Grazing Impacts on Soil Physical, Chemical, and ...

20 Grazing Impacts on Soil Physical, Chemical, and Ecological Properties in Forage Production Systems Miguel A. Taboada, Gerardo Rubio, and Enrique J. Chaneton

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xtensive systems of animal husbandry (e.g., meat and dairy cattle, sheep, goats, and so on) are mainly based on the direct grazing of grasslands, pastures, fodder crops, and crop residues by livestock. Grazing effects on soil properties of forage production systems follow direct and indirect pathways. Direct effects relate to animal trampling and excretion, while indirect effects are mediated by changes in vegetation structure and function. Figure 20|1 shows a simplified model of a grazing system, integrating both direct and indirect effects of livestock on soil physical properties, as proposed by Greenwood and McKenzie (2001). Grazing animals affect soil and vegetation properties through the action of treading, defoliation, and excretal returns. This may result in soil compaction and/or poaching damage to the pasture, which may recover as a function of the effectiveness of abiotic (e.g., wetting and drying cycles) or biotic (e.g., roots and worms activity) mechanisms of structural resilience. In addition, defoliation and excretal returns impact carbon and nutrient cycling in soil, which involve particular responses by soil organisms. Soil pores are the habitat for microbes and meso- and microfauna (Young and Ritz, 2005), which in turn are responsible for biotic mechanisms of soil structural resilience. In this chapter, we review the various effects of grazing on soil physical properties, carbon, and nutrient cycling in forage production systems. Where possible, we also discuss some management implications focusing on possible ways of minimizing negative grazing effects as well as recovering desirable soil properties for sustainable production.

Soil Physical Impacts of Livestock Grazing

Several review articles have been published over the last decades on the soil physical impacts of grazing (e.g., Gifford and Hawkins, 1978; Willatt and Pullar, 1983; Greenwood and McKenzie, 2001; Drewry, 2006), and most of their main conclusions are summarized in this section. Grazing impacts on soil physical properties are caused by defoliation and treading. Both actions cannot

Miguel A. Taboada, Soil Institute, CIRN, INTA, Division of Fertility and Fertilizers, Agronomy Faculty, University of Buenos Aires and CONICET, Nicolás Repetto and De Los Reseros, 1686 Hurlingham, Buenos Aires Province, Argentina; Gerardo Rubio, Division of Fertility and Fertilizers, Agronomy Faculty, University of Buenos Aires, and INBA-CONICET; Enrique J. Chaneton, Division of Ecology, Agronomy Faculty, University of Buenos Aires, and IFEVA-CONICET, San Martín Avenue 4453, C1417 DSE Buenos Aires, Argentina. *Corresponding author ([email protected]; [email protected]). doi:10.2136/2011.soilmanagement.c20 Copyright © 2011. American Society of Agronomy and Soil Science Society of America, 5585 Guilford Road, Madison, WI 53711, USA. Soil Management: Building a Stable Base for Agriculture. Jerry L. Hatfield and Thomas J. Sauer (ed.)

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study, an area grazed year-round by cattle for more than a century (average stocking rate 0.5 cattle ha–1) was compared with an adjacent, 4-ha exclosure protected from grazing for 7 yr. Within the exclosure, accumulation of litter and standing dead and live plant biomass reduced increases of topsoil temperatures during the day. Such increases did occur in the grazed area, because of its low ground cover and short canopy height, which resulted from greater cover by short grasses and low-growing forbs as compared with the exclosure (Sala et al., 1986). Similar Fig. 20|1. Interactions among soil, vegetation, and changes in topsoil temperatures livestock in grazing systems (modified from Greenwere found in semiarid South Afriwood and McKenzie, 2001). can rangelands by Snyman and du Preez (2005). Increased topsoil temperatures caused by grazing led to be easily decoupled in field studies (Curll higher soil water evaporation rates (Table and Wilkins, 1983), but their various conse20|1). In the Flooding Pampa case, soil quences deserve separate analyses. evaporation rates in the grazed area were 3.6 to 10.2 higher than in the ungrazed area, showing the effect of mulching by plant litChanges Imposed by Defoliation ter in soil water conservation. and Decreased Cover High evaporation rates under grazing The intensity of defoliation has greater (Table 20|1) may promote the accumulation effects on net herbage accumulation than of salts in the topsoil because of the upward treading (Curll and Wilkins, 1983). Grazing movement of water and soluble salts from exerts more pressure on the most palatable salinized deep horizons (Lavado and plant species, which may decrease dependTaboada, 1987; Alconada et al., 1993). An ing on grazing system and exclusion example of this process is shown in Fig. periods, changing grassland floristic com20|2, after results from Lavado and Taboada position, and vertical canopy structure (1987). Salts did not accumulate at the soil (Sala et al., 1986). Defoliation increases the surface in the absence of grazing because proportion of bare soil, and thus daily tema thick layer of litter reduced evaporative perature fluctuation and maximum daily losses and thus decreased the upward flux temperatures. Such effects are illustrated in of water and salts. As a result, only minor Table 20|1 based on results from a natural salinity fluctuations were detected in the grassland in the Flooding Pampa of ArgenA and BA horizons of a nitric soil (typic tina (Lavado and Taboada, 1987). In this Natraquoll). The episodic salinization of

Table 20|1. Topsoil temperature and soil evaporation rate in two summer dates (modified from Lavado and Taboada, 1987). Soil temperature range Months

GE†

G†

Evaporation rate GE† Mean

−1 —————————°C ————————— mm d

G† CV

Mean

CV

%

mm d−1

%

January

22.8–25.5

26.4–32.7

0.19

93.12

2.13**

43.89

February

17.7–23.9

21.5–30.4

0.39

81.65

1.79**

30.41

** Highly significant differences between treatments (P < 0.01). † GE, grazing exclusion; G, grazed land.

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Grazing Impacts on Soil | MiguelChapter A. Taboada | Authors et al.

