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This work has succeeded in quantifying the major changes in forest cover in south-central Chile, and in assessing the substantial forest loss that has taken place over the last 25 years in the study area. The research applies standard remote sensing methods in conjunction with spatial pattern analyses to an area of southern-hemisphere temperate forest never previously studied.

The analyses conducted illustrated how forest loss was strongly associated with an increase in area of exotic-species plantations. The changes in pattern, in turn, were related to increased forest fragmentation, which was re- flected by significant modification of size, density, edge, isolation, connectivity, and core area of forest patches. The successful description of pattern change accompanying deforestation and forest fragmentation provides a critical component of habitat analysis.

At a local level, these changes may result in the elimination, displacement, or enhancement of species populations. Additionally, the identification of these patterns is important to facilitate future landscape management and monitoring actions in this type of forest. Proactive management with a focus on biodiversity conservation and sustainable use is urgent. Although an analysis of the effects of fragmentation on the species and genetic levels of biodiversity was outside the scope of the present study, the description of landscape spatial pattern provides a basis for future research investigating such impacts.

This research was made possible by funding from BIOCORES (Biodiversity conservation, restoration, and sustainable use in fragmented forest landscapes) project ICA-CT-2001-10095; Bursary for young researchers from developing countriesEuropean Union (administered by UFZ, Germany); the UNEPWCMC Chevening Scholarship in Biodiversity; and the FORECOS Scientific Nucleus P01-057-F (MIDEPLAN). The manuscript was improved by comments from three anonymous reviewers.

appendix 1

appendix 1

source: Rapid deforestation and fragmentation of Chilean Temperate Forests
by Cristian Echeverria a,b, * ,1 , David Coomes a , Javier Salas c , Jose´ Marı´a Rey-Benayas d , Antonio Lara b , Adrian Newton

Forest cover change

forest change

The native forest of Rio Maule-Cobquecura has undergone high degrees of forest clearance during the three decades analysed here, compared to many other forested landscapes. Initially, the forest was severely deforested and degraded due to logging for fuelwood and clearance for cultivation (San Martı´n and Donoso, 1997). Forest losses recorded for other temperate forests, assuming that they were calculated using the compound-interest-rate formula, have generally been lower.

A forest loss rate of 0.53% was estimated for the Klamath-Siskiyou ecoregion, USA, and an overall (cumulative) reduction of forest cover by 10.5% was recorded over the period 1972–1992 (Staus et al., 2002).

In western Oregon, deforestation rates by clearcutting between 1972 and 1995 varied from 0.5% to 1.2% with almost 20% of the total forest impacted (Cohen et al., 2002). Similarly, in other areas of western Oregon, between 1972 and 1988 the rate of deforestation by primarily clearcutting was 1.2% of the entire study area including the wilderness area (Spies et al., 1994). In the central Sikhote-alin Mountains of the Russian Far East, an 18.3% total reduction in conifer forest and a 7.4% total reduction in hardwood forest cover were reported between 1973 and 1992 due mainly to disturbance by timber harvesting (Cushman and Wallin, 2000).

A rate of 0.6%, much lower than that determined for the period 1990–2000 in the present study, was found for the years 1986 to 1996 in the Napo region of western Amazonia (Sierra, 2000). A higher rate of 6% was determined for lowland deciduous forest in eastern Santa Cruz, Bolivia in the middle 1990s (Steininger et al., 2001), thought to be one of the highest deforestation rates reported anywhere in the world. Higher forest loss rate has been reported in an aerial photography-based study conducted for Nothofagus alessandrii forests, a threatened endemic tree to Chile, within the present study area (Bustamante and Castor, 1998). Using also the compound-interest-rate, this forest type declined at a rate of 8.15% between 1981 and 1991. If this rate remains constant, a total extinction of N. alessandrii forest is predicted by the authors in the next decade.

