Salinas Chávez, Eduardo; Middleton, John. 1998. La ecología del paisaje como base para el desarrollo sustentable en América Latina / Landscape ecology as a tool for sustainable development in Latin America. http://www.brocku.ca/epi/lebk/lebk.html
LANDSCAPE ECOLOGY, DEFORESTATION, AND FOREST FRAGMENTATION: THE CASE OF THE RUIL FOREST IN CHILE.
Audrey A. Grez1, Ramiro O. Bustamante2 , Javier A. Simonetti2 & Lenore Fahrig3
1 Facultad de Ciencias Veterinarias y Pecuarias, Universidad de Chile, Casilla 2, Correo 15, La Granja, Santiago, Chile. e-mail: email@example.com
2 Facultad de Ciencias, Universidad de Chile, Casilla 653, Santiago, Chile. e-mail: firstname.lastname@example.org, email@example.com
3 Ottawa-Carleton Institute of Biology, Carleton University, Ottawa, Canada K1S 5B6. e-mail: firstname.lastname@example.org
SUMMARY: Deforestation and fragmentation of native forests is dangerously increasing around the world. These processes change landscape structure, affecting the persistence of populations. Here, we analyzed the deforestation and fragmentation of the ruil forest, a temperate and endemic formation restricted to 100 km of the coastal range of Central Chile. Between 1981 and 1991, more than 50 % of the forest has been lost, mainly due to the expansion of plantations of Monterey pine. Currently, the 352 ha of the ruil forest conforms an "archipielago" with many small, regular fragments and few large, irregular ones, relatively isolated, and surrounded by a matrix of pine plantations. Only 45 ha are under formal protection. This situation is critical because the direct and indirect effects of deforestation and fragmentation imply, in the short term, the loss of species and a whole and unique ecosystem. Therefore, today forest management is unsustainable. In Chile, landscape ecology should be further developed and applied for a more sustainable regional management which conciliates the interests of both the forest industry and conservationists.
RESUMEN: La deforestación y fragmentación de los bosques nativos es un fenómeno cada dia más frecuente en el mundo. Estos procesos generan cambios en la estructura del paisaje, lo que afecta la persistencia de las poblaciones. Aquí analizamos la deforestación y fragmentación del bosque de ruil, un bosque templado y endémico, restringido a 100 km de la cordillera de la costa de Chile Central. Entre 1981 y 1991 más del 50 % del bosque ha sido deforestado, principalmente por la expansión de plantaciones de pino. Hoy el bosque de ruil está reducido a 352 ha y constituye un archipielago con muchos fragmentos pequeños y regulares y unos pocos grandes e irregulares, relativamente aislados entre si. Sólo 45 ha están protegidas para conservación. Esta situación es crítica puesto que los efectos directos e indirectos de la deforestación y fragmentación del bosque implican la pérdida a corto plazo de especies y de un ecosistema único en el mundo. El manejo actual del bosque no es sustentable. La ecología del paisaje debería ser mayormente desarrollada en el pais para que, a traves de su aplicación, ayude a resolver estrategias de manejo a nivel regional, que concilien intereses de la industria forestal y de los conservacionistas.
There is growing concern over the effects of large-scale human activities on the functioning of natural systems. However, our understanding of how natural systems function has been based on ecological research which has traditionally been conducted on very small spatial scales, for practical reasons (Kareiva & Andersen 1988). The need for larger-scale understanding of ecological systems has led to the relatively new field of landscape ecology, the study of how landscape structure affects the abundance and distribution of organisms.
Landscape structure is the spatial pattern of habitats in a landscape, where a landscape is a mosaic of habitat patches. Note that the scale and description of a landscape will depend on the scale of movement and habitat associations (respectively) of the organism under consideration. There are two components to landscape structure, "composition" or the amounts of the various types of habitat in the landscape and "configuration" or the spatial arrangements of those habitats (Turner 1989, Dunning et al. 1992).
Destruction of forest around the world is a major factor in landscape structural change. Both components of landscape structure are afected by deforestation. Composition of the landscape is changed as mature forest is removed and replaced with native or planted young forest, crop land or grazing land. In addition, configuration is changed as the remaining forest becomes fragmented into a larger number of smaller forest patches.