Soil Management Practices the topsoil may substantially decrease grassland forage production. It can be controlled by increasing ground cover through changes in grazing regime, by introducing longer rest periods or changing from continuous to rotational grazing. In the Flooding Pampa, Alconada et al. (1993) showed that salt ascension by capillarity from groundwater may be effectively limited through the retention of runoff water by earth banks. Interestingly, this practice also caused the replacement of unpalatable halophytic grasses by more palatable and productive hygrophilous species (Alconada et al., 1993).

Changes Imposed by Trampling and Treading

The mechanical impact of animal hooves on soil surface can be a severe disturbance on topsoil structure through the external forces applied by treading and trampling (Gifford and Hawkins, 1978; Willatt and Pullar, 1983 Greenwood and McKenzie, 2001; Drewry, 2006). This mechanical impact changes the form and stability of soil aggregates, which results in changes in bulk density, pore size distribution, and soil strength, among other properties. Different factors may change aggregate form and stability in grazed soils. Figure 20|3 shows a proposed conceptual model that integrates those factors in a dynamic equilibrium affected by opposed disintegration and regeneration forces.

Influence of Animal Type, Soil Properties, and Vegetation Disintegration forces are due to animal treading and

Fig. 20|2. Soil salinity measured as electrical conductivity of saturation extracts in A and BA horizons of a Typic Natraquoll in the Flooding Pampa, Argentina, under long-term grazing exclusion and continuous grazing (after Lavado and Taboada, 1987).

Fig. 20|3. Conceptual model describing factors causing variations in topsoil structural form and stability in grazed systems. Solid lines denote direct influences, dotted lines indirect influences.

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trampling, and their impact depends on animal type, soil bearing capacity, vegetation, and grazing system. These forces are counteracted by regeneration forces. The abiotic or biotic nature of this regenerative process depends on soil type, and recovery times on climatic, vegetation, and grazing system factors. The pressure exerted by grazing animals on soil is a function of the animal’s mass, foot size, and kinetic energy. Table 20|2 shows static pressures exerted by different kinds of livestock. These data show that pressures exerted by sheep and cattle average 66 kPa and 138 kPa, respectively, when standing. Despite sheep exerting lower static pressure than cows, herbivory and trampling by sheep may lead to landscape degradation in semiarid and desert areas (Rostagno 1989; Golodets and Boeken, 2006; Li et al., 2008). Most animal movement probably occurs during grazing, causing the amount of treading by each animal to depend on forage availability and grazing period. Cattle tread more on bare ground than on ground covered by tussocks, probably due to a preference to walk on even surfaces (Balph et al., 1989). The distance traveled by grazing animals is also affected by the distance to water points and quality

of feed (Greenwood and McKenzie, 2001; Fensham and Fairfax, 2008). Soil water content largely determines surface bearing capacity, and then controls structural damages by treading (Greenwood and McKenzie, 2001). Table 20|3 summarizes the most expected soil structural effects resulting from animal treading with different soil water conditions. For soils trampled when dry, structural deterioration results from aggregate crushing by animal hooves. This leads to the prevalence of smaller aggregates in the upper horizons (Warren et al., 1986; Taboada et al., 1999). When moist, animal trampling compresses the soil beneath the hoof (Scholefield et al., 1985) and collapses the larger soil pores (Warren et al., 1986; Taboada and Lavado, 1993). This results in increases in bulk density by compaction in the soil surface (Willatt and Pullar, 1983; Greenwood and McKenzie, 2001). Grazing compaction often results in reduced infiltration rates and low saturated hydraulic conductivity (Gifford and Hawkins, 1978; Willat and Pullar, 1983). Greenwood et al. (1997) concluded that the loss of porosity caused by grazing was related to a decrease in the number and/or continuity of pores greater than 1.2

Table 20|2. Comparative weight, foot area, and static pressure of grazing animals (modified from Greenwood and McKenzie, 2001).

Sheep

Mass

Total foot area

Static pressure

kg

cm

kPa

40–54

55–84

48–83

2

Cattle

306–612

264–460

98–192

Horses

400–700

736

54–95

Goats

40

55

60–73

Table 20|3. Expected damages produced by animal treading at different soil water conditions, affected soil depth, and changes in soil physical properties. Soil water conditions during treading

Dry

Moist

Saturated

Expected damage

aggregate crushing

compaction

puddling and poaching no effect

Ponded

Depth

4.5%) silty loam and clay loam topsoils (up to 10 cm). Soil macroporosity (>30 mm) and infiltration rate showed the highest improvement percentages after grazing cessation (up to 127%). Visual hoof damages from soil puddling and deformation were reduced to a half of the initial value within 87 to 165 d, depending on the characteristics of the soil. Physical deterioration of soil to about 5-cm depth can be naturally ameliorated by the burrowing activities of macroinvertebrates

associated with dung deposition (Herrick and Lal, 2005). Sometimes, natural recovery forces are strong enough not to require cessation of grazing, as in flooded grassland soils in Argentina (Taboada and Lavado, 1993; Taboada et al., 1999, 2001). Recovery generally demands many years in drier climates (Gifford and Hawkins, 1978; Greenwood and McKenzie, 2001). Gifford and Hawkins (1978) reported for rangelands in the USA that infiltration rates might be still increasing 13 yr after grazing stopped. Other authors also found long periods (up to several years) of recovery (Braunack and Walter, 1985; Steffens et al., 2008). These long recovery periods require grazing exclusion for several years, which is only practically possible under “cut-andcarry” or nil-grazing systems, which are difficult to implement under extensive grazing systems. Faster responses are however possible in temperate climates, because of the action of frequent wetting and drying cycles, freezing and thawing cycles, and vigorous pasture root growth.

Grazing Impacts on Soil Carbon Dynamics Carbon Budget in Grazing Systems

More than two-thirds of the carbon stored in grasslands is located below ground in soil organic matter pools (Parton et al., 1987; Burke et al., 1989). Grassland carbon stocks are primarily determined by climatic factors; total carbon increases with precipitation as a result of increased primary production, and decreases with increasing temperature due to increased decomposition (Burke et al., 1989; Sala et al., 1996). Large grazers remove 20 to 75% of the aboveground net primary production (Milchunas and Lauenroth, 1993; Oesterheld et al., 1999). By diverting carbon away from the plant detrital pathway, grazing animals may reduce the energy supply to soil decomposers (Cebrian, 1999; Wardle and Bardgett, 2004). Thus, grazing exerts a major influence on grassland carbon cycling, affecting not only transfers among vegetation and soil compartments, but also ecosystem input and output flows (Fig. 20|4). In the long run, these alterations may have important consequences for the capacity of managed grasslands to store carbon and 307

apply also to nutrient cycling, which is the focus of the next section.

Grazing Impacts on Carbon Pools

Fig. 20|4. Diagrammatic model of the carbon budget in a grazing system showing major carbon pools and transfer flows. In this schematic, microbial decomposers are included in the soil labile organic matter compartment. Solid arrows depict ecosystem inputs and transfers among compartments; broken arrows denote gaseous losses. NPP, net primary productivity; R, respiration; C, consumption; D, detritus production; E, excretion. Grazing may directly or indirectly control all these fluxes.

contribute to the regulation of the global carbon cycle (Sala et al., 1996; Steinfeld and Wassenaar, 2007). In this section, we discuss the effect pathways whereby grazers alter the carbon budget of grasslands. We begin by discussing grazing impacts on carbon stocks and flows, and then move on to review the mechanisms for grazing effects on litter decomposition, a key process in the maintenance of soil fertility and forage production. In doing so, we consider how changes in plant functional composition influence soil functioning. Lastly, we address some management implications of vegetation–soil feedbacks for restoring carbon stocks in degraded grazing lands. To a large extent, these arguments

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The amount of CO2 fixed by plants through net primary production (NPP) can follow different pathways (Fig. 20|4). Carbon is translocated from above- to belowground plant organs and back to shoots depending on season and grazing pressure. Herbivores consume a sizeable fraction of the aboveground NPP; the rest of the carbon is accumulated as plant biomass and eventually enters above- and belowground detrital routes, undergoing decomposition by soil microorganisms. Grazers also transfer carbon into the soil organic matter via waste deposition. Soil organic matter can be divided into different compartments according to their turnover rates (Parton et al., 1987, 1988); here, soil microbes are included in a “labile” carbon pool with rapid (1–25 yr) turnover times. A small fraction of recalcitrant detritus, largely of root origin, slowly accumulates in a “passive” soil pool with a slow turnover time (>200 yr), which represents a net carbon sink (Chapin et al., 2002). Most carbon losses from grasslands occur through plant, herbivore, and soil respiration (Fig. 20|4). Yet livestock also releases substantial amounts of methane (CH4) through digestive fermentation (Steinfeld and Wassenaar, 2007). A static view of the ecosystem allows one to assess the effects of grazing on carbon distribution among plant and soil compartments. Such studies typically involve comparisons between grazed grassland and ungrazed exclosures and are subjected to several sources of variation, including management, age of exclosure, site productivity, vegetation composition, soil type, and topography (Milchunas and Lauenroth, 1993; Stohlgren et al., 1999). Nevertheless, taken as a whole, exclosure studies offer a global picture of the direction and magnitude of grazing effects on ecosystem function (Milchunas and Lauenroth, 1993). Whereas in mesic grasslands grazing decreases the amount of carbon stored in aboveground vegetation, except where woody species encroachment shifts the dominant life form (Asner et al., 2004), its impact on root mass is far more variable and context dependent (Milchunas and Lauenroth, 1993). On the other hand, livestock grazing has been reported to decrease


Soil Management Practices (Marrs et al., 1989; Frank et al., 1995; Altesor et al., 2006), increase (Dormaar et al., 1994; Chaneton and Lavado, 1996; Schuman et al., 1999), or even produce no observable change in soil organic carbon (Henderson et al., 2004). A few examples suffice to illustrate the variability of grazing effects on soil carbon pools. In mixed prairies of the North-Central USA Frank et al. (1995), found that moderately grazed sites contained 15% less soil carbon (107-cm depth) than grazing exclosures. The effect disappeared for heavily grazed sites, where increased cover by shallow-rooted grasses would have compensated for soil carbon losses (Frank et al., 1995). In similar grassland, Schuman et al. (1999) reported an increase of soil carbon in the rooting zone (0–30 cm) under heavy stocking rates. In the Flooding Pampa of Argentina, cattle grazing did not affect soil organic carbon (10-cm depth) in an upland site, but increased it by 28% in a lowland site (Chaneton and Lavado, 1996), suggesting that topographic position modulated grazing effect on carbon cycling. Moreover, grazing effects may depend on the soil pool being sampled. For several native grasslands in Uruguay, Altesor et al. (2006) found more organic carbon at the surface soil (0–5 cm) layer in grazed compared with ungrazed plots, although they also detected an 8% reduction in carbon stock for deeper layers (60–100 cm) of grazed sites. This pattern may suggest that increased root production of grazed vegetation (Doll, 1991; Altesor et al., 2006) may counteract grazing-induced reductions in aboveground detrital inputs to soil. Lack of consistent grazing effects on soil carbon has also been commonplace in some grassland regions. Henderson et al. (2004) sampled nine locations across the Canadian Great Plains and detected no trend in root mass, subsurface litter, or soil particulate carbon in response to grazing. Stohlgren et al. (1999) found no definite, overall grazing effect on soil carbon for several Rocky Mountain grasslands, although negative and positive effects occurred within different management units. Neutral grazing effects may well be associated with the short time scale inherent to exclosure studies, given the slow turnover rate of major soil carbon pools (Milchunas and Lauenroth, 1993; Piñeiro et al., 2006). A different trend emerges for arid rangelands; there

reductions in soil carbon due to overgrazing appear to be the norm (Schlesinger et al., 1990; Abril and Bucher, 2001; Asner et al., 2004; Neff et al., 2005). Indeed, in a global analysis of grazing impacts across systems (200–800 mm yr –1), Milchunas and Lauenroth (1993) found that soil organic matter changes were similarly divided between negative and positive, even though grazing decreased aboveground NPP irrespective of habitat.

