The role of ectomycorrhizal fungi in mediating nutrient cycling in temperate and boreal forest ecosystems

~laura dev 2013~

Introduction

In systems where nutrients are limiting, such as boreal and temperate forests, mycorrhizae play a particularly important role in mediating the nutrient cycling of the ecosystem. They are allowed this control because they are brokers of certain limiting resources, which they are adept at scavenging from the soil (Read, Leake, & Perez-moreno, 2004). All tree species form mutualistic associations with mycorrhizal fungi (Simard, 2009), and for a given tree species there are many fungal species with which it can form associations.

Mycorrhizae act as an extension of the root system into the bulk soil, increasing the absorbing length of roots (Chapin, Matson, & Mooney, 2002) and allowing a greater volume of soil to be exploited. The fungal hyphae can also bridge the zone of nutrient depletion that tends to form around roots, and enter soil pores that are too small for plant roots to access (Godbold, 2004). In exchange for the provision of nutrients, plants invest upwards of 30% of their carbon (C) in supporting mycorrhizal symbionts.

Underground networks of mycorrhizae can connect different plant individuals and even species within a forest (Simard, 2009; Simard et al., 2012). Carbon, nutrients, and water can be transferred among the individuals connected by a common mycorrhizal network according to source-sink gradients. These networks have been shown to facilitate plant species coexistence, regeneration after disturbances, and the persistence of seedlings growing in shaded areas. The presence of mycorrhizal networks not only impacts plant productivity and species composition, but also has great influence on the amount of carbon that is able to be stored in the soil. This review will discuss ways in which mycorrhizae influence the functioning of temperate and boreal ecosystems and how mycorrhizal presence serves to tighten the cycling of nutrients within nitrogen-limited ecosystems. It will conclude with a discussion of recent methodological advances used to study mycorrhizae and a few possible directions for future research.

Discussion

Nutrient cycling in forest ecosystems is highly localized within the root zone of plants, and is driven by the release of nitrogen (N) and phosphorous (P) from organic sources such as amino acids, proteins, and leaf litter by enzymes secreted by microorganisms, including mycorrhizae (Schimel & Bennett, 2004). It was previously thought that the mineralization of nutrients from the organic polymer form was only done by free-living decomposers in the soil, but ectomycorrhizae have shown a widespread ability to carry out this function as well (Read & Perez-moreno, 2003). Indeed, ectomycorrhizae are critical for N recycling and acquisition in boreal and temperate forests (Schlesinger & Bernhardt, 2013), directly mobilizing nutrients from organic sources by excreting extracellular phosphatase and cellulase enzymes (Lambers, Chapin, & Pons, 2008). These nutrients are absorbed by the fungal hyphae and may then be transferred to the symbiotic plant hosts (Perez-moreno & Read, 2000).

The ability of ectomycorrhizal fungi to mineralize organic nutrients was most clearly demonstrated in a study of birch (Betula pendula) seedlings inoculated with the ectomycorrhizal fungus Paxillus involutus. Seedlings were planted in transparent observation chambers in factorial treatments of mycorrhizae inoculation and litter addition (Perez-moreno & Read, 2000). The presence of the mycorrhizae reduced litter nutrient concentrations, particularly P, indicating that the fungus was able to mineralize organic nutrient supplies. Concurrently, the birch seedlings with mycorrhizal inoculation treatments had significantly increased biomass production and tissue nutrient concentrations with litter present compared to treatments without litter, implying that the organic compounds were passed from the mycorrhizae to the tree seedlings.

The ability of mycorrhizal fungi to mobilize recalcitrant nutrient sources allows plants to survive in systems where rates of N mineralization would not otherwise be fast enough to meet plant requirements. In the pygmy forests of California, the soil is extremely infertile, acidic, and N-limited. The Pinus contorta trees in this environment have formed a tight nutrient cycle with their fungal symbionts, Amanita muscaria. The plants immobilize N in their tissues by producing high amounts of tannins (Northrup, Yu, Dahlgren, & Vogt, 1995). Tannins are very recalcitrant complexes, which lock up N in a form that is then only available to plants via mineralization by ectomycorrhizal fungi. When the tannin-rich litter is deposited in the soil the tannins adsorb to soil surfaces and are then able to be taken up by the plants’ own mycorrhizal symbionts. In this way the trees have adapted to maximize their chances of recovering the N lost in their own litter.

In environments with slow rates of mineralization, fungal associations are particularly important for plants to acquire adequate nutrient supplies (Godbold, 2004). Mycorrhizae are more abundant and have greater N mineralizing activity in areas with fewer N inputs and less nitrification, such as the more northerly boreal forests compared with temperate forests (Read & Perez-moreno, 2003). Further north, the soil is often more acidic, with fewer nutrient resources and therefore less mineralization in the absence of the fungi. This has been demonstrated using N isotopes to illuminate pathways of N transfer (Hobbie, Macko, & Williams, 2000; Hogberg et al., 1996). Natural abundances of 15N are altered when they pass through different transformations. When transferred from mycorrhizae to plants, the lighter 14N isotopes are preferentially passed to plants, causing depletion in foliar δ15N signatures (Hogberg et al., 1996). When this material is returned to the soil as plant litter, it results in deeper soil layers having greater 15N enrichment and surface layers being more depleted in 15N. An experiment in coniferous and broadleaved forests in central and northern Europe showed this effect to be most pronounced in N-limited forests (Hogberg et al., 1996). Hobbie et al (2000) got a similar result from a study conducted in Alaskan boreal forests. They found that trees in older sites with low N concentrations had lower foliar δ15N.

Both studies indicate that a greater proportion of plant nutrition was obtained by way of fungal symbionts when N was scarcer. This trend was also correlated with a decline in total foliar N concentrations, likely due to increased allocation to belowground C (Hobbie, Macko, & Shugart, 1999). This evidence suggests that as N availability increases, foliar N concentrations also increase, while mycorrhizal fungi decline. Therefore, boreal forests tend to have greater dependence on mycorrhizae for the mineralization of organic nutrient sources, and their associated fungi often have increased abilities for polymerase function.

Mycorrhizal networks can serve to transfer carbon, nutrients, and water to younger trees to help with growth and establishment (Simard, 2009). For example, seedling survival and growth rate in Douglas fir (Pseudotsuga menziesii) is improved when connected to mycorrhizal networks containing large trees (Simard et al., 2012), though the mature trees themselves have competitive effects on seedlings. This creates an area of maximum seedling performance in a circle around mature trees, but out of the way of competitive effects (Buscardo et al., 2012). These networks can prevent nutrient leaching from the system by taking up nutrients and distributing them where most needed based on source-sink gradients.

