latter to the spatial configuration of the habitat. The ability of the Fitzroya restoration plantations to maintain the genetic structure and variability of the present remnants has not been yet studied. Future research on the experimental plantation in Chaqueihua is needed to address this very relevant topic, not only for the conservation of Fitzroya, but for the theory and practice of ecological restoration in general. Acknowledgements We acknowledge the support and enthusiasm of Mr. Alfredo Núñez, his wife Mrs. Elba and his son José (recently deceased). This paper is dedicated to them. We are also thankful to Dr. A. Newton (Bournemouth University, U.K.) for his advise and introducting us to the topic of ecologic restoration. We thank Professor Claudio Donoso (U. Austral de Chile) for developing the knowledge on Fitzroya propagation and plantation. We are also thankful to all the colleagues and students from Chile and abroad who have visited the plantation and have made valuable suggestions and comments. Funding was provided by grants from the European Union (SUCRE and BIOCORES (IC18-CT97-0146 and 2002/022/E/DID/F respectively) de la Unión Europea. Additional funding was provided by the Millenium Scientific Initiative, Ministry of Planning, Chile (FORECOS P04-065-F). Understanding biodiverity loss: an overview on forest fragmentation in South America 8 140 References Allnut, T; Newton, A; Lara, A; Premoli, A; Armesto, J; Vergara, R. & Gardner, M. (1999) Genetic variation of Fitzroya cupressoides (alerce), a threatened South American conifer. Molecular Ecology. 8: 975-987 CONAF, CONAMA, Universidad Austral de Chile. (1999) Catastro y Evaluación de Recursos Vegetacionales Nativos de Chile. Informe Regional Décima Región. Santiago de Chile. Di Castri, F. & Hajek, E. (1976) Bioclimatología de Chile. Ediciones Universidad Católica de Chile, Santiago. Donoso, C. (1993a) Bosques templados de Chile y Argentina. Editorial Universitaria S.A. Santiago de Chile. Donoso, C. (1993b) Producción de semillas y hojarasca de las especies del tipo forestal alerce (Fitzroya cupressoides) de la Cordillera de la Costa de Valdivia, Chile. Revista Chilena de Historia Natural 66: 53-64. Donoso, C; Escobar, E; Castro, H; Zúñiga, A. & Grez, R. (2000) Sobrevivencia y crecimiento de alerce (Fitzroya cupressoides Mol. (Johnston)) en plantaciones experimentales en la Cordillera de la Costa de Valdivia. Bosque 21: 13-24. FAO-UNESCO. (1971) Soil Map of the World. South America, IV, Paris, pp. 193. Fraver, S; González, M.E; Silla, F; Lara, A. & Gardner, M. (1999) Composition and structure of remnant Fitzroya cupressoides forests of Southern Chile’s Central Depression. Journal of the Torrey Botanical Society 126: 49-57. Lara, A. & Villalba, R. (1993) A 3620-year temperature record from Fitzroya cupressoides tree-rings in Southern South America. Science. 260: 1104-1106. Lara, A; Aravena, J.C; Wolodarsky-Franke, A; Cortés, M. & Fraver, S. (2003) Fire regimes and forest dynamics in the lake district in south-central Chile. T.T. Veblen., W. Baker, G. Montenegro. & T.W Swetnam. Fire and Climatic Change in the Temperate Ecosystems of the Western Americas, Springer-Verlag, New York, USA, pp. 322-342. Oldfield, S; Lusty, C.& MacKinven, A. (Compilers) (1998) The World List of Threatened Trees. World Conservation Press, Cambridge, UK. Pezoa, LS. (2003) Recopilación y análisis de la variación de las temperaturas (período 1965 2001) y las precipitaciones (período 1931-2001) a partir de la información de estaciones meteorológicas de Chile entre los 33° y 53° de latitud sur. Tesis de grado Escuela de Ingeniería Forestal. Universidad Austral de Chile. Premoli, A; Vergara, R; Souto, C; Lara, A. & Newton, C. (2003) Lowland valleys shelter the ancient conifer Fitzroya cupressoides in the Central Depression of southern Chile. Journal of the Royal Society of New Zealand. 33: 623-631. Silla, F; Fraver, S; Lara; Allnutt, T.R. & Newton, A. (2002) Regeneration and stand dynamics of Fitzroya cupressoides (Cupressaceae) forest of southern Chile´s Central Depression. Forest Ecology and Management 165: 213-224. Stanturf, J.A. & Madsen, P. (2004) (Eds) Restoration of boreal and temperate forests. Boca Raton London, New York, Washington, DC. Tosso, J. (1985) Suelos Volcánicos de Chile. Instituto de Investigaciones Agropecuarias, Santiago, Chile. Veblen, T; Delmastro, R. & Schlatter, J. (1976) The conservation of Fitzroya cupressoides and its environment in Southern Chile. Environmental Conservation 3: 291-301. Veblen, T; Burns, B.R; Kitzberger, T; Lara, A. & Villalba, R. (1995) The ecology of conifers of Southern South America. N.J. Enright & R.S. Hill (eds). Ecology of the Southern Conifers, Melbourne University Press, Melbourne, Australia, pp. 120-155. Wolodarsky-Franke, A. & Lara, A. (2005) The role of ‘‘forensic’’ dendrochronology in the conservation of alerce (Fitzroya cupressoides ((Molina) Johnston)) forests in Chile. Dendrochronologia 22: 235-240. Understanding biodiverity loss: an overview on forest fragmentation in South America 9 141 Conservation approaches in the Atlantic Forest of Argentina: from eco-region to single-species M. J. Pacha1, M. S. Di Bitetti2, G. Placci3, E. Carabelli1, A. Paviolo2, C. D. De Angelo 2, M. Jaramillo1 1 Fundación Vida Silvestre Argentina, Programa Selva Paranaense, Av. Cordoba 464- Puerto Iguazu CP 3370, Misiones, Argentina 2 National Research Council of Argentina and Asociación Civil Centro de Investigaciones del Bosque Atlantico (CeIBA), Yapeyú 23, (3370) Puerto Iguazú, Misiones, Argentina 3 Consultant - Constitución 237 – 5800 – Río Cuarto, Córdoba, argentina Abstract Conservation efforts worldwide have evolved in recent decades from a single- species approach to consider whole areas or ecor-egions. Theoretical approaches need to be put in practice on the ground to discover their benefits and challenges. For this reason we analyse eco-regional, landscape