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Litter decomposition sustains ecosystem productivity and can provide a feedback to climate change through changes in the rate of CO2 return to the atmosphere. Climate change is altering litter decomposition rates through a variety of mechanisms (Dukes & Field, 2000; Fierer et al., 2005; Feng et al., 2008), but little is known about how quickly these rates are changing or the relative importance of the various chemical components of litter driving this response (Cornelissen et al., 2007; Cornwell et al., 2008; Salinas et al., 2011). Understanding the sensitivity of turnover rates and the stability of biochemical components in litter is critical, as these factors influence the amount and chemical composition of organic matter in soil (Crow et al., 2009; Kramer et al., 2012; Wickings et al., 2012).
Biochemical compounds in plant tissues vary in their susceptibility to decomposition, on a spectrum from labile (decaying quickly; e.g. carbohydrates and amino acids) to relatively recalcitrant (decaying slowly; e.g. lignins (Boerjan et al., 2003; Floudas et al., 2012), tannins (Lorenz et al., 2007) and cuticular matrix (Hu et al., 2000)). The recalcitrant compounds are enriched in plant litter as a result of resorption of labile compounds during senescence, and can potentially influence ecosystem functions by modulating the decomposition rates. According to kinetic theory, recalcitrant compounds with high activation energies should be relatively sensitive to temperature (Bosatta & Agren, 1999; Davidson & Janssens, 2006), implying that warming could disproportionately accelerate the decomposition of recalcitrant compounds. This effect of climate on decomposition is not only governed by the kinetics of biochemical reactions, but is also modulated by other abiotic and biotic interactions, many of which have received little attention. For example, climate might have a comparatively small effect on the decomposition of labile compounds, as these substrates can be utilized by a broader spectrum of microorganisms that operate across a wider range of climatic conditions. Further, when labile compounds are abundant, bacterial communities may outcompete fungal communities (McGuire & Treseder, 2010) for non-carbon substrates, thus slowing the decomposition of recalcitrant compounds irrespective of climate. The quality of the litter matrix (as indicated by the proportions of labile and relatively recalcitrant compounds) could thus alter the direct effect of climate on decomposition by affecting microbial community composition (Fig. 1). Apart from the climatic controls, decomposition and associated microbial transformations also depend on how the carbon in the substrates is partitioned between microbial biomass production and microbial respiration (carbon use efficiency, CUE; Manzoni et al., 2012), which, in turn, depends on the nutritional quality of the substrates (Keiblinger et al., 2010). Thus, there are several plausible mechanisms by which the chemical composition of litter matrix might modulate the climate sensitivity of recalcitrant compounds. However, this type of litter chemistry-mediated effect of climate on decomposition remains relatively unexplored.
Figure 1. Conceptual diagram illustrating how litter quality mediates the effect of climate on litter decomposition. Climate may influence litter decomposition indirectly by affecting the chemistry of litter at its formative stage (Tharayil et al., 2011), and by mediating changes in microbial community, physiology and enzyme activity (Allison et al., 2010). The dotted lines represent these indirect effects of climate on decomposition. Climate can also affect the rate of decomposition directly by altering the activation energy of compounds in litter (Davidson & Janssens, 2006). Our study highlights that the effect of climate on decomposition may be modified by litter chemistry (chemistry of heteropolymers and the proportion of labile and recalcitrant compounds) via changes in microbial community composition and physiology.
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Previous studies have attempted to elucidate the effects of climate on decomposition using litter from different plant species which varies in its composition of heteropolymers (Preston et al., 2009), thus achieving different ratios of labile to relatively recalcitrant compounds. However, the structural identity of plant heteropolymers differs across different species (Ralph et al., 2004; Dixon et al., 2005) and among individuals of the same species as a result of genetic and environmental variation (Vanholme et al., 2010; Tharayil et al., 2011). Most heteropolymers are functionally/operationally defined, their biological reactivity determined by the type and number of monomers, the linkages connecting these monomers (Nierop et al., 2006; Schweitzer et al., 2008; Talbot et al., 2012) and the overall polydisperse matrix, all of which vary significantly within and between species, and could potentially influence processes associated with litter decomposition. For example, the presence of acylated monolignols, by imparting hydrophobicity to lignin (del Rio et al., 2007), could delay its mineralization. Similarly, in tannins, prodelphinidins which contain tri-hydroxy B-rings, have greater enzyme inactivation and protein complexation capacity than procyanidins, which contain di-hydroxy B-rings (Nierop et al., 2006). Thus, studies that use litter from different species to characterize the influence of litter quality on the climate sensitivity of decomposition compare litter types that differ not only in the relative concentrations of compounds (Austin & Ballare, 2010), but also in the molecular structures of their heteropolymers, which also influence decomposition rates. This conflation of the concentration and structure of chemical compounds in litter limits the interpretation of results.