At the scale of the individual forest patch, forest loss and fragmentation can have a broad range of effects on population survival, ecological interactions and biodiversity (Fahrig & Grez 1996). As patches become smaller, populations tend to be more vulnerable to extinction because of demographic, environmental or genetic risks (Gilpin 1987, Goodman 1987). When patches become isolated, with no connections between them, migration of organisms may be precluded (Kareiva 1987). Small patches also have a higher edge/interior ratio. For forest interior species, this represents a loss of habitat greater than predicted by the reduction in patch size alone (Wilcove et al. 1986, Williams-Linera 1990). The magnitude of such effects depends on the landscape-scale pattern of forest loss which will determine the number of the remaining patches, their sizes and shapes, the distances among them, and the nature of the surrounding "matrix" habitat. To date there are few empirical studies that connect pattern of forest loss and fragmentation with ecological processes at the landscape scale (Groom & Schumaker 1993).
The main question for ecologically sustainable forest management is then: how much forest must be preserved to ensure persistence of the various species that inhabit it, and the maintenance of its ecological processes? A secondary one is: can conservation of these species and processes be enhanced by restoring forest habitats and/or by altering the spatial relationships among the remaining forest patches, e.g., by reducing forest fragmentation?
To answer these questions, one must analyze the separate effects of changes in forest amount and configuration at the landscape scale (McGarigal & McComb 1995, Fahrig 1997). However, due to logistical problems of studying multiple landscapes, most researchers conduct patch-scale studies and extrapolate the results to the landscape scale (Dodd 1990, Robinson et al. 1992, van Apeldoorn et al. 1992, Celada et al. 1994, Hunter et al. 1995). When making such extrapolations it is important not to confound implications of forest loss and forest fragmentation. For example, one can make landscape-scale predictions about effects of forest loss from patch-scale effects of isolation, since patch isolation is typically a function of forest amount in a landscape (Fig. 1). Landscape-scale predictions based on patch size effects are not as clear since patch size can be reduced by habitat loss alone or by a combination of habitat loss and fragmentation (Fig. 1).
In this chapter we describe the first stages of a study documenting the patch-scale ecological effects of forest patch size and isolation in the ruil forest in Chile. These effects will be extrapolated to make predictions about landscape-scale effects of forest loss and fragmentation in this area, and to discuss possible ways for a more sustainable management.
Chile is located between 17° and 55°S, encompassing a wide variety of biomes, ranging from the Atacama desert in the northern portion of the country to cool temperate forests and cool steppes in the southern extreme. Furthermore, there are significant west-east environmental gradients shaped by the presence of the Andean Cordillera, rendering a high diversity of environments and vegetation types (e.g., Gajardo 1994).
Chile has a high biodiversity, not in terms of species numbers (app. 30,000), which clearly is lower than in tropical regions, but in the high level of endemism. Over 50% of the vascular flora, 34% of hymenopterans, and 77% of the amphibian species are endemic (Simonetti et al. 1995). A significant fraction of this richness is associated with the temperate forests (Armesto et al. 1996).
Forest lands cover 17,9 million ha in Chile, of which 9.6 million are native forests, 6.7 milllion ha are bare lands or sparsely vegetated, and 1.6 million ha are forest plantations. Over 83% of plantations are the exotic Monterrey pine (Pinus radiata) (PAF-Chile 1993). Native forests and plantations play a significant role in the Chilean economy.
The Chilean economy is an open one, with exports of products and services as a mayor commodity, including forest products. From 1980 to 1990, the forest industry grew at an annual rate of 6%. By 1992, exported forest products amounted to US$ 1.1 billion, accounting for 10% of the total Chilean exports (Núñez 1992, Infor 1994). Exports by the year 2010 are expected to reach US$ 3 billion. Such growth may not be sustainable. The development of the forest industry is threatening native forests due to the incresing tendency to replace them by plantations of exotic species, particularly in south-central Chile, the region where the highest diversity and endemism is attained in the temperate forests (Donoso & Lara 1996, Lara et al. 1996). Forest exploitation has resulted in the alteration, degradation or substitution of 120,000 ha of native forest annually (Lara et al. 1996, Lara & Veblen 1993). Most destruction of native forests is for fuelwood production, industrial production, the export of roundwood and chips, conversion to forest plantations, largely Monterey pine, and transformation into pasturelands. Such habitat transformation is threatening to reduce both biome and species richness. In fact, 52% of Chilean threatened species of vertebrates and vascular plants are from temperate regions, including endemic species such as Gomortega keule, the single species representative of the Gomortegaceae family. Forest types, such as the ruil forest, is structured by another endemic species, Nothofagus alessandrii, the most primitive of Nothofagus species (Simonetti & Armesto 1991, San Martín & Donoso 1996).