and single- species approaches to conservation in the Upper Parana Atlantic Forest of Argentina, Brazil and Paraguay and explore their interdependency. Eco-regional approaches concentrated on the creation of a Biodiversity Vision to guide conservation actions while landscape approaches were carried out through Forest Landscape Restoration initiative. Finally, a single-species approach was concentrated on the jaguar (Panthera onca) population. Challenges in this region include ambitious targets in a tri-national context, community involvement and stakeholder participation and cultural shifts to appreciate more the importance of the Atlantic Forest. To be effective, conservation efforts in any area need to include a combination of different coarse and fine filter approaches. Introduction Conservation efforts have evolved in the last decades to stop the worldwide loss of species and habitats and the degradation of the remaining fragments of original habitats. But conservation is proving to be more complicated that we once thought with changing policies, theoretical and technological innovations that need to be accommodated. The central questions in this quest have been about what exactly we are proposing to conserve, how we are proposing to do it and where the efforts should be focused (Redford et al 2003). In order to clarify these issues, conservation organisations around the world have tried to define both their targets and the scope of their interventions. The conservation targets represent those entities (species, ecosystems, ecological processes) whose long-term persistence the conservation efforts is attempting to ensure. On the other hand the conservation approach represents the type of actions that are carried out to preserve species, ecosystems or ecological processes. The approach is carried out in a defined spatial scale, the scope, and different conservation approaches have evolved during the last decades reflecting changing values placed on components of nature by different elements of society. The conservation approaches did not replace one another in a temporal sequence, but rather have tended to accumulate in a process of redefinition, absorption and addition. Species have been initially the target of conservation in the western world first because of concerns with the loss of the resource and then as objects worth protecting for the intrinsic value. Big charismatic mammals such as pandas, tigers and whales became international Understanding biodiverity loss: an overview on forest fragmentation in South America 1421 symbols for conservation. There are two types of targets species: first are those that somehow are particular to an area and which its survival is in jeopardy. These types of species, are the so- called threatened, endangered or imperilled species. The second type consist of species that by the very nature of their life history requirements, infer insights into the conservation needs of other species in the region. Examples of the latter are flagships, umbrella and focal species. For a thorough review of these categories see Groves (2003). Several conservation organisations around the world still have a strong focus on species: Conservation International (CI) identified species as the most prominent and readily recognizable form of biodiversity and identified “hotspots”, areas featuring exceptional concentrations of endemic species (Myers et al 2000). The second conservation target considered has been the ecosystem. This term has been widely defined and used in many different ways by ecologists but there has been a lack of precision causing considerable confusion in professional discourse (Groves, 2003). Ecosystems as defined by Whittaker (1975) as “the sum of the community and its environment treated together as a functional system that transfers and circulate energy and matter”. The concern for ecosystems began as a result of loss of the resources they provide and later by their intrinsic value. At the moment, the value of the ecosystems is weighted mainly by the assets that they provide to humans: ecosystem services providing mainly goods, clean water, air and fertile soils. This conservation approach has been practised for centuries by communities around the world, as they have managed ecosystems and natural resources by imposing rules that regulate their use. An example of such conservation management practices (forest management in Japan during the XX century) is nicely described by Diamond (2005) and by indigenous people in Mexico (Toledo et al. 2003). Most recently (in the 1990s), NGOs have expanded conservation efforts to include as their targets several ecosystems, species and also ecological processes that interact at ecological and evolutionary scales. An example of this approach is the one adopted by World Wildlife Fund (WWF), re-defining ecosystems and expanding its scale with the concept of eco-region which is defined as a relatively large unit of land or water that contains a distinct assemblage of natural communities sharing a large majority of species, dynamics, and environmental conditions. A terrestrial eco-region is characterized by a dominant vegetation type, which although not universally present in the region, is widely distributed and gives unifying character to it. This large-scale approach is considered adequate because they (a) protect major driving ecological and evolutionary processes that create and maintain biodiversity; (b) maintain populations of the species that need the largest areas; (c) include biogeographically related communities for representation analysis; and (d) help determine the best places to invest conservation efforts (Dinerstein et al, 2003). For these reasons, WWF has focused its attention on critical eco-regions, the Global 200 (WWF 2000), that constitute a set of eco- regions selected among all terrestrial, marine, and freshwater habitats around