Future directions

Recent technological advances have greatly improved our ability to study the role of mycorrhizae in nutrient cycling. For example, culture-independent DNA analysis has been key in identifying fungal symbionts to genus or species, which was previously impossible (Simard et al., 2012). Advanced microscopy techniques have also been helpful in elucidating mycorrhizal physiology. Still, of the 5000 species that can form mycorrhizal associations, only a small number of these have been studied because of the difficulty of culturing them. In situ studies will be important for studying fungi that are not culturable (Read & Perez-moreno, 2003).

In light of projected changes in climate, it will be necessary to estimate the importance of mycorrhizal-mediated nutrient cycles for soil C storage. Since mycorrhizae increase the proportion of C in the soil relative to N and P concentrations, they contribute to soil carbon retention (Read & Perez-moreno, 2003). Warmer temperatures may disrupt the tight cycling of boreal forest nutrient cycling by diminishing the performance of extracellular enzymes, which function optimally at low temperatures. This could cause systems to lose more N and decrease their ability to store C. Increased rates of N deposition will also affect the function and abundance of mycorrhizal networks. Since mycorrhizal associations are favored by low-nutrient systems, increased N loads could decrease dependence on mycorrhizae.

Conclusion

Symbiotic mycorrhizae control nutritional processes, productivity, and species composition in boreal and temperate forests. These fungi provide a shortcut in nutrient cycling by accelerating the mineralization of organically-bound nutrients and providing them to plants faster than would be possible via diffusion in the soil (Lambers et al., 2008). Though mycorrhizal associations are critical for plant survival in systems that are limited by nutrients, their importance declines as N availability increases. Complex, belowground mycorrhizal networks influence the sink-source balance of carbon in the system by mediating storage and transfer of carbon and nutrients, thereby determining to a large extent the quality of the plant tissues that are returned to the soil in the form of litter (Read et al., 2004). Understanding how mycorrhizal networks are regulated will greatly improve our understanding of forest ecosystem processes.

Literature Cited

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Hobbie, E. A., Macko, S. A., & Shugart, H. H. (1999). Insights into nitrogen and carbon dynamics of ectomycorrhizal and saprotrophic fungi from isotopic evidence. Oecologia, 118(3), 353–360.

Hobbie, E. A., Macko, S. A., & Williams, M. (2000). Correlations between foliar d15N and nitrogen concentrations may indicate plant-mycorrhizal interactions. Oecologia, 122(2), 273–283.

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Northrup, R. R., Yu, Z., Dahlgren, R. A., & Vogt, K. A. (1995). Polyphenol control of nitrogen release from pine litter. Nature, 377(21 September), 227.

Perez-moreno, J., & Read, D. J. (2000). Mobilization and transfer of nutrients from litter to tree seedlings via the vegetative mycelium of ectomycorrhizal plants. New Phytologist, 145(2), 301–309.

Read, D. J., Leake, J. R., & Perez-moreno, J. (2004). Mycorrhizal fungi as drivers of ecosystem processes in heathland and boreal forest biomes. Canadian Journal of Botany, 1263, 1243–1263.

Read, D. J., & Perez-moreno, J. (2003). Mycorrhizas and nutrient cycling in ecosystems: A journey towards relevance? New Phytologist, 157(3), 475–492.

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How does the timing of precipitation affect ecosystem processes in arid and semi-arid grasslands?

~laura dev 2012 (Chapter 1 of M.S. Thesis)~

Introduction

Precipitation is a strong driver of ecosystem structure and function among all terrestrial biomes. Grassland systems, particularly those that are water-limited, can be very sensitive to temporal changes in precipitation as well (Knapp and Smith 2001), and the distribution of precipitation in relation to growing season temperatures may determine some distinguishing ecosystem properties (Prentice et al. 1992). While mean annual precipitation is most often used to describe ecosystem water relations, it is becoming increasingly clear that the temporal dynamics of precipitation, such as seasonal distribution and the size and frequency of events can significantly modify ecosystem response to total precipitation quantity (Swemmer et al. 2007). I will review the differences between ecosystem properties under climates characterized by primarily summer precipitation (coupled temperature and precipitation) versus climates characterized by dry summers and wet winters (decoupled), and will also explore how climate change will potentially alter ecosystem processes under future predicted precipitation regimes.

Future climate models unanimously report changes in the temporal distribution of precipitation worldwide, although there is great uncertainty surrounding the direction of these changes (Christensen et al. 2007). Nonetheless, it is expected that in many places these changes will take the form of increases in extreme events and time between events, as well as shifts in the seasonality of precipitation. In arid and semi-arid environments, temporal changes in precipitation may result in more dramatic changes in ecosystem functioning than both increasing temperature and increasing CO2 levels, and is expected to interact with these factors as well (Weltzin et al. 2003). Therefore, there is some urgency in synthesizing how the temporal distribution of precipitation affects ecosystem processes, which has been the subject of much recent research.

Although precipitation most directly influences soil moisture, the indirect effects on plant and microbial communities may be more relevant (Heisler and Weltzin 2006). The timing of precipitation can directly influence abiotic soil processes such as drainage, infiltration, evaporation, soil temperature, and water availability for uptake by plants (Austin et al. 2004). These processes in turn affect the biotic processes of production, carbon and nutrient cycling, decomposition and microbial, plant, and animal species composition. This paper will specifically focus on the effects of precipitation timing on biotic processes in arid and semi-arid grassland ecosystems.

Review

Primary Production

It is a well-established relationship that at regional scales aboveground net primary productivity (ANPP) increases with precipitation quantity (Sala et al. 1988). However, much variation in ANPP cannot be explained by variation in precipitation quantity alone. Temporal variation in ANPP is linked with precipitation variability, but is also constrained by production potential (Knapp and Smith 2001). On global scales total precipitation and therefore production potential tends to be inversely correlated with variability in precipitation inputs (as measured by CV). The highest ANPP variability occurs in grassland biomes because they generally have intermediate production and intermediate precipitation variability. ANPP variability in very arid biomes is constrained by low production potential, whereas ANPP variability in mesic biomes tends to be low because there is relatively low precipitation variability. Grasslands are therefore the most responsive to changes in precipitation inputs, though they tend to be more responsive to increases rather than deficits in precipitation. This may be due to drought adaptations in many grassland species. Regression models of grassland ANPP are greatly improved by including precipitation variability measures as explanatory variables in addition to precipitation amount (Nippert et al. 2006).

At the local level, these broad-scale relationships are often weaker. Even among sites with the same dominant species, interannual variation in ANPP can be explained by different mechanisms (Swemmer et al. 2007). ANPP variation at mesic sites tends to be more dependent on the length of time between precipitation events, indicating low tolerance for drought stress. Meanwhile, ANPP at arid sites can be more constrained by event size and number. In these sites water is likely limiting, and increases in event size and number can increase production up to a saturating point over which the precipitation cannot be converted into new growth. Therefore, mean soil water content is insufficient to explain annual and seasonal variation in ANPP, which is often more responsive to temporal soil moisture availability (Knapp et al. 2002).