Lower rates of forest loss have been reported in other aerial photography-based studies conducted in the temperate forests of southern hemisphere, using the absolute-loss-rate formulae. For the pre-Andes of the Maule Region of Chile, Olivares (2000) reported deforestation rates in two study areas ranging from 0.5% to 1.4% between 1987 and 1996. For the Coastal Range of Chile, Lara et al. (1989) determined deforestation rates for the period 1978–1987 from 2.0% to 3.5% in the coast of the Maule (including Rio Maule-Cobquecura) and Bio-Bio regions respectively – closely comparable to the period 1990– 2000 analysed for Rio Maule-Cobquecura (3.64%). Conversion to exotic-species plantations and clearance for agricultural land played a major role in these totals (Olivares, 2000; Lara et al., 1989).

For the entire Maule region, a lower annual forest loss of 0.58% was determined between 1994 and 1999, using aerial photographs at difference mapping scales (Conaf and Uach, 2000). Further south, in northern provinces of the Lake region, an overall reduction of 18,100 ha, corresponding to an annual deforestation of 0.3% between 1995 and 1998 was obtained using remote sensing and aerial photographs (Conaf et al., 1999b).

Different types of data and mapping scales have been used at different measurement times in the studies mentioned above, which might have had an effect on the estimation of forest loss. The present study, in contrast, used a consistent type of data over a longer period. An important increase in the area of shrublands was detected by 1990 as a result of the elimination of native forests and arboreus shrublands during 1975–1990. Agricultural and pastureland lands also expanded slightly by 2000 with the reduction of native forests and shrublands in flat areas situated under 200 m elevation.

source: Rapid deforestation and fragmentation of Chilean Temperate Forests
by Cristian Echeverria a,b, * ,1 , David Coomes a , Javier Salas c , Jose´ Marı´a Rey-Benayas d , Antonio Lara b , Adrian Newton


The present study constitutes the most extensive analysis of deforestation and forest fragmentation ever conducted in Chile. The results demonstrate how changes in spatial patterns of the forested landscape may be assessed using multitemporal satellite data, as has been achieved recently in other areas of the world (Tommervik et al., 2003; Rees et al., 2003).

Assessment of methods and results

As classifications of satellite imagery into land cover types are never completely accurate, analyses of forest loss and landscape patterns of the present study are affected by errors in the classifications. According to the confusion matrices (Appendix 1), image accuracy tended to improve as the image date became more recent. This trend may be related to image quality, which is better in TM and ETM+ images. For the original image, the accuracy assessment showed that the native forest cover is slightly overestimated which can affect the analyses derived from the 1975 map. However, this might be related to the availability of concurrent land cover maps for the earliest images.

For instance, the accuracy of the 1975 map was assessed using aerial photograph-based maps for the year 1978. This difference in time may have caused a higher error commission than error omission, as an important area classified as native forest in 1975 had been converted to pine plantations or shrublands by 1978. This rapid conversion is more accentuated in the earliest images as this is when considerable land cover changes occurred in the study area. The high percentages of overall accuracy of the images revealed that the supervised classification, which was strongly supported by ground-based information, provided a suitable identification of land cover types in each of the satellite scenes processed.

Although the classification conducted for the oldest scene (MSS 1975) suffered from the disadvantage of limited ground validation, a cautious interpretation considering field control points that had not changed over time enabled high values of accuracy to be obtained, and clearly identifiable categories of land cover in the classification. In order to manage these errors propagated through the analyses of change and patterns, some measures were adapted from Brown et al. (2000).