Grazing Alteration of Major Carbon Flows

A dynamic view of input and output flows may help to understand how soil carbon stocks respond to grazing in different systems. Aboveground NPP increases along a precipitation gradient from arid to humid grasslands (Milchunas and Lauenroth, 1993; Oesterheld et al., 1999). Stocking rates and biomass consumption by livestock also increase with precipitation, but in an exponential fashion (Oesterheld et al., 1998, 1999). As a result, grazing impacts on vegetation composition are greater in humid than in dry grasslands, and this is correlated with increasing reductions in aboveground NPP by grazing (Milchunas and Lauenroth 1993). Trends in total root mass are less clear-cut; grazing effects are generally positive but seem unrelated to aboveground NPP changes across ecosystems (see Milchunas and Lauenroth 1993). Thus, grazing is expected to reduce the carbon flow from standing vegetation to surface litter and to the soil organic matter (Fig. 20|4), although larger quantities of plant carbon may be allocated belowground in grazed grassland (Doll, 1991; Schuman et al., 1999). Microbial decomposition of litter entering the labile soil pool is the main process whereby carbon is released to the atmosphere (Fig. 20|4). Decomposition rates vary markedly between shoot and root litter substrates, the latter being less degradable (Semmartin et al., 2004, 2008). Grazing may accelerate or retard aboveground litter breakdown through various mechanisms (see below), with the direction of the effect depending on site productivity and plant community turnover (Bardgett and Wardle, 2003). In contrast, grazing effects on root decomposition are often negligible (Semmartin et al., 2008). This suggests that carbon release per unit of the largest litter 309

pool may not be substantially affected by livestock grazing. We now synthesize the above discussion in a hypothetical model of grazing effects on the amount of organic carbon stored in the rooting zone (i.e., the layer in which most roots are located). The model suggests that the relative effect of grazing on soil carbon may shift from negative in arid systems, to positive in mesic through humid habitats (Fig. 20|5). Yet, in mesic-tohumid systems, soil carbon may be either increased or decreased, depending on the balance between root production (input) and above- and belowground litter decomposition (output). Both of these processes will be influenced by the functional traits of plant species that come to dominate with grazing, and by site conditions (Bardgett and Wardle, 2003; Garibaldi et al., 2007). At the opposite end of the gradient, negative grazing effects on soil carbon prevail (Fig. 20|5). A large body of evidence suggests that reduced litter inputs from herbivory (Oesterheld et al., 1999; Abril and Bucher, 2001), increased erosion (Asner et al., 2004; Neff et al., 2005), and photodegradation of litter (Austin and Vivanco, 2006) may lead to significant reductions in soil carbon stocks in

arid rangelands. Note that this model does not allow for within-site heterogeneity in soil carbon pools, e.g., “fertility islands” in arid lands (Schlesinger et al., 1990).

Net Carbon Balance: A Long-Term Perspective

Long-term changes in input and output fluxes may not only alter topsoil carbon content, but may also be transmitted to deeper layers, affecting storage in less active pools, and the grassland net carbon balance (Chapin et al., 2002). To account for grazing impacts on whole-grassland carbon dynamics, a long-term (>100 yr) perspective is needed (Burke et al., 1989). Recently, Piñeiro et al. (2006) tackled this issue for the Río de la Plata grasslands in southern South America, using the CENTURY model (Parton et al., 1987, 1988), which simulates long-term dynamics of soil carbon and nitrogen pools. They evaluated the impact of livestock grazing since its introduction to the region some 400 years ago. They estimated changes in grassland carbon stocks in vegetation and soil pools by comparing model outputs between c.1600 and 1970. Total carbon in vegetation and surface microbes was estimated to have decreased an average of 32% (n = 11 sites) after livestock introduction (1600: 1804 g C m–2 vs. 1970: 1223 g C m−2), while carbon stored in soil pools declined by 21% (1600: 10,082 g C m−2 vs. 1970: 7967 g C m−2). The largest decrease (32%) was in the “slow” organic carbon pool (turnover ~25 yr). As a result, grazing redistributed soil carbon to the “passive” pool from 39% to 47%. These changes partly reflected the imbalance between input and output carbon fluxes (Table 20|4). Both NPP inputs and respiration losses decreased with domestic grazing but total respiration decreased more. A greater proportion of NPP was released through herbivore respiration, which reduced transfers Fig. 20|5. A summary model of the relative effect to soil microbes, decreasing soil resof grazing on soil organic carbon in the rootpiration (Table 20|4). The increase in ing zone, along a habitat moisture gradient from carbon losses through herbivore respidry to humid grasslands. Values above (positive) ration, however, was smaller than the and below (negative) the dashed line indicate absolute decrease in carbon fixation by increases and decreases in soil carbon as a NPP (Piñeiro et al., 2006). result of grazing, respectively. The stippled area Most importantly, livestock introrepresents cross-site variability in soil carbon duction augmented gaseous N output response within habitat types. (mainly as NH3) from animal excreta. 310


Soil Management Practices Table 20|4. Major carbon flows (g C m−2 yr−1) estimated before and after the introduction of livestock grazing in Río de la Plata grasslands (after Piñeiro et al., 2006). 1600

1970

Change %

Net primary productivity

889

670

Detrital inputs to soil

884

562

Herbivore respiration

5

107

Soil respiration

884

571

−24.6 −36.4 +2040 −35.4

These N losses were associated with higher C:N ratios of all system compartments, suggesting that N availability, rather than carbon, constrained organic matter accumulation in soils (Piñeiro et al., 2006). These results illustrate the remarkable control that long-term livestock grazing may exert on the size of major elemental pools and flux rates in rangelands.