As climate changes, precipitation event sizes and time between events are expected to increase outside of the range of historic variability, and examining records for ANPP and precipitation relationships may not allow for extrapolations to future precipitation regimes (Nippert et al. 2006). Experimentally increasing the variability in precipitation of grasslands outside of the historic range tends to decrease ANPP, though not in all systems (Knapp et al. 2002, Fay et al. 2003, Nippert et al. 2006). In one study, reduction in ANPP due to increasing variability was found to be roughly equal to results obtained from a 30% reduction in annual precipitation (Fay et al. 2003). This could be related to decoupling the water supply and evaporative demand, putting more stress on the plants and decreasing root activity. However, increased variability has been shown to increase root to shoot ratios meaning that total production might remain constant while allocation is shifted belowground.

Time lags in production response to variation in precipitation inputs add further complexity, and can occur on both monthly and annual time scales (Ma et al. 2010). Monthly aboveground production is related to both precipitation and temperature of the preceding months, whereas interannual variations in ANPP are correlated with the previous-year precipitation, the effects of which can be stronger if certain conditions such as drought persist for multiple years. Seasonally, this means that higher winter and spring precipitation will result in more productive early growing seasons with dry summers and falls decreasing late growing season production.

Production can also vary depending on the seasonality of precipitation. Systems whose temperature and precipitation peaks are temporally decoupled (decoupled systems) tend to be responsive to different inputs than systems whose precipitation inputs peak during the summer (coupled systems). Production in decoupled arid systems, such as the Colorado Plateau, has been found to be more sensitive to winter drought than summer drought, probably because plants there rely on winter precipitation for the majority of their ANPP and do not appear to use summer water inputs (Schwinning et al. 2005b). Summer monsoon rainfall is not as consistent as winter precipitation, so plants in decoupled arid climates may not have adapted to use these growing-season inputs. However, in some less arid systems the previous year’s summer precipitation is a greater driver of early season production than winter precipitation (Morecroft et al. 2004), meaning that a decoupling of precipitation and temperature patterns could result in overall decreases in annual productivity if the decreases in late season biomass production are not equally compensated by increases in early season production.

The seasonality of precipitation can also affect how much water is lost from the system. Systems with decoupled precipitation and temperature regimes tend to have lower evapotranspiration rates than coupled systems because there is less evaporative demand during the wet season. Site-level differences in ANPP response to precipitation inputs may be attributed to differences in soil characteristics. For example, in arid grassland systems, soils with a coarser texture tend to lose less water to evapotranspiration due to deeper penetration, and thus have higher ANPP (Sala et al. 1988, Austin et al. 2004). However, as systems become more mesic this relationship reverses, and coarser soils actually lose more water to leaching.

Carbon Cycling

In addition to production and the implications that this has for carbon (C) sequestration, the timing of water inputs can also have great effects on respiration and decomposition, with significant consequences for ecosystem C cycling. Since decomposition is fostered by coinciding warm and wet conditions, systems whose temperature and precipitation patterns are coupled tend to experience faster rates of decomposition. In turn, shifts in precipitation that create sufficiently warm and wet conditions for a greater proportion of the year are likely to result in increasing decomposition rates and loss of terrestrial soil organic carbon (SOC) stores (Aanderud et al. 2010). However, in systems that are naturally coupled, adding winter moisture tends to have a greater impact on CO2 flux than adding summer precipitation, causing a net C gain with increased winter precipitation (Chimner et al. 2010). This result indicates that changes in the quantity of dry-season precipitation can have greater effects than changes in the amount of wet-season or total precipitation. If winter precipitation takes the form of snow, insulating effects on soil temperatures may increase CO2 loss due to increased microbial respiration (Walker et al. 1999), though this has not been found to be significant in all systems (Chimner and Welker 2005).

There may also be a lagged effect of seasonal variation in precipitation on C flux. For example, increasing winter moisture in the mixed grass prairie of WY resulted in greater CO2 flux during midsummer (Chimner and Welker 2005). This is likely due to increases in soil moisture and gross photosynthesis, and this result is enhanced following summer droughts (Chimner et al. 2010). In contrast, in arid systems, drought in the previous wet season can decrease CO2 flux after a rewetting event, resulting in a net accumulation of C in the system (Potts et al. 2006). This is due to an increase in the efficiency with which the system uses this water pulse. In these particularly arid climates, decreased substrate and soil nitrogen availability following drought likely constrain microbial activity after a rewetting pulse. Although primary production is also constrained by drought, at the ecosystem level the decrease in microbial respiration seems to be greater than the decrease in photosynthesis. This suggests that plant and microbial communities have differing responses to rainfall variability.

In coupled systems, with mostly summer precipitation, CO2 fluxes tend to be largest during the early summer, though increased variability of precipitation significantly reduces flux during this time of year (Harper et al. 2005). In decoupled systems, with relatively dry summers, early season respiration is mostly attributed to plant growth, and once the soils dry down respiration stops until pulses of precipitation stimulate large increases in heterotrophic respiration and ecosystem C loss (Xu et al. 2004). Large increases in microbial respiration after rewetting could be due to soil aggregates breaking up and releasing more available labile C (Harper et al. 2005). Indeed, sites with higher labile organic C, such as high production sites and those with fine-grained soil, tend to be more sensitive to water pulses, resulting in increased C losses due to heterotrophic respiration (Austin et al. 2004, Xu et al. 2004). However, after several subsequent rain events this respiration pulse has been shown to decline, indicating that all labile substrates may have been exposed during the preceding events (Zhang et al. 2010).

The size of the respiration pulse after rewetting is dependent on species composition (Norton et al. 2008), how dry the soil is initially, and the size of the rain event (Xu et al. 2004). The size of the rain event is generally positively correlated with soil moisture and CO2 exchange (St Clair et al. 2009), and small pulses may be too small to affect soil moisture or stimulate physiological activity in the form of CO2 flux and primary production. The minimum threshold amount of rainfall to stimulate activity can vary throughout the growing season. A case study in Mongolia (Hao et al. 2009) determined that during the early season, when the canopy was not as filled in, small events (<5mm) could stimulate production, but once leaf area index (LAI) increased later in the season, events smaller than 5mm did not penetrate deeply in the soil, and could only stimulate CO2 loss through microbial respiration. However, this threshold differs among systems, and can be influenced by the decoupling of nutrient supply and demand as well as rooting depth of plants.

Less frequent events tend to increase cumulative CO2 loss significantly more than more frequent events of the same size (Zhang et al. 2010). In fact, a high frequency of rainfall events can actually increase C storage in the system. Microbial biomass responds inversely to the frequency of rain events, possibly contributing to greater C loss during infrequent rewetting events. Nonetheless, experimentally increasing rainfall variability, creating more extreme but less frequent events, has been shown to result in significant net decreases in annual CO2 flux (Harper et al. 2005). This is probably because decreases in CO2 flux during prolonged dry periods outweigh the temporary increases after rewetting. Respiration responses are related most strongly to surface soil moisture (Qi et al. 2010), so the drying of the top layer during dry intervals, and the deeper percolation of larger events will cause a much lower mean surface soil moisture during the growing season.

