SCOPE 21 -The Major Biogeochemical Cycles and Their Interactions 

7

The Effects of Deforestation on Air, Soil, and Water

P. M. VITOUSEK
 
Abstract
7.1 Introduction
7.2 What is Deforestation?
7.3 The Effects of Deforestation
7.3.1 Effects on Temperature, Water and Erosion
7.3.2 Effects on Carbon
7.3.3 Effects on Nitrogen
7.3.4 Effects on Phosphorus 
7.3.5 Effects on Sulphur
7.3.6 Reforestation
7.4 Rates of Response to Deforestation
7.5 Interactions of C, N, P, and S
7.5.1 Element Supply
7.5.2 Ratios of Element Availability
7.5.3 Specific Effects
7.5.4 Interactions Off of the Site
7.6 Conclusions
Acknowledgements
References 

ABSTRACT

Deforestation causes increased losses of carbon, nitrogen, phosphorus, and sulphur from terrestrial ecosystems. Where deforestation is followed by conversion to other than forest land uses, the effects of deforestation are magnified. The major causes of organic carbon losses are harvest, burning of forest residue, accelerated decomposition, decreased production of wood and roots, and erosion. Nitrogen and sulphur are lost by the same pathways, and additionally by leaching of nitrate and sulphate to stream-water and ground-water and by the anaerobic production of N- and S-containing gases. Phosphorus is lost primarily through harvest and erosion. More than half of the C and N and somewhat less P and S can be lost in sites where forest land is converted to other uses.

Losses of these elements following deforestation are most rapid in sites with high decomposition rates, especially in the tropics and on fertile soils. The interactions of the C, N, P, and S cycles affect losses of any element through nutrient limitations to biological transformations, ratios of element availability, which cause either biological mobilization or immobilization, and anion/anion interactions in the soil solution.

7.1 INTRODUCTION 

Deforestation has provided a major focus for process-level (cf. Hesselman, 1917), budgetary (Cole and Gessel, 1965; Likens et al., 1970), and modelling (Aber et al., 1979) studies of ecosystem-level nutrient cycling and flux. The cycles of carbon, nitrogen, phosphorus, and sulphur have received particular attention, in part because bicarbonate, organic anions, sulphate, and (in disturbed forests or agricultural sites) nitrate are the most important anions in the soil solution (Johnson and Cole, 1980). As such, their concentrations and mobilities control the losses of cations as well as anions to stream-water and ground-water. Additionally, the erosion of organic matter and phosphorus to streams and lakes contributes to aquatic production and eutrophication. More recently, the need to evaluate forest ecosystems as net sources or sinks for atmospheric CO2 (Woodwell et al., 1978; Broecker et al., 1979), oxides of nitrogen (Crutzen and Ehhalt, 1977), and sulphur gases (Eaton et al., 1978; Rice et al., 1981) has become apparent.

In this paper, I examine the major effects of deforestation on water flux through ecosystems and on carbon, nitrogen, phosphorus, and sulphur transformations and losses. This examination will necessarily be a rather general survey. I evaluate some of the processes causing major differences between forest types in the pattern of their responses to deforestation. Finally, I examine how interactions among the carbon, nitrogen, phosphorus, and sulphur cycles control large scale responses to deforestation.

7.2 WHAT IS DEFORESTATION?

A wide range of forest land use practices can be termed `deforestation.' Two important practices that strongly differ in intensity are forest clear-cutting and forest land conversion. In clear-cutting all tree stems over some minimum diameter are cut and stem wood is removed, and then the site is either replanted with tree seedlings, or natural revegetation is allowed to occur. The major variants of this practice include whole-tree harvest and complete forest removal (in which other parts of the trees in addition to just stem wood are removed) and slash burning (in which the debris remaining on site after logging is burned).

Forest land conversion involves the removal of trees (as above) followed by the conversion of the land to agriculture, pasture, development, or some other non-forest use. The intensive harvesting of forest lands for fuel-wood in much of the world fits into this category. The effects of forest land conversion (especially to agricultural use) are generally more severe than those of clear-cutting. The shifting cultivation system which is widely practiced in the tropics fits between these extremes. Essentially, it involves the temporary conversion of forest to agriculture, followed by the natural re-establishment of forest cover.

7.3 THE EFFECTS OF DEFORESTATION 

7.3.1 Effects on Temperature, Water and Erosion

The conversion of forest land to other uses decreases above- and below-ground biomass on a site. Shading of the soil surface is thus decreased, and soil temperature increases (Stone, 1973; Harcombe, 1977). Additionally, plant uptake and transpiration of soil water and mineral nutrient uptake are usually decreased for at least 2-3 years even in sites that rapidly regrow to forests (Marks and Bormann, 1972; Gholz, 1980; Boring et al., 1981). With reduced evapotranspiration, water flux through the soil is increased (Figure 7.1), and so losses of nutrients through leaching to ground-water and stream-water can be increased.

Figure 7.1 The water cycle in an undisturbed forest (above) and a deforested site (below). The width of the arrows is proportional to the amount of water following each path; the system represented is a relatively wet forest

While the soil of a deforested site is thus on the average warmer and wetter than a forest soil, the extremes in temperature and moisture levels are also increased. When subjected to direct solar radiation, the upper few cm of forest floor or mineral soil can dry to moisture contents well below those in undisturbed forest (Likens et al., 1978). Similarly, re-radiation from bare surface soil on clear nights can cause ground-level frosts in midsummer in boreal forests (C. O. Tamm, personal communication). Surface soils are thus subjected to extremes of heating and cooling and wetting and drying in deforested sites.

A consequence of these changes in temperature and moisture is an increase in rates of decomposition and nutrient mineralization in deforested sites (Dominski, 1971; Stone, 1973; Stone et al., 1979). The forest floor decomposes rapidly (Covington, 1976; Bormann and Likens, 1979), and without forest regeneration will eventually disappear. The combination of increased decomposition (which consumes oxygen) and wetter soils (which slow oxygen diffusion) may also increase the occurrence of anaerobic microsites within the soil.