The replacement of native forests has been more intensive in Central Chile, the area that accounts for 85% of forests plantations (PAF-Chile 1993). This substitution has been particularly intensive in areas such as the coastal range of the VII Administrative Region, where 31% of native forest was replaced by Monterrey pine in just nine years, from 1978 to 1987 (Lara et al. 1996). This is precisely the region where the scarce remnants of the ruil forests survive, which are increasingly encroached upon by enlarging plantations of exotic trees (San Martin & Donoso 1996).
The contribution of forest products to the National Gross Product has ignored the depreciation of native forests. The degradation of the native forests represents a severe loss of natural capital including potential loss of biodiversity and ecosystem processes that may render unsustainable the current type of forest industry (Núñez 1992, Christensen et al. 1996, Lara et al. 1996).
LANDSCAPE ECOLOGY IN CHILE
Landscape ecology is a recently and scarcely developed discipline in Chile, pioneered by E.R. Fuentes in the 80s. Fuentes and collaborators adopted this approach for the understanding and management of the landscape structure of Central Chile (Fuentes & Hajek 1979, Fuentes & Prenafeta 1988, Fuentes et al. 1984a, b, 1989, 1990, Fuentes 1990, 1994). Largely, they employed the proccess-to-pattern approach to infer how ecological processes associated with human activities shaped the the current landscape (Fuentes & Hajek 1979, Fuentes 1990).
Through a series of experiments, Fuentes et al. (1984b, 1989, 1990) attemped to explain how seed dispersal, seedling herbivory, shrub browsing by goats, livestock and introduced rabbits along with herb competition shaped landscape composition and configuration. In such a way, they explained how the original Prosopis woodland, formerly dominant in the Intermediate Depression to the north of Santiago, turned into a Acacia caven savanna. The conversion was largely due to the beneficial effect of seed dispersal by cattle on Acacia but not on Prosopis (Fuentes et al. 1989). Similarly, they explained the lack of recruitment of native sclerophyllous trees due to dessication, competition with annual herbs and hebivory by introduced animals of seedlings, which changed the vegetation from a complex mixture of sclerophyllous trees to an almost monospecific landscape dominated by the winter decidous tree A. caven (Fuentes et al. 1990). Fuentes (1990) synthesized the main types of human impacts responsible for the landscape changes, concluding that woodcutting, overgrazing and land clearing have changed the original dense woodland vegetation to a more shrubby and scattered one, triggering severe erosion problems and reducing the potential recovery of the vegetation
Paleoecological studies carried out in contexts other than landscape ecology, to develop understanding of vegetation dynamics and human impacts upon the biota, have yielded significant understanding of current landscape structure (e.g., Simonetti & Cornejo 1990). In central Chile, reduction in vegetation cover has been a pervasive effect of human settlement over the last 10,000 years. Triggered by the land-clearing associated with the adoption and expansion of agriculture and wood-cutting for fuel and construction, deforestation and forest alteration occurred in a patchy fashion accross the region long before Spanish arrival. This heterogeneous use of the land rendered a mosaic of vegetation patterns which also affected the structure and long term dynamics of the fauna. Therefore, current landscape structure has been partly shaped by the long term human interaction with natural resources (Simonetti & Cornejo 1990)
Although some recient definitions of landscape ecology do not limit the organism or spatial scale involved (Wiens 1992, Pickett & Cadenasso 1995), most studies conducted so far in Chile that deal explicity with landscape ecology are based on human scales and human-induced problems. That is, in fact, the most typical approach of landscape ecology (e.g., Forman 1995). An exception in Chile is the analysis of the effects of habitat heterogeneity on the dynamics of herbivorous and predatory insects, in which the studied landscape was scaled to match the scale of insect movements (less than a hectare) (Grez 1997).