Reductions in respiration and CO2 uptake by plants are both greater in response to variability in soil moisture than to mean soil water content, indicating that precipitation variability may be more important than total precipitation in terms of ecosystem C exchange (Knapp et al. 2002). However, total annual precipitation does modify the magnitude of both respiration and photosynthesis responses, with lower precipitation systems showing greater sensitivity to soil drying (St Clair et al. 2009). Decreases in C flux are likely mediated by plant responses, with lower root mass, root respiration, and root exudation under more variable rainfall regimes, which in turn reduce C inputs into the rhizosphere (Harper et al. 2005). Interestingly, it has been shown that soils experiencing increased variability in precipitation were less sensitive to temperature than under ambient rainfall conditions, meaning that the timing of precipitation may be important in modifying effects of warming.

Nutrient Cycling and Microbial Community

Because nitrogen (N) availability is tightly coupled with soil moisture, changes to water inputs can have strong effects on N cycling, and many of these effects can be attributed to changes in microbial communities and populations. During extended dry periods, N tends to build up in the system due to decreased diffusion and increases in low C:N substrates from microbial deaths (Norton et al. 2008). Rewetting therefore leads to a pulse of available N to plants. Phosphorus can also become more available after a water pulse. Snow can impact N cycling as well, increasing rates of N mineralization and net nitrification (Walker et al. 1999).

Water pulses lead to increased N fluxes with more losses due to volatilization and leaching, as well as higher rates of mineralization due to increased microbial activity (Austin et al. 2004). This mechanism is supported by observed increases in microbial biomass during the first 8 hours after a water pulse, after which it declines (Norton et al. 2008). Plant species composition seems to mediate different rates of decline following the initial microbial growth period. Plant species composition can affect soil nitrification processes as well, with certain species accelerating nitrification rates and gaseous N2O emissions following water pulses. Mineralization pulses are especially large in finer textured soils with larger labile C and N pools. Coarser textured soils on the other hand tend to lose more N due to leaching, so the mineralization pulses are not as large. When precipitation falls primarily during the winter, nutrient supply and plant demand are decoupled, causing even greater N losses due to leaching (Austin et al. 2004).

The timing of precipitation may affect microbial community composition. Although much remains unknown about this, we do know that different environmental conditions, particularly with regard to soil moisture, can favor different types of soil microbes. Fungi are more tolerant of desiccation and have higher N-use efficiency than bacteria, and thus, tend to be more abundant during the dry season, which can lead to slower rates of N mineralization (Austin et al. 2004). However, summer droughts and winter warming have been shown to affect mycorrhizal fungal associations and result in increased colonization of plant roots, but decreased hyphal densities (Staddon et al. 2003). While the decline in hyphal density is likely a direct effect of decreased water availability, the other effects may be indirect effects of changes in plant species composition caused by drought and warming.

Plant Community Composition

Changes in precipitation regimes may affect plant diversity and species ranges, as well as alter competitive dynamics between species. While the effect of changes in competitive abilities depends on how strongly competition is affected, as well as the sensitivity of the species in question (Levine et al. 2010), if competitive advantages change along with climate, rapid shifts in community structure and function could occur. Furthermore, changes in precipitation may interact with other climatic changes such as increasing temperatures and atmospheric CO2 concentrations to further drive changes in species composition (Tietjen et al. 2009). This could have dramatic effects on biodiversity of animals as well as plants, if habitat or forage quality is decreased, and may also feedback to affect local and regional climate if albedo and evapotranspiration are altered significantly (Weltzin et al. 2003). Alterations in community structure that increase the abundance of woody species may result in cascading effects on ecosystem processes such as nutrient and carbon cycles and vegetation dynamics. Increases in shrubs can result in much higher resource heterogeneity in the landscape, with resources accumulating beneath the woody plants, and the spaces between losing nutrients rapidly, causing resource islands around the shrubs (Reynolds et al. 1999).

Supplementary watering during winter and summer has revealed the differing effects of seasonality of precipitation on plant communities (Robertson et al. 2010). In systems with naturally coupled temperature and precipitation patterns, increases in winter precipitation may facilitate increases in diversity and shifts in species composition, as winter precipitation is most important for recruitment. However, in systems with naturally decoupled precipitation and temperature, native perennial species rely most strongly on winter precipitation, which can moderate the impact of summer drought on species composition (Morecroft et al. 2004). Summer precipitation on the other hand tends to increase growth and density of existing species, suggesting that it may have a more stabilizing effect on productivity. The response of individual species to changes in precipitation patterns is not always predictable, as the response to any given change is determined by a complex interplay among functional growth form, reproductive strategy, photosynthetic pathway, phenology, and competition.

Competition plays an important role in differentiating species niches, and in environments with high soil moisture variability there are several adaptations that species can employ to differentiate themselves from competitors. These different strategies combined with high levels of inter-annual water variability may be what allows arid and semi-arid systems to maintain surprising levels of species diversity (Chesson et al. 2004), as the annual timing of resource availability can determine which strategy may be favored during a given year. There are two components to competition in a variable water environment, as outlined by the two-phase resource dynamics hypothesis (Goldberg and Novoplansky 1997). These are the pulse period, in which species that are able to most quickly extract soil water will be favored, and the interpulse period, which is the dry period in which drought tolerance traits become most important to survival. The strategy that a plant uses may be geared toward one or another of these phases, and include phenological adaptations such as setting seed early before dying (annuals), becoming dormant (perennials), and timing germination to make use of early resources; they also include physiological adaptations such as minimizing water loss through storage and C4 or CAM photosynthetic pathways, or tolerating drought through utilizing more stable water sources (Chesson et al. 2004).

Partitioning of resources can occur both spatially and temporally, though the two may be strongly correlated. For instance, the two-layer model suggesting that grasses and woody species use water sources from different vertical soil layers (Walter 1971) also implies that they utilize water from different temporal precipitation events (Golluscio et al. 1998). Shrubs, trees, and other deep-rooted species tend to make use of deeper water sources, which are recharged during winter precipitation events. Meanwhile, shallow rooted grass species generally make use of water in the top layers of soil that falls during the growing season (Kulmatiski et al. 2010). While it has been verified that grasses and woody species do not often compete strongly for water (Golluscio et al. 1998), isotopic analysis has shown that their water use can sometimes overlap significantly (Schwinning et al. 2005a, Kulmatiski et al. 2010), and shrubs in particular can be very flexible in their vertical use of water resources (Dodd et al. 1998).

The size of the precipitation event can also affect which species are able to utilize the water. Unless there is a particularly large event, summer precipitation tends to only penetrate into the top layer of soil where it is either taken up by plants or evaporated (Golluscio et al. 1998). Although grasses are always able to respond to large rainfall events during the growing season, they cannot use all of the water when it percolates deeper than their roots. Shrubs on the other hand are only able to respond to large rainfall events when the soil is already dry at their rooting depth, otherwise their growth is not water-limited, and it is only during particularly dry seasons that this condition would coincide with phenologically active periods. Because of this, in decoupled environments deeper-rooted shrubs tend to be the most drought-adapted species (Morecroft et al. 2004, Golluscio et al. 2009).

There is evidence that there is a tradeoff between the dependence on winter precipitation of deeper-rooted plants, and the ability to use summer precipitation of the more shallow rooted plants, indicating that the latter group may be more opportunistic in their water use (Golluscio et al. 2009). Interestingly, this means that even in quite arid environments plants are not necessarily able to make use of large increases in annual precipitation if it occurs during the summer months. Because summer monsoon rains in arid systems tend to be much more uncertain than winter precipitation, species in decoupled climates may not have evolved adaptations to increase summer rain use efficiency (Schwinning et al. 2005b). This means that deeper rooted species such as shrubs are relatively insensitive to pulses or temporary increases in water supply, while grasses can have greater growth responses (Kochy and Wilson 2004). Moreover, grasses in stands of shrubs also exhibit a lack of response to increasing water availability, implying that shrubs may serve to stabilize water supply for other species. This may be due to hydraulic lift, in which water from deeper sources is relocated to surface soil via the root system of shrubs, where it can then be utilized by other plants (Caldwell et al. 1998).

Increasing winter precipitation tends to have the expected effect of favoring deeper-rooted, cool-season plants both observationally (Snyder and Tartowski 2006) and experimentally (Chimner et al. 2010), while summer droughts result in decreases in perennial grass cover (Morecroft et al. 2004). In some arid decoupled systems, however, herbaceous species actually respond negatively to shifting precipitation toward the summer, with decreases in biomass, cover, and density (Bates et al. 2006). Shrubs on the other hand do not respond strongly in terms of cover or density, but instead can produce more reproductive structures with later water availability. This effect is likely due to the early phenology common to places with very dry summers (Bates et al. 2006). Species in arid systems tend to enter dormancy early in the season and do not use summer precipitation to support the next year’s growth (Schwinning et al. 2005b). Therefore, shifting timing of precipitation toward the summer essentially causes winter drought, which decoupled systems tend to be much more sensitive to than summer drought. Some species, however, have adapted a flexible phenology that can take advantage of water inputs when they become available (Reynolds et al. 1999).

Resource partitioning among different growth forms may differ depending on life stage. In savanna systems, for example, trees use different depths of water during different life stages (Weltzin and McPherson 1997). During the recruitment phase, tree seedlings utilize even shallower water than grasses, which may increase recruitment rates within established grassy areas. As they mature, tree seedlings increasingly reach to deeper water sources, until after around two years of age their depth surpasses that of grasses. Shrubs show a similar change in resource partitioning in which the growth of shrubs is much more closely linked to fluctuations in water inputs during the early stages of development, after which it become more insensitive (Reynolds et al. 1999). This means that shifts in the timing of precipitation and therefore the depth of soil water sources, may impact species differently during different life stages, and could alter the stability of the coexistence between grasses and woody species.

Although decoupled systems tend to favor woodier species, coupled systems tend to favor C4 (warm season) grasses and CAM plants (Winslow et al. 2003). This is because these photosynthetic pathways have better water-use efficiency at high temperatures than C3 (cool-season) species (Amundson et al. 1994). When precipitation and temperatures are coupled, both C3 and C4 grass phenologies tend to be linked closely with water availability (Niu et al. 2005). However when precipitation and temperature are decoupled, C3 plants are more active and show the greatest response to water inputs during the fall, while C4 plants are more active and more responsive to water inputs during the mid-summer. This may be due to seasonal differences in whether water or temperature is the limiting factor to growth. C4 plants in general are not as affected by interspecific competition as C3 plants, indicating that C4 plants might be better competitors for water due to higher water-use efficiency (Niu et al. 2005).

Plant species diversity has been shown to increase in response to more variable precipitation (Knapp et al. 2002). Communities tend to shift toward more drought tolerant composition because increased variability in precipitation generally leads to more drought stress due to the top layer of soil drying during longer intervals between events. Meanwhile, more extreme events may lead to greater soil recharge in the deeper water layers and less overall evaporative water loss (Tietjen et al. 2009). Indeed, deeper-rooted forbs tend to increase production in response to larger precipitation events, while grasses tend to favor more frequent small events (Robertson et al. 2009).

More extreme but less frequent rainfall events can also differentially affect the response of C3 and C4 plants at the physiological level (Fay et al. 2002). C3 forbs tend to be more responsive to soil moisture, with stomatal conductance and photosynthetic efficiency increasing with soil moisture. C4 grasses on the other hand do not respond as much to soil moisture, possibly because of a higher innate water use efficiency and drought tolerance. Fay et al. (2003) found the dominant species in their system to be even less responsive than other C4 graminoid species. Coupled systems that receive most of their precipitation during summer monsoons tend to have more variable precipitation patterns in general, thus, tolerance of high variability may be what gives these dominant grass species a competitive edge.

Altering the timing and distribution of precipitation events has also been shown to impact the spread of invasive species, which may be conferred competitive advantages over native species by utilizing soil water from different depths or at different times during the growing season (Kulmatiski et al. 2006). For example, though increasing winter precipitation in an already decoupled system did not result in strong responses from native species, Bromus tectorum, an invasive annual plant, did respond positively to this treatment (Bates et al. 2006). A similar experiment in a coupled system found that increases in snow aided recruitment of the invasive taprooted forbs Centaurea diffusa, Gypsophila paniculata, and Linaria dalmatica, without response from native species (Blumenthal et al. 2008). These same invaders were seldom seen without the addition of snow.

Isotopic analysis of water use by exotic and native plants show that increasing invasion under altered rainfall is likely due to the different timing of water extraction by the exotic invaders (Kulmatiski et al. 2006). Exotic annual grasses tend to use water early in the growing season preempting the activity of native plants, while forbs with deep taproots are able to use deep soil water after the native plants have already senesced in the later part of the growing season. The presence of invasive species such as B. tectorum can also act as a feedback in the ecosystem, accelerating rates of carbon and nitrogen cycling, and thereby creating conditions more favorable for its own growth (Norton et al. 2008). These consequences are more pronounced when there are frequent summer rain events, and under these conditions the changes in ecosystem properties can be quite significant.

Discussion

Precipitation and its effects on soil moisture mediate most ecosystem processes, and differences in precipitation regimes globally are responsible for many of the ecosystem features that distinguish biomes from one another (Prentice et al. 1992). The timing and distribution of precipitation is an important factor in driving these differences, although it is only relatively recently that the effects of precipitation timing have been explored in a mechanistic way. Since precipitation regimes are expected to experience changes worldwide, it is important that the implications of these changes on ecosystem processes are incorporated into models, yet there are still gaps in our understanding. This is partly due to extreme complexity, as water plays a role in nearly all aspects of ecosystem functioning. This makes predicting the effects of precipitation changes quite challenging. Nonetheless, there are still some important avenues of research that will greatly improve our understanding of the subject. I will discuss some potential directions for future studies, as well as the strengths and limitations of the main approaches used to investigate these questions.

While some notable multi-factor climate studies have been conducted, the interactions between timing of precipitation and other global change processes are still relatively unknown. Warming, CO2 increases, grazing, land-use change, changes in fire regimes and nutrient deposition will all likely interact with precipitation timing, and should be investigated further through cross-site observation, multi-factor experimentation, and modeling. One potential interaction between warming and precipitation timing is that more winter precipitation may fall as rain rather than snow. Some studies suggest that the form of precipitation can be important, though the effects are unclear (Walker et al. 1999). Isotopic analysis may be an important tool in distinguishing the effects of rain versus snow by tracking the flow of water from each form through the ecosystem.

There is also a dearth of information regarding belowground processes. Very few studies investigate how microbial communities will respond to shifts in precipitation timing, though the importance of these communities in mediating ecosystem processes is becoming increasingly apparent. I suggest that future studies include measurements of microbial populations and community composition, as well as decomposition dynamics. While it is known that edaphic differences affect response to precipitation inputs, there is still little beyond the inverse texture hypothesis in the way of elucidating these effects. More needs to be known about how soil properties interact with precipitation timing, as well as how precipitation timing can feed back to alter these soil properties.

There are four main approaches used by scientists to investigate the effects of the timing of precipitation on ecosystems:

  1. Observational studies that compare ecosystem properties across climatic gradients. These studies are useful in identifying broad patterns that can be better generalized across systems, particularly when they are global in scope. However, while this method can be effective for understanding intrinsic differences between systems with differing precipitation regimes, it cannot provide much insight into how these systems might respond to changes in precipitation.
  2. Observational studies that compare ecosystem responses to natural seasonal and annual variability in precipitation. These studies can be an elegant way of investigating short-term ecosystem responses to precipitation inputs, but are not as useful for looking at effects of sustained shifts in precipitation. They are also constrained by only looking at the historic range of variability in precipitation, which climate models predict will be surpassed in many systems.
  3. Experiments that manipulate water availability through supplemental watering, snow fences, or rainout shelters, and investigate the response of ecosystem processes to these treatments. These studies are quite useful in identifying the mechanistic basis of ecosystem responses, however, the nature of these experiments is that they are usually single-system investigations, and are often only in place for a few years due to time and money constraints. I suggest that more coordinated efforts be established to investigate long-term, multi-system responses to certain precipitation manipulations with a uniform experimental design. Because precipitation change is expected to interact with other simultaneously occurring global changes, multi-factor climate experiments will be necessary to understand these interactive effects.
  4. Ecosystem models that incorporate hydrological and vegetation responses to altered precipitation. In general, these models can be very useful tools and will only improve as mechanistic knowledge is gained through experimentation and observation. However, models that fail to incorporate the hydrological differences associated with temporal precipitation patterns, and simply use mean annual precipitation as a black box for water availability may be drastically oversimplifying a key ecosystem driver that affects almost all ecosystem processes either directly or indirectly. Models may also be improved by including feedbacks that large-scale vegetation shifts may have on climate.

Conclusion

The timing of precipitation affects belowground microbial communities as well as net photosynthesis, which together determine decomposition dynamics, C and N cycling, and productivity. As precipitation regimes shift due to climate change, altering these processes alone could result in changes to ecosystem functioning, but in concert with expected changes in vegetation could cause system flips and rapid restructuring of ecosystems. The implications of altered precipitation regimes can be quite complex and difficult to predict because these processes can have cascading effects on species interactions, soil properties, production, climate, fire regime, and nutrient cycling. Improvements in our understanding of these complexities will be gained through coordinated efforts in experimentation and synthesis.

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Trait-based approaches to grassland plant community assembly and responses to grazing

Introduction

Out of all the species on earth, why is it that communities in nature assemble into the patterns of species coexistence and abundance that we observe, many of which repeat themselves over and over again across the landscape? It has long been a goal of community ecology to determine the rules that govern which species can coexist in a given community, and what determines the structure of that community as indicated by species abundances (Elton 1946, Preston 1948, Hutchinson 1959, Cody and Diamond 1975). Alhough many notable ecologists have attempted to find generalizable rules regarding community assembly (Diamond 1975, Weiher and Keddy 1995), community ecology’s holy grail remains elusive (Lavorel and Garnier 2002), and some would say that it does not truly exist (Lawton 1999, Simberloff 2004). Still, we do know that abiotic factors such as habitat, resource availability and disturbance, as well as biotic factors such as dispersal, competition and predation both contribute to patterns of community assembly observed in nature (Chesson 2000, Grime 2006).

In this paper I will give a brief overview of processes driving community assembly, as well as the historical context behind modern theory. I will then discuss the more recent emergence of trait-based approaches to examining community assembly, with a focus on plant assemblages. This approach is widely thought to be the best way forward for finding general laws in community ecology (Weiher and Keddy 1995, McGill et al. 2006). I will then demonstrate how recent studies using trait-based approaches can be applied specifically to investigate how grassland plant assemblages respond to grazing.

 

Overview of community assembly processes

The regional species pool is comprised of all potential species that could colonize a given site based on dispersal processes. Species richness is limited by colonization from the species pool, and successful immigrants tend to show trait-based species sorting along environmental gradients (Houseman and Gross 2011). Environmental constraints provide a filter on the regional species pool based on which species are able to survive under the given conditions in the absence of competitors (Diaz et al. 1998). This filter tends to cause convergence among species traits, in which species that are able to colonize have certain similarities that enable them to survive in the given environment (Chesson 2000, Messier et al. 2010).

Though environmental filters are important for determining species presence, they tend to be weaker than the biotic filters that determine abundance (Cingolani et al. 2007). However, this may depend on the productivity of the site. There tends to be more abiotic filtering at low productivity sites, since many species cannot establish even in the absence of competitors. In high productivity sites, biotic filtering is stronger as many species are only able to establish in the absence of competitors (Houseman and Gross 2011).

