Fine root decomposition and soil carbon
Interest in the processes contributing to stabilized organic matter in forest soils derives both from its role in maintaining soil fertility and increasingly its role as a large stock of global carbon. Three principal sources of forest soil C can be distinguished: aboveground detritus, belowground detritus, and inputs from living roots (rhizosphere C flux). Although aboveground plant detritus usually comprises the largest proportion of total C input to forest soils, evidence from soil food web studies indicates the dominant role of below-ground inputs in powering forest soil food webs (Pollierer et al. 2007; Gilbert et al. 2014), and biochemical tracers also indicate that roots are principal contributors to forest SOM (Nierop 1998). This pattern may be due in part to the differential placement of these inputs (i.e., on soil surface vs within soil matrix; Sokol et al 2019b).
The role and mechanisms of fine root detritus in supplying stabilized forest S0M have received limited study. We traced 13C from two fine root types (order 1–2 vs 3–4) into several soil fractions in four acid forest soils over seven years to better understand fine root detrital dynamics. We hypothesized that after seven years the decomposition of fine roots would approach an asymptote, by analogy to “limit values” noted for leaf litter in temperate and boreal forests (Berg et al. 1996). However, we observed significant reduction in total 13C recovery in the final year (Fig. 2) at which point recovery averaged 8.9% across the four soils and two root orders. This extent of decomposition greatly exceeds the limit value of about 20% remaining, measured for sugar maple leaf litter after six years of decay (Lovett et al. 2016). Thus, the extent of decay for sugar maple fine roots is greater than for leaf litter in these acid forest soils; after six years residual humified leaf litter apparently contributes more C to surface soil horizons than fine root detritus contributes to mineral S0M. However, the long-term stability of leaf litter derived SOM probably differs markedly from that of root litter.
After the first year no significant differences were detected in 13C recovery between order 1–2 and order 3–4 fine roots. Notably, the differences in substrate quality between these root types were modest (Table 2). Goebel et al. (2011) observed similar decay rates for order 1 to 4 fine roots whereas other studies have reported significant differences in decay rates between lower and higher order roots; however, the direction of these difference varied among studies, some reporting higher and others lower (Fan and Guo 2010; Biedler and Pritchard 2017) decay rates for lower order fine roots. Also, these studies relied on buried root litter bags which have known artifacts (Dornbush et al. 2002). Our results indicate that the small differences observed in substrate quality probably do not significantly influence the longer term (7 year) decay and incorporation of fine root detritus into SOM.
Soil properties and stabilization of root litter carbon
Our study was designed to explore how differences in soil properties might influence stabilization of fine root detrital C. The four soil horizons chosen for study differed markedly in texture, structure, SOC, pH, and metal (Al, Fe) content (Table 1). After seven years the only significant difference in overall 13C recovery that we observed among these four soils was the lowest recovery in soil SP B2. This spodic horizon from an “iron podzol” that developed on very coarse glacial outwash was distinctive in having the lowest content of occluded microaggregates (Table 3); however, the overall low recovery was mostly associated with low values for the particulate organic matter fraction. The small differences in total recovery among the other three soils is suggestive of more subtle influences of physical and chemical properties on the stabilization of root litter derived SOM.
Our study provides some insights into the process of incorporation of fine root detritus into SOM. The traditional concept that resistance to microbial decomposition was afforded biochemically has been replaced by recent work ascribing SOM stabilization primarily to physical mechanisms that protect inherently labile organic matter from degradative microbial exoenzymes (Dungait et al. 2012). Two interrelated mechanisms contribute to this stabilization of SOC: sorption on mineral colloid surfaces (phyllosilicates, metal oxides) (Kogel-Knabner et al. 2008), and protection within soil microaggregates (Six et al. 2004, Lehmann et al. 2007). The latter are thought to form because of binding by and between organic matter and soil minerals and may afford physical protection by retarding the entry of biota, enzymes, oxygen, and nutrients (Totsche et al. 2018; Biesgen et al. 2020). Resistance to oxidative degradation of organic matter sorbed to surfaces of mineral colloids may result from strong binding, especially within the confines of microaggregates (Churchman et al. 2020; Kleber et al 2022). These mechanisms help to account for the inherently low turnover of mineral-associated organic matter in soils.
Soil aggregates are an intimate mixture of organic matter, soil minerals, and microorganisms and their by-products. Many studies have examined aggregate formation and stability, especially in agricultural soils because of impacts of tillage and cropping on soil C (Six et al. 2004). The current paradigm is that microaggregates (< 250 µm in size) are a relatively stable mixture of secondary minerals and highly processed detritus and by-products of microbial decay. These microaggregates are bound together into less stable macroaggregates (> 250 µm) by fresher organic matter (Golchin et al. 1994), and disruption of macroaggregate structures can lead to more rapid microbial turnover of the organic components (Six et al. 2004).
The four soils we studied were highly structured, as defined operationally by the slaking procedure, with high proportions of macroaggregates and microaggregates, the latter being either both free or occluded within the macroaggregates (Table 3). After one year of fine root decomposition, we recovered the 13C label in all the mineral soil aggregate fractions at readily detectable levels, ranging from about 0.5–3% of the added 13C. Presumably this recovery represented mostly soluble organic C adsorbed on soil colloids as well as some particulate organic matter, including microbes. Most of these products of initial decay of fine root detritus were recovered within macroaggregates, including coarse particles, as well as microaggregates and silt + clay (Fig. 2). During the second year of decay a decline in percent recovery in these fractions was observed despite the likely continuing supply of labeled root decay products. This result suggests that some of the mineral-associated organic matter (MAOM), including that found in aggregates and sorbed onto silt and clay, remained available for microbial decomposition or possibly desorption. Kaiser and Guggenberger (2003) argued that much of the sorbed SOM remains labile, and Saidy et al (2013) reported desorption of about 10% of sorbed DOC from various secondary clay minerals. We previously observed a significant reduction in sorbed C derived from sugar maple leaf litter in one of the soils studied here (IN A1; Fahey et al. 2011). The initially high recovery in macroaggregates, followed by the reduction in years 2 and 3 also supports the role of fresh root litter in macroaggregate formation (Six et al. 2001) as well as the relatively low stability of these structures (Rabbi et al. 2014).