In the present study, the error was minimised by applying improved topographic correction algorithms in the image processing and by aggregating some of the sub-categories of land cover types. Additionally, different ground-based surveys conducted from previous works and additional field visits were used to support the image classifications. Also, the error was minimised by filtering the classified image to remove small patches of less than four 30 m-pixels (equivalent to 0.36 ha).

source: Rapid deforestation and fragmentation of Chilean Temperate Forests
by Cristian Echeverria a,b, * ,1 , David Coomes a , Javier Salas c , Jose´ Marı´a Rey-Benayas d , Antonio Lara b , Adrian Newton


Accuracy assessment

chileAccording to the rules suggested by PCI geomatics for ranges of separability values of class signatures, a ‘‘good’’ signature separability was achieved for each image classified, the average Bhattacharrya distance ranging from 1.979 to 1.987. Overall agreement of classification was 82.7% for the 1975 MSS, 83.3% for 1990 TM image, and 84.9% for the 2000 ETM+ image (Appendix 1). The lowest values of accuracy corresponded to shrubland, arboreus shrubland, and native forest categories. These three formations are very similar in their spectral signatures (Bhattacharrya distance less than 1.9) because they correspond to stages in a continuous succession process, which may easily produce misclassifications between the categories assigned in the training site and those classified by the algorithm. Supporting data from other studies and control points were important to distinguish these stages of forest succession.

Major changes in land cover

Changes in land cover were analysed using the area statistics (Table 1) derived from land cover type maps (Fig. 2). The estimated cover of native forests decreased from 119,994 ha in 1975 to 39,002 ha in 2000. In other words, 67% of the native forests existing in 1975 had been replaced by other land cover types by 2000. In 1975 the native forests were distributed throughout the study area. Twenty five years later, these formations were restricted to small patches, sparsely distributed


across the landscape (Fig. 2). Conversely, exotic-species plantations increased from 5% land area in 1975 to 17% in 1990; by 2000 this land cover type was the dominant cover on the map, comprising 36.6% of total land area.

In 1975 exotic-species plantations were located principally in three specific areas within the study area. By 2000, the plantations had rapidly expanded across the landscape, reaching sites at different altitudes and aspects. Shrublands and arboreus shrublands were the dominating vegetation formations in the cover maps of 1975 (33%) and 1990 (58%). Ten years later, these formations decreased to represent 34% of total area, and the exotic-species plantation became the dominant land cover type at that time.

During the first time interval, 36% of the native forest area was converted to shrubland and arboreus shrublands, 29% to exotic-species plantations, and 31% remained as native forest. During the second time interval, a substantial area (50%) of native forest was replaced by exotic-species plantations whereas only 7% was converted by logging to shrublands and arboreus shrublands and 36% remained as native forest. More than half (53%) of the native forest cover existing in 1975 has gradually been converted into exotic-species plantations by 2000; another substantial area (40%) was transformed into shrublands or arboreus shrublands.

From 1975 to 2000, agricultural and pasture lands exhibited a slight increase (Table 1). However, in 1990 the area occupied by this cover type declined to 14% as a consequence of the conversion of land cover types (including pasture lands) to exoticspecies industrial plantations after the promulgation of the law on forestation in 1976. Conversely, shrubland presented an increase from 1976 to 1990 due to the clearance of secondary forests. In 2000, most of this shrubland appeared to be covered by exotic-species plantations.


By overlapping the forest native cover of each year, it was observed that between 36% and 44% of native forest was derived from regeneration of shrubland into secondary forest between time intervals respectively. However, it was noted that between 80% and 90% of this category corresponded to small patches whose area was less than 5 ha. During the whole study period, the annual deforestation rate was 3240 ha year 1 , equivalent to 4.5% year 1 using the compound-interest-rate formula. Most of the forest loss was concentrated in the first 15 years of the study period, at a deforestation rate of 5.06% year 1 , corresponding to 4257 ha year 1 . Between 1990 and 2000, the rate decreased slightly to approximately 3.64% year 1 , equivalent to 1713 ha year .

Variation in forest patch size

One of the basic symptoms of forest fragmentation is the increase in number of smaller patches (Fig. 3). In the Rio Maule-Cobquecura considerable changes were found in the distribution of forest patch size between time intervals (Fig. 3). By 1975, 44% of the forest area was concentrated in a large patch between 20,000 and 100,000 ha; the remaining forest area occurred in isolated patches of less than 10,000 ha, with almost half occurring in very small patches of less than 100 ha. In 1990, 59% of the total area of native forests occurred in patches of less than 100 ha. By 2000, this percentage increased to 69% and only 3% had a size greater than 1000 ha .