Biotic interactions are most important in determining species abundances rather than presence. For example, strong competitors will outcompete weak competitors for a given resource supply, and will be found in higher abundance (Mouillot et al. 2007). This idea relates to niche theory, in which species are thought to occupy distinct ecological niches in terms of function and resource utilization within an n-dimensional space of environmental conditions (Hutchinson 1959). Traditionally, niches are viewed in pairwise species interactions, where the fundamental niche, or set of environmental conditions that could be occupied by the species is restricted to the realized niche, or conditions in which the species actually does occur, due to competition with a second species with an overlapping niche (Cody 1991). In this way, species differentiate their resource use to minimize niche overlap.

This process can select for divergence among species traits, in which species that are more functionally dissimilar are better able to coexist (Macarthur and Levins 1967, Stubbs and Wilson 2004, Kraft et al. 2008). These niche differences represent tradeoffs among multiple traits, and tend to have a stabilizing effect on community structure since they cause species to limit the growth of their own populations more than the growth of their competitor’s populations (Adler et al. 2007, Levine and HilleRisLambers 2009). Because the abiotic and biotic filters function simultaneously with opposite effect, the results can sometimes appear to cancel each other out (Mouillot et al. 2007).

In theory, species should be able to perform optimally at the center of their realized niche (Macarthur and Levins 1967). However, in natural communities species abundance is not necessarily highest at this supposed performance optimum (McGill et al. 2006). Instead, plants often share preference for benign environmental conditions and high resource availability. In these cases, abundance is determined by growth strategy. The strategy employed represents a tradeoff between fast growth rates under favorable conditions and tolerance of unfavorable conditions. Among plant assemblages, this shared niche preference may actually be more common than preference for distinct niches (Wisheu 1998).

Diametrically opposed to niche theory is Hubbell’s unified neutral theory (Hubbell 2001).   In recent years, neutral theory has gained attention with the idea that neutral processes alone can explain many patterns of community assembly (Bell 2001, Volkov et al. 2003). This theory operates under the assumption that all species are functionally equivalent, and processes such as dispersal, historical contingence, and other sources of demographic stochasticity are responsible for community patterns rather than niche differences. However, while this assumption may be true in some areas (Hubbell 2005), it has been shown in many studies that while neutral processes may play some role, they are not the only drivers of community assembly (Fargione et al. 2003, Harpole and Tilman 2006, Mouillot et al. 2007, Levine and HilleRisLambers 2009).

With this evidence there is a push to move away from viewing niche and neutral processes as a mutually exclusive dichotomy, and toward the more useful endeavor of quantifying the relative importance of these two mechanisms in a given system (Adler et al. 2007, Schamp et al. 2008, Stokes and Archer 2010). Niche processes are defined as those that stabilize interactions between species with large differences in fitness, whereas neutral processes are defined as weak stabilizing mechanisms acting on species of similar fitness. The relative contributions of fitness differences and stabilizing mechanisms can be determined by examining the relationship between per capita growth rates and relative abundance in a community (Adler et al. 2007).

 

Historical Context

In the past, community assembly has been largely focused on which combinations of species are able to coexist. Jared Diamond’s classic work (Diamond 1975) categorized species on islands of New Guinea into different types of incidence functions based on the probability of the species occurring on an island containing a given number of species. He observed that only certain combinations of species occurred together on the same island, and he categorized pairs that never coexisted as “forbidden combinations”. However, these rules were almost entirely pattern-based and lacked any mechanistic basis.

Diamond’s work received criticism for not taking into account whether the patterns he observed were significantly different from what one would expect by a random distribution of species (Connor and Simberloff 1979). In response to this criticism, the use of null models, simulated from random sorting of species, have become widely used. These null models are useful tools to verify that communities assemble in a non-random fashion. In this way they have helped provide support for the existence of niche processes, and have helped to illuminate mechanisms causing divergence and convergence among species traits (Weiher and Keddy 1995, Grime 2006).

In order to find rules that could be more applicable across systems, ecologists have tried to move away from simply looking at pair-wise species interactions. Instead, species are often classified into functional groups based on responses to environmental factors and effect on ecosystem functioning. For example, grassland plant species could be categorized as C3 graminoids, C4 graminoids, shrubs, forbs, or legumes. Functional groupings led Fox and Brown to come up with new assembly rules, which specify that communities containing equal numbers of species from each functional group will be favored, and therefore more commonly found in nature (Fox and Brown 1993).

Grouping by function is still not a perfect solution for finding general rules, as there is a great deal of ambiguity surrounding the appropriate way to aggregate species into functional groups (Petchey and Gaston 2006). For example, plant species could be classified by photosynthetic pathway, reproductive strategy, or by growth habit. Transferring continuous trait data into categorical data also presents a difficulty, since it requires a degree of subjectivity in making what is sometimes an arbitrary decision. Furthermore, there seems to be no consensus on the best way to measure a community’s functional diversity, as represented by differences in function among the species involved (Mouillot et al. 2005, Ricotta 2005). This is why there has been much focus lately on using a trait-based approach for investigating community assembly (Keddy 1992, McGill et al. 2006, Shipley et al. 2006).

 

Trait-based approaches

Many ecologists are now asserting that assembly rules will be generalizable only if based on traits, and that species-based assembly rules can only be site-specific and they rely on unsure taxonomy (Weiher and Keddy 1995). Trait-based approaches to community assembly utilize quantifiable measurements of traits that are linked to performance rather than grouping species into named categories. This approach provides a more mechanistic look into assembly processes.

Functional traits can pertain to physiology, morphology, or life-history characteristics, but it is important that they strongly affect the performance of the organism. In the case of plant assemblages, these traits could be related to resource acquisition (i.e. rooting depth, tissue stoichiometry), photosynthetic rates (i.e. CO2 intake per leaf dry mass), or reproductive strategy (i.e. seed production, tillering rates), among other things. In this way, traits can act as a common currency among species, allowing comparisons across communities in which the taxa are not necessarily shared (McGill et al. 2006, Shipley et al. 2006). These species traits are then related to a given performance currency, such as growth rate (r2) or productivity. Because it is assumed that trait differences alter species performance in communities, this approach is inherently in opposition to neutral theory.

Trait-based approaches are most useful when studied across environmental gradients, as they help to determine which traits correspond most strongly to given environmental factors, and therefore can provide more general insight to community assembly under varying abiotic conditions (McGill et al. 2006). However, while many traits may contribute to performance, it is unrealistic to measure all of them. It is necessary for researchers to select the most important contributors to performance based on the gradient of interest. This can also be accomplished by reducing a suite of traits that represent a given tradeoff into a single axis of variation. A prime example of this would be the leaf economics spectrum, in which several different leaf traits correspond to a tradeoff between conservative growth strategies that limit tissue loss, and extravagant growth strategies that maximize resource acquisition (Wright et al. 2004).

One key assumption of trait-based approaches is that trait variation within a species is less than trait variation among species. While this assumption is generally safe to make, intraspecific differences can account for a significant amount of variation between communities as well (Jung et al. 2010). Intraspecific varitation can drive both convergence due to habitat filtering and divergence through niche differentiation. Including intraspecific trait variation in models can help detect and illuminate these processes (Jung et al. 2010).

