The effects of forest conversion on annual crops and pastures:

Agriculture, Ecosystems and Environment 69 (1998) 17±26

The effects of forest conversion on annual crops and pastures: Estimates of carbon emissions and plant species loss in a Brazilian Amazon colony S. Fujisakaa,*, C. Castillab, G. Escobara, V. Rodriguesc, E.J. Veneklaasa, R. Thomasa, M. Fishera a

Centro Internacional de Agricultura Tropical (CIAT), A.A. 6713, Cali, Colombia b International Centre for Research in Agroforestry (ICRAF), Kenya c Empresa Brasileira de Pesquisa Agropecuaria (EMBRAPA), Brazil

Abstract The municipality of Theobroma in Rondonia, Brazil, covers 2165 km2, of which 43% was deforested by 1993. Between 1973 and 1993, the national government continued to improve highway BR364 connecting the area to Brazil's south-central region and established a colony in Theobroma. During this period, 98% of the deforestation occurred. Some 1800 settler families continue to convert forest into pasture in a system based on the slash-and-burn agriculture and dual-purpose cattle production. Trends in carbon emissions and plant species losses during the 20-year history of Theobroma were analyzed by combining observed shifts in land-use types with estimates for the carbon stocks and plant species richness of each of these types. Carbon stocks declined from about 200 t haÿ1 in the forest to 28 t haÿ1 in the pasture and of 326 plant species encountered in the forest, only 20 remained in pastures (along with 66 species not found in forests). The effects of converting more than 93 000 ha of forest into other uses over 20 years include approximate losses of 14 million tons of C to the atmosphere and substantial losses of plant species. Land use alternatives that would store more C include agroforestry and ± given the strong incentives for settlers to convert lands into pasture ±improving pasture management or developing silvopastoral systems. Plant species conservation may be improved with proposed ways to add private value to the forest. # 1998 Elsevier Science B.V. All rights reserved. Keywords: Plant biodiversity; Carbon emissions; Slash-and-burn agriculture; Tropical deforestation

1. Introduction The two main sources of greenhouse gas emissions are fossil fuel consumption and agricultural practices. *Corresponding author. Tel.: 00 57 2 4450 000; fax: 00 57 2 4450 073; e-mail: [email protected] 0167-8809/98/$19.00 # 1998 Elsevier Science B.V. All rights reserved. PII S0167-8809(98)00091-7

One third of the emissions from agricultural practices are associated with land clearing, principally deforestation in the tropics (Duxbury et al., 1993). Tropical deforestation is thought to release about 1.5±3.0 Gt C yrÿ1 as CO2 or about 10±14% of global C emissions (Houghton, 1991 and Gifford, 1994). For all the greenhouse gas emissions (CO2, CH4, NOx and chloro-

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¯uorocarbons), land clearing is thought to comprise about 8% of the total (Duxbury et al., 1993). Globally, the Amazon represents the world's largest remaining tropical forest and suffers the highest deforestation rates. Analysis of satellite images indicates that up to 10% of the 4 million km2 of Brazil's originally forested Amazon was cleared by 1991 (Fearnside, 1993 and Moran et al., 1994). Government policies to build roads, establish colonies and provide incentives for cattle ranching (Fujisaka et al., 1996) have spurred deforestation of the Brazilian Amazon (which holds 55% of the Amazon). Within the area, deforestation has been highest (in terms of percentage of original forest lost) in the states of Maranhao, Tocantins, Mato Grosso and Rondonia (Fearnside, 1993), while recent rates have been highest in Rondonia and Mato Grosso (Anderson, 1990). Theobroma is a municipality and a colony of Rondonia along the highway BR364, which now connects the area to Cuiaba and the south-central region of Brazil and, to the west, to Acre as far as Rio Branco. Established some 25 years ago, the colony covers 2165 km2 and was 43% deforested by 1993. Colonists ®rst clear the forest for slash-and-burn agriculture then largely establish pastures rather than fallow and eventually re-crop lands. Colonists do not practice shifting cultivation as such. Theobroma represents 0.4% of the area of the state of Rondonia, which (statewide) was 17% deforested by 1993. The climate of this region in Rondonia is of the Am type (KoÈppen classi®cation) with an average temperature of 248C and annual rainfall of 2000±2250 mm (RADAMBRASIL, 1978). The months of June, July and August form a distinct dry season. Soils of the sampled sites are Oxisols and Ultisols with low fertility and high aluminium saturation (RADAMBRASIL, 1978). This paper examines two of the environmental impacts of deforestation ± estimated C emissions and decreases in plant diversity. Data gathered were C stocks under different land uses, plant species under the same land uses and colonists' land uses. Secondary data on deforestation in the municipality were also obtained. This paper estimates total C emissions, discusses the implications of plant diversity losses and offers some potential alternatives that would store more C and decrease deforestation and the resulting species losses.