Spatial configuration of native forest cover

The highest density of patches was recorded in 1990 with 1.65 fragments of native forest per 100 ha, subsequently decreasing to 1.36 in 2000. During the first period, the native forests were mainly affected by severe fragmentation (increasing largest forest patch (Table 2), ranging from 7% in 1975 to 0.2% of the total area in 2000. Continuous areas of quality habitats decreased following the introduction of disturbed fragments into the matrix. These modifications of the landscape were also characterised by the presence of more patch edges (Table 2).


source: Rapid deforestation and fragmentation of Chilean Temperate Forests
by Cristian Echeverria a,b, * ,1 , David Coomes a , Javier Salas c , Jose´ Marı´a Rey-Benayas d , Antonio Lara b , Adrian Newton


Remote sensing data

forestTo analyse changes in forest area and spatial pattern over time, a set of three Landsat scenes were acquired for the years 1975 (MSS), 1990 (TM), and 2000 (ETM+). In order to carry out a quantitative comparison of the images in the present study, the original 79 m MSS raster grids were resampled to the resolution of the TM and ETM+ raster grids (30 m) (Staus et al., 2002; Steininger et al., 2001). This process was made using the re-project algorithm of PCI (2000). The fine grain used here (30 m) allowed the identification of non-forest areas within forest patches, which is an important spatial attribute to identify in areas affected by high-grading (selective felling). The presence of small patches is an important attribute to quantify in forest fragmentation analysis, but can only be assessed using high-resolution imagery (Millington et al., 2003). Although some of these small fragments may be important for the conservation of some species (Grez, 2005), a minimum mapping unit of greater than 5 pixels were used in this study. This enabled differences in data quality produced by the resampling of the MSS images to be minimised.

Pre-processing of the satellite data

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). Geometric correction was performed using the ‘‘full processing’’ module in PCI Geomatics. This consisted of a transformation of each image using both GCPs (ground control points) and a 2nd order polynomial mathematical model. ETM+ images were spatially corrected in order to use them as a basis to correct the MSS and TM images. The satellite images were georeferenced separately to vector maps by locating approximately 55–65 corresponding GCPs in each image and the reference map.

Road networks, based on topographic maps digitalized in 1970s, were used to correct the 1976 and 1975 images. The geometric accuracy ranged from 0.10 to 0.39 pixels, corresponding to 3– 11.7 m. Atmospheric correction was applied to all the scenes transforming the original radiance image to a reflectance image (Cha´vez, 1996). The topographic correction was performed for each scene using the method proposed by Teillet et al. (1982) in order to remove shadows in hilly areas.

Image classification

Four resources were available to aid image classification. ‘‘Catastro’’ is a GIS-based data set of thematic maps derived from aerial photographs and satellite imagery between 1994 and 1997 (Conaf et al., 1999a). This data set provided detailed information on land use and forest types (including dominant tree species, forest structure, and degree of disturbances) at 1:50,000 scale considering a minimum mapping unit of 6.25 ha. Catastro was used both to define the land cover types for the present study and for the image classification of the 2000 ETM+ scene.

A second set of data comprised 11 digital aerial photographs at 1:115,000 taken in 1999 (Conaf and Uach, 2000), which were also used to the image classification of the ETM+ image. A third data set corresponded to forest cover maps generated from aerial photograph at 1:60,000 between 1978 and 1987 (Lara et al., 1989). These maps were used for the image classification and for the accuracy assessment of the earliest images. A four reference group corresponded to 65 control points sampled in field visits between 2001 and 2002.

