Litter decomposition is one the of most crucially recognized biochemical C sequestration processes in grassland ecosystems, globally. Alteration of plant species diversity in grasslands are associated with modifications seen in litter quality and decomposer communities and current understanding the effects of litter decomposition and associated C availability at the plant-soil interface is not well understood (Lu et al. 2017). Grasslands cover roughly 40% of the Earth’s above sea-level land surface and possess upwards of 20% of the soil organic carbon (SOC) stock worldwide (Schuman et al.
2002). Annual litter fall and derived decomposition processes play integral roles in the development and turnover of soil organic matter (SOM) in grassland ecosystems (Liski et al. 2002).
Climate, litter quality and related decomposer communities have traditionally been considered as drivers, regulating the rate of litter decomposition (Bradford et al. 2015). Studies express discrepancies surrounding observations describing interactions between litter quality and soil decomposers that have both positive and negative effects on litter decomposition (Ayers et al. 2009; St. John et al.
2011; Perez et al. 2013); few details honor how such interactions actually affect decomposition and C availability in soils (Lu et al. 2017).
Recent studies express how variations in plant residue decomposition rates depend greatly on litter quality and can be measured by N, lignin and polyphenol concentration where decomposition processes proceed more rapidly in materials high in N and low in lignin and other polyphenolic compounds (Swift et al. 1979; Pastor et al. 1987; Chapin 1995; Hobbie 1996; Aerts and De Caluwe 1997). High concentrations of phenols (i.e. lignin, tannins) control decomposition and nutrient release rate due to polyphenols performing decomposition processes slower than carbohydrate polymers (Melillo et al.
1982; McClaugherty and Berg 1987; Dominguez 1994; Vitousek et al. 1994; Norhup et al. 1995).
Plant species traits determine litter quality and are thought to control litter decomposition rate variability between species (Cornwell et al. 2008), due to differences in chemical composition. Cleveland et al. (2014) observed that decomposition of a species’ litter is consistently correlated with that species’ C strategy. Consequently, C compounds (i.e. non-structural carbohydrates, phenols) that are more accessible and less recalcitrant C (condensed tannins, lignin) can promote litter decomposition (Hattenschwiler and Jorgensen 2010); legumes (i.e. Lespedeza cuneata) with high polyphenol (i.e. tannins) concentrations can vary in their rate of decomposition depending on the age of the individual (Kalburtji et al. 1999), where seedlings have low tannins relative to mature legume stands (i.e. Lespedeza cuneata tannin content 1st year growth
Climate, soil nutrient content, decomposer community and chemistry of the litter influence litter decomposition rate (Swift et al. 1979; Aerts 1997). Commonly, leaf litter high in N concentration is favored by bacteria and fungi and decompose plant material rapidly (Melillo et al. 1982) and interspecific differences in leaf litter quality can affect decomposition rates, which in turn can impact soil processes; plants will generally reinforce existing patterns of nutrient availability (Wedin and Tilman 1990; Hobbie 1992). While differences between species’ litter quality is common, species-specific differences in growth can also influence decomposition patterns and nutrient cycling (Hobbie 1996). Invasive plants often maintain higher concentrations of leaf N (Vitousek et al. 1987; Vitousek and Walker 1989; Witkowski 1991; Baruch and Goldstein 1999; Nagel and Griffin 2001) and consequently are expected to decompose more rapidly and release more N to the soil than native species.
When N acquisition derived from decaying invasive plant species litter is greater than native plant-derived N, N availability at the soil surface may increase and rates of nutrient cycling in invaded areas may rise (Vitousek and Walker 1989; Witkowski 1991). Thus, differences in leaf-level properties of invasive and native plants can have large impacts on ecosystem processes throughout invaded communities (Levine et al. 2003; Ashton at al. 2005). Soil microorganisms play a vital role in litter decomposition and partitioning C between CO2 evolution and the C sequestration into temporary storage pools in soils. Litter decomposition is greatly influenced by the activities of soil decomposers. Bacteria and fungi comprise more than 90% of the soil microbial biomass and are commonly observed in decomposition efficiency (Rinnan and Baath 2009). Considerable progress has been made in understanding how these factors influence the rate of litter decomposition and the formation of SOM (Stevenson 1994) yet the role of litter decomposition, C sequestration and the placement in trophic food webs that microinvertebrates influence is still largely understudied.
While understanding general consequences and impacts invasive and native plant species can have on an ecosystem, it is vague if impacts are a consequence of invasion by exotics in general, or if impacts are a consequence of chemical and physical properties of the litter from particular invaders chosen for a study. It may also be possible that invaders could alter ecosystem properties through indirect effects that cannot be tied to the direct impacts of differences in the chemistry or physical traits of the invader. For instance, invaders may alter the soil environment in ways that lead to changes in the community of soil decomposing organisms (Belnap and Phillips 2001; Kourtev et al. 2002). Direct or indirect, these alterations to the soil environment may facilitate further invasion of exotic plant species (Simberloff and Von Holle 1999). In order to manage invaded areas more effectively, it is essential that the effects of invasive species on decomposition processes is understood more definitively.