In sites with even a slight slope, another consequence of deforestation is an increase in erosion and particulate transport. The delivery of soil to stream courses is increased because: (i) the wetter soil after deforestation is both heavier and less cohesive, and thus more subject to both soil creep and more rapid slope failure; (ii) the decay of tree roots reduces the cohesiveness of the soil and increases both soil creep and the probability of debris avalanches (Swanson et al., 1981) ; and (iii) the decrease and eventual disappearance of the forest floor alters the infiltration rate of the soil, allows raindrop impact on the mineral soil, and can thus increase surface run-off. Once material reaches streams, the increased stream flows in deforested sites are able to transport more and larger particulates downstream (Bormann et al., 1974). The relationship between stream flow and particulate transport often has an increasing exponential form, so the capacity to transport particles increases more rapidly than increases in peak stream flows. Where deforestation leads to agricultural land use, higher rates of erosion will be maintained indefinitely (Ritchie et al., 1974; Rapp, 1975).

7.3.2 Effects on Carbon

The carbon cycle in a natural forest and a deforested site are contrasted in Figure 7.2. The most important consequence of deforestation is a substantial decrease (well over 50%) in total organic carbon (above and below ground, living and dead) in a deforested site. This decrease has several important causes:

  1. The removal (by harvest) of organic carbon for wood or paper products. Most of our knowledge about this flux is based on national statistics on the rate of carbon removal in merchantable stems, and the information appears to be relatively good in many of the developed countries (Armentano and Ralston, 1980). Less information is available on the longer term fate of the organic carbon harvested. The amount of carbon in each class of forest products (i.e., firewood, building material, paper products) and the mean residence time before oxidation to CO2 of the material in each class is essential to an evaluation of the importance of forest harvesting in the global CO2 budget (Armentano and Hett, 1979). It is likely that, on the average, harvested material has a shorter turn-over time than it would have had it not been harvested.
  2. The combustion of residue left after deforestation. Where fire is used for land clearing and conversion or as a silvicultural practice, a large amount of organic carbon is rapidly released as C02. Substantial CO emissions also occur during such fires (Crutzen et al., 1979; Crutzen, Chapter 3 this volume), and an unknown but probably substantial amount of recalcitrant elemental carbon (charcoal) is produced (Seiler and Crutzen, 1980).
  3. Accelerated decomposition. After deforestation, the warmer, wetter soil conditions accelerate the decomposition of residues left from land clearing, the forest floor (Covington, 1976; Bormann and Likens, 1979), and soil organic carbon. Even with immediate revegetation, considerable losses of forest floor organic carbon (up to 60%) can occur early in succession (Covington, 1976). Most of this carbon is probably lost as CO2, although some may be incorporated into the mineral soil (at least temporarily).
  4. Lack of replacement of organic carbon. The amount of organic carbon in the soil declines under continuous cultivation (Haas et al., 1957). This decline is partially due to accelerated decomposition of the more labile fraction of native soil organic matter, but another important cause is the smaller amount and greater lability of organic matter added to the soil by crops as opposed to forests.
  5. Erosion of organic carbon (often complexed with clay particles). The organic carbon removed by erosion may be redistributed to lower-lying areas within the terrestrial system (McCallan et al., 1980), in which case its turnover is probably little affected. Alternatively, it may be transferred to lacustrine or marine sediments, where its turn-over time is probably increased.

Figure 7.2 The carbon cycle in an undisturbed forest and in a deforested site shortly (23 years) after deforestation

Other losses of carbon from deforested systems include CH4 flux (probably somewhat increased in the warmer, wetter soils of deforested sites), the leaching of dissolved organic carbon to stream-water and ground-water and the leaching of carbonate species. Bicarbonate is the most abundant anion in river-water (Garrels and MacKenzie, 1971) and in many soils (Johnson et al., 1977), and bicarbonate concentrations and losses can be increased by deforestation (Cole et al., 1975; Snyder et al., 1975). While this increase could significantly increase cation mobility through soils (Johnson and Cole, 1980), it represents a relatively minor flux of carbon. Losses of dissolved organic carbon can also be increased by deforestation (Sollins and McCorison, 1981), but this is also likely to be a minor flux.

Overall, the major effect of deforestation on carbon cycling is the large decrease in the organic carbon pool on the deforested site. Much of this decrease makes a net contribution (from that site) to the atmosphere and subsequent pools.

This conclusion appears straightforward, and it has been clearly stated previously (Delcourt and Harris, 1980). It is important enough, however, and it has been obscured enough (cf. Brown et al., 1980) to warrant a more extended development. Old-growth forest ecosystems, when averaged over sufficient space or time to include natural disturbances (White, 1979), make little or no net contribution to atmospheric CO2. They take up in photosynthesis about as much as they lose in plant and decomposer respiration, although any small patch of forest is likely to be either a net source or a sink. The same is true (without the spatial heterogeneity) of agricultural fields or pastures on land that has been in agriculture for a long time, as long as the plant or animal harvest is consumed reasonably rapidly. The transition between the two, though, involves the loss of most of the organic carbon in living biomass, most of the forest floor, and much of the organic carbon in mineral soil. Some of this organic carbon may go into wood used in construction, some may go into lacustrine or marine sediments, and an unknown amount of partially oxidized organic carbon is retained with the soil. The reminder is released as CO2 and makes a net contribution to atmospheric CO2.

The same argument applies to shifting cultivation systems and even to commercial forestry. If an area of primary rain-forest is converted to a shifting cultivation mosaic, the total amount of organic carbon in the area at equilibrium is decreased. At this equilibrium the landscape as a whole may be neither a source nor a sink, and any fallow patch (of regrowing forest) may be much more of a sink than it was when it was in primary forest (Lugo, 1980). The difference in equilibrium (or mean) organic carbon between the primary forest and the shifting cultivation system represents a loss of carbon to the landscape, however. Unless this amount is tied up in building materials, charcoal, or aquatic sediments, it represents a loss of CO2 to the atmosphere. Similarly, where forest harvests are more frequent than natural disturbances, the mean organic matter content of forest land is decreased in the long run by forest harvesting.

Viewed in this way, the tropical forests must now be a source for CO2 under almost any set of assumptions. Temperate forests may presently be a sink for atmospheric CO2 (their total organic carbon is apparently increasing (Armentano and Ralston, 1980). Viewed over the longer term, however, the North American forests have lost organic carbon over the time since European settlement.

