Fiscalía Regional Metropolitana Oriente 1. Elementos de contexto:

Recently biodiversity has become a major feature in conservation science, where it is often considered key in determining areas to spend resources. At present woodland biodiversity in the UK is in decline due to high rates of forest loss and degradation as a result of over- exploitation in some cases, neglect in others, and conversion of forest to other land uses (Newton et al., 2009b; Turner et al., 2003). Here ‘biodiversity’ is referred to as species and certain characteristics of species, primarily their distribution and number within a given area. In addition the use of biodiversity is meant more broadly to mean species assemblages and ecological communities – groups of interacting and interdependent species (Turner et al., 2003). Scale is an implied and key component of this definition (McElhinny et al., 2005). Scientifically sound management requires frequent and spatially detailed assessments of species numbers and distributions (Turner et al., 2003). Such measurements can be prohibitively expensive to collect directly. Underlying this assessment of ‘condition’ is the assumption that certain key environmental parameters, which can be detected, will drive the distribution and abundance of species across landscapes and determine how they occupy

habitats (Turner et al., 2003). McElinny et al. (2005) state that the quantification of diversity should be made through the identification of the structural attributes for a forest stand. An effective and efficient biodiversity surrogate measure needs to be formed from an array of different structural variables, some of which were identified as: foliage arrangement; canopy cover; tree diameter; tree height; tree spacing; tree species; stand biomass; understorey vegetation and deadwood.

Most management programmes to conserve biodiversity focus upon the creation of protected areas. Conservation status is typically based on the assessment of landscape and ecosystem level features such as habitat loss, habitat fragmentation, the size and number of large habitat blocks, the degree of protection, and current potential threats (Noss, 1999). Loss of biodiversity and ecosystem integrity is often experienced in conjunction with a number of factors illustrating degenerative trends, for example: old forests have been replaced with younger forests and plantations; structurally complex stands have been replaced by simplified ones; large well connected patches have been replaced with smaller, more isolated patches; natural fires have been suppressed; many miles of road have been built in what were unbroken landscapes.

There are two general approaches to the measurement of biodiversity (Turner et al., 2003). The first is the direct measurement of individual organisms, species assemblages, or ecological communities. This depends upon the scale of measurement, for example surveying species occurrence or absence from a sample region. Alternatively, there is the indirect approach to monitoring biodiversity through the reliance on environmental parameters as proxies. Consider for example, many species are restricted to discrete habitats, such as woodland or grassland which can be identified at a broader scale. By combining information about the habitat requirements of species with maps of landcover, estimates of potential species ranges and patterns of species richness are available (Turner et al., 2003). Lindenmayer et al. (2006) state that spatial connectivity between habitats should be maintained, in addition to conservation of landscape heterogeneity and stand structural complexity, in order to better guide biodiversity conservation, which again engages the idea of the interplay between different scales.

indicators and criteria have been proposed to assess the sustainable management of forests; however their scientific validity remains uncertain. Because the effects of forest disturbance, such as logging, are often specific to particular species, sites, landscapes, regions and forest types, management through the use of indicator species, focal species, or threshold levels of vegetation cover are argued to be of limited generic value, controversial and difficult to select dependent upon the species. In many cases, attributes of a species’ population, for example demographics, would be more useful in validating indicators rather than as indicators themselves (Noss, 1999).

Species loss is predominantly driven by habitat loss, and thus the overarching goal of conservation management must be to prevent this. The conservation of forest biodiversity will depend on the maintenance of habitat across a range of spatial scales. Newton et al. (2009a) outline several key principles which must be considered in the scope of monitoring for this objective. The first consideration is that of forest loss and fragmentation. Deforestation is typically accompanied by substantial forest fragmentation. The changes are associated with a decrease in percentage area of forest patches and an increase in isolation of those patches. It is necessary to consider the factor of connectivity. This is the linkage of habitats, communities and ecological processes at multiple spatial and temporal scales (Noss, 1999). Connectivity influences key biodiversity conservation processes, such as population persistence, recovery and disturbance, the exchange of individuals and genes in a population, and the occupancy of habitat patches (Lindenmayer et al., 2006; Newton et al., 2009a). The characteristics of habitat edges are influenced by patterns of land use surrounding forest fragments and can have a major impact on biodiversity by affecting ecological processes such as dispersal, establishment, survival, and growth (Fuller, 2012; Newton et al., 2009a). Edge effects influence a variety of processes, including seed rain, seed germination, removal and predation, tree growth, animal movement and avian nest predation (Murcia, 1995). Newton et al. (2009a) identified that edge effects were influenced by human disturbance within the forest fragments, such as collection of firewood and livestock browsing. Indeed substantial forest biodiversity loss can occur due to human disturbance within the fragments themselves, through activities such as logging of timber, fuel wood cutting, livestock browsing, the development of infrastructure and fire setting.