Other studies have dealt with the human perception of landscapes (Filp et al. 1983, Fuentes et al. 1984a, Muñoz-Pedreros et al. 1993). Although these studies are not within the definition of landscape ecology, they may provide useful information for landscape management. Landscape units are differentially valued by Chileans. In southern Chile, native forests are valued higher than any other landscape component. Native forests were valued almost three times higher than sparse shrublands. Interestingly, the exotic Monterey pine plantations were valued 1.5 times higher than sparse native vegetation, suggesting that people may be reacting more toward vegetation cover or phisiognomy rather than to species origin (Muñoz-Pedreros et al. 1993). Similarly, in central Chile, people prefer landscapes with high woody vegetation, although no distinction is made between native and exotic species (Fuentes et al. 1984a).
Studies of the reciprocal effects of spatial pattern on ecological processes (Turner 1989, Kareiva & Wennergren 1995) are scarce in Chile. That is, the emphasis has been on the causes of landscape modifications but not on the ecological consequences of these modifications. One exception is Willson et al. (1994), who reported how smaller patches of southern temperate forest support a reduced biodiversity and abundance of native birds. The precise processes accounting for these biodiversity and abundance patterns such as immigration, emigration or survival rates, remain to be determined. The pattern-to-process approach should be further developed since one of the aims of landscape ecology is to elucidate if the present (already modified) landscapes provide adequate conditions for maintaining ecological processes and biodiversity, and if not, how to improve the situation through habitat restoration. This is the approach we address in the next section.
CASE STUDY: THE RUIL FOREST:
The ruil forest (sensu San Martín & Donoso 1996) is a temperate forest, currently distributed exclusively along 100 km of the coastal range of central Chile (35° - 36° S, Fig. 2)., restricted to southwest facing slopes from 160 to 440 masl. The dominant tree species are the caducifolious broad-leaved trees Nothofagus alessandrii (ruil), and Nothofagus glauca, which coexist with perennial trees such as Cryptocarya alba, Aextoxicon punctatum, Gevuina avellana, Peumus boldus and Lithraea caustica (San Martín et al. 1984). A conspicuous feature of this forest is the abundance of vines including Lapageria rosea, Boquila trifoliata, Lardizabala biternata and Proustia pyrifolia (San Martín & Donoso 1996).
Since the end of XIX century, the ruil forest has been intensively deforested and fragmented (Donoso
& Landaeta 1983) due to the expansion of timber production and exports based mainly on plantations of Monterey pine (Armesto et al. 1992). Currently, the landscape is a mosaic with scattered fragments of the ruil forest, surrounded by a matrix almost exclusively composed by plantations at different stages of development of Monterey pine.
Based on Bustamante & Castor (1997), here we describe the deforestation rate of the ruil forest for the period 1981-1991; the number, areas and shapes of remnant fragments, and the distances among them. These fragment attributes were estimated from aerial photographs 1:10000 taken in 1991 along the whole range of the forests present distribution. They were digitalized using the GIS program ARC/INFO.
In 1981, the estimated area of the ruil forest was 824.8 ha (Donoso & Landaeta 1983). In 1991 the forest area was reduced to 352.2 ha. Thus, in a period of ten years a total of 472.2 ha of forest has been lost at a rate of 8.14% per year, based on the algorithm used by Dirzo & García (1992). To our knowledge this is one of the highest rates of deforestation reported in the literature (JA Simonetti & RO Bustamante, unpublished).
By 1991, the ruil forest was distributed in 185 fragments, averaging 1.9 ha and with a size distribution clearly skewed. Most of the fragments were just 1 - 2 ha and less than 5% of fragment were larger than 12 ha (Fig. 3). Fragments were relatively isolated with no corridors between them. The average nearest-neighbour distance among fragments was 55 m (range: 10 - 620 m). The distance distribution was also skewed, with more than 50% of fragments near 20 m apart, and less than 10 % over 100 m apart (Fig. 4). The smaller fragments were the most isolated ones (Bustamante & Castor 1997).
The average shape index of fragments was 1.83 (range: 1.07 - 7.32) with the shape distribution skewed to the lowest Si values; that is, most of the fragments tended to be regular (Fig. 5). However, the largest fragments were also the most irregular ones (Bustamante & Castor 1997).