2. Methods 2.1. Carbon Carbon was measured above and below ground in forest, newly cleared and burned cropland, fallows of 2±3 years and in pastures (largely of Brachiaria decumbens and Brachiaria brizantha with a mean age of about 10 years). At each of the two sites and for each of the different land uses, ®ve 100 m2 (425 m) transects were sampled. For aboveground C measurements, live and dead trees, logs, understorey and litter were sampled. Live tree biomass was estimated based on the diameter of trees (>2.5 cm) at breast height using a published allometric relationship (Brown et al., 1989). Carbon stocks in standing or fallen logs were estimated based on length, mean diameter and a density of 0.4 g cmÿ3. Understorey C was measured for two 1 m2 quadrats within each transect. Surface litter was collected from the same transects to the depth of the mineral soil (Anderson and Ingram, 1993). The biomass of surface litter was corrected following ashing. It was assumed that all the biomass pools contained 48% C (Castilla, 1992 and Yvan Biot 1996, personal communication). Below-ground pools included roots and soil organic C. Samples were taken at 0±20 and 20±40 cm depths. A 202020 cm sample was taken from each litter quadrat; a 2 kg sub-sample was taken and roots removed from the remainder by washing over a 1 mm sieve. Roots were oven dried and ashed and C estimated. Soil sub-samples were air dried, sieved through a 2 mm mesh and measured for C by wet oxidation (Anderson and Ingram, 1993). Large structural roots were not included in the sampling and could be a source of some C underestimation. Soil organic contents were corrected for bulk density taken from undisturbed samples. 2.2. Plant species and community change Species composition was determined for the following communities: forest, the ®rst year of cropping after forest clearing, fallowed ®elds and pasture (again, of Brachiaria sp. and of an approximate 10year mean age). Six transects were sampled for each land use except pastures. In each transect, ten 4 m2 plots were sampled with a 5 m distance between each

S. Fujisaka et al. / Agriculture, Ecosystems and Environment 69 (1998) 17±26

and the number of individuals per species recorded. The point-quadrat method (Mueller-Dombois and Ellenberg, 1974) was used to sample pastures in ®ve 20 m transects on eight ®elds, sampling every 20 cm, yielding percentage cover per species. A local forest expert (matero) helped identify species to local and, in many cases, scienti®c name. Botanists at the herbariums of the Fundacao de Tecnologia do Estado do Acre (FUNTAC) and the Universidad Federal del Estado de Acre (UFAC), Rio Branco, helped obtain additional scienti®c names. Nomenclature follows the Index Kewensis (Royal Botanic Gardens, 1993). Of all the species, 88% could be identi®ed to family, 53% to genus and 28% to species levels. Species were classi®ed as trees, shrubs (species with maximum height <2 m), herbaceous plants (nonligni®ed), vines, palms and a few `others' (including Musaceae and Bambusaceae). 2.3. Land use In 1994, 74 Theobroma farmer settlers were interviewed regarding, among other subjects, land use. Data were reviewed, cross-checked, entered and tabulated. Published secondary data on deforestation for the whole municipality showed close agreement with

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farmer-reported rates and quantities of forest conversion and were used to estimate total C emissions over the 20 years covered. 3. Results 3.1. Carbon The forest plots stored 200 t haÿ1 C, with 75% in living trees aboveground, 16% in the form of organic soil C and 4% in the roots. With slashing and burning (for annual crop production) C declined to a total 76 t haÿ1, with 45% in the form of soil organic C, 38% in the form of standing or fallen dead logs and 7% in the roots. Total C in fallowed plots (105 t haÿ1) represented 53% of what was found in the forest, with the largest proportion of C stored in aboveground live trees, soil C and dead understorey biomass. Finally, pasture with 28 t haÿ1 C represented only 14% of that stored in forest, with almost 20 t haÿ1 (69%) in the form of organic soil C (Table 1). Overall, most of the C stored in forest, cropped ®elds and fallows was in the aboveground biomass. Pastures, with the least total C, held the largest proportion in the soil. Total soil C remained

Table 1 Carbon stocks (t haÿ1) by land use, Theobroma, Rondonia, Brazil. Means and standard deviations based on ten transects (five in the case of annual crops) of 100 m2 each Land use Aboveground Trees: Understorey: Charcoal Belowground Roots:

Total Soil: Total

Live Dead Live Dead Total Fine (0±20 cm) Fine (20±40 cm) Coarse (0±20 cm) Coarse (20±40 cm) (0±20 cm) (20±40 cm) Total

Forest (nˆ10)

SD

Annual crops (nˆ5)

SD

148.8 2.4 2.8 4.9 0.0 158.9

80.0 4.0 2.0 2.1 ± 82.0

0.0 28.6 2.4 5.0 1.0 37.0

± 9.0 3.9 1.6 1.4 11.7

37.5 16.5 3.5 10.5 0.0 68.0

16.9 12.7 5.1 4.2 ± 15.5

0.0 6.5 0.9 0.2 0.0 7.6

± 12.7 0.3 0.1 ± 12.7

2.8 2.0 1.6 2.0 8.5 22.8 9.7 32.4 199.8

1.2 3.0 0.3 1.4 4.9 4.5 3.6 7.4 80.9

2.6 0.4 1.8 0.3 5.1 26.1 7.9 34.0 76.1

2.6 0.3 0.07 0.01 2.8 15.4 4.6 19.2 27.6

1.5 0.4 0.7 0.3 2.9 24.1 10.2 34.3 105.3

1.5 0.4 0.5 0.3 2.4 7.2 3.3 7.7 20.3

0.8 0.2 0.0 0.0 1.1 15.0 4.5 19.6 28.3

0.7 0.2 ± ± 0.7 6.5 3.1 7.9 15.0

Source: adapted from C. Castilla, unpublished.

Fallow (nˆ10)

SD

Pasture (nˆ10)

SD

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S. Fujisaka et al. / Agriculture, Ecosystems and Environment 69 (1998) 17±26

Fig. 1. Carbon stocks under four different land uses in Theobroma, Rondonia, Brazil. Different letters indicate significantly different values (P<0.05, Duncan test).

unchanged from forest to annual crops to fallow, but dropped signi®cantly with conversion to pasture. Total aboveground C decreased signi®cantly with conversion to annual cropping, increased with fallowing (not signi®cantly) and dropped with conversion to pasture. Total root C showed a linear decline from forest to pasture (Fig. 1). 3.2. Plant species A total of 326 plant species were encountered in the forest, 180 in cropped ®elds, 227 in fallowed plots and 86 in pastures. Of the original 326 plant species found in the forest, however, only 92 (28%) were found in the annual crop ®elds, 106 (33%) in the fallowed ®elds and only 20 (6%) in the pastures. `New' pioneer and

weedy species invaded lands converted from forest to other uses: 88 species found in the cropped ®elds were not found in the forest; similarly 121 species found in the fallow and 66 species found in the pasture were not found in the forest (Fig. 2). The 196 tree species encountered in the forest accounted for 60% of all forest plant species. Numbers of these same tree species declined to 46 in the cropped ®elds, 55 in fallows and dropped to only 6 in the pastures. At the same time, 37 non-forest tree species were found in the cropped ®elds, 54 in fallows and, ®nally, 13 in the pastures. The forest herbaceous plant species declined from 33 in the forest to 2 in the pasture and the forest shrubs from 19 in the forest to 1 in the pasture. Non-forest herbaceous species reached 27 in the cropped ®elds, 26 in fallows and 34 in the

S. Fujisaka et al. / Agriculture, Ecosystems and Environment 69 (1998) 17±26

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Fig. 2. Numbers of species by plant type and land use, Theobroma, Rondonia, Brazil; (a) species found in the forest, continuing in other land uses; (b) species appearing in other land uses; (c) total number of species.

pastures; while non-forest shrubs included 11 in the cropped ®elds, 15 in fallows and 10 in the pasture. Finally, some forest species were lost ± some 20% of all forest plant species are found only in the forest: 40

(20%) of the forest trees, 7 vines, 7 herbaceous plants, 3 palms and 3 shrubs or a total of 60 species appeared only in the forest and not in any of the other land uses (Fig. 2). The most common trees found in the pastures

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Table 2 Farmer-settler land use, Theobroma, Rondonia, Brazil, 1993±94. Means and standard deviations for a sample of nˆ74 Land use

Mean area (ha)