the world through a science-based ranking effort. To identify the most outstanding examples, this ranking is based on a comparative analysis of biodiversity data throughout the world, using the eco-regions as the units of analysis. The Global 200 includes representations of all major habitat types in each major biogeographic unit. The objective of this ranking is to prioritize conservation actions throughout the world. In the 1990s, landscapes also became conservation targets as scientists recognized the relevance of large-scale patterns, disturbances and the importance of landscape heterogeneity (Wiens, 2000). This approach also emphasizes the explicit spatial structure of entire, heterogeneous areas or mosaics and their dynamic structure over time (Opdam & Wiens, 2002). Landscapes represent an intermediate scale from species to eco-regions and could be regarded as areas where processes and patterns can be more easily identified and managed. Landscape ecology provided the theoretical basis for this new approach where biophysical and societal causes and consequences of landscape heterogeneity and Understanding biodiverity loss: an overview on forest fragmentation in South America 1432 degradation are explored (Hobbs, 1997). In the landscape approach, human beings are necessary components of the systems of what can be called biocultural landscapes (Antrop, 2000). The connection of conservation with human activities is a recent phenomenon in conservation organizations through the recognition of the impact and the importance of human activities to achieve sustainable development and the rising concern for human rights and their plight for the rural poor. Underlying these positions is the perspective of a holistic approach where humans are themselves a part of nature and that their activities can foster or stop conservation of natural habitats and thus conservation without humans does not make sense (Naveh, 2000). One example of this approach is the Forest Landscape Restoration programme set originally by WWF, which seeks to restore environmental functions in degraded forests, restore benefits that forest bring to people, reduce the vulnerability of forests and biodiversity to threats such as climate change, pests/disease and forest fires and address root causes of forest loss and degradation (Dudley et al., 2005). The programme fosters this by strengthening the relationship between rural development, forestry and other natural resource management and conservation approaches. It is a collaborative and participatory process involving several stakeholders from government agencies, indigenous people, local communities, NGOs, land owners and companies which all have vital roles in the process. Other conservation targets mentioned by Redford et al (2003) include scenery and biodiversity. The former includes places with monumental natural areas for the public enjoyment. This has been in the form of large protected areas seeking to protect spectacular landscapes. This was the main approach considered when the first natural parks were created in the late XIX and early XX century in North America, Australia and Argentina. Biodiversity also became a common conservation target mainly in the last 15 years and has been widely adopted though sometimes not defined and misused. When specified, biodiversity is often defined in different ways, depending on the interest of the group but what is clear is that biodiversity occurs at multiple spatial scales and multiple levels of biological organizations. CI's Hotspots approach (Mittermeier et al., 1998; Myers et al., 2000) is aimed at increasing our ability to preserve the largest amount of the global biodiversity (in this case mainly understood as species richness) and thus, is the one that more clearly identifies biodiversity as its main target. All these theoretical considerations of conservation approaches need to be applied on the ground to assess their effectiveness to achieve real conservation results. The aim of this paper is to consider three conservation approaches with different scope: eco-region, landscape and species in an endangered region, the Atlantic Forests of South America and to assess their effectiveness and the challenges that they present. The Atlantic Forest Eco-region complex The Atlantic Forests extends from Brazil, Paraguay, and Argentina (Figure 1), is one of the WWF Global 200 Ecoregions, being one of the most endangered tropical rainforests in the world, with just 7.4 percent of its original 1,713,535 square kilometers of native forest cover remaining. In spite of its highly fragmented condition (Figure 1) the Atlantic Forest remains one of Earth's most biologically diverse ecosystems, containing about 7% of the world´s species (Quintela 1990 in Cullen et al. 2001; Galindo-Leal and de Gusmão Câmara 2003). In the Atlantic Forest, there is a high proportion of endemic species: 52% of its tree species, 74% of its bromeliad species, 80% of the primate species, and 92% of its amphibians are endemic (Mittermeier et al. 2001). Many of these species are now threatened with extinction. The Atlantic Forests Eco-region Complex has also been identified as a Biodiversity Hotspot, by Conservation International as one of its 25 Hotspots Myers et al. 2000). Understanding biodiverity loss: an overview on forest fragmentation in South America 1443 In this paper we are going to concentrate in the southern most eco-region of this complex: the Upper Paraná Atlantic Forest eco-region (UPAF) where the original forest covers the largest area (471,204 km2) of all the eco-region complex, extending from the western slopes of the Serra do Mar in Brazil to eastern Paraguay and the Province of Misiones of Argentina (Figure 1). With a subtropical climate and nutrient-rich soils, the characteristic vegetation of the UPAF is a semi-deciduous subtropical forest. Variations in the local environment and type