Here, we used two litter types that differed in the relative proportions of labile and recalcitrant compounds, but had heteropolymers with similar molecular structure, to characterize the overall mass loss and compound-specific response of plant litter decomposition to the combined effects of warming and altered precipitation. We hypothesized that the influences of warming and altered precipitation on the decomposition of litter would depend on the relative abundance of labile compounds in the litter. We predicted that labile compounds would limit the sensitivity of decomposition to climate, such that the decomposition of litter with a higher proportion of labile compounds (as indicated by a high carbohydrate carbon : methoxyl carbon (CC : MC) ratio; Mathers et al., 2007) would be less responsive to warming and precipitation changes. Further, we predicted that warmer and wetter climates would accelerate the decomposition of recalcitrant compounds only in litter with a lower proportion of labile compounds (low CC : MC ratio).
To hold the structural identity of heteropolymers relatively constant across litter types, we used stem litter of Polygonum cuspidatum (Japanese knotweed; syn. Fallopia japonica, Reynoutria japonica) from a single clonal (genetically identical) population. We achieved a relative difference in proportions of labile and recalcitrant compounds in this litter by collecting samples at two stages after senescence. We collected newly senesced litter (referred to hereafter as NL) and old litter (referred to hereafter as OL; this comprised 1-yr-old senesced stems that had been decomposing upright for a year in P. cuspidatum stands with no direct contact with soil), which we expected to have a lower proportion of labile compounds than NL.
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Overall, our results suggest that the effect of climate on litter decomposition depends on the proportion of labile compounds in the litter. The higher proportion of labile compounds in NL prevented climate from affecting the mass loss rates for 2 yr. After 3 yr of decomposition, although the mass loss varied similarly with climate in both litter types, warmer, wetter conditions accelerated the degradation of recalcitrant matrices only in litter that was initially low in labile compounds (OL). Thus, the proportion of labile compounds regulated both the overall mass loss dynamics and the decomposition dynamics of the recalcitrant compounds. Our results also emphasize that, after 3 yr of decomposition, although mass loss was similar in OL and NL, the chemical composition of the remaining litter in OL diverged with climatic treatments. We also found that precipitation treatments affected mass loss more strongly than warming treatments, and this effect persisted even when c. 80% of the mass had been lost from both litter types.
Effects of warming and altered precipitation on rates of litter decomposition
Although the climate treatments significantly affected the mass loss of OL within the first 124 d of field incubation, their effects on NL could only be detected later – after 805 d of field incubation (Fig. 2b,c). Thus, in the earlier stages of decomposition, litter with a smaller proportion of labile compounds (OL) was more sensitive to climate than litter with more labile compounds (NL). As all the heteropolymers in both litters were structurally similar, we attribute this delay in the appearance of climatic effects on decomposition of NL to its greater initial abundance of labile compounds. During the early stages of decomposition, microbes target unprotected and high-energy-yielding substrates, such as cellulose and hemicellulose (Berg & McClaugherty, 2008). As most of these labile compounds can be utilized by a wide range of microbial communities (Jones et al., 2009; McGuire & Treseder, 2010), having different climatic and other environmental tolerances (Pett-Ridge & Firestone, 2005; Cruz-Martinez et al., 2009), the potential for climate to affect the utilization of these compounds may be limited. Irrespective of climate, in the presence of labile carbon, bacteria are superior competitors for non-carbon compounds and may competitively exclude the slower growing fungi that specialize in the decomposition of relatively recalcitrant compounds (Fontaine et al., 2003; Moore et al., 2003; Waldrop & Firestone, 2004; McGuire & Treseder, 2010). In addition, fungal communities may have more specific climatic requirements (McGuire et al., 2010; Hawkes et al., 2011). This difference in response of bacterial and fungal communities to climate and labile carbon could explain the initial confinement of climatic influence on decomposition to OL. Recently, it has been shown that, in soil, the microbial utilization of labile substrates that do not require enzymatic breakdown is insensitive to warming (Frey et al., 2013). In addition, as most of the initial depolymerization of heteropolymers is facilitated by microbial exo-enzymes, greater moisture should accelerate the mineralization of heteropolymers by increasing diffusion rates, facilitating enzyme–substrate interactions. This could explain the greater mass loss from OL exposed to ambient and wet precipitation (relative to dry) in the initial stages of decomposition. In NL, a similar response to precipitation appeared only during the later stages of decomposition (at 805 d), probably after the advanced degradation of labile compounds. Previous studies have shown that the decomposition of recalcitrant compounds, such as lignin, starts only after low-molecular-weight labile compounds have been degraded (Berg & McClaugherty, 2008; Adair et al., 2008; Schneider et al., 2012). Our study demonstrates that this initial decomposition of labile compounds is relatively insensitive to climate, and hence the effect of climate on tissue decomposition should be more pronounced only after the relative depletion of labile substrates. Our finding that litter quality determines the climate sensitivity of litter decomposition suggests that the traditional conceptual model for decomposition, in which climate and litter quality act as two separate factors affecting litter decomposition, should be modified to include an interaction (Fig. 1).
Climate has been considered to have a negligible influence on decay rates during the later stages of litter decomposition (Berg & Meentemeyer, 2002), as the influence of nutritional constraints increases (Prescott, 2010). However, we found a contrasting pattern: the decomposition of NL initially did not respond to the climate treatments and, in later stages, ambient and wet treatments accelerated the mass loss of both NL and OL relative to the dry treatment, even after 1140 d, when c. 80% of the initial mass had been lost. Our results thus suggest that sufficient moisture can remove some of the constraints on decomposition in the later stages of decay, resulting in greater mass loss.
In this study, precipitation treatments regulated litter mass loss and decay rates more strongly than did warming treatments (Fig. 2b,c). As moisture availability is normally much higher in mesic systems such as ours than in arid and semiarid systems, one might expect only a modest response of ecosystem processes to moisture relative to the response to temperature. The Boston area's climate has strong seasonality to temperatures, and the magnitude of the BACE warming treatments was relatively small relative to seasonal temperature swings (Fig. S2a). Although the region experiences relatively constant average monthly precipitation, there was considerable variability among months and years during this experiment (Fig. S2b). Soils can go through pronounced drying and wetting cycles during the growing season, and the BACE treatments strongly affected soil moisture (e.g. Hoeppner & Dukes, 2012; Suseela et al., 2012). We would expect the litter wetting and drying cycles to be even more rapid and more dramatic. Our results suggest that the precipitation treatments altered the degree to which moisture levels constrained the degradation of compounds in litter during the warmer seasons, when temperatures were optimal. At the same study site (BACE), we found that precipitation treatments had a stronger effect on soil microbial respiration than did warming (Suseela et al., 2012).