Annually, more than 90 gigatons of terrestrial plant biomass become incorporated into the ever-growing dead organic matter pool throughout ecosystems globally (Cebrian 1999) and C and N cycling efficiency is directly linked through the rate in which litter is decomposed (Parton et al. 2007; Berg and McClaugherty 2008) and ultimately the resulting net C storage in soils (De Deyn et al. 2008). Lower biodiversity, seen through species loss across trophic levels, significantly alters decomposition rates (Gesser et al. 2010; Handa et al. 2014) observed to be equally or more detrimental compared global environmental change effects (i.e. N deposition, rising CO2) (Hooper et al. 2012).
Species diversity ranges of leaf litter have been observed to alter litter decomposition (Hattenschwiler et al. 2005; Handa et al. 2014), but the effect of decomposition of replicated leaf litter in non-native ranges is less studied. As below ground roots account for the majority of plant biomass in grassland ecosystems, and therefore a major constituent of C in grassland soils (Rasse et al. 2005), soil C accumulation and storage also increase resulting plant biodiversity and increased standing biomass (Fornara and Tilman 2008; Steinbeiss et al. 2008; Adair et al 2009; Cong et al. 2014; Lange et al. 2015); soil C is the end product of litter production and decomposition. Therefore, in managed grasslands where above-ground biomass is removed (i.e. mowing, grazing, prescribed fire), root litter production and decomposition is paramount (Oram et al. 2018).
The rate of decomposition is determined through two main pathways: the soil environment and the litter quality (i.e. species composition of litter mixture, litter mixing effects) (Swift et al. 1979; Aerts 1997; Parton et al. 2007). Above-ground plant diversity has been observed to alter decomposition, influencing factors in both in these pathways. However, studies report positive (Hector et al. 2000; Cong et al 2015), weak-negative (Knops et al. 2001; Fornara et al. 2009; Chen et al. 2017) or insignificant effects on decomposition (Scherer-Lorenzen 2008) when comparing different soil environments to standard litter decomposition (Oram et al. 2018). These observations explain that while plant diversity may not effect decomposition consistently throughout all soil types, limitations to decomposition seem to be site specific.
Diverse communities are observed to have greater canopy cover (Spehn et al. 2005) that can result in lower temperatures at ground-level (Verheyen et al. 2008) and in top soils (Rosenkranz et al. 2012); herbaceous plant community can also affect microclimate in immediate areas. In early spring, decomposition can be especially slow due to reduced activity of decomposers. Soil moisture has also been found to promote decomposition (Prescott 2010), but plant diversity and soil type still provide a variety of results. Plant diversity can also influence soil biota by increasing decomposer abundance (Eisenhauer et al 2011) and activity (Balvanera et al. 2006), microbial biomass (Eisenhauer et al. 2010) and microbial activity (Lange et al. 2015), signaling that decomposition increases with plant diversity.
Researching how plant diversity simultaneously influences soil abiotic and biotic factors, and the relative importance of these factors to decomposition are needed to more effectively predict the effect of plant diversity on decomposition through changes in the soil environment. Studies have suggested that soil invertebrates have the potential to modify the outcome of competitive interactions between invasive and native plant species, while plants have the potential to modify the composition of the soil invertebrate community (Hunter and Price 1992). Comparing soil invertebrate community structure and decomposition rates between areas with an invasive and native species can also provide important insights into the mechanisms that allow for invasion, as well as the outcome of interactions between soil invertebrates, plants and resulting decomposition.
Observing shifts in litter quality across a diversity of gradients (examining native litter decomposing in its home environment), plant diversity reportedly has positive (Hector et al. 2000; Cong et al 2015), weak-negative (Knops et al. 2001; Fornara et al. 2009; Chen et al. 2017) or insignificant effects on litter decomposition (Scherer-Lorenzen 2008). Litter quality may be influenced by plant diversity through shifts in species abundance which can influence litter traits, diversity or litter mixing effects. As a result, decomposition can be negatively affected by plant diversity due in part to increase grass presence and the associated increase in C:N ratio in diverse communities.
Outcomes of the plant diversity and decomposition relationship could be inconsistent due to functional traits of an individual on ecosystem process compared to its taxonomic identity (Diaz and Cabido 2001; Scherer-Lorenzen 2008). Trait-based approaches by investigating species-specific characteristics are now frequently used to explain ecosystem process; plant nutrient uptake and growth can be predicted by leaf traits (Wright et al. 2004; Reich 2014), and root traits (Roumet et al. 2016). The root economic spectrum is not ubiquitous to all ecosystems, which may reflect a disconnect between root traits and functioning in certain ecosystems (Weemstra et al. 2016).
Traits can also inform outcomes of plant-soil interactions. For instance, leaf traits can explain the composition of soil food webs (Orwin et al. 2010; de Vries et al. 2012b). Plant-soil feedbacks have been explained by traits of leaves (Baxendale et al. 2014) and root (Cortois et al. 2016). Combining leaf and root traits has been shown to predict population biomass (Schroeder-Georgi et al. 2015), and explain community biomass and net biodiversity effects (Roscher et al. 2012). Plant functional traits can explain variation in soil C storage across biomes