As emphasized by McGill et al. in their 2006 paper, trait-based approaches to community assembly allow researchers to ask a modified suite of research questions compared with traditional species-based approaches. Instead of looking at species diversity, we can look at variation in traits between and within communities; instead of asking what environments species occur in, we can look at what traits are important in determining the range of environmental conditions in which a species can survive; instead of looking at pair-wise competition between species, we can look at traits that confer competitive dominance (McGill et al. 2006). These questions provide a framework with which to tackle the unenviable task of finding general assembly rules for communities.

 

How do grassland plant communities respond to grazing?

Plants employ two main strategies to cope with grazing: defense (avoidance) and regrowth (tolerance), and the preferred strategy may be related to site resource environment, grazing intensity, or history of grazing (Vandermeijden et al. 1988). Avoidance strategies, which serve to make plants less palatable, are conservative growth strategies that often indicate that tissue replacement is expensive (Herms and Mattson 1992). This strategy is characterized by traits including higher stem to leaf ratios, short stature, physical defenses such as spines or thorns, thick and/or waxy leaves, and presence of recalcitrant or toxic compounds in tissues. Tolerance strategies involve rapid regrowth from defoliation and are characterized by allocation to resource acquisition traits. These include higher nitrogen (N) content in leaf tissues, higher specific leaf area (SLA), and more allocation to root growth (Caldwell et al. 1981).

Many recent studies have used trait-based approaches to investigate how grassland plant assemblages respond to grazing by assessing these community traits corresponding to grazing avoidance and tolerance strategies (Diaz et al. 2007). Trait based approaches in measuring plant response to grazing have been found to be more informative than simply looking at species responses (Pakeman 2004). This is particularly relevant because communities that favor avoidance strategies tend to have low palatability and forage quality, whereas tolerance strategies often result in higher forage quality. However, it is not well understood under what conditions each strategy is favored, and how grazing can cause negative or positive feedbacks in terms of forage quality. Here I will synthesize results from recent studies that have investigated plant traits associated with these two strategies in response to grazing. I will specifically look at how these responses vary across gradients of resource environment, grazing intensity, and evolutionary history of grazing. I will also discuss how herbivore selectivity corresponds to these traits. The traits I have selected to focus on are: height, SLA, tissue N content, leaf toughness, and root:shoot ratio.

Many studies have shown that the resource environment and climate can mediate the plant response to grazing. This may be because environmental filtering from the regional species pool selects for a limited set of species that are more suited to certain grazing responses (De Bello et al. 2005). These responses have been particularly linked to annual precipitation and site fertility. Plants in arid environments have been found to exhibit more avoidance traits, such as short height and tougher leaves, whereas in more mesic sites have been found to exhibit more tolerance traits such as higher nitrogen (N) content in leaf tissues, indicating faster growth (Adler et al. 2004, Zheng et al. 2011). Similar results were found on soil fertility gradients (Rusch et al. 2009), indicating that low-resource environments and grazing avoidance may induce convergent traits. However, this response has been shown to differ over larger spatial and temporal scales (Sandel et al. 2010).

Plant strategy may also be contingent on grazing intensity. Studies have found that higher stocking rates are correlated with decreases in plant height and palatability (Pakeman 2004), with more allocation to defenses, and decreased allocation to reproduction and vegetative growth (De Miguel et al. 2010), as well as increases in root: shoot ratio (Evju et al. 2009). These studies indicate that at higher intensities of grazing, avoidance strategies may be most favored.

However, results from another study which held different levels of grazing intensities constant for a period of 15 years show the opposite to be the case (Cruz et al. 2010). In this study, higher grazing intensity was associated with high SLA, indicating that plants allocating to resource-acquisition and fast growth rates were favored. This discrepancy may be explained by taking into account intraspecific trait variation. In one such study, as grazing intensity increased, palatable species showed less allocation to resource acquisition, while unpalatable species increased in abundance as well as resource acquisition traits (Chen et al. 2005). This could lead to an increase in community-level acquisition traits (associated with tolerance) while the dominant strategy is still avoidance. Again, the dominant community response may depend on environmental filters on the regional species pool, and may interact with climatic factors (Zheng et al. 2011).

Regions that have a long evolutionary history of grazing may contain species that have had time to evolve better defenses (Diaz et al. 2007). This was found to be the case in a study comparing Patagonian grasslands to the sagebrush steppe of the western United States (Adler et al. 2004). They found that plant communities in Patagonia, which is known to have a longer history of grazing, tended to have lower forage quality (indicating defenses), while the sagebrush steppe communities tended to have taller stature plants. Decreased plant height seems to be a universal response to grazing, irrespective of climate variation (Diaz et al. 2007), which may be an indication of increased allocation to roots (Evju et al. 2009). Although belowground responses have not been well-studied (May et al. 2009). This study indicates that plant community traits such as height could be used as a proxy for evolutionary history of grazing.

These comparisons are based on the assumption that avoidance strategies are actually effective. Although it is commonly thought that reduced height is a grazing avoidance strategy (Coley et al. 1985), trait-based studies are increasingly finding that short stature may not be an indication of grazing avoidance. Diaz et al. (2001) found that species that respond positively to grazing tend to be shorter compared to species that respond negatively to grazing (Diaz et al. 2001). A study examining herbivore selectivity found that grazers actually seemed to preferentially eat shorter plants, as well as those with tougher leaves (Cingolani et al. 2005).

While SLA, which is an indicator of growth rate of the plant, was not correlated with herbivore selectivity at a species level, on a community level sheep tended to select for higher SLA communities (Cingolani et al. 2005). Also, plants with higher SLA (faster growing) tend to respond positively when grazed, compared with those with low SLA (Diaz et al. 2001, Cingolani et al. 2005). These studies found that grazing tends to reduce height, increase SLA, and increase plants preferred by sheep, indicating a positive feedback between grazing and forage quality, as indicated by both SLA and sheep preference. This is consistent with other studies indicating that sheep prefer plants with higher nutrition quality, but that selectivity was not a strong driver of the community response to grazing (Evju et al. 2009).

 

Conclusion

Although trait-based approaches to questions of community assembly may not be the answer to finding general laws in ecology, they provide an approachable framework for illuminating contingencies that hold true across disparate systems. In this paper I have demonstrated how trait-based approaches have been used to address the question of what conditions cause grazing to induce tolerance versus avoidance strategies in plant assemblages. This is a complex question that has not been usefully addressed in species-based studies. Results from the studies I have discussed may be generalizable across systems, but they are still highly contingent on site resource environment, history of grazing, and grazing intensity. Other factors such as herbivore selectivity, seasonality, scale, and regional species pool may also prove to be important contingencies that should be investigated further.

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