Grazing Effects on Soil Microbes and Decomposition

Human activity has been historically associated with large changes in herbivore loads. A pervasive consequence of managed grazing systems is the alteration of ecological feedbacks between aboveand belowground processes (Wardle and Bardgett, 2004), which parallel grazing-induced changes in vegetation structure and composition (Milchunas and Lauenroth, 1993; Chaneton et al., 1996; Asner et al., 2004). Up to this point, we included decomposers in the labile soil carbon pool. We now discuss responses of soil organisms to livestock grazing to provide for a mechanistic understanding of grazing effects on soil carbon cycling. Recent studies have found variable effects of grazing on litter decomposition. In these experiments, different litter types are incubated in a common environment or, less often, in grazed vs. ungrazed sites.

As a general trend, data show marked differences in decomposition rate among litter species, as well as strong effects of incubation site, but patterns do not always conform to a grazing effect. Litter from grazing-promoted species may either decompose faster or more slowly than litter from grazingreduced species (Garibaldi et al., 2007). In a study comparing rangelands along a 900mm rainfall gradient, Semmartin et al. (2004) found that species increasing with grazing in a humid site decomposed faster than those decreasing with grazing. However, this pattern was reversed for semiarid and arid grasslands. On the other hand, shoot decomposition rates were higher in grazed grassland than in exclosures, but this was not true for root-derived litter (Semmartin et al., 2008). This suggests that historical grazing effects on the soil environment for aerial litter decomposition may play a crucial role in carbon dynamics. Large grazers can affect soil microorganisms and the processes they regulate through direct and indirect pathways involving changes in the quantity and quality of

Fig. 20|6. Direct and indirect mechanisms whereby grazing affects soil biota and litter decomposition. Numbers and line patterns indicate different effect pathways: direct pathway, through excretion and trampling (1, dotted lines); indirect pathways, through changes in plant productivity (2, dashed lines), plant resource allocation (3, chained lines), and plant community composition (4, solid lines). Litter quantity and quality are key intermediary factors of indirect pathways, comprising both shoot and root residues. 311

resources entering the soil (Bardgett and Wardle, 2003). Figure 20|6 depicts several key, interrelated mechanisms whereby grazing influences the soil biota and carbon cycling rates during litter decomposition. There can be positive, negative, or neutral effects of grazing on decomposition depending on the relative importance of these mechanisms. First, grazing animals directly alter the quality of resource inputs to soil by returning soluble organic matter in feces (Fig. 20|6), which represent a nutrient-rich, albeit localized, carbon source that stimulates microbial biomass (Bardgett and Wardle, 2003). Trampling by livestock may also alter the physical environment for decomposers (see sections above). This includes disturbance of biological soil crusts, which help to stabilize surface soil carbon in arid lands (Neff et al., 2005). Second, grazing controls the quantity of plant litter produced through its impact on above- and belowground NPP (Fig. 20|6). Long-term negative effects on NPP (see Milchunas and Lauenroth, 1993) may drive significant decreases in microbial biomass, perhaps to the extent of reducing carbon mineralization. Stimulation of aerial or root productivity (Doll, 1991) at moderate stocking rates, particularly in productive habitats, would cause the opposite effect by alleviating carbon limitation of soil microbes (Bardgett and Wardle, 2003). In addition, decreased plant cover by animal consumption and treading can drastically modify soil temperatures and moisture content (Lavado and Taboada, 1987) in ways that could affect microbial communities (Young and Ritz, 2005). Soil cover may further determine the extent of carbon release through physical degradation of litter by UV solar radiation, especially in arid zones (Austin and Vivanco, 2006). A third pathway by which grazing affects soil decomposers is by modifying the patterns of resource allocation of individual plants (Bardgett and Wardle, 2003; Fig. 20|6). Shoot herbivory in highproductivity grasslands on fertile soils often increases leaf nutrient content, thus enhancing litter quality and decomposition (Semmartin et al., 2008). Across terrestrial ecosystems, plant nutritional quality is positively correlated to the percentage NPP consumed by herbivores, and also to the percentage detrital production decomposed annually (Cebrian and

312

Lartigue, 2004). Thus, plant tissue quality may be regarded as an indicator of grazing control on plant biomass, detritus accumulation, and ecosystem carbon storage (see Cebrian, 1999). Further, herbivory may alter root exudation. Carbon allocated to root exudation may increase in grazed plants, which may then stimulate microbial biomass and activity in the rhizosphere, driving a feedback mechanism that accelerates nutrient return to plants and productivity (Hamilton and Frank, 2001; Bardgett and Wardle, 2003). However, the significance of this mechanism in managed grasslands remains to be tested. Finally, changes in plant community composition associated with grazing may strongly affect soil microbial communities and decomposition (Bardgett and Wardle, 2003). Plant species composition may influence decomposers through various pathways (Fig. 20|6). This mechanism operates over successional time scales, and should be most important when compositional changes involve a shift in life-form dominance, such as an increase in forb species at the expense of perennial C4 grasses (Garibaldi et al., 2007). It may lead to a reduction in average community litter quality when unpalatable, low-quality plants take over, or otherwise may increase litter quality inputs to soil when palatable, high-quality species remain dominant under grazing (Bardgett and Wardle, 2003; cf. Garibaldi et al., 2007). Invasion by exotic species plays a central role in these community changes. It appears that invasive plants generally accelerate carbon cycling, and may also increase plant and soil carbon stocks (Liao et al., 2008). However, the effect of woody invasions on soil organic carbon was found to shift from positive to negative along a decreasing precipitation gradient (Jackson et al., 2002). These authors showed that in drier grasslands carbon losses offset gains in vegetation due to woody species establishment. Furthermore, grazing has been found to alter the functional composition of the soil microbial community. In some systems, decomposition is dominated by bacteria in grazed sites, whereas it becomes dominated by fungi in ungrazed or lightly grazed areas. Hence, a fast carbon cycle would prevail in grazed sites as more energy goes through the “bacterial channel,” while a slow cycle


Soil Management Practices would dominate in the absence of grazing, as more energy is processed by the “fungal channel” (Wardle, 2005).