Information on the history of land cover change for the points visited was also collected for the interpretation of the images, particularly for the earliest ones. Owing to the availability of these ground-based data sets, a supervised classification was the method chosen to classify the three Landsat scenes. The statistical decision criterion of Maximum Likelihood was used in the supervised classification to assist in the classification of overlapping signatures, in which pixels were assigned to the class of highest probability. The selection of training sites was done considering representation of all digital categories of radiance according to the numeric values (spectral signature) and colour composites (Chuvieco, 1996).

Some of these training areas were consistently delineated in each scene in order to minimise classifi- cation errors when performing change detection (Luque, 2000). Signature separability was assessed by the Bhattacharrya distance which is used to analyse the quality of training sites and class signatures before performing the classification. Accuracy assessments of the MSS (1975) and TM (1990) images were conducted using aerial photograph-based land cover maps developed by Lara et al. (1989) for the years 1978 and 1987.

Two new sets of 369 and 360 points were used for this purpose for each image respectively. The points were overlain on the reference land cover maps and assigned to the respective class. Confusion matrices were constructed to compare the class identified for each sample point with the land covers derived from the satellite images (Appendix 1). The accuracy of the ETM+ image was assessed by ground-truthing of 226 points visited between 2002 and 2003 (Appendix 1).

Land cover types

The following basic categories of land cover were identified from each image: (1) agricultural land (three sub-categories), (2) shrubland, (3) arboreus shrubland (an intermediate successional stage between shrubland and secondary forest with dominance of sclerophyllous species), (4) secondary forests (composed mainly of Nothofagus species such as N. obliqua, N. glauca and N. alessandri), (5) exotic-species plantation (mainly Pinus radiata), (6) young or new plantation, (7) wetland, (8) bare ground, (9) urban areas, and (10) water bodies. All forest cover in the study area was classified as secondary forests, owing to the absence of primary forest formations in the landscape examined.

source: Rapid deforestation and fragmentation of Chilean Temperate Forests
by Cristian Echeverria a,b, * ,1 , David Coomes a , Javier Salas c , Jose´ Marı´a Rey-Benayas d , Antonio Lara b , Adrian Newton

Study area

fig1The Rio Maule-Cobquecura study area covers approximately 578,164 ha of land located in the Coastal Range of the Maule and Bio-Bio regions in south-central Chile (Fig. 1). The studied landscape partially includes the municipalities of Constitucio´n, San Javier, Quirihue and Cobquecua, and all the area of the municipalities Empedrado, Chanco, Pelluhue, and Cauquenes. The area is characterised by rainfall concentrated during the winter that leads to dry summers from September to April with little cloud cover and high luminosity. The natural forest is mainly dominated by secondary forest of Nothofagus species (N. obliqua and N. glauca) (Fagaceae) and sclerophyllous species including Acacia caven (Mimosaceae), Quillaja saponaria (Rosaceae), and Maytenus boaria (Celastraceae).

Also, many endangered species such as N. alessandri, Pitavia punctata (Rutaceae), and Gomortega keule (Gomortegaceae) occur in the study area. Forest clearance on a significant scale began with the arrival of European colonizers in the XVI–XVIIth century. Then, in middle of the XXth century a boom in the cultivation of wheat crops resulted in the elimination of extensive forest areas in the Coastal Range of the study area (San Martı´n and Donoso, 1997). In recent decades, the use of native forests for firewood has led to extreme degradation of this forest resource (Olivares, 2000).

According to field observations and different surveys, most of the native forests in the study area correspond to highly degraded secondary forest (Donoso and Lara, 1995). This type of fragmented forest is severely impoverished in commercially valuable timber species as a result of selective logging and fire wood extraction. Some of the remnant forest fragments are areas that were clear-felled for shifting cultivation and subsequently abandoned (Lara et al., 1997).

source: Rapid deforestation and fragmentation of Chilean Temperate Forests
by Cristian Echeverria a,b, * ,1 , David Coomes a , Javier Salas c , Jose´ Marı´a Rey-Benayas d , Antonio Lara b , Adrian Newton


chileHabitat 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) 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 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 Ko¨hler, 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 temporal evaluation of forest change based on satellite imagery linked to fragmentation analysis is becoming a valuable set of techniques for assessing the degree of threat to ecosystems (Armenteras et al., 2003; Franklin, 2001; Imbernon and Branthomme, 2001; Luque, 2000; Sader et al., 2001).