Forest stand structural complexity embodies various stand attributes in addition to how they are spatially arranged within stands. Such factors contributing to stand structural complexity can include, for example, the diameter and age-class distributions of individual trees within the stand, in relation to those trees which survive or are removed. Additional factors are the spatial distribution of structural elements within the stand, the presence of large living trees, the presence of deadwood, the presence of gaps within the stand and age since formation. The vertical heterogeneity created from multiple or continuous canopy layers and horizontal heterogeneity including foliage density, canopy openness, and horizontal patchiness of profile types are of importance, for example these attributes have links to light penetration, and providing resources to various animal species (Lindenmayer et al., 2006; Noss, 1999). This structural complexity is critical for forest biodiversity conservation because it allows organisms to persist where they would otherwise be eliminated and facilitates a more rapid return of logged and regenerated stands to a suitable habitat condition for species which have been displaced (Lindenmayer et al., 2006).

Tree species richness can be influenced by management activities and disturbance events. Newton et al. (2009a) report that the total area, core area, edge length and proximity of forest fragments were all negatively associated with mean species richness of pioneer species, and positively associated with richness of forest interior species. Patch size appears to be the most important attribute influencing different measures of species composition. Forest fragmentation can also affect genetic variation within forest species, by influencing processes of gene flow, inbreeding and genetic drift.

Velland et al. (2007) indicate that past disturbance and management may cause community species composition across sites to become more or less homogenous. The author gives the example of the alpha and beta diversity of forest plants growing on former agricultural fields in contrast to older (ancient) forests in North-America and Europe. The presence of a number of ecological filters to colonisation may exist at any stage of the colonisation process for new sites. For example, recent forests may show reduced beta diversity if habitat specialists were less successful colonists than generalist species. Strong relationships between species isolation and species richness have been reported. Velland et al. (2007) present results

effectively decoupling species composition from environmental gradients (Devictor et al., 2008).

Noss (1999) also reports a number of impacts which may occur following human modifications to the environment and those related to climate change. Road construction for example has been correlated highly with disturbance levels, and in the case of road construction for the harvesting of timber, with habitat destruction. The invasion of exotic species and their dispersal by vehicles and equipment can alter community composition, cover and biomass levels. Increased air pollution, including low-level ozone, acid precipitation and particulates has impacts upon biomass increments and tree productivity due to changes in soil pH. and nutrient content. Direct damage to leaves and other tissues can occur. Global warming is also recognised as a modifier of forest condition status due to changes in temperature and moisture abundance. Forest systems may experience changes in biomass levels, productivity and species distributions as a result (Read et al., 2009).

It should be noted that the increasing use of forests for recreation, such as the activities of hiking, hunting, camping, etc. will alter the status of a forest site. Disturbance may take the form of displacing wildlife, footpath erosion, vegetation damage, changes in ground-level vegetation density and condition, and exotic species invasions (Noss, 1999).

It is recognised that ungulate herbivores can have a profound impact on the vegetation and soils within forests (Fuller and Gill 2001). Damage caused through twig browsing and bark peeling is an increasing problem in many European countries (Reimoser et al., 1999). Browsing and grazing from wild ungulates have always played a role in determining the structure and dynamics of natural ecological systems both in terms of their present day influence on the functioning of those ecological communities and as a powerful selection pressure in the original development of such systems (Putman, 1996).

Damage caused by large ungulates, such as deer, can occur at many levels, for example through trampling, feeding on the fruit or germinating seedlings, reducing the seed source and hampering natural regeneration, fraying, and through browsing or bark stripping of older trees that have survived the recruitment stage (Putman, 1996; Reimoser et al., 1999). Heavy grazing pressure can also result in dramatic changes in the composition and relative abundance of species of the woodland floor, even reducing diversity (Kirby, 2001). Grazing

does have positive effects via the maintenance of the heterogeneity of structure that many conservation managers seek to mimic: such as the opening of clearings, treading-in of seeds into the ground and their dispersal, and reducing canopy shade to permit the existence of shade-intolerant species (Putman, 1996; Reimoser et al., 1999). Further variability may be caused by the feeding of ungulates in one place and dunging in another, creating discontinuities in nutrient flows which may be detrimental or beneficial. It should be noted that wild herbivores, particularly deer, increase in population rapidly, often to the detriment of woodland habitats due to overgrazing (Quine et al., 2011). A long term study into Denny Wood in the New Forest, UK, a semi-ancient woodland, concluded that regeneration of tree species ceased after 1964, principally due to heavy grazing and browsing by deer and ponies (Mountford et al., 1999).

It can be argued that damage to forests is not solely dependent upon the number of ungulates present in an area, but rather a combination of environmental factors. Such factors include: the forest type, size of available area, availability of cover, habitat structure, and distance to preferred forages, etc. Therefore, more damage can be anticipated in areas where the ‘attractiveness’ of an area is high, but forage availability is low (Putman, 1996).