of soil, that allow for the occurrence of different plant communities - gallery forests, bamboo forests, palmito (Euterpe edulis) forests, and araucaria forests. Most of the remaining forests have been exploited for timber, and some are second growth forests recovering from deforestation. Forest fragments are thus composed of both primary and secondary forests at different stages of succession. The UPAF plays an important role in the conservation of watersheds, with the Upper Paraná, the Upper Uruguay and the Iguazú rivers crossing the area. The UPAF, is located over a large portion of the largest groundwater reservoir in the world - the Guaraní Aquifer (Holz and Placci 2003). Figure 1. The 15 Ecoregions of the Atlantic Forests Global 200 Ecoregion Complex. The Upper Parana region is highlighted in yellow. The degradation of the UPAF is in closely associated with the growth of the human population and its activities. In this ecoregion there is a population of 35 million people, most of them in Brazil, followed by Paraguay and Argentina. Forest conversion to other land uses, especially agriculture, unsustainable exploitation of the native forest, infrastructures (two of the largest hydroelectric dams in the world are located in this region: Itaipú and Yacyretá) and poaching are the main causes of the degradation of this eco-region . Different approaches to conservation are in place in this eco-region that will be explored in the following sections. Understanding biodiverity loss: an overview on forest fragmentation in South America 1454 Conservation at different levels Eco-regional conservation: building a common Vision In recent years the conservation community has been promoting the design and implementation of biodiversity conservation actions at larger scales. WWF has embraced this approach, focusing conservation planning and action on eco-region. Since most ecological and evolutionary processes that sustain biodiversity occur at these larger scales, WWF has determined that eco-regions are the best units to design and implement biodiversity conservation actions. One of the key elements needed to implement WWF's approach on eco-region conservation is a Biodiversity Vision. A Biodiversity Vision is a planning tool, usually in the form of a document, aimed at guiding biodiversity conservation activities in the eco-region. A Biodiversity Vision sets a number of biodiversity conservation goals based on widely- accepted principles of conservation biology, and identifies critical areas to be either conserved, managed, or restored in order to meet those goals (Di Bitetti et al., 2003). These areas are identified through a science-based process that relies on the best available biodiversity data and socioeconomic information. Through this process, WWF developed a Biodiversity Conservation Landscape that is represented in a map illustrating how the eco- region would look in 50-100 years if we are successful in conserving biodiversity. This Biodiversity Conservation Landscape is a central piece of the Biodiversity Vision, and its representation in a map helps to focus conservation activities on those areas and to set specific actions hat would render the best results for biodiversity conservation. Underlying the Biodiversity Vision is a series of complex analyses aimed at designing a Biodiversity Conservation Landscape that if implemented will accomplish the conservation goals. From 2000 to 2003, WWF and Fundacion Vida Silvestre Argentina (FVSA) have led a tri-national participatory process involving more than 30 local organizations representing multiple sectors and disciplines. A more detailed description of the whole process can be found in Di Bitetti, Placci and Dietz (2003). The starting point to create a tri-national Biodiversity Vision is to first identify the conservation goals that need to be accomplished in the following 50-100 years. For the UPAF these were: 1. The conservation of blocks of natural forest large enough to be resilient to short-term and long-term environmental changes. 2. The maintenance of viable populations of all native species in their natural patterns of abundance and distribution, and with the genetic diversity necessary to meet environmental challenges. 3. The maintenance of healthy ecological processes and selective factors such as disturbance regimes, hydrological processes, nutrient cycles, and biotic interactions, including predation. 4. The representation of all native biological communities and seral stages across their natural range of variation within a Biodiversity Conservation Landscape. For the analyses, we used various overlays of maps representing the distribution of the different biological and socio-economic variables. Three separate but interdependent analyses were critical to arrive at the final Biodiversity Conservation Landscape: a landscape units analysis, a fragmentation analysis and a threat and opportunities analysis. These analysis were then combined to produce a biodiversity potential map that represented a broad-scale cost-benefit analysis for conservation. These steps ended up in the final map representing the Biodiversity Conservation Landscape (Figure 2). These analysis were carried out using a Geographic Information System (Arc View 3.2) that provided a critical Understanding biodiverity loss: an overview on forest fragmentation in South America 1465 tool for conducting the analyses and visually describing different layers of information in various maps. Figure 2: Process of Development of the Biodiversity Conservation Landscape. Modified from Di Bitetti et al 2003. The first step involved the identification of individual landscape units. A landscape unit was used identified areas that contain a set of species, communities, or ecological processes that differs from other such landscape units. Each landscape unit usually has a characteristic climate, soil type, and set of