A number of deforestation studies have been conducted in tropical forests (Imbernon and Branthomme, 2001; Sader et al., 2001; Skole and Tucker, 1993; Steininger et al., 2001; Turner and Corlett, 1996) and, in particular, in the Amazon, which is now considered as the most studied region in the world by some researchers (Jorge and Garcia, 1997; Laurance, 1999; Laurance et al., 2000; Pedlowski et al., 1997; Ranta et al., 1998; Sierra, 2000). Conversely, few studies of deforestation and fragmentation have been reported for temperate forests (Staus et al., 2002), particularly in the southern hemisphere.

Chile has the largest temperate forest area in South America and more than half of the total area of temperate forests in the southern hemisphere (Donoso, 1993). Most of these forests are distributed along the Coastal and the Andean Range of Chile from 35 to 56 S and extend to a total of 13.4 million ha in the country (Conaf et al., 1999a). The temperate forest of Chile has been classified a biodiversity hotspot for conservation (Myers et al., 2000) and has also been included among the most threatened eco-regions in the world in the Global 200 initiative launched by WWF and the World Bank (Dinerstein et al., 1995).

In these forests, a 34% of the plant genera are endemic (90% monotypic) (Armesto et al., 1997). However, Chile’s temperate forests are being harvested to supply the increasing global demand for wood and paper products. A substantial amount of forest has also been lost due to the conversion of native forests to pasturelands, human-set fires, high grading (selective felling) and other logging practices (Lara et al., 2000). Although some of the ecological consequences of forest fragmentation have been studied in Chile (Bustamante and Grez, 1995; Donoso et al., 2003; Vergara and Simonetti, 2004; Willson et al., 1994), integrated spatial and temporal analyses have not been conducted. Although some attempts have been made in Chile to estimate the rate of deforestation (Lara et al., 1989; Olivares, 2000) or to assess land cover change (Conaf et al., 1999b; Sandoval, 2001), these have been undertaken at local scales, over short time intervals (no more than 10 years), or using different types of data to compare forest cover over time, which confers some methodological limitations. Longterm analyses of the spatial patterns of deforestation and fragmentation of temperate forest ecosystems at the landscape scale have not yet been reported either in Chile or elsewhere in the southern hemisphere.

The purpose of this study is to contribute to the understanding of the patterns of deforestation and fragmentation in the temperate forests of Chile at the landscape level. In particular, we examined the patterns of land-cover change and the changes in the spatial configuration in the Maulino temperate forest over time and space by using satellite scenes acquired at different time intervals. In this study, we hypothesised that there has been a substantial loss of native forest as a result of an increase in area of exotic-species plantations. Also, we anticipated that this forest loss was related to fragmentation of Maulino forests due to changes in the spatial configuration in terms of size, shape and degree of isolation of forest patches. This work is the first step to understand the potential ecological effects of fragmentation and the proximate drivers and causes of deforestation, which will be addressed in other studies.

source: Rapid deforestation and fragmentation of Chilean Temperate Forests
by Cristian Echeverria a,b, * ,1 , David Coomes a , Javier Salas c , Jose´ Marı´a Rey-Benayas d , Antonio Lara b , Adrian Newton


forest chileThe temperate forests of Chile are classified a biological ‘‘hotspot’’ as a result of their high species diversity and high endemism. However, they are being rapidly destroyed, with significant negative impacts on biodiversity. Three land-cover maps were derived from satellite imagery acquired over 25 years (1975, 1990 and 2000), and were used to assess the patterns of deforestation and forest fragmentation in the coastal range of south-central Chile. Between 1975 and 2000, there was a reduction in natural forest area of 67% in the study area, which is equivalent to an annual forest loss rate of 4.5% per year using a compound-interest-rate formula.