7.3.3 Effects on Nitrogen

The nitrogen cycle in an undisturbed forest and a deforested site are contrasted in Figure 7.3. Nitrogen cycling in undisturbed forests is relatively closed; the internal soilplantmicro-organism cycle in most forests involves 1030 times more nitrogen than annual nitrogen inputs or outputs (Rosswall, 1976), and annual inputs in aggrading forests generally exceed annual outputs (Likens et al., 1977). When this soilplantmicro-organism cycle is disrupted by deforestation, however, a large fraction of the organic nitrogen present in an ecosystem can be lost. Deforestation interrupts the cycle both by preventing plant uptake of nitrogen and by increasing the rate of nitrogen mineralization (Vitousek and Melillo, 1979).

The five major causes of organic carbon losses are also important in causing nitrogen losses from deforested sites. Additionally, the leaching of nitrogen in the form of nitrate to stream-water and ground-water can be a significant pathway of nitrogen loss. Nitrate is produced in large amounts in some disturbed forests (Likens et al., 1970; Vitousek and Melillo, 1979), and it is often the most important anion in agricultural soils  (Nye and Tinker, 1977). The nitrate anion is relatively mobile in the soil solution, and nitrate and associated cations are thus easily leached through most soils. Some tropical soils could retain large amounts of nitrate by anion adsorption, however (Kinjo and Pratt, 1971; Bartholomew, 1977).

Figure 7.3 The nitrogen cycle in an undisturbed forest and a deforested site. The system represented in a relatively fertile site before and 23 years after deforestation. Dashed arrows represent possible alternative pathways for losses of nitrogen from the site 

There is considerable uncertainty over the most important pathways of nitrogen losses in deforested ecosystems, especially in some tropical sites for which losses of as much as 1400 kg N/ha from the upper 30 cm of soil in the 2 years following clearing and burning have been estimated (Nye and Greenland, 1964; Bartholomew, 1977). The nitrogen mineralized following deforestation may: (i) be held within the site (temporarily or permanently) by microbial immobilization, clay fixation, and other processes (Vitousek et al., 1979); (ii) be lost to the atmosphere through ammonia volatilization, N2O production during nitrification (Bremner and Blackmer, 1978), or denitrification to N2O or N2 (Firestone et al., 1980); or (iii) be leached from the site as dissolved organic nitrogen or nitrate. N2O production during nitrification is probably a small proportion of nitrification (Bremner and Blackmer, 1978), but the relative magnitude of losses by denitrification and nitrate leaching in a range of ecosystems is largely unknown. The presence of warmer, wetter soils in deforested sites should increase rates of both nitrification and denitrification.

In general, deforestation causes a decrease in the amount of organic nitrogen on a site. Losses by harvest, erosion, volatilization, and leaching are all important in at least some sites. As discussed below, nitrate production and loss (to streams and to the atmosphere) is most rapid in relatively fertile sites. If forest regeneration is prevented, however, nitrate production will eventually occur on almost any site (Wiklander, 1981; Vitousek et al., 1982). The major environmental concerns are: (i) that plant growth on deforested sites will be slowed because of the nitrogen lost following deforestation; (ii) that the NOx released during burning or the N2O produced during nitrification or denitrification will adversely affect atmospheric chemistry (Crutzen, chapter 3, this volume); and (iii) that nitrate leached to stream-water and ground-water could affect downstream ecosystems and even human health (Magee, 1977).

7.3.4 Effects on Phosphorus

The effects of deforestation on phosphorus are summarized in Figure 7.4. The overall effects (Haas et al., 1961) are quite different from those on carbon and nitrogen for several reasons. Phosphorus does not ordinarily undergo oxidationreduction reactions in terrestrial ecosystems; it is present in both plants and soil as phosphate. The phosphate ion is relatively immobile in soilsit is immobilized by micro-organisms, precipitated with calcium in circumneutral and basic soils, and precipitated with iron and aluminium in acid soils. Accordingly, phosphorus losses upon deforestation occur primarily in forest product removal and erosion, although substantial leaching losses have been observed after deforestation in sites on quartz sand parent material (Herrera, personal communication).

Phosphorus cycling in deforested sites may also be different from carbon and nitrogen in that decomposition and gross nitrogen mineralization are rather closely linked, but phosphorus mineralization is carried out in large part by extracellular enzymes produced by phosphorous-requiring organisms (McGill and Cole, 1981).

Figure 7.4 The phosphorus cycle in an undisturbed forest and a deforested site

Most importantly, phosphorus has no significant gaseous form. Once it is either depleted or converted to biologically unavailable forms on a site, phosphorus inputs from atmospheric deposition are generally too low to maintain high levels of available phosphorus. Phosphorus is added to the biological cycle through rock weathering, but very old, deeply leached soils become depleted in weatherable phosphorus (Walker and Syers, 1976).

The most significant impact of deforestation on phosphorus cycling is an increase in particulate phosphorus transport to streams and lakes, where it can drive increased production and eutrophication (Schindler, 1977). The reduction of soil phosphorus content that accompanies deforestation may be important in reducing the fertility of old soils, particularly in the tropics.

Figure 7.5 The sulphur cycle in an undisturbed forest and a deforested site. The fluxes to the atmosphere represented here are largely speculative

7.3.5 Effects on Sulphur

Some of the possible effects of deforestation on sulphur cycling and losses are summarized in Figure 7.5. Perhaps because sulphur cycling is rarely limiting to forest growth under natural conditions, sulphur cycling has received relatively less study than the cycles of carbon, nigrogen, and phosphorus.

The same processes cause losses of sulphur in a deforested site as cause losses of nitrogen (Bettany et al., 1980). Harvest and fire remove sulphur rapidly, and the erosion of organic sulphur-containing compounds can be significant. Sulphate leaching is perhaps the most important form of sulphur loss, but soil and stream-water sulphate concentrations decline precipitously, in at least some deforested sites. Several mechanisms could cause this decrease (Bormann and Likens, 1979).

The sulphur cycle in a deforested site is different from that of nitrogen in other ways as well. Sulphate is more tightly held by anion adsorption, particularly by acid sesquioxides in older soils (Couto et al., 1979; Johnson et al., 1980). Additionally, only reduced organic sulphur (carbon-bonded sulphur) is mineralized with organic carbon and nitrogen at accelerated rates following deforestation. The mineralization of ester-sulphates, which are the major form of organic sulphur in many sites, may be uncoupled from decomposition in the same way as phosphate mineralization is (McGill and Cole, 1981).