Response of recalcitrant compounds to warming and altered precipitation
Although similar percentages of mass were lost from NL and OL after 1140 d of decomposition (Fig. 2a), OL experienced faster decomposition of the recalcitrant matrix (alkyl carbon and lignin) per unit of degradation of labile matrix (Fig. 4a). In addition, the normalized (1060 cm−1, carbohydrates) DRIFT spectra of OL exposed to climatic treatments after 1140 d revealed a significant reduction in the intensity of lignin in WHt relative to DHt (Fig. 4a,c; 1700–1596 cm−1). Similarly, the decrease in the intensity of DRIFT peaks in the fingerprint region (1500–1200 cm−1) of OL exposed to a warmer and wetter climate suggested that there had been greater decomposition of non-carbohydrates per unit of carbohydrate (Figs 4a, 6). The lignin degradation in OL exposed to WHt was further evident from the relative increase in condensed and cross-linked lignin, as indicated by the ratios of the peak intensities at 1510 and 1631 cm−1 (Fig. 8; 42% increase relative to DHt; Mann et al., 2009). Condensed and cross-linked lignin is composed of carbon–carbon linkages, which are more difficult for microbes to break down than less condensed lignin, with its greater proportion of labile β-O-4 linkages (Boerjan et al., 2003; Talbot et al., 2012). The higher relative abundance of condensed and cross-linked lignin in the residue of OL exposed to WHt indicates advanced decomposition relative to OL in the DHt treatment. This greater decomposition of recalcitrant compounds in WHt is further supported by the 13C NMR analysis (Fig. 7), where OL from the WHt treatment had the lowest CC : MC ratio, indicating that decomposition had advanced further than in DHt litter. Previous studies have shown steady declines in the CC : MC ratio with advancing decomposition (Almendros et al., 2000; Blumfield et al., 2004; Mathers et al., 2007). The CC : MC ratio is a more robust indicator of the decomposition of woody residues than the alkyl C : O-alkyl C ratio (Baldock et al., 1997; Mathers et al., 2007). After both litter types had decomposed for 3 yr, this effect of the climate treatments on the decomposition of the recalcitrant matrix was visible only in OL, suggesting that the warmer, wetter conditions accelerated the decomposition of recalcitrant compounds only after substantial degradation of a labile structural matrix. As decomposition progresses, the cellulose and hemicellulose matrices that are less protected are more easily degraded, increasing the relative abundance of lignocellulosic material (Berg & McClaugherty, 2008). Thus, in the advanced stages, during the decomposition of lignocellulose, we would expect the degradation of cellulose to be tightly associated with the degradation of lignin, which would result in an increase in the decomposition of lignin per unit degradation of carbohydrate, as observed in OL at 1140 d. We did not observe the above compound-specific degradation pattern in OL exposed to ambient precipitation treatments (Fig. 4a), underscoring the importance of greater moisture availability in the facilitation of the decomposition of recalcitrant compounds. The climatic treatments did not instigate a preferential decomposition of recalcitrant compounds from NL. The greater abundance of relatively labile compounds could have altered the sequence of compounds decomposed in NL, thus nullifying the effect of climate. Our finding that labile compounds in litter decrease the climate sensitivity of decomposition of recalcitrant matrix is novel and suggests that, within a species, indices of litter quality (the proportion of labile to recalcitrant compounds) may signal the degree to which climatic factors can affect soil carbon sequestration. After 3 yr of decomposition, both litter types had lost similar amounts of mass (Fig. 2a), but the chemistry of the residual OL varied with climatic treatments (Fig. 4a), whereas the chemical composition of NL exposed to different climatic treatments was similar (Fig. 4b). These results indicate that initial litter quality can influence whether the chemistries of residual litter converge (Wallenstein et al., 2013) or diverge (Wickings et al., 2012) over time under different climates.
Figure 8. Intensity ratio of diffuse reflectance infra-red Fourier transform (DRIFT) peaks between 1510 and 1631 cm−1 of 1-yr-old (OL) Polygonum cuspidatum litter after 1140 d, showing the relative enrichment of condensed and cross-linked lignin in the wet treatment (triangles) relative to the dry (squares) and ambient (circles) treatments (higher ratios signify greater proportions of condensed and cross-linked lignin). The higher relative abundance of condensed and cross-linked lignin in the residue of OL exposed to WHt (wet + high warming) indicates advanced decomposition relative to OL in the DHt (dry + high warming) treatment. Values represent means (n = 3) ± SE. Ppt, precipitation; W, warming.
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The results from this study indicate that the effect of climate on litter decomposition depends on the quality of litter; litter with a greater initial proportion of labile compounds was less sensitive to warming and altered precipitation. Similarly, the warmer and wetter climate treatment preferentially accelerated the decomposition of recalcitrant compounds in the more recalcitrant litter. The significant effects of precipitation on mass loss and chemical composition, even in the late stages of litter decomposition, reveal the potential of climate to alter the amount and quality of carbon in plant litter available for sequestration. These results emphasize that litter chemical composition has an overriding effect on the climate sensitivity of decomposition; thus, litter quality may regulate litter-derived carbon sequestration under future climates.