Forest fragmentation was associated with a decrease in forest patch size, which was associated with a rapid increase in the density of small patches (<100 ha), and a decrease in area of interior forest and in connectivity among patches. Since the 1970s, native forest loss was largely caused by an expansion of commercial plantations, which was associated with substantial changes in the spatial configuration of the native forests. By 2000, most native forest fragments were surrounded by highly connected exotic-species plantations. The assessment of forest loss and fragmentation provides a basis for future research on the impacts of forest fragmentation on the different component of biodiversity. Conservation strategies and land use planning of the study area should consider the spatial configuration pattern of native forest fragments and how this pattern changes over time and space at landscape level.

source: Rapid deforestation and fragmentation of Chilean Temperate Forests
by Cristian Echeverria a,b, * ,1 , David Coomes a , Javier Salas c , Jose´ Marı´a Rey-Benayas d , Antonio Lara b , Adrian Newton

Impact of dominant tree dynamics on site index curves



Pinus banksiana Lamb

Pinus banksiana Lamb

Site index curves were modeled for two species of different shade tolerance, black spruce (Picea mariana (Mill.) B.S.P.) and jack pine (Pinus banksiana Lamb.), from an extended network of permanent sample plots (PSP) that covers periods of time varying from 10 to 30 years, in the province of Quebec. A data set reserved for validation allowed us to compare the site index curves derived from PSPs with published site index curves fitted to temporary sample plots (TSP) and stem analyses (SA). For both species, the site index curves calibrated from PSPs and TSPs behave similarly as they have comparable average bias and accuracy.

The major difference is seen with the SA curves that strongly overpredict the dominant height growth of the PSPs. The similar pattern of change of site index curves calibrated from TSP and PSP data reinforces their validity as both types of curves were calibrated with independent data sets and methodologies. The differences observed between SA and PSP curves were likely produced by the dynamics of dominant height related to tree mortality and change in social status. For both species, approximately one tree out of five (22% for black spruce and 16% for jack pine) was replaced every 10 years in the tree group that was used to estimate dominant height. Consequently, the trajectory of dominant height through time for a particular plot is saw-toothed, the size of the ‘‘teeth’’ being, among other things, a function of stand regularity, as measured by an evenness index. Due to this tree replacement dynamic, stand dominant height curves are also more rapidly asymptotic than those of individual trees Continue reading “Impact of dominant tree dynamics on site index curves” »

Dynamic growth model for Scots pine (Pinus sylvestris L.) plantations in Galicia



Pinus sylvestris L In this study we developed a dynamic growth model for Scots pine (Pinus sylvestris L.) plantations in Galicia (north-western Spain). The data used to develop the model were obtained from a network of permanent plots, of between 10 and 55-yearold, which the Unidade de Xesti ´on Forestal Sostible (Sustainable Forest Management Unit) of the University of Santiago de Compostela has set up in pure plantations of this species of pine in its area of distribution in Galicia. In this model, the initial stand conditions at any point in time are defined by three state variables (number of trees per hectare, stand basal area and dominant height), and are used to estimate stand volume, classified by commercial classes, for a given projection age. The model uses three transition functions expressed as algebraic difference equations of the three corresponding state variables used to project the stand state at any point in time. In addition, the model incorporates a function for predicting initial stand basal area, which can be used to establish the starting point for the simulation.

This alternative should only be used when the stand is not yet established or when no inventory data are available. Once the state variables are known for a specific moment, a distribution function is used to estimate the number of trees in each diameter class, by recovering the parameters of the Weibull function, using the moments of first and second order of the distribution (arithmetic mean diameter and variance, respectively). By using a generalized height–diameter function to estimate the height of the average tree in each diameter class, combined with a taper function that uses the above predicted diameter and height, it is then possible to estimate total or merchantable stand volume. Continue reading “Dynamic growth model for Scots pine (Pinus sylvestris L.) plantations in Galicia” »