The effects of deforestation on the net exchange of sulphur with the atmosphere is difficult to quantity. A large fraction of atmospheric inputs of sulphur to terrestrial ecosystems come in the form of SO2 absorbed by foliage and to a lesser extent sulphate-containing aerosols impacted on vegetation (Galloway and Whelpdale, 1980). Deforestation would reduce these inputs. At the same time, more rapid decomposition rates and warmer, wetter soils could increase the frequency of anaerobic conditions and thus increase sulphate reduction and H2S volatilization, and possibly the production of volatile organic sulphur compounds (Trudinger, 1979). If deforestation caused the eutrophication of downstream aquatic systems, sulphate reduction could also be enhanced in the sediments of those systems.

The overall effects of deforestation on sulphur cycling thus include a decrease in the sulphur pool size in soils and vegetation, usually without serious consequences for the fertility of the site. The delivery of sulphate and other soluble sulphur compounds to aquatic systems is relatively little affected by deforestation. However, deforestation does cause a decrease in sulphur uptake from the atmosphere and probably an increase in gaseous losses of sulphur to the atmosphere.

7.3.6 Reforestation

Reforestation of long term agricultural and pasture land essentially reverses the patterns discussed above for all of the elements. The total pool of each element in organic matter in the site increases, and losses to downstream ecosystems and to the atmosphere decrease (Vitousek and Reiners 1975; Bormann and Likens, 1979). The rate of reforestation varies widely in different sites (Tamm et al., 1974), and the overall rate of accumulation of all four elements can be controlled by limited supplies of any one of the elements (as discussed below).

The source of carbon for reforestation is atmospheric CO2. Nitrogen comes from fixed nitrogen in the atmosphere and from biological nitrogen fixation. Nitrogen fixation often peaks in early or mid-succession, especially in primary succession or on drastically disturbed sites (Gorham et al., 1979). Sulphur can come from precipitation, absorption or interception of atmospheric sulphur by regrowing vegetation, and sulphur minerals on the site. Phosphorus comes almost entirely from minerals on the site.

7.4 RATES OF RESPONSE TO DEFORESTATION

Given enough time without forest regeneration, most ecosystems follow the patterns outlined above. They do so, however, at vastly different rates in different sites. The patterns discussed above thus provide useful statements of direction, but they are virtually useless for evaluating the timing of response to deforestation in any real system.

Studies of nitrate losses from forest ecosystems following clear-cutting and other disturbances provide a clear illustration of differences in timing of response. Some sites have large, rapid losses of nitrate to stream-water and ground-water following cutting; others respond little if at all to an essentially similar treatment (Vitousek and Melillo, 1979). In experimental studies with trenched plots in which plant regrowth was prevented, elevated nitrate losses below the rooting zone were eventually observed in 18 out of 19 sites (Vitousek et al., 1982). Some responded very rapidly, however, while others had delays of up to 3 years before elevated nitrate losses to below the rooting zone occurred. Sites which never show elevated nitrate losses may yet lose nitrate; delays of more than 10 years have been observed in commercial clear-cuts in Sweden (Wiklander, 1981). Similarly, large amounts of carbon and nitrogen disappear from surface soils immediately after land clearing and cultivation in West Africa (Nye and Greenland, 1964), but it takes decades for changes of similar magnitude to occur following cultivation in the mid-western United States (Haas et al., 1957).

What are the major causes of differences among sites in the rate of response to deforestation? Any answer to this question is necessarily partly speculativesome of the important differences have not been studied directly. The differences that are likely to prove important, however, include the following.

  1. The rate of decomposition. The rate of leaf litter decomposition in undisturbed forests can be predicted reasonably well from calculated actual evapotranspiration (Meentemeyer, 1978). Decomposition rates are high (the exponential decay constant K 4) in tropical forests (Nye, 1961; BernhardReversat, 1977) and much lower (K 0.25) in some boreal forests (Berg and Staff, 1980), and the breakdown of labile soil organic matter should also be more rapid in tropical sites. Litter inputs to the forest floor vary less among forests from different latitudes than do decomposition rates, so more dead organic matter is accumulated in boreal sites than in tropical sites (Jordan, 1971; Schlesinger, 1977). The much more rapid decomposition rates and relatively small organic matter pools in tropical forests cause much more rapid losses of a substantial portion of a site's organic carbon after deforestation (Nye and Greenland, 1964). The loss of elements associated with organic carbon (nitrogen and carbon-bonded sulphur) may lag behind carbon losses, but they too should be lost more rapidly where decomposition rates are higher.

  2. Site fertility. Both the maximum rate of nitrate losses from experimental trenched plots and length of delays in those losses are strongly correlated with site fertility in a range of temperate coniferous and deciduous forests (Vitousek et al., 1979). Litter with a much wider carbon: nitrogen ratio (and thus with a much greater capacity for microbial immobilization of nitrogen) is produced in infertile sites. While litter in tropical sites appears to be generally nitrogen-rich and coniferous forests generally nitrogen-poor (Figure 7.6), all of the biomes include sites representative of the `fertile' and `infertile' patterns of nutrient circulation (Herrera and Jordan, 1981). Additionally, the net mineralizability of the organic nitrogen produced is less in infertile sites than in fertile sites (Vitousek et al., 1982). The loss of carbon may not be too dissimilar between a nitrogen-rich site and a nitrogen-poor site following deforestation, but nitrogen losses are lower and long-delayed in the nitrogen-poor site.

  3. Climate and soils. The rate and pathway of element losses varies with the amount of water entering and percolating through the soil. Soil texture also exerts an important effect on soil water retention, aeration, and erosion, thus affecting the pathways of element loss in deforested systems.

    Soil age and minerology affect element losses somewhat more subtly. The old, deeply leached soils of the tropics and sub-tropics have substantial anion adsorption capacities, and can retain phosphate and sulphate quite effectively. In some sites, anion adsorption is sufficient to hold even the relatively weakly adsorbed anions nitrate and chloride against leaching (Singh and Kanehiro, 1969; Kinjo and Pratt, 1971). The retention of sulphate and especially nitrate in the subsoil would reduce the impacts of deforestation on downstream ecosystems.