species. Given the lack of complete or sufficient biological information available to define and map all ecological communities, climatic, altitude, and topographic information were used as proxies for developing a biological model. Using these three layers of information, 18 separate landscape units were identified. The second step involved the identification of native forest fragments with the highest potential for achieving conservation goals. For this fragmentation analysis, a map of forest fragments obtained from satellite images was used. Forest fragments were ranked according to a Fragment Importance Index developed to indicate the relative contribution of forest fragments to biodiversity conservation. The index was based on four variables: fragment size, fragment size after excluding a buffer zone of 500 m (an indirect measure of edge effects), distance to nearest fragment, and altitudinal range within the forest fragment. The third step was a threats and opportunities analysis, where the objective was to map areas that represent critical threats and important opportunities for biodiversity conservation. Land use information provided a critical basis for assessing conservation opportunities and threats. The threat variables used in the analysis included: distance to cities, agriculture, cattle ranching, and rural population density. Threat variables were weighted differently according to the degree of threat each poses to biodiversity conservation. Opportunity variables that were used included: the distance from a strictly protected area, the proximity to a river (assuming that rivers in this eco-region constitute potential biological corridors), and zones of planned conservation. Variables were weighted according to their relative impact on biodiversity conservation. The current status of forest cover and the representation of the different landscape units within the protected area system was analyzed using the landscape units map in combination with the forest fragments map and the protected areas map. This gave an idea of how well represented each landscape unit was in the actual landscape, and guided Understanding biodiverity loss: an overview on forest fragmentation in South America 1476 decisions on how to improve representation of those underrepresented landscape units in the final Biodiversity Conservation Landscape (Figure 3). Combining the fragment importance index map with the threats and opportunities map, a biodiversity conservation potential map was constructed that illustrates where the areas with the highest biodiversity conservation potential are located in the eco-region. Using this biodiversity conservation potential map as the basic layer of information, a Biodiversity Conservation Landscape was defined (Figure 3). The resulting Biodiversity Conservation Landscape is composed of three main types of areas: core, strategic and sustainable use areas. This biodiversity conservation landscape was designed with the jaguar as the focal species, supported by the idea that due to its umbrella effect, most species will benefit from its conservation and core areas (> 10,000 ha) were defined as areas that could maintain at least a minimum breeding jaguar unit (at least one adult male plus one or more adult females). With this in mind the Vision focused on the objective of connecting the Core Areas through corridors and establishing buffer zones around Core Areas, priority areas, and corridors. Figure 3: Biodiversity Conservation Landscape of the Upper Parana Atlantic Forest eco-region. Defining the Biodiversity Conservation Landscape was just the beginning. The implementation of this Biodiversity Conservation Landscape requires a series of actions at different time frames and spatial scales. The Vision is a useful tool to identify actions, milestones and time frames needed to achieve the Biodiversity Conservation Landscape. Six actions were defined in this exercise that includes the effective management of protected areas, the creation of new ones, the implementation of sustainable use areas, restoration of degraded forests, public support and constant monitoring and adaptive management of the Vision. The strength of this planning exercise is that it helps focusing resources on those areas or targets that are critically important in order to achieve conservation results. It also helps different conservation organizations to reach a consensus among them on important and common goals not as a result of negotiations (or a voting process) but as a result of an science-based objective process. Understanding biodiverity loss: an overview on forest fragmentation in South America 1487 The main challenge for the implementation of the “Vision” is that actions are too ambitious i.e. 1,226,175 ha of new protected areas and 2,684,722 ha of sustainable use areas will need to be created. Furthermore, the area is share by countries with different language, different socio-political history and different economic priorities. Also there is restricted funding for these purposes and no financial plan was established. Thus coordination and implementation of this ambitious plan is extremely challenging. Landscape restoration and community-based conservation approach The focus of the Forest Landscape Restoration (FLR) approach in the UPAF of Argentina is the Andresito colony that is located in northeastern Argentina and it was identified in the Biodiversity Vision as an important corridor area (Figure 4) . This area is critical for maintaining the connectivity of forest cover between two World Heritage Sites, the Iguaçu National Park in Brazil and the Iguazu National Park in Argentina, and other two Provincial Parks in Argentina (Urugua-í and Foerster). Andresito colony is surrounded by this strictly protected forest; a continuous patch of native forest of more than 360,000 ha that constitutes the largest block of strictly protected areas in