SD

% of total

Forest Cleared: Pasture Fallow Annual crops Perennials Total Total

41

43.3

54

20 6 5 4 35 76

28.4 7.3 3.8 4.9 29.6 59.6

26 8 7 5 46 100

% of cleared ± 57 17 14 12

± 100

were the pioneers Eugenia sp., Inga buorgoni and Vismia guianensis. 3.3. Land use At the time of their 1993±94 crop season, farmer settlers in Theobroma had parcels of a mean 76 ha, of which 41 ha (54%) was still forested and 35 ha (46%) had been cleared. Of the cleared area, 20 ha (57% of cleared area) were in pasture, 6 ha (17%) in fallow, 5 ha (14%) sown to annual crops and about 4 ha (12%) in perennial crops, largely coffee (Table 2). 4. Discussion and conclusions The effects of deforestation include emission of greenhouse gases, loss of biodiversity, loss of cultural diversity, degradation of land and water resources and possible impacts on the regional climate. The research focused on CO2 emissions and loss of plant diversity. Assuming that most of the differences in C measured in forest compared to cropped ®elds, fallows and pastures represent balances between C lost to the atmosphere and amounts stored in these non-forest land uses, total C emissions over time in Theobroma can be estimated with the addition of data on the municipality-wide deforestation. For a total area of 216 500 ha, deforestation increased from less than 2000 ha (0.8%) in 1973, to almost 7000 ha (3%) in 1978, 55 000 ha (25%) in 1987 and 93 000 ha (43%) in 1993 (Pereira da Silva and Pedroso Guimaraes, 1996). Notably, farmer interview data (46% of land cleared

by the 1993±94 cropping season) ®t reasonably well with the reported data (43% cleared by 1993). If we consider the proportions of the current land uses in Theobroma (Table 2) and their respective C stocks (Table 1), we can estimate the amount of C lost through deforestation: EC ˆ …total area  Cforest † ÿ …areaLU  CLU † where EC is the total C emission in tons for the municipality, Cforest is the C stock of forest (t haÿ1), areaLU are the areas (in ha) currently in use as forest, annual crop, fallow, pasture and perennials respectively and CLU are the C stocks of these land-use types (t haÿ1). The C stock value used for the perennials (42 t haÿ1) is based on the samples from coffee plantations using the same methodology (C. Castilla, unpublished data). The amount of C thus released would be about 150 t haÿ1 and the amount stored would average about 50 t haÿ1. As such, by 1993 the 93 211 ha deforested would represent C emissions of almost 14 million tons over 20 years for the municipality of Theobroma alone (Table 3 and Fig. 3). The emission of 14 million tons of C over 20 years, or 700 000 t yrÿ1, is equivalent to the C emission over the same time of 325 000 automobiles (assuming a car travels 32 000 km yrÿ1 and consumes an average of 10 km lÿ1 or 3330 l yrÿ1 of gasoline, which are typical ®gures for many urban areas in the northern hemisphere). The calculation is based on gasoline weight of 0.77 kg lÿ1, C in gasoline is 84.2%, such that each automobile burns 2.6 t yrÿ1 of gasoline and exhausts 2.15 t yrÿ1 of C. As expected, variability of the C estimates was large (see standard deviations in Table 1). The source of this Table 3 Deforestation (ha)a and estimated tons of carbon released into the atmosphereb, Theobroma, Rondonia, Brazil, 1973±1993 Year

Area deforested

% of total

C released (million t)

1973 1978 1987 1993

1690 6810 54600 93200

0.8 3.0 25.0 43.0

0.25 1.0 8.2 14.0

a

Source: Pereira da Silva and Pedroso Guimaraes (1996). Assumes C stocks and proportions of areas per land use as estimated in this study. b

S. Fujisaka et al. / Agriculture, Ecosystems and Environment 69 (1998) 17±26

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Fig. 3. Area deforested and estimated carbon emissions, Theobroma, Rondonia, Brazil, 1973±1993.

