~laura dev 2012 (Chapter 1 of M.S. Thesis)~
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.
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.
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.
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.
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:
- 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.
- 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.
- 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.
- 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.
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.
Adler, P. B., J. HilleRisLambers, and J. M. Levine. 2007. A niche for neutrality. Ecology Letters 10:95-104.
Adler, P. B., D. G. Milchunas, W. K. Lauenroth, O. E. Sala, and I. C. Burke. 2004. Functional traits of graminoids in semi-arid steppes: A test of grazing histories. Journal of Applied Ecology 41:653-663.
Bell, G. 2001. Ecology – neutral macroecology. Science 293:2413-2418.
Caldwell, M. M., J. H. Richards, D. A. Johnson, R. S. Nowak, and R. S. Dzurec. 1981. Coping with herbivory – photosynthetic capacity and resource-allocation in 2 semi-arid agropyron bunchgrasses. Oecologia 50:14-24.
Chen, S. P., Y. F. Bai, G. H. Lin, Y. Liang, and X. G. Han. 2005. Effects of grazing on photosynthetic characteristics of major steppe species in the xilin river basin, inner mongolia, china. Photosynthetica 43:559-565.
Chesson, P. 2000. Mechanisms of maintenance of species diversity. Annual Review of Ecology and Systematics 31:343-+.
Cingolani, A. M., M. Cabido, D. E. Gurvich, D. Renison, and S. Diaz. 2007. Filtering processes in the assembly of plant communities: Are species presence and abundance driven by the same traits? Journal of Vegetation Science 18:911-920.
Cingolani, A. M., G. Posse, and M. B. Collantes. 2005. Plant functional traits, herbivore selectivity and response to sheep grazing in patagonian steppe grasslands. Journal of Applied Ecology 42:50-59.
Cody, M. L. 1991. Niche theory and plant-growth form. Vegetatio 97:39-55.
Cody, M. L. and J. M. Diamond. 1975. Ecology and evolution of communities. Belknap Press of Harvard University Press, Cambridge, Massachusetts, USA.
Coley, P. D., J. P. Bryant, and F. S. Chapin. 1985. Resource availability and plant antiherbivore defense. Science 230:895-899.
Connor, E. F. and D. Simberloff. 1979. The assembly of species communities – chance or competition. Ecology 60:1132-1140.
Cruz, P., F. L. F. De Quadros, J. P. Theau, A. Frizzo, C. Jouany, M. Duru, and P. C. F. Carvalho. 2010. Leaf traits as functional descriptors of the intensity of continuous grazing in native grasslands in the south of brazil. Rangeland Ecology & Management 63:350-358.
De Bello, F., J. Leps, and M. T. Sebastia. 2005. Predictive value of plant traits to grazing along a climatic gradient in the mediterranean. Journal of Applied Ecology 42:824-833.
De Miguel, J. M., M. A. Casado, A. Del Pozo, C. Ovalle, P. Moreno-Casasola, A. C. Travieso-Bello, M. Barrera, N. Ricardo, P. A. Tecco, and B. Acosta. 2010. How reproductive, vegetative and defensive strategies of mediterranean grassland species respond to a grazing intensity gradient. Plant Ecology 210:97-110.
Diamond, J. M. 1975. The assembly of species communities. Pages 342-444 in M. L. Cody and J. M. Diamond, editors. Ecology and evoloution of communities. Belknap Press of Harvard University Press, Cambridge, Massachusetts, USA.
Diaz, S., M. Cabido, and F. Casanoves. 1998. Plant functional traits and environmental filters at a regional scale. Journal of Vegetation Science 9:113-122.
Diaz, S., S. Lavorel, S. McIntyre, V. Falczuk, F. Casanoves, D. G. Milchunas, C. Skarpe, G. Rusch, M. Sternberg, I. Noy-Meir, J. Landsberg, W. Zhang, H. Clark, and B. D. Campbell. 2007. Plant trait responses to grazing – a global synthesis. Global Change Biology 13:313-341.
Diaz, S., I. Noy-Meir, and M. Cabido. 2001. Can grazing response of herbaceous plants be predicted from simple vegetative traits? Journal of Applied Ecology 38:497-508.
Elton, C. 1946. Competition and the structure of ecological communities. Journal of Animal Ecology 15:54-68.
Evju, M., G. Austrheim, R. Halvorsen, and A. Mysterud. 2009. Grazing responses in herbs in relation to herbivore selectivity and plant traits in an alpine ecosystem. Oecologia 161:77-85.
Fargione, J., C. S. Brown, and D. Tilman. 2003. Community assembly and invasion: An experimental test of neutral versus niche processes. Proceedings of the National Academy of Sciences of the United States of America 100:8916-8920.
Fox, B. J. and J. H. Brown. 1993. Assembly rules for functional-groups in north-american desert rodent communities. Oikos 67:358-370.
Grime, J. P. 2006. Trait convergence and trait divergence in herbaceous plant communities: Mechanisms and consequences. Journal of Vegetation Science 17:255-260.
Harpole, W. S. and D. Tilman. 2006. Non-neutral patterns of species abundance in grassland communities. Ecology Letters 9:15-23.
Herms, D. A. and W. J. Mattson. 1992. The dilemma of plants – to grow or defend. Quarterly Review of Biology 67:283-335.
Houseman, G. R. and K. L. Gross. 2011. Linking grassland plant diversity to species pools, sorting and plant traits. Journal of Ecology 99:464-472.
Hubbell, S. P. 2001. The unified neutral theory of biodiversity and biogeography. Princeton University Press, Princeton, NJ.