the UPAF. However, expansion of agricultural, ranching and forestry lands has led to the fragmentation of this connection and the loss of forest goods and services for the local people. Figure 4: Location of Andresito colony in the Upper Parana Atlantic Forest eco-region. The challenge in Andresito is to stop deforestation and forest degradation due to the logging of native tree species for timber, the growing of yerba mate (Ilex paraguayensis, a local tea that requires not only land for its crop but lumber for the drying process), cattle ranching and tobacco plantations; and to increase connectivity among protected areas and to establish buffer zones around the protected areas while improving local people’s living standards. Understanding biodiverity loss: an overview on forest fragmentation in South America 1498 To achieve these objectives, since 2002 FVSA works with local farmers towards a more natural and sustainable landscape while providing them with an economic activity that increases or preserves the forest cover. Activities that have been carried out include: 1) a fine scale land-use analysis over the last 20 years in the municipality of Andresito (forest cover, property size, potential corridors, etc.) (Guerrero Borges et al., 2005); 2) an on-going participatory land-use planning process with key stakeholders; (Jaramillo et al., 2005) 3) restoration pilot projects to evaluate economic and environmental efficiency of different forest restoration techniques; 4) studies on natural regeneration of abandoned lands; and 5) identification and development of new economic alternatives in co-operation with the landowners. This last objective is related with the sustainable use of palmito (Euterpe edulis) a native understory palm tree that due to forest conversion, has become very scarce and has a high price in the food market. The aim is to sustainably harvest palmito with a management plan to increase the forest value and productivity of this non-timber forest product as well as the landowners’ income. For this end the Agro-ecological Cooperative of Peninsula Andresito was formed with more than 26 farmers and a canning factory was built to encourage the sustainable harvesting of palmito hearts and other non-timber forest products. In all stages of this approach is necessary that all community sectors – local and provincial government, farmers, teachers, aboriginal community members, NGOs, tourism and forestry entrepreneurs, among others – participate and commit to the process. In trying to adopt the landscape approach to conservation, challenges are related with community and social issues that can foster or stop the whole process. Andresito is a new community that was formed in the 1970s as part of a government plan to stop illegal immigration and invasion of lands by Brazilians and Paraguayans. In this colonization plan farmers were encouraged to “improve” the land by logging 75% of the native forest area provided. People from different parts of the Misiones province and also Brazilians were benefited, this created a complex mix of social and cultural backgrounds that precluded the integration of the community. Also land distribution was not even and most landholders had small to medium sized farms with properties ranging from 5 to 200 ha. In these small farms, there is a high level of poverty and most of the farmers have not finished primary education. There are several challenges when applying this approach to obtain conservation results, as they are sometimes lost during the complicated social process. The first challenge is to assure an active and participatory process within the community. In general, the participatory process takes significant time to become a solid and sustainable practice. The time frame is usually considered a weakness of the approach; however it provides longer term results in contrast to projects with high inputs and short project duration. We believe that is preferred a longer low-budget intervention than a large budget to be executed in a short period of time. Creating confidence with stakeholders takes time. The second challenge is how to involve the private and government sectors so they can appreciate the advantages of community- driven approaches to sustainable development and diversification of productive practices. Thirdly, it is difficult to engage land owners to conservation practices when there have been government incentives in the past where forests were regarded as waste-land. Land owner need compensation from society to preserve forests as they have a direct impact in water quality and quantity, clean air, climate change and scenic beauty. Payments for ecosystem services can be a solution to conserve forests in the UPAF. The final challenge is that actions in the landscape approach are taking place at a minor scale; therefore, to scale up the process is very difficult. Understanding biodiverity loss: an overview on forest fragmentation in South America 1509 Species level: preserving jaguars Despite the low representation of the UPAF in Brazil (where 2% of the original forest cover remains) and Paraguay (12%), in the province of Misiones, Argentina there are good possibilities for conservation and restoration of the native forest, as it has the largest remaining continuous block with almost 50% of its original cover. Still it provides habitat for the original set of large vertebrates of the whole Atlantic Forest, including large predators such as harpy eagles (Harpia harpija), crested eagles (Spizaetus tyranus and Spizaetus ornatus), pumas (Puma concolor), ocelots (Lepardus pardalis), and jaguars (Panthera onca), and large herbivores, such as tapirs (Tapirus terrestris), three species of brocket deer (Mazama sp.), and two species of peccaries (Tayassu sp.). The area has also been identified as a “Jaguar Conservation Unit”(Sanderson et al., 2002). The jaguar, was used as a focal species