is probably a combination of the natural variation in site conditions and differences in land management. While this does indicate that more samples are required for better estimates of absolute quantities of C, the differences between land uses are large enough to show the major changes upon deforestation. The calculations of municipality-wide changes in C stocks assume that the transects sampled are representative for all forests, pastures, cropped ®elds and fallows, respectively. In the case of pastures and fallows, C stock may change a little with age. The 10-year-old pastures sampled are considered to have passed the initial phase in which organic matter residues of the forest diminish and organic matter from the grasses is added to the soil. In fallows, especially aboveground, changes occur at much faster rates. However, fallows make up a small (and decreasing) part of the landscape and the ones sampled are considered representative for the range of fallows currently encountered in Theobroma. Therefore the sampling of more pastures and fallows of different ages would probably have a minor effect on the overall estimate of the deforestation impact on C stocks. By limiting sampling to 40 cm depth, undoubtedly a complete accounting of soil C was not made. Although most root activity occurs near the soil surface and root abundance declines sharply with depth, roots are found down to great depth, especially if rainfall is seasonal (Nepstad et al., 1994). Deep sampling will

yield higher estimates of belowground biomass and soil organic matter and this will tend to increase the differences between forest and non-forest agroecosystems because trees are expected to root deeper. Unlike the loss of C upon forest conversion, loss of plant species is not proportional to the size of the area converted. Most genetic diversity can be conserved in a region where a large share of the land is deforested because many species will survive in the remaining fragments and secondary forests (if the latter are left as fallow for long enough periods). Even a 10% loss of species richness, however, constitutes an important number of species in absolute terms. The process of extinction continues for several years after forest clearing because the regeneration requirements of some of the species can no longer be met. Species disappearing ®rst are primary forest specialists that occur at low population densities. The remaining ¯ora both is less species-rich and loses its uniqueness, with more pioneer or early secondary species, which usually have broad geographical distributions. In a study in another Brazilian forest colony, far more useful species (i.e., for human use) were found in the primary forest rather than in other land uses (Fujisaka et al., 1998). As expected, both C stocks and numbers of plant species decrease as forest is converted to crops or pasture and both show recuperation during fallowing (Fig. 4). The trends suggest that near-forest levels of

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Fig. 4. Total carbon stocks and number of plant species by land use, Theobroma, Rondonia, Brazil.

stored C and diversity of plant species could be restored in systems in which lands are allowed to recover after cropping with longer term fallows (unlike the almost exclusively `young' fallows encountered in Theobroma), but that conversion to pasture represents a high-impact land use. Alternatives that would decrease rates of deforestation, conserve biodiversity and/or decrease C emissions range from policy changes to improved local resource management strategies. Brazilian policies over the last decade have been less supportive of road building, colonization and incentives supporting cattle ranching (and land speculation) in the Amazon, leading (apparently) to declines in rates of deforestation (Moran et al., 1994). A Brazilian law requiring settlers to maintain 50% of their lands in forest has not yet been effectively enforced and few prospects exist for slowing deforestation via such policies. Because assets like C and biodiversity are of global importance, the sequestration of C and conservation of ¯ora and fauna can be seen as environmental services to the world (Turner, 1993) best ®nanced by the world community. Global environmental markets (Pearce and Moran, 1994) are beginning to develop: power companies in developed countries invest in forestry to compensate for CO2 emitted by their plants; and pharmaceutical enterprises pay for the rights to use species as a source for new products. Such economic

transactions require sound valuation of these environmental resources and services. Although several new valuation methods exist (Kramer et al., 1992), few have been tested in developing countries or tropical forest environments and less so for non-market goods. A recent contingent valuation study in the Peruvian Amazon (Smith et al., 1997) suggests that trading of C through a global market could include incentives at the local level for more sustainable land use (with more environmental bene®ts). Improved local resource management options, which may be able to sequester more C and better maintain forest biodiversity, include agroforestry and silvopastoral systems. Agroforestry systems include mixtures of perennial crops, indigenous fruit and nuts with market value and (usually fast-growing) trees for timber production. Agroforests in Rondonia combining Brazil nut (Berthollethia excelsa), cupuacu (Theobroma grandi¯orum) and peach palm (Bactris gasipaes) were estimated to recover up to 120 t haÿ1 C in 40 years (Gazal Yared et al., 1993). Silvopastoral systems combining palms or fruit trees with grazed herbaceous legumes are similarly expected to sequester more C than Brachiaria monocultures. Because incentives continue to support cattle ranching and conversion of forest to pasture, ways to promote improved pasture management need to be developed. Pastures, often considered the least desir-