Hubbell, S. P. 2005. Neutral theory in community ecology and the hypothesis of functional equivalence. Functional Ecology 19:166-172.
Hutchinson, G. E. 1959. Homage to santa-rosalia or why are there so many kinds of animals. American Naturalist 93:145-159.
Jung, V., C. Violle, C. Mondy, L. Hoffmann, and S. Muller. 2010. Intraspecific variability and trait-based community assembly. Journal of Ecology 98:1134-1140.
Keddy, P. A. 1992. Assembly and response rules – 2 goals for predictive community ecology. Journal of Vegetation Science 3:157-164.
Kraft, N. J. B., R. Valencia, and D. D. Ackerly. 2008. Functional traits and niche-based tree community assembly in an amazonian forest. Science 322:580-582.
Lavorel, S. and E. Garnier. 2002. Predicting changes in community composition and ecosystem functioning from plant traits: Revisiting the holy grail. Functional Ecology 16:545-556.
Lawton, J. H. 1999. Are there general laws in ecology? Oikos 84:177-192.
Levine, J. M. and J. HilleRisLambers. 2009. The importance of niches for the maintenance of species diversity. Nature 461:254-U130.
Macarthur, R. H. and R. Levins. 1967. Limiting similarity convergence and divergence of coexisting species. American Naturalist 101:377-&.
May, F., V. Grimm, and F. Jeltsch. 2009. Reversed effects of grazing on plant diversity: The role of below-ground competition and size symmetry. Oikos 118:1830-1843.
McGill, B. J., B. J. Enquist, E. Weiher, and M. Westoby. 2006. Rebuilding community ecology from functional traits. Trends in Ecology & Evolution 21:178-185.
Messier, J., B. J. McGill, and M. J. Lechowicz. 2010. How do traits vary across ecological scales? A case for trait-based ecology. Ecology Letters 13:838-848.
Mouillot, D., N. W. H. Mason, and J. B. Wilson. 2007. Is the abundance of species determined by their functional traits? A new method with a test using plant communities. Oecologia 152:729-737.
Mouillot, D., W. H. N. Mason, O. Dumay, and J. B. Wilson. 2005. Functional regularity: A neglected aspect of functional diversity. Oecologia 142:353-359.
Pakeman, R. J. 2004. Consistency of plant species and trait responses to grazing along a productivity gradient: A multi-site analysis. Journal of Ecology 92:893-905.
Petchey, O. L. and K. J. Gaston. 2006. Functional diversity: Back to basics and looking forward. Ecology Letters 9:741-758.
Preston, F. W. 1948. The commonness, and rarity, of species. Ecology 29:254-283.
Ricotta, C. 2005. A note on functional diversity measures. Basic and Applied Ecology 6:479-486.
Rusch, G. M., C. Skarpe, and D. J. Halley. 2009. Plant traits link hypothesis about resource-use and response to herbivory. Basic and Applied Ecology 10:466-474.
Sandel, B., L. J. Goldstein, N. J. Kraft, J. G. Okie, M. I. Shuldman, D. D. Ackerly, E. E. Cleland, and K. N. Suding. 2010. Contrasting trait responses in plant communities to experimental and geographic variation in precipitation. New Phytologist 188:565-575.
Schamp, B. S., J. Chau, and L. W. Aarssen. 2008. Dispersion of traits related to competitive ability in an old-field plant community. Journal of Ecology 96:204-212.
Shipley, B., D. Vile, and E. Garnier. 2006. From plant traits to plant communities: A statistical mechanistic approach to biodiversity. Science 314:812-814.
Simberloff, D. 2004. Community ecology: Is it time to move on? American Naturalist 163:787-799.
Stokes, C. J. and S. R. Archer. 2010. Niche differentiation and neutral theory: An integrated perspective on shrub assemblages in a parkland savanna. Ecology 91:1152-1162.
Stubbs, W. J. and J. B. Wilson. 2004. Evidence for limiting similarity in a sand dune community. Journal of Ecology 92:557-567.
Vandermeijden, E., M. Wijn, and H. J. Verkaar. 1988. Defense and regrowth, alternative plant strategies in the struggle against herbivores. Oikos 51:355-363.
Volkov, I., J. R. Banavar, S. P. Hubbell, and A. Maritan. 2003. Neutral theory and relative species abundance in ecology. Nature 424:1035-1037.
Weiher, E. and P. A. Keddy. 1995. Assembly rules, null models, and trait dispersion – new questions front old patterns. Oikos 74:159-164.
Wisheu, I. C. 1998. How organisms partition habitats: Different types of community organization can produce identical patterns. Oikos 83:246-258.
Wright, I. J., P. B. Reich, M. Westoby, D. D. Ackerly, Z. Baruch, F. Bongers, J. Cavender-Bares, T. Chapin, J. H. C. Cornelissen, M. Diemer, J. Flexas, E. Garnier, P. K. Groom, J. Gulias, K. Hikosaka, B. B. Lamont, T. Lee, W. Lee, C. Lusk, J. J. Midgley, M. L. Navas, U. Niinemets, J. Oleksyn, N. Osada, H. Poorter, P. Poot, L. Prior, V. I. Pyankov, C. Roumet, S. C. Thomas, M. G. Tjoelker, E. J. Veneklaas, and R. Villar. 2004. The worldwide leaf economics spectrum. Nature 428:821-827.
Zheng, S. X., Z. C. Lan, W. H. Li, R. X. Shao, Y. M. Shan, H. W. Wan, F. Taube, and Y. F. Bai. 2011. Differential responses of plant functional trait to grazing between two contrasting dominant c3 and c4 species in a typical steppe of inner mongolia, china. Plant and Soil 340:141-155.