to design a Biodiversity Conservation Landscape because of its involvement in a key ecological process: predation and its top-down regulation effect and its umbrella effect (see Miller and Rabinowitz 2002). The core areas of the Biodiversity Conservation Landscape were identified based on the territorial requirements of a adult males of this species. One of the activities of FVSA-WWF´s Action Plan for the UPAF is to monitor the jaguar population. This is important not only as a way to assess the conservation status of the jaguar (and as a result of its umbrella effect, an indirect assessment of the status of biodiversity) but also to validate the Biodiversity Conservation Landscape and/or to suggest changes to its design and implementation. Since 2002, with the support of WWF and other NGOs, Fundación Vida Silvestre Argentina in partnership with scientists are implementing The Jaguar Project. The aim of this project is to know the conservation status of the jaguar population of the UPAF. This is achieved through a two-pronged approach. First, we are building an updated and very detailed map of the current presence of jaguars in all the forest remnants of the UPAF, a process that involves a network of volunteers (>150 people) trained to collect information on jaguar presence (feces and plaster molds of feline tracks). Second, we are conducting camera trap surveys in different areas within the green Corridor of Misiones to estimate jaguar, puma and ocelot population densities (see Di Bitetti et al., 2006a, 2006b) Recent results show that the jaguar population has abruptly declined in the last 10 years and that jaguar densities for Misiones are extremely low. In fact, the population values represent the lower density estimated for jaguars in all the Neotropical sites already studied (Di Bitetti et al., 2006, Paviolo et al., 2006a). Extrapolating these numbers to areas where jaguars are still present (94,000 km2) in Misiones and nearby areas of Brazil, it is estimated a total population of less than 100 adult individuals (most probably around 50 adult animals). Alarmingly none of the protected areas of the region can by itself ensure jaguar conservation due to their low population density. The possibility for this species survival is closely related to the safeguarded connection between the protected areas and the existing remnants in the three countries so as to ensure free movement of the species (Paviolo et al., 2006a). This information is being used to develop a jaguar population viability analyses (PVA) which is based on a population model agreed upon on a participatory processes and workshops with experts from Brazil, Paraguay and Argentina Basic population information is essential to understand population status of jaguars, but to achieve effective conservation of the species it is necessary to focus on the causes of the decrease of the population. Not surprisingly, the abundance of prey species is clearly related to hunting pressure and the degree of implementation of the protected areas. Jaguars are being killed to unsustainable levels, not only by farmers and cattle ranchers as a result of retaliation (for being considered dangerous or because they sometimes prey on livestock and other domestic animals) but mostly by poachers that shot them when hunting for game Understanding biodiverity loss: an overview on forest fragmentation in South America 15110 animals (Paviolo et al., in prep). Ocelot and pumas show similar population trends as those shown by jaguars, being more abundant in the best protected areas and showing lower densities in areas with high poaching levels and/or where the forest is being converted to other uses or unsustainably harvested. As it was shown, community play an important part in the conservation of jaguars. There is a need to change cultural perceptions and to educate landowners of the alarming situation of the jaguar population and the real threat they pose to human activities (and the way to avoid conflicts). For this, a communication campaign was designed with different stakeholders aimed at informing the public and raising awareness of the level of threat and extinction risk of jaguars. FVSA have created partnerships with other organisations like the National Parks Administration and the Ecology Ministry of the Misiones province. Together a brochure was produced (“Yaguareté, un símbolo de la selva”), that will be used and distributed in provincial schools, and other public and private institutions. Also a school contest is underway to engage children and teenagers of more than 1000 schools in the province of Misiones in active conservation actions. The challenge in the single-species conservation approach for the jaguar are twofold: first to change cultural practices that are jeopardizing the persistence of jaguars in the UPAF and second, at a more large-scale, change land use changes and practices that have fostered the unsustainable use of the native forest and the intense fragmentation of the remaining habitats. Discussion In the race to conserve nature, several approaches have emerged worldwide in different organisations. We explored different approaches that are being currently used to preserve the UPAF: large-scale eco-regional, landscape and single-species. These were targeted at different levels and spatial scales. The eco-regional approach and the Biodiversity Vision was essential to identify important areas for biodiversity conservation and to help conservation planning in a complex area like the Atlantic Forest that is shared by three countries with different cultural and economic backgrounds. Systematic conservation planning is the following step where goals and objectives can be achieved in a certain temporal scale. However, the full implementation of the Biodiversity Conservation Landscape described in the Vision can be too ambitious and impossible to achieve in the current political