  4. Slope. The greatest impact on downstream ecosystems caused by deforestation probably comes about through erosion. Although erosion can be important even on relatively gentle slopes (Pimentel et al., 1976; Juo and Lal, 1979), its greatest impact is on steep, unconsolidated slopes.

7.5 INTERACTIONS OF C, N, P, AND S

Deforestation affects the interactions of the carbon, nitrogen, phosphorus, and sulphur cycles on three levelselement supply, ratios of element availability, and specific interactions.

7.5.1 Element Supply

The most important interaction in the cycling and loss of carbon, nitrogen, phosphorus, and sulphur in deforested ecosystems is simply that many of the major transformations of these elements are biological transformations. As such, they are carried out by organisms which require relatively large amounts of each element to grow and carry out any transformation, and a low supply of any of these elements can inhibit transformations for all of them.

Figure 7.6 Nitrogen circulation in litterfall and the dry weight:N ratio of that litterfall in a range of forest ecosystems. `T' represents tropical forests, `D' deciduous forests, `N' temperate symbiotic N-fixing forests, `C' coniferous forests, and `M' Mediterranean-type ecosystems. Redrawn from Vitousek (1982)

The best-known role of element supply in affecting biological transformations is the control of rates of carbon fixation and organic matter accumulation by the availability of nitrogen and/or phosphorus. Numerous fertilization studies demonstrate that rates of vegetation regrowth and reforestation are often limited by these nutrients. The supply of phosphorus can similarly affect rates of nitrogen fixation (Gorham et al., 1979).

Other relatively well-documented examples where a transformation of one element is inhibited by the lack of another include the suppression of decomposition by low nitrogen supply (Gadgil and Gadgil, 1978), and the suppression of nitrogen mineralization (Jones and Richards, 1977) and nitrification (Purchase, 1974) by low levels of available phosphorus.

7.5.2 Ratios of Element Availability

The best-studied example of the importance of the relative availability of different elements in controlling the rate of a transformation in terrestrial ecosystems is the effect of the carbon:nitrogen ratio on net nitrogen mineralization (Black, 1968). Where material with a wide carbon:nitrogen ratio (such as that from the nitrogen-poor sites in Figure 7.6) is decomposed, decomposers have an abundance of organic carbon (energy) relative to nitrogen (protein). Accordingly, they release carbon as CO2 but retain nitrogen within their biomass, and they generally even remove and incorporate available nitrogen from the soil (Aber and Melillo, 1980; Berg and Staff, 1981). When the carbon:nitrogen ratio is lower, both nitrogen and carbon are released by decomposers. The addition of easily oxidized organic carbon to microscosms with a high rate of nitrogen release stops net nitrogen release (Johnson and Edwards, 1979; Robertson, 1980). Thus, while gross nitrogen mineralization (the breakdown of organic nitrogen compounds) may be closely associated with the rate of decomposition (McGill and Cole, 1981), net nitrogen mineralization (the release of ammonium) is controlled by the relative availability of organic carbon and nitrogen. The effect of deforestation on net nitrogen mineralization will thus depend on the relative availability of carbon and nitrogen in the deforested site.

A similar kind of control over rates of phosphorus and sulphur mineralization has been proposed (Gosz et al., 1973). Alternatively, if McGill and Cole (1981) are correct that phosphorus and ester-sulphate mineralization occurs primarily by extracellular enzymes produced by P- or S-deficient organisms, then both gross mineralization rates and microbial immobilization of P and S should be greatly increased in sites with excess available organic carbon.

Another important element/element interaction could affect the total anion concentration (and thus anion and cation leaching) in the soil solution of a deforested site. Nitrification results in the formation of a mobile anion and hydrogen ions (Figure 7.3). The hydrogen ions produced can replace cations on exchange sites, leading to nitrate and cation leaching (Nye and Greenland, 1960; Likens et al., 1969). Alternatively, some of the hydrogen ions produced can combine with bicarbonate, producing carbonic acid and CO2. In this case the total anion concentration of the soil solution is buffered, as some of the increase in nitrate concentration is offset by a decrease in bircarbonate concentration.

7.5.3 Specific Effects

More specific effects of one element on the cycle of another are difficult to document. It has been suggested that elevated nitrate concentrations can inhibit sulphur-oxidizing bacteria (Likens et al., 1970), thus decreasing sulphate losses from deforested systems. Alternative explanations for the decline in sulphate losses observed in deforested sites are possible, however (Bormann and Likens, 1979).

7.5.4 Interactions Off of the Site

Not all of the important element/element interactions resulting from deforestation would be expected to occur on the deforested site. One important consequence of deforestation is its impacts on downstream ecosystems.

Deforestation causes an increase on the erosional loss of phosphorus and organic carbon. The phosphorus delivered to aquatic systems can drive the eutrophication of lakes, with attendant hypolinmetic oxygen deficits. The organic carbon added could also contribute to the development of anaerobic conditions in lake sediments. The productivity of coastal marine waters could be increased from nitrate lost from deforested sites (Wollast, chapter 14, this volume).

The increase in productivity and anaerobic conditions in lake sediments may cause a somewhat increased preservation of organic carbon, representing a small net sink for atmospheric CO2 (Broecker et al., 1979). More importantly, the increased area of anaerobiosis will increase the rate of sulphate reduction and H2S volatilization (Trudinger, 1979), even though sulphate delivery to aquatic systems may be little affected by deforestation.

7.6 CONCLUSIONS 

Deforestation causes a substantial decrease in the total carbon and nitrogen pools in terrestrial ecosystems. Losses of phosphorus and sulphur occur to a lesser extent. Much of the carbon lost from forest ecosystems eventually enters the atmosphere as CO2. Deforestation is a net source for CO2 wherever: (i) the mean organic carbon pool over the landscape for the land use which follows deforestation is less than that in the undisturbed forest; and (ii) the sum of the organic carbon contained in wood used in building materials, charcoal, and organic matter in aquatic sediments is less than this decline in the terrestrial carbon pool. These conditions are probably usually met. There is also a loss of soil structure and fertility resulting from the loss of soil organic carbon in deforested sites.