S. Fujisaka et al. / Agriculture, Ecosystems and Environment 69 (1998) 17±26

able land use after forest clearing, have now been recognized as a viable option in deforested regions (Fearnside, 1996). Initial decreases of soil C of the order of 20% were reported following conversion of forest to pasture for at least the ®rst few years (Chone et al., 1991). In the present study, greater decreases of up to 40% were recorded (Table 1). However, longer term pastures in the Amazon have been reported to contain similar or more soil C than forested soils at a depth of 20±30 cm (Chone et al., 1991; Feigl et al., 1995; Moraes et al., 1996). Growing evidence indicates that well-managed, improved, deep-rooting pasture species can substantially increase the organic matter content of soils (Serrao et al., 1979; Liegel, 1992; Fisher et al., 1994). Fisher et al. (1994), reported sequestration of 3±14 t C haÿ1 yrÿ1 by grasses introduced from Africa; perhaps more importantly, over 75% of this C accumulated below 20 cm ± that is, it is probably less prone to oxidation and hence loss during subsequent changes in land use. Cerri et al. (1994), estimated that losses of soil C in Manaus over 8 years to 20 cm depth were 40 t haÿ1 and that a Brachiaria humidicola pasture could return 118 t haÿ1 to the soil over the same time period. However, soil C to 40 cm in Theobroma, far from increasing, fell by 13 t haÿ1 in pastures compared to forest. Cerri et al. (1994), also ignored the loss of C from the decay of tree roots, assuming that root turnover was the same in both pastures and forest. This is uncertain in that the present data showed that as much as 8.5 t haÿ1 of C in root material in the 0±40 cm layer were lost over the 10 years after conversion from forest. The key may be that the pastures in Manaus were described as `unfertilized [but] well-managed'. Clearly, management of pastures is a critical issue: many of the pastures from which the present data came were in various states of degradation, evidenced in part by their weedy nature. If introduced pastures are managed productively and are not burned it also seems likely that the soil C equilibrium level can be at least maintained and possibly raised substantially (Greenland, 1995). Although the amounts of C that could be accumulated under pasture do not totally compensate for losses via deforestation, efforts to improve pastures and associations of trees and pastures are well justi®ed from an economic and environmental perspective (Sombroek et al., 1993; Thomas et al., 1995; Fisher et al., 1998).

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To the extent that the municipality of Theobroma serves as an example, the impacts of converting forest to other uses (especially pasture) are substantial in terms of C emissions and plant species loss. Fortunately, however, much of the plant diversity may be maintained in remaining forests and in secondary forests (although research is needed on losses of forest specialists occurring at low densities). Also positively, relatively higher amounts of C can be sequestered in pastures with the use of more intensively managed, introduced grasses, possibly combined with forage legumes and/or trees; and agroforestry systems may be able to decrease the net C losses among alternative agricultural land uses. Finally, the future may bring better valuation of environmental resources and services, followed by increases in either payments by the global community or trading in global markets for these resources and services, which are of greater global than local bene®t. References Anderson, A., 1990. Deforestation in Amazonia: dynamics, causes and alternatives. in: Anderson A. (Ed.), Alternatives to Deforestation: Steps toward Sustainable Use of the Amazon Rainforest. Colombia University Press, New York. Anderson, J.M., Ingram, J.S.I., 1993. Tropical Soil Biology and Fertility: A Handbook of Methods. Cab International, UK. Brown, S., Gillespie, A.J.R., Lugo, A.E., 1989. Biomass estimation methods for tropical forests with applications to forest inventory data. For. Sci. 35, 881±902. Castilla, C., 1992. Carbon dynamics in managed tropical pastures: The effect of stocking rate on soil properties and on above- and below-ground carbon inputs. Ph.D. thesis, North Carolina State University, Raleigh, NC, Unpublished. Cerri, C.C., Bernoux, M., Blair, G.J., 1994. Carbon pools and fluxes in Brazilian natural and agricultural systems and the implications for the global CO2 balance. Transactions of the 15th World Congress of Soil Science, Acapulco, Mexico, 10± 16 July 1994, Volume 5a, pp. 399±406. Chone, T., Andreux, F., Correa, J.C., Volkoff, B., Cerri, C.C., 1991. Changes in organic matter in an oxisol from the central Amazonian forest during eight years as pasture, determined by 13 C composition. in: Berthelin, J., (Ed.), Diversity of Environmental Biogeochemistry. Elsevier, New York, pp. 307±405. Duxbury, J.M., Harper, L.A., Mosier, A.R., 1993. Contributions of agroecosystems to global climate change. in: Agricultural Ecosystem Effects on Trace Gases and Global Climate Change. ASA Special Publication 55, Madison, WI, pp. 1±19. Fearnside, P.M., 1993. Deforestation in Brazilian Amazonia: The effect of population and land tenure. Ambio 22(8), 537±545.

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