Soil Carbon Stocks: Management Implications

Overall, there is increasing empirical evidence suggesting that grazing effects on soil decomposers and the processes they regulate may vary predictably with habitat moisture and productivity. Positive grazing effects on decomposition would be common in fertile systems with high herbivore loads, whereas negative effects may be most common in low fertility systems that sustain low grazing pressures (Bardgett and Wardle, 2003). These contrasting ecosystem dynamics would also apply to early- vs. late-successional grasslands, which are alternatively dominated by fast-growing species with high litter quality and slow-growing plants with poor litter quality, respectively (Bardgett and Wardle 2003; Wardle, 2005). This broad trend may have important implications for the restoration of degraded rangelands, a major worldwide concern (Asner et al., 2004). Indeed, plant functional composition may determine the rate of carbon accumulation in soil, as recently shown by Fornara and Tilman (2008). These authors documented a 500% increase in soil carbon stored to 1-m depth after 12 yr of succession in mixed, C4 grass–legume communities, while communities lacking either plant group were much less efficient at storing carbon. They suggest that complementary resource use between grasses and legumes leading to higher total root biomass accrual was the mechanism responsible for the observed effect (Fornara and Tilman 2008). Hence, restoration of soil fertility and basic ecosystem functions (i.e., forage production) in overgrazed rangelands may need to consider the functional composition of the successional vegetation.

Grazing Impacts on Soil Nutrient Cycling

Grazing animals strongly affect the dynamics of nutrients through the soil– plant–animal system in many ways (cf. Fig. 20|6). They return to soil around 60 to 95% of the nutrients contained in the forage they ingest. These nutrients are usually returned

to the soil in a different chemical form compared with that of nutrients they absorbed by plants, thus significantly affecting the rates of nutrient cycling (Chaneton et al., 1996; Bardgett and Wardle, 2003). Grazers return nutrients as dung and urine forming localized excretal patches. Nutrient partitioning between dung and urine depends on the specific element and on other factors, notably the nutrient composition of the diet. Excretal patches cover only a small proportion of the pasture area, which increases the spatial heterogeneity of nutrient inputs to soil and availability to plants. Impacts of grazing are also evident in nutrient losses, which are concentrated in excretal patches and comprise both gaseous and leaching outputs. In this section, we discuss aspects specific to the influence of grazing animals on nutrient cycling. We focus on N and P, which are the main limiting nutrients to forage production in most grazing systems worldwide.

Nutrient Budgets in Grazing Systems

The demand for nutrients by grazing animals is relatively stable. For example, values of 25 g N kg−1 and 8 g P kg−1 retained by cattle adult body mass, 30 g N kg−1 and 8 g P kg−1 in the calf body mass, and 6 g N kg−1 and 1 g P kg−1 in the milk mass are generally accepted (ARC, 1980). However, it is very difficult to make generalizations on the amount of nutrients removed by grazers because they are largely determined by the total amount of ingested forage. The consumption of forage is intrinsically variable because it is regulated simultaneously by the availability of forage, the farming management and the rate of forage utilization, among other factors. Forage availability is highly variable because it depends on unpredictable factors such as rainfall. As a consequence, the total amount of nutrients consumed by grazing animals, and hence the role of grazing on nutrient budgets, may be difficult to predict accurately (Batabyal, 2007). For example, in a Flooding Pampa grassland under low stocking rate, Chaneton et al. (1996) found that the N added from rainfall (7 kg yr−1) was sufficient to compensate for the N exported in animal body mass. In contrast, a continuous loss of P (0.4 kg ha−1 yr−1) was estimated for the same nonfertilized grassland. Haygarth et al. (1998) compared the P budgets for two 313

contrasting grassland farming systems: a dairy farm and an extensively managed hill sheep farm. They observed a high rate of soil P accumulation (26 kg ha–1 yr–1) in the dairy farm, which received an annual input of 16 kg P per ha−1 from P fertilization. This surplus was related to the feed concentrates, which represented around 27 kg ha−1 yr−1. The low stocking rate sheep farm, contrastingly, showed an almost equilibrated P budget (+0.28 kg ha−1 yr−1). In their study of Río de la Plata grasslands, Piñeiro et al. (2006) estimated that, in the long term, grazing altered the N cycling by accelerating returns via animal depositions and increasing N outputs by volatilization and leaching. As a result, soil N storage diminished 19% and N outputs increased 67% after four centuries of simulated domestic livestock grazing.