The nitrogen lost from deforested sites is lost a little more slowly than the carbon, except in relatively nitrogen-rich forests. Much of the loss follows oxidation to nitrate; nitrate can be lost to the atmosphere as N2 or N2O (with an environmental impact if it is N2O) or to stream-water or ground-water as nitrate (again with an environmental impact). Nitrogen availability to plants is generally increased shortly after deforestation, but the continued loss of nitrogen in deforested sites may eventually reduce the availability of nitrogen to levels below those in the forest (Smith and Young, 1974), necessitating substantial nitrogen fertilization.

Phosphorus is lost primarily through harvest and erosion, and the P eroded from deforested sites can adversely affect downstream ecosystems. The resulting phosphorus depletion in the soil is likely to prove most important in old, deeply leached soils like those in much of the tropics. Losses of sulphur to the atmosphere may be important within deforested sites or downstream from them.

Many of the biological transformations that lead to losses of these four elements are biological transformations, and any of these biological processes can be inhibited by an inadequate supply of any one of these elements. Additionally, the relative availability of these elements can control their net release or immobilization.

ACKNOWLEDGMENTS

I thank R. Herrera, W. A. Reiners and an anonymous reviewer for their helpful criticisms of this manuscript.

7.7 REFERENCES

Aber, J. D., Botkin, D. B., and Melillo, J. M. (1979) Predicting the effects of different harvest regimes on productivity and yield in northern hardwoods, Can. J. Forest Res., 9, 10-18.

Aber, J. D., and Melillo, J. M. (1980) Measuring the relative contributions of organic matter and nitrogen to forest soils, Can. J. Bot., 58, 416-121.

Armentano, T. V., and Hett, J. (eds) (1979) The Role of Temperate Zone Forests in the World Carbon CycleProblem Definition and Research Needs, U.S. D.O.E. Symposium CONF-7903105, Springfield, Virginia, National Technical Information Service.

Armentano, T. V., and Ralston, C. W. (1980) The role of temperate zone forests in the global carbon cycle, Can. J. Forest Res., 10, 53-60.

Bartholomew, W. V. (1977) Soil nitrogen changes in farming systems in the humid tropics, in Ayanaba, A., and Dart, P. J. (eds) Biological Nitrogen Fixation in Farming Systems of the Tropics, New York, Wiley, 27-42.

Berg, B., and Staff, H. (1980) Decomposition and chemical changes in Scots Pine needle litter. II. Influence of chemical composition, in Persson, T. (ed.) Structure and Function of Northern Coniferous Forests: An Ecosystem Study, Ecol. Bull. (Stockholm), 32, 375-390.

Berg, B., and Staff, H. (1981) Leaching, accumulation, and release of nitrogen in decomposing forest litter, in Clark, F. E., and Rosswall, T. (eds) Nitrogen Cycling in Terrestrial Ecosystems: Processes, Ecosystem Strategies, and Management Implications, Ecol. Bull. (Stockholm), 33, 163-178.

Bernhard-Reversat, F. (1977) Recherches sur les variations stationelles des cycles biogéochemiques en fórêt ombrophile de Côte d'Ivoire, Cah. ORSTOM, ser Pédol, 15, 75-189.

Bettany, J. R., Saggar, S., and Stewart, J. W. B. (1980) Comparison of the amounts and forms of sulphur in soil organic matter fractions after 65 years of cultivation, Soil Sci. Soc. Amer. J., 44, 70-75.

Black, C. A. (1968) Soil-Plant Relationships, New York, Wiley.

Boring, L. R., Monk, C. D., and Swank, W. T. (1981) The role of sucessional species in nutrient conservation on a clearcut Appalachian watershed, Ecology 62, 1244-1253..

Bormann, F. H., and Likens, G. E. (1979) Pattern and Process in a Forested Ecosystem, New York, Springer-Verlag.

Bormann, F. H., Likens, G. E., Siccama, T. G., Pierce, R. S., and Eaton, J. S. (1974) The export of nutrients and recovery of stable conditions following deforestation at Hubbard Brook, Ecol. Monogr., 44, 255-277.

Bremner, J. M., and Blackmer, A. M. (1978) Nitrous oxide: emission from soils during nitrification of fertilizer nitrogen, Science, 199, 295-296.

Broecker, W. S., Takahashi, T., Simpson, H. J., and Peng, T. H. (1979) Fate of fossil fuel carbon dioxide and the global carbon budget, Science, 206, 409-418.

Brown, S., Lugo, A. E., and Liegel, B. (eds) (1980) The Role of Tropical Forests on the World Carbon Cycle, U.S. D.O.E. Symposium CONF-800350, Springfield, Virginia, National Technical Information Service.

Cole, D. W., and Gessel, S. P. (1965) Movements of elements through forest soil as influenced by tree removal and fertilizer additions, in Youngberg, C. T. (ed.) ForestSoil Relationships in North America, Corvallis, Oregon State University Press, 95-104.

Cole, D. W., Crane, W. J. B., and Grier, C. C. (1975) The effect of forest management practices on water chemistry in a second-growth Douglas-fir ecosystem, in Bernier, B., and Winget, C. F. (eds) Forest Soils and Land Management, Quebec, Les Presses de L'Université Laval, 195-208.

Couto, W., Lathwell, D. J., and Bouldin, D. R. (1979) Sulphate sorption by two oxisols and an alfisol of the tropics, Soil Sci., 127, 108-116.

Covington, W. W. (1976) Forest floor organic matter and nutrient content and leaf fall during secondary succession in northern hardwoods, Doctoral Thesis, Yale University, New Haven, Connecticut.

Crutzen, P. J. Atmospheric interactionshomogeneous gas reactions of C, N, and S containing compounds, Chapter 3, this volume.

Crutzen, P. J., and Ehhalt, D. H. (1977) Effects of nitrogen fertilizer and combustion on the stratospheric ozone layer, Ambio, 6, 112-117.

Crutzen, P J., Heidt, L. E., Krasnec, J. P., Pollock, W. H., and Seiler, W. (1979) Biomass burning as a source of atmospheric gases CO, H2, N2O, NO, CH3Cl, and COS, Nature, 253-256.

Delcourt, H. R., and Harris, W. F. (1980) Carbon budget of the south-eastern U.S. biota: Analysis of historical change in trend from source to sink, Science, 210, 321-322.

Dominski, A. S. (1971) Nitrogen transformations in a northern-hardwood podzol on cutover and forested sites, Doctoral Thesis, Yale University, New Haven, Connecticut.