and economic scenario. High political and economic uncertainty is a common phenomena and lack of funding can halt the whole process as well as corruption, poor information availability, weak or pseudo participatory process of problem analysis and decision making and weak governance. The almost lack of law enforcement capacity in developing countries, like the ones involved in the UPAF, render almost impossible to create environmental policies based on a new regulatory framework to produce large-scale and long-term changes in the eco-region. Environmental education, communication campaigns, advocacy work are effective tools that could produce cultural changes and create new socio-cultural scenarios in which conservation take place. The Vision should be understood as a prioritization devise where different levels of urgency and importance are given to certain geographic areas, species or activities. One organization cannot effectively act on the whole area, but the Vision serves as a guide of where to start putting resources and efforts to achieve conservation results. The success on implementing the vision will depend on the development of “on the ground” projects executed at spatial and temporal scales in accordance with the capacity of each organization. The coordination of activities among different NOGs and governmental agencies is also essential to start Understanding biodiverity loss: an overview on forest fragmentation in South America 15211 implementing the Vision. As a live document, the Vision will be revised and actualised as the reality changes. The landscape approach put into practice in Andresito includes a combination of various disciplines: biology, sociology and economy in trying to find solutions with the rural poor which is over exploiting the natural resources because of a lack of alternatives and the dominance of international markets in their production. Here, the challenge is more focused on aspects connected with skills by the conservation organisations to increase community participation in decisions affecting natural resources management. The practice of conservation has traditionally and mainly being pursued by naturalists, biologists and the like. This community based-approach requires a much higher level of interdisciplinary, and it should particularly be enriched by people with a background in the social sciences. Improving the well being of the local people is an important and necessary goal of any community-based conservation project, but one of the main challenges in a community-based conservation project, like the FLR project in Andresito, is not to loose the conservation goals that gave origin to the project. Ideally, every single landscape of this planet should be managed under a landscape planning scenario that integrates conservation with people livelihood but some landscapes are more important for conservation than others. The biodiversity vision is a key step in the process of selecting the priority landscapes in which to implement conservation project at the landscape level. Finally, the species-level approach focusing on the jaguar population also shows the intricate relationships between culture, conservation and policies. Not only are we talking about conservation of a species, but also about changes in people perceptions, practices and political pressure to achieve the integration of economic and environmental issues in the political agenda. This is possible because the jaguar is not only an umbrella species that helps conservation planning at large scales, but also because it is a flagship and charismatic species, which creates opportunities for people to become involved in conservation (Miller and Rabinowitz, 2002) These different approaches used in UPAF can be compared with the coarse-filters/ fine- filters approaches to conservation; focus on targets that occur at coarser spatial scales, such as eco-regions, will conserve many targets that occur at finer spatial scales (Noss, 1996). The Biodiversity Vision of UPAF can be considered a coarse-filter approach. However, some targets such as rare and endangered species will pass through this coarse filter and will need to be conserved through individual, fine-filter approaches. It is important when surrogate targets, like jaguar populations, are used to understand the underlying targets of conservation actions. From all these approaches it is clear the conservation project scope expansion from species to entire eco-regions with the appreciation that the single-species approach is insufficient to achieve conservation at higher orders of organization. In this sense, is only the larger spatial scales that will permit conservation actions to be directed to areas where ecological and evolutionary processes are likely to operate (Redford et al 2003). Many approaches have multiple targets and biodiversity occurs at multiple spatial scales and multiple levels of biological organization, different approaches are needed to conserve natural resources. Each of these approaches is not exclusive but complementary. For example, the Biodiversity Vision of UPAF used the jaguar as a focal species to identify the core areas for conservation. The jaguar project is mainly focusing on the establishment and implementation of conservation areas in the UPAF. Without a well-implemented system of strictly protected areas there is no future for the jaguar and the regional biodiversity at large. In the process of creating of the Biodiversity Conservation Landscape of the UPAF, we could identify critical corridors and areas for sustainable use. Social indicators and threats and opportunities Understanding biodiverity loss: an overview on forest fragmentation in South America 15312 analyses were critical to identify these areas and the FLR project in Andresito is being implemented in one of such important areas. 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