Thus, the ruil forest today forms an "archipielago" with many small, regular fragments and few large, irregular ones, surrounded by a matrix of pine plantations. The ruil deforestation has consequences for biodiversity conservation at both the ecosystem and species level. The ruil forest as an ecosystem is heading toward extinction. If the current rate of deforestation remains unabated, even ignoring deleterious effects other than area reduction, the ruil forest as a recognizable biome will disappear within the next decade due to the extinction of many species associated with this forest. In fact, currently the ruil forest contains 18 % of the 11 endangered tree species of Chile (Benoit 1989). Among these, Pitavia punctata and N. alessandrii, are the most representative of the forest.
Despite being a unique ecosystem severely threatened, only 45 ha of the ruil forest are under formal protection in the Chilean System of Protected Areas. This figure is striking as almost 18% of Chile is under protection, and over 90% of such coverage is in the region of temperate forest (Simonetti & Armesto 1991). Given the tendency to replace native forest by plantations, the reduction of the ruil forest to just these 45 ha is, albeit pessimistic, a potential scenario. Based on area-species curves, it is feasible to speculate on the magnitude of species extinctions if such a forest reduction occurs (Reid 1992). Assuming that the current 352 ha of ruil forest is a single unit (a very conservative assumption), an attrition to 45 ha means an 87% reduction in area, that could trigger the extinction of 36 to 73 species of current 146 plant species composing the ruil forest. Even if extinctions were local, with species surviving elsewhere in Chile, such a massive change in composition conveys the extinction of the ruil forest as a biome. Some extinctions though, could be global, as some species such as P. punctata and N. alessandrii are restricted to the ruil forest. If the ruil forest is to be conserved, the survival of all remnants, largely those in productive unprotected lands, seems mandatory.
The specific configuration of forest remnants may have additional negative effects. The small fragments and the larger and irregular ones may be particularly prone to biotic and abiotic influences from the matrix (Murcia 1995). In the ruil forest, these effects may be particularly evident at early stage of development of pine plantations where structural differences between the matrix and the forest are striking. Abiotic effects preclude the regeneration of shade-tolerants plants such as C. alba, P. boldus and G. avellana. In this scenario, shade-intolerants species, such as N. alessandrii an N. glauca, could be favoured by habitat fragmentation. This is a paradoxical case where an endangered species, such as N. alessandrii, may be favoured by fragmentation. Nevertheless, this favourable effect may be hampered due to the high deforestation rate, which is reducing the absolute amount of edge habitat in the landscape.
Monterey pine presents a double threat to the ruil forest. This is an invasive species (Rejmánek 1996), intruding on the fragments of ruil forest (San Martin et al. 1984). Due to its higher ability to obtain water (Lara et al. 1996), Monterey pine could outcompete native trees. Monterey pine is also fire-prone (Lara & Veblen 1993). In the VII Administrative Region where the ruil forest is located, 10 % of the plantations were burned between 1983 - 1990, with an average of 13 ha per fire (Sáiz 1990). Since the ruil forest is embebed in a pine matrix, any fire in plantations may obliterate the ruil remnants.
The extinction probability of populations in archipielagoes, like the ruil forest, is higher than in a single remant of similar area (Burkey 1995). Thus, distance among fragments and kind of matrix are two important variables determining isolation and extinction probabilities. Distance among fragments of the ruil forest (55 m in average) may not be critical for wind-dispersed seed if the matrix were an open area (Serra et al. 1986). However, the presence of pine plantations between fragments may be an effective barrier for seed dispersal (Fenner 1985). Thus, 35% of the native species of trees which are wind-dispersed, such as N. alessandrii, may have additional problems for population survival (Bustamante & Castor 1997). Pine plantations may not be a barrier for vertebrate-dispersed species such as C. alba, P. boldus and G. avellana, due to the relative high vagility of dispersers (i.e., foxes, birds) (Bustamante et al. 1992).
Forest removal also modifies other biotic interactions such as pollination (Murcia 1996). As forest loss reduces the abundance of native pollinators (Aizen & Feinsinger 1994a, b), it is reasonable to expect detrimental effects on the reproductive output for these plants that depend on them, adding more risk to the persistence of the forest. This could be the case for 78 % of tree species of the ruil forest which requires biotic pollination (Bustamante & Castor 1997).