Eaton, J. S., Likens, G. E., and Bormann, F. H. (1978) The input of gaseous and particulate sulphur to a forest ecosystem, Tellus, 30, 546-551.

Firestone, M. K., Firestone, R. B., and Tiedje, J. M. (1980) Nitrous oxide from soil denitrification; factors controlling its biological production, Science, 208, 749-751.

Gadgil, R. L., and Gadgil, P. D. (1978) Influence of clearfelling on decomposition of Pinus radiata litter, N. Z. J. For. Sci., 8, 213-224.

Galloway, J. N., and Whelpdale, D. M. (1980) An atmospheric sulphur budget for eastern North America, Atmos. Environ., 14, 409-417.

Garrels, R. M., and MacKenzie, F. T. (1971) Evolution of Sedimentary Rocks, New York, Norton.

Gholz, H. L. (1980) Production and the role of vegetation in element cycles for the first three years on an unburned clearcut watershed in western Oregon, Ecol. Soc. Amer. Bull., 61,149 (Abstract).

Gorham, E., Vitousek, P. M., and Reiners, W. A. (1979) The regulation of element budgets in the course of terrestrial ecosystem succession, Rev. Ecol. Syst., 10, 53-84.

Gosz, J. R., Likens, G. E., and Bormann, F. H. (1973) Nutrient release from decomposing leaf and branch litter in the Hubbard Brook Forest, New Hampshire, Ecol. Monogr., 43, 173-191.

Haas, H. J., Evans, C. E., and Miles, E. F. (1957) Nitrogen and carbon changes on Great Plains soils as influenced by cropping and soil treatments, U.S.D.A. Technical Bulletin 1164.

Haas, H. J., Grunes, D. L., and Reichman, G. A. (1961) Phosphorus changes in Great Plains soils as influenced by cropping and manual applications, Soil Sci. Soc. Amer. Proc., 25, 214-218.

Harcombe, P. A. (1977) Nutrient accumulation by vegetation during the first year of recovery of a tropical forest ecosystem, in Cairns, J., Dickison, K. L., and Herricks, E. E. (eds) Recovery and Restoration of Damaged Ecosystems, Charlottesville, Virginia, University of Virginia Press, 347-378.

Herrera, R., and Jordan, C. F. (1981) Nitrogen cycle in a tropical Amazonian rain forest: the caatinga of low mineral nutrient status, in Clark, F. E., and Rosswall, T. (eds) Nitrogen Cycling in Terrestrial Ecosystems: Processes, Ecosystem Strategies, and Management Implications, Ecol. Bull. (Stockholm), 33, 493-505.

Hesselman, H. (1917) Om vissa skogsföryngringsåtgarders inverkan på saltpeterbildningen i marken och dess betydelse für barrskogen föryngring, Medd. Stat. skogsförkningsanst, 13-14, 923.

Johnson, D. W., and Cole, D. W. (1980) Anion mobility in soils: relevance to nutrient transport from forest ecosystems, Env. Internat., 3, 79-90.

Johnson, D. W., Cole, D. W., Gessel, S. P., Singer, M. J., and Minden, R. B. (1977) Carbonic acid leaching in a tropical, temperate, subalpine, and northern forest soil, Arc. Alpine Res., 9, 329-343.

Johnson, D. W., and Edwards, N. T. (1979) The effects of stem girdling on biogeochemical cycles within a mixed deciduous forest in eastern Tennessee. II. Soil nitrogen mineralization and nitrification rates, Oecologia, 40, 259-271.

Johnson, D. W., Hornbeck, J. W., Kelly, J. M., Swank, W. T., and Todd, D. E. (1980) Regional patterns of soil sulphate accumulation: relevance to ecosystem sulphur budgets, in Shriner, D. S., Richmond, C. R., and Lindberg, S. E. (eds) Atmospheric Sulphur Deposition: Environmental Impact and Health Effects, Michigan, Ann Arbor Science, 507-520.

Jones, J. M., and Richards, B. N. (1977) Effect of reforestation on turnover of 15 N-labelled nitrate and ammonia in relation to changes in soil microflora, Soil Biol. Biochem., 9, 383-392.

Jordan, C. F. (1971) A world pattern in plant energetics, Am. Sci., 59, 425-433.

Juo, A. S. R., and Lal, R. (1979) Nutrient profile in a tropical alfisol under conventional and no-till systems, Soil Sci., 127, 168-173.

Kinjo, T., and Pratt, P. F. (1971) Nitrate adsorption. I. In some acid soils of Mexico and South America, Soil Sci. Soc. Amer. Proc., 31, 722-725.

Likens, G. E., Bormann, F. H., and Johnson, N. M. (1969) Nitrification: importance to nutrient losses from a cutover forested ecosystem, Science, 163, 1205-1206. 

Likens, G. E., Bormann, F. H., Johnson, N. M., Fisher, D. W., and Pierce, R. S. (1970) Effects of forest cutting and herbicide treatment on nutrient budgets in the Hubbard Brook ecosystem in New Hampshire, Ecol. Monogr., 40, 23-27.

Likens, G. E., Bormann, F. H., Pierce, R. S., Eaton, J. S., and Johnson, N. M. (1977) Biogeochemistry of a Forested Ecosystem, New York, Springer-Verlag.

Likens, G. E., Bormann, F. H., Pierce, R. S., and Reiners, W. A. (1978) Recovery of a deforested ecosystem, Science, 199, 492-496.

Lugo, A. E. (1980) Are tropical forests sources or sinks of carbon? in Brown, S., Lugo, A. E., and Liegel, B. (eds) The Role of Tropical Forests on the World Carbon Cycle, U.S. D.O.E. Symposium CONF-800350. Springfield, Virginia, National Technical Information Service, 1-18.

McCallan, M. E., O'Learly, B. M., and Rose, C. W. (1980) Redistribution of Caesium-137 by erosion and deposition on an Australian soil, Aus. J. Soil Res., 18, 119-128.

McGill, W. B., and Cole, C. V. (1981) Comparative aspects of organic C, N, S and P cycling through soil organic matter during pedogenesis, Geoderma, 26, 267-286. 

Magee, P. N. (1977) Nitrogen as a health hazard, Ambio, 6,123-125.

Marks, P. L., and Bormann, F. H. (1972) Revegetation following forest cutting: mechanisms for return to steady state nutrient cycling, Science, 176, 914-915.