Nutrient Recycling by Grazing Animals

Proportion of Nutrients Returned to Soil

Although it is difficult to make quantitative generalizations on nutrient budgets in grazing systems, some common qualitative patterns emerge if we analyze the fate of nutrients after they are ingested by livestock. In ungrazed systems, nutrients taken up by plants are returned to the soil largely through the litter deposited on plant senescence. These nutrients may be re-used by plants and soil microorganisms. When livestock is introduced to the system, only a fraction of the plant nutrients are returned to the soil via the plant detrital pathway (cf. Fig. 20|4). Most nutrients go through the more complex path in the animal before they are returned to the soil. The amount of nutrients that are returned to the soil depends largely on the type of animal and the farming management (Mohamed Saleem, 1998). Depending on the grassland productivity and herbivore management, the rate of forage utilization will determine the grazing productivity and nutrient cycling. Only a small fraction of the ingested nutrients are retained in animal products. Dahlin et al. (2005) reported that only 10 and 13% of the ingested N and P are retained by the animal in low-input grazed pastures, so that the percentage of ingested nutrients

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returned to the soil was about 90 and 87% for N and P, respectively. These figures are lower if dairy cows are considered. In such a case, a substantial part of the ingested and retained nutrients is exported in the milk. For example, Haynes and Williams (1993) calculated that as much as 10 and 25% of the ingested N and P are contained in the exported milk. If these proportions are coupled with the nutrients retained by the animal, the percentage of ingested nutrients returned to the soil lowers to 80 and 65%, for N and P, respectively. The partitioning of nutrients between dung and urine differs broadly among nutrients and grazing systems. Some nutrients are excreted in a rather equivalent proportion between dung and urine, others preferentially in dung, and others in urine. Nitrogen is excreted in a rather equivalent proportion in both dung and urine. Sodium, Cl, and S are also partitioned significantly in both ways. Phosphorous is excreted mainly in dung, with reported values ranged around 95% of the total deposition (Haynes and Williams, 1993). Only a small amount of P is found in urine, although the proportion can increase if the diet is rich in this element. Calcium, Mg, Zn, Cu, Fe and Mn are also predominantly excreted in the solid feces. Potassium is the only nutrient predominantly excreted in urine (Haynes and Williams, 1993).

Forms of Nutrients Returned to Soil

Grazing may have a positive effect on the nutrient cycling of the soil. Compared with above- and belowground dead plant materials, animal excretions have higher concentrations of nutrients than the nondigested materials. More than half of the N in the urine is constituted by urea, which is rapidly hydrolyzed to ammonium in the excretal patch. This ammonium is highly susceptible to volatilization, causing gaseous N losses, which can be small (Chaneton et al., 1996) or large (Piñeiro et al., 2006), relative to internal nutrient flows, depending on the time-frame of analysis. The remaining proportion of urine-N is predominantly constituted by amino acids and peptides. In an elegant work, Clough et al. (1998) studied the fate of the 15N labeled urine applied to four soils. Plant N uptake was the predominant sink of urine-N, accounting from 22 to 31% of the applied N. Nitrogen retained in


Soil Management Practices the soil ranged from 20 to 24%, almost all in organic compounds. Leachate accounted from 13 to 32% whereas measured gas loss was less than 4% in all cases. Phosphorous is returned to soil mainly in inorganic forms, dicalcium phosphate being the predominant P form in dung. Results from Sharpley and Moyer (2000) show consistently more inorganic P (63%) that organic P in dairy manure. Most of the inorganic P (around 80%) was water extractable. The total P in feces increases as the P content of the diet increases, but particularly the amount and proportion of water-soluble inorganic P compounds. For instance, Dou et al. (2002) found that increasing the proportion of P in the diet from 3.4 to 6.7 g P kg−1 feed dry matter caused an increase in the water-soluble fraction of fecal P of 2.9 to 10.5 g kg−1 fecal dry matter, representing 56 and 83% of acid digest P. The water-soluble fraction of fecal P is the most susceptible fraction to loss in the environment, which suggests that higher P contents in the diet lead to an increase in the susceptibility to environmental losses (Sharpley and Moyer, 2000). In such cases, Dou et al. (2002) suggest a measure of 2 g P kg−1 fecal dry matter, as inorganic P in a 1-h extract, to be a simple fecal P indicator that meets lactating cow requirements with minimal excess. The other fecal P fractions, including organic P and less soluble inorganic P, are less affected by dietary P concentration (Dou et al., 2002). Organic P forms in the dung contain inositol hexaphosphates, among other compounds. The proportion of organic P in the animal excretions is commonly lower than the proportion of organic P in the ingested forage. This means that a significant mineralization of ingested organic P is produced during passage through the animal. Since plants absorb mainly soil P inorganic forms, grazing animals are expected to exert a positive effect on P cycling. Accordingly, Aarons et al. (2004) observed that dung P contributed only to increases in available inorganic P, with extractable organic P unchanged in the short term.

Grazing Effects on Plant Litter Quality

Litter quality can be significantly affected by changes in the grassland species

composition promoted by grazing (Bardgett and Wardle, 2003). Published evidence reveals that the pattern of response is far from universal and depends largely on the functional groups involved in the species replacement that is promoted by grazing. In humid grassland, it was found that grazing promoted a change in the botanical composition with an increase of forb species and a decrease of the more palatable graminoids (Semmartin et al., 2004). The forbs produced litter with higher N and P contents and higher N mineralization rates than the graminoids, and this coincided with higher soil N availability in grazed than in ungrazed sites (Garibaldi et al., 2007). In contrast, Ritchie et al. (1998) observed that grazing decelerated the cycling of N and lowered soil N availability by reducing the proportion of highly palatable legumes. Therefore, the particular group of species that are diminished and promoted by grazing must be taken into account when analyzing the consequences of plant–herbivore interactions for the rates of nutrient cycling in grasslands.