The current situation of the ruil forest is critical due to the multiple direct and indirect effects of forest deforestation and fragmentation. This is not an isolated case. Latinamerica is one of the main reservoirs of natural forests in the world, but deforestation (and probably fragmentation) is dangerously increasing. In fact, 25 - 30 % of the forest cover in Latin America was lost between 1850 and 1985, and half of that reduction ocurred after 1960 (Houghton et al. 1991). The dominant image emerging from this deforestation is forest fragments surrounded by a structuraly completely different matrix such as grasslands. The ruil case has an heuristic value as forest remnants are surrouned by forests, challenging the current knowledge regarding edge effects, the meaning of corridors and the matrix itself, and offering an opportunity to advance ecological knowledge relevant to ensuring the persistence of an unique ecosystem and its component species.
It seems that land use in central Chile is not sustainable. Sustainability implies economical, ecological and socio-cultural issues. Even when pine plantations may offer a profitable economical income (under current market interests), this benefit is reached at the expense of socio-cultural and ecological aspects. From a socio-cultural point of view, extensive forest plantations increase poverty and unemployent as plantations demand low workforce. The increasing local unemployment has triggered the emigration of peasants (Lara & Veblen 1993, Unda et al. 1997). Furthermore, the loss of native forest because of an inappropiate management is considered by local people to be one of the main environmental problems of the region (Hajek et al. 1990). From an ecological point of view, land management is definitively unsustainable. We have no evidence that Monterey pine is degrading the land where it is planted, but as discussed above, this exotic species is the main reason for ruil forest loss and fragmentation, and ultimately for its current endangered status.
The critical situation of the ruil forest deserves an active conservation strategy. Going back to the former two questions dealing with ecological sustainability of forest management, we have the following propositions for the ruil forest. First, deforestation must be stopped as soon as possible. Furthermore, since there are many species that are already endangered in the ruil forest, even if no more forest is cut, it might be appropiate to suggest that forest restoration and not just conservation is necessary to ensure survival of the existing species. Several existing species may be already committed to extinction due to prior habitat loss and fragmentation (Tilman et al. 1994). The time lag between deforestation and species extinction offers an opportunity to restore habitats before loosing species that otherwise will vanish. Restoration should be combined with changes in the spatial configuration of remnants, thus reducing forest fragmentation. Isolation effects must be minimized by increasing the connectivity of the landscape, through corridors. This should not be difficult if we consider the average distance among fragments of 50 m. Larger but irregular fragments should be modified toward more circular-shaped ones, by adding native species between the lobes. This will increase the forest surface and reduce the negative edge effects. Also the invasion of Monterey pine inside remnants must be prevented by setting up buffer zones without pine but with other non invasive plants or by actively removing pine seedlings and saplings from remnants (Bustamante & Grez 1995).
Clearly, the fate of the ruil forest depends upon the survival of the remnants outside the minute protected area where it exists today. This challenge, faced by many species, is a challenge to the forest industry and conservationists in order to agree in land management practices suitable to ensure both the development and economic success of forest plantations and the maintenance of the Chilean natural heritage (Simonetti 1997).
Landscape ecology offers conceptual, empirical and methodological elements for a regional land management. In fact, the landscape scale has been considered highly promising for planning a sustainable environment and for nature conservation (Forman 1990, Hansson & Angelstam 1991). However, in Chile the discipline is poorly developed, with no permanent undergraduate or graduate formation in the area (Grez et al. 1995). Therefore, it is understandable that there has not been a systematic application of this discipline for the long lasting sustainable management of natural resources in spite of its great potential (Simonetti & Armesto 1991). However, landscape ecology alone will not solve all the problems associated with land management. Ecosystem management at a landscape level needs the coordination of the whole community including those with sociological, ecological and economical priorities. Most of the time this is a hard task, due to the different expectations of the groups involved with the system to be managed. But that should not be an excuse for doing nothing. We have to begin with a mutidisciplinary work for landscape management before it is too late to avoid extincions and even recover our native biodiversity.
AKNOWLEDGMENTS: Our work in landscape ecology has been partially developed under the auspice of Fondecyt 1173-92 and 1970853 to AA Grez. R Bustamante has been supported by the Programa Bosques Nativos, Departamento de Investigación y Desarrollo, Universidad de Chile, and L Fahrig has been supported by a research grant from the Natural Sciences and Engineering Council of Canada.
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