Meentemeyer, V. (1978) Macroclimate and lignin control of litter decomposition rates, Ecology, 59, 465-472.

Nye, P. H. (1961) Organic matter and nutrient cycles under moist tropical forest, Plant Soil, 13, 333-345.

Nye, P. H., and Greenland D. J. (1960) The soil under shifting cultivation, Commonwealth Bureau of Soils, Harpenden, England, Technical Bulletin No. 51. 

Nye, P. H., and Greenland, D. J. (1964) Changes in the soil after clearing tropical forest, Plant Soil, 21, 101-112.

Nye, P. H., and Tinker, P. B. (1977) Solute Movement in the SoilRoot System, Oxford, Blackwell.

Pimentel, D., Terhune, E. C., Dyson-Hudson, R., Rochereau, S., Samis, R., Smith, E. A., Denman, D., Reifschneider, D., and Shepard, M. (1976) Land degradation: effects on food and energy resources, Science, 194,149-155.

Purchase, B. S. (1974) The influence of phosphate deficiency on nitrification, Plant Soil, 41, 541-547.

Rapp, A. (1975) Soil erosion and sedimentation in Tanzania and Lesotho, Ambio, 4, 154-163.

Rice, H., Nochumson, D. H., and Hidy, G. M. (1981) Contributions of anthropogenic and natural sources to atmospheric sulphur in parts of the United States, Atmos. Environ., 15, 1-9.

Ritchie, J. C., Spraberry, J. A., and McHenry, J. R. (1974) Estimating soil erosion from the redistribution of fallout 137Cs, Soil Sci. Soc. Amer. Proc., 38, 137-139. 

Robertson, G. P. (1980) The characterization and control of nitrification in primary and secondary succession, Doctoral Thesis, Indiana University. Bloomington, Indiana. 

Rosswall, T. (1976) The internal cycle between vegetation, micro-organisms, and soils, in Svennson, B. H., and Söderlund, R. (eds) Nitrogen, Phosphorus, and SulphurGlobal Cycles, SCOPE Report No. 7, Ecol. Bull. (Stockholm), 22,157-167. 

Schindler, D. W. (1977) Evolution of phosphorus limitation in lakes, Science, 195, 260-262.

Schlesinger, W. H. (1977) Carbon balance in terrestrial detritus, Ann. Rev. Ecol. Syst., 8, 51-81.

Seiler, W., and Crutzen, P. J. (1980) Estimates of gross and net fluxes of carbon between the biosphere and the atmosphere from biomass burning, Climatic Change, 2, 207-247.

Singh, B. R., and Kanehiro, Y. (1969) Adsorption of nitrate in amorphous and kaolinitic Hawaiian soils, Soil Sci. Soc. Amer. Proc., 29, 681-683.

Smith, S. J., and Young, L. B. (1974) Distribution of nitrogen forms in virgin and cultivated soils, Soil Sci., 120, 354-360.

Snyder, G. G., Haupt, H. F., and Belt, G. H., Jr. (1975) Clearcutting and burning alter quality of streamwater in northern Idaho. U.S. D.A. Forest Service Research Paper INT-168, Intermt. Forest and Range Experimental Station.

Sollins, P., and McCorison, F. M. (1981) Changes in solution chemistry after clearcutting in an old-growth Douglas-fir watershed, Water Resources Research, 17, 1409-1418.

Stone, E. (1973) The impact of timber harvest on soil and water, in President's Advisory Panel on Timber and Environment Report, Washington, D.C., U.S. Government Printing Office, 427-467.

Stone, E. L., Swank, W. T., and Hornbeck, J. W. (1979) Impacts of timber harvest and regeneration on stream flow and soils in the eastern deciduous region, in Youngberg, C. T. (ed.) Forest Soils and Land Use, Fort Collins, Colorado State University Press, 516-535.

Swanson, F. J., Fredriksen, R. L., and McCorison, F. M. (1981) Material transfer in a western Oregon forested watershed, in Edmonds, R. L. (ed.) Analysis of Coniferous Forest Ecosystems in the Western United States, Stroudsberg, Pennsylvania, Dowden, Hutcheson, and Ross (in press).

Tamm, C. O., Holmen, H., Popovic, B., and Wiklander, G. (1974) Leaching of plant nutrients from soils as a consequence of forestry operations, Ambio, 3, 211-221. 

Trudinger, P. A. (1979) The biological sulphur cycle, in Trudinger, P. A., and Swaine, D. J. (eds) Biogeochemical Cycling of Mineral-Forming Elements, Amsterdam, Elsevier Publishing Company, 293-313.

Vitousek, P. M. (1982) Nutrient cycling and nutrient use efficiency, Am. Nat. 119, 553-572.

Vitousek, P. M., and Melillo, J. M. (1979) Nitrate losses from disturbed forests: patterns and mechanisms, Forest Sci., 25, 605-619.

Vitousek, P. M., and Reiners, W. A. (1975) Ecosystem succession and nutrient retention: a hypothesis, BioScience, 25, 376-381.

Vitousek, P. M., Gosz, J. R., Grier, C. C., Melillo, J. M., Reiners, W. A., and Todd, R. L. (1979) Nitrate losses from disturbed ecosystems, Science, 204, 469-474. 

Vitousek, P. M., Gosz, J. R., Grier, C. C., Melillo, J. M., and Reiners, W. A. (1982) A comparative analysis of potential nitrification and nitrate mobilization in forest ecosystems, Ecol. Monogr., 52, 155-177.

Walker, T. W., and Syers, J. K. (1976) The fate of phosphorus during pedogenesis, Geoderma, 15, 1-19.

White, P. (1979) Pattern, process, and natural disturbance in vegetation, Bot. Rev., 45, 229-299.

Wiklander, G. (1981) Rapporteur's comment on clearcutting, in Clark, F. E., and Rosswall, T. (eds) Nitrogen Cycling in Terrestrial Ecosystems: Processes, Ecosystem Strategies, and Management Implications, Ecol. Bull. (Stockholm), 33, 642-647. 

Woodwell, G. M., Whittaker, R. H., Reiners, W. A., Likens, G. E., Delwiche, C. C., and Botkin, D. B. (1978) The biota and the world carbon budget, Science, 199, 141-145.

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