Spatial Redistribution of Soil Nutrients

In their excellent review, Haynes and Williams (1993) found a range of 8 to 12 urinations per day for cattle and 18 to 20 for sheep. Each urination event ranged from 1.6 to 2.2 and 0.10 to 0.18 L for cattle and sheep, respectively. In regard to the number of defecations, these authors reported a range of 11 to 16 and 7 to 26 for cattle and sheep, respectively, with a range of 1.5 to 2.7 and 0.03 to 0.17 kg per defecation. These quantities are influenced by environmental and farming conditions. For instance, the number of urinations is largely affected by water intake and water content of the diet. The surface covered by a single urination event is 0.16 to 0.49 m2 for cattle and 0.03 to 0.05 m2 for sheep. In the case of dung, the area covered by a single defecation event ranges from 0.05 to 0.09 m2 for cattle and 0.008 to 0.025 m2 for sheep (Haynes and Williams 1993). This great range for sheep is because the fecal depositions are in the form of pellets that are often distributed over a wide area. Excretion patches are distributed nonuniformly over the soil surface, which may determine two scales of spatial heterogeneity. At a broader scale, soil nutrient 315

concentrations are particularly high at stock camps (i.e., areas within the paddock where animals congregate spontaneously), around water sources and fences, beneath trees, and areas where supplemental feed is provided. Dahlin et al. (2005) found an average transfer to stock camps of 32 and 22% (percentage of ingested nutrients) for N and P, respectively. This translocation increases the risk of nutrient losses, which is aggravated by the fact that those areas are usually trampled and, consequently, forage production and nutrient uptake is small. In such sense, stock camps are highly enriched with nutrients and constitute “´hot spots”´ for nutrient losses, especially mobile nutrients, such as N. Less mobile nutrients, such as P, are retained by the soil and may become available to the herbage in subsequent growing seasons (Oenema et al., 2006). In the remaining areas of the paddock there is a pattern of spatial heterogeneity at a smaller scale. This is primarily due to the deposition of excreta in patches rather than evenly over the field. A grazed grassland paddock consists of a mixture of aggregates of grazed and ungrazed microsites. Excretal patches are old, fresh, isolated, overlapped, or mixed urine and dung depositions covering 40% of the total area (Lantinga et al., 1987). West et al. (1989) found that grazing markedly affected the pattern of spatial variability of N and P, which is highly dependant on the stocking rate. For example, Lavado et al. (1996) attributed the lack of spatial variability on soil N and P in a native grassland devoted to calf production to the low livestock density, but also to the size of the paddock (800 ha) and to the distance from the study site to the water source (1200 m). Therefore, it is expected that nutrient cycling will display high spatial variability in grazed grasslands. This spatial heterogeneity creates technical and logistical problems that make the study of nutrient dynamics difficult. For this reason, modeling is a natural alternative for the study of nutrient cycling in grazed pastures (Hutchings et al., 2007).

Management Strategies to Increase Nutrient Use Efficiency

Rotational grazing is an effective practice to increase forage utilization efficiency, minimize daily variations in intake and

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quality feed, and allocate pasture to animals more efficiently, based on nutritional needs. With large pastures, the animals decide where, when, and how long to graze. Rotational grazing reduces the size of the pasture and transfers these decisions from animal to manager and usually results in more efficient use of the pasture resources and nutrient cycling. Rotational grazing contributes to minimizing camping and to spreading out animal depositions over large areas and thus reduces heterogeneity and nutrient losses (Oenema et al., 2006). The components of a good rotational grazing system are a balanced forage system, an electrical fencing system, and spatially distributed water and shade supply. An interesting alternative to farms that use both stable and direct grazing is the prototype system for the realization of very high environmental standards that has been developed in the Netherlands (Oenema et al., 2006). The system consists of grazing for a restricted number of hours alternated with periods in the stable for supplementation. This practice reduces the fraction of dung and urine deposited on the soil, thus reducing field heterogeneity. The excreta collected in the stable can be distributed through appropriate application technology to the pasture at the right amount and spatial distribution.

Concluding Remarks We have discussed how grazing affects soil physical properties, carbon, and nutrient cycling in forage production systems. These effects vary across the different climates, soil types, grazing systems, and livestock types. Nonetheless, some broad generalizations can be pointed out: 1. Grazing changes soil physical properties by defoliation and treading. Defoliation decreases ground cover, which increases topsoil temperature and evaporation rates, and causes salinization in saltaffected areas. Animal treading often results in shallow compaction when soil is trampled moist, and causes puddling when soil is trampled wet. Whereas soil infiltration rate is highly sensitive for detecting these changes, particularly in


Soil Management Practices bare ground areas, the effect of grazing on soil water content is hard to predict. 2. Structural damage by grazing can be restored by natural regenerative processes, which may take many years in dry climates. Faster responses are possible in temperate climates, because of the action of frequent wetting and drying cycles, freezing and thawing cycles, and vigorous pasture root growth. 3. The balance between soil carbon inputs and outputs in grazed systems varies widely among systems, making it difficult to make generalizations about the effect of grazing on soil carbon dynamics. However, the size of major carbon pools and fluxes in rangelands can be strongly controlled by grazing. Grazing alters litter decomposition and the functioning of the soil microbial community, and these effects are to a large extent mediated by plant community turnover. Several lines of evidence suggest that grazing effects on soil carbon and nutrient cycling may vary predictably along habitat moisture/ fertility gradients. 4. The partitioning of nutrients between dung and urine differs among nutrients and grazing systems. Nitrogen is excreted in a rather equivalent proportion in dung and urine, whereas most P returns to soil in dung paths. Grazing alters N and P cycling by accelerating returns via animal depositions, and by increasing N outputs through volatilization and leaching. 5. Rotational grazing is an effective practice to increase forage utilization efficiency and reduce the fraction of dung and urine deposited on the soil, thus reducing field heterogeneity. However, rotational grazing has less clear consequences on the recovery of soil physical properties and carbon stocks. 6. Soil protection by vegetation is most important in semiarid and desert grasslands, where decreased infiltration rates by grazing are closely associated with bare ground areas. This objective can be attained by excluding grazing for several years (rest periods). However, restoration of soil organic matter in degraded arid rangelands would take exceedingly long (and economically unsustainable) periods of livestock exclusion.

Acknowledgments We thank the Consejo Nacional de Investigaciones Científicas y Técnicas (CONICET), Agencia Nacional de Promoción Científica y Tecnológica (FONCYT grant scheme), and University of Buenos Aires for funding our research during the preparation of this chapter.

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