During the later stages of fine root decomposition (3 to 7 year) we observed steady or increasing percent recovery in the MAOM fractions (Fig. 2), indicating that any decomposition of root-derived MAOM was counterbalanced by continuing supply from microbial processing of detrital residue. Most striking was the large increase of recovery in microaggregates, especially those occluded within macroaggregates, in year 6 and 7. This result strongly supports the concept that microaggregate formation is influenced by organic matter derived from fine root decay.
Total recovery of fine root detrital 13C declined significantly in year 7 across all the soils; this decline was primarily associated with declining recovery in the low-density particulate organic matter fraction. At the same time, recovery in occluded microaggregates increased significantly. Although the cause of this striking pattern is uncertain, it is notable that warm season (May-October) precipitation in year 7 was by far the highest during the entire study, exceeding the long-term average by about 20%. We speculate that exceptional wetting caused natural slaking of macroaggregates along with incorporation of the label into new, water-stable microaggregates during this summer cycle (cf., Amézketa 1999; Menon et al. 2020).
Based on the differences in soil properties and in % recovery in various fractions among the four soils, we can speculate on the mechanisms influencing the incorporation of fine root detritus into MAOM in these acid forest soils. After seven years, recovery of root-derived C in MAOM differed significantly among the four soils (Fig. 3). Most striking was the high recovery in occluded microaggregates in the A horizon of a higher pH soil (pH = 5.5) that coincidentally had the highest metal oxide content (IN A2; Table 1). The source of the metal oxides in this soil is uncertain but periodic poor drainage conditions may cause anoxic conditions, reduction of the iron, and re-oxidation and re-precipitation as iron oxides under subsequent oxic conditions. This process is possible at the higher pH that facilitates re-oxidation of reduced iron (Cornell and Schwertmann 2003). Very high recovery was observed in multiple replicates; for example, five of eight replicates of this soil exceeded 3% recovery in this fraction in year 7. This high recovery was mirrored by high C concentration in this fraction (6.97%). Significantly lower total 13C recovery was observed in this fraction in the two spodic horizons with low pH and more moderate concentrations of metal oxides (Table 1).
Although macroaggregates were abundant in the B2ir horizon soil (SP B2; Table 3), by year 7 incorporation of root-derived C into these macroaggregates was very low, and especially for microaggregates occluded within them (Fig. 3). Notably, these microaggregates had low carbon concentrations in this soil (4.83%) compared with the Bs horizon soil (9.93%). Thus, the contrast in pH coincided with differences in the incorporation of root litter C into microaggregates between these soils. Notably, the SP B2 soil also had the highest recovery in the free < 53 µm fraction, possibly indicative of alternative fates for root litter C in MAOM.
The Inceptisol A horizon collected from the incubation study plot (IN A1) had much higher clay content than the Adirondack soils (12% vs 4%; Table 1), yet isotope recovery in the < 53 µm fractions of this soil was similar to the others (Fig. 3) On one hand this is surprising because the dominant phyllosilicate mineral in the IN A1 soil is illite (M. McBride, pers. comm.), which has a higher capacity for DOC sorption than kaolinite (Saidy et al. 2013) and vermiculite, the dominant clay minerals in the Adirondack soils (Kahle et al. 2004). On the other hand, organic matter binding to physllosilicates occurs via cation bridging and exchange (Kleber et al. 2021), and perhaps the low concentration of exchangeable Ca in this acidic soil limits sorption.
After seven years of fine root decay about 4% of the added 13C was recovered in low-density particulate organic matter, representing about half of the total recovery; most of the remainder was recovered in the MAOM fractions. Just how stable is the new C in these fractions? Perhaps the remaining low-density particulate organic matter is humified and analogous to surface soil humic matter. According to current concepts most of the MAOM, especially that contained in microaggregates, and strongly sorbed to metal oxides via ligand exchange, is probably very stable (Kleber et al. 2021). In sum, our results indicate that a comparatively small proportion of fine root derived C (< 3%) contributes to the stable stock of MAOM in these acid forest soils. This observation supports the recent evidence of Sokol et al. (2019a) that living root inputs (rhizosphere C flux) predominate over above- and below-ground detritus in formation of stable SOM in forest soils.
Comparisons with living root inputs and leaf litter
Parallel studies of sugar maple fine roots, litter and SOM provide information relevant to our interpretations. At a nearby site on the same Mardin series soil type as IN A1, Yavitt et al. (2015) traced 13C from roots of sugar maple saplings into soil aggregate fractions; note that this C would include some fine root turnover as well as abundant rhizosphere C flux (Phillips and Fahey 2005). After three years, 21.5% of the label from living tree roots was recovered in soil (compared with 13.1 ± 2.1% for root detritus in the present study[ Fig. 1). The distribution of this root-derived 13C among soil aggregate fractions also differed qualitatively from the root litter in the present study (Table 4). In particular, after three years a much lower proportion of total recovery was found in microaggregates derived from root litter (6.7%) compared with the sapling root system label (24.2%). This observation also accords with the conclusion of Sokol et al. (2019a) that C inputs from living roots (rhizosphere C flux) are most readily converted into MAOM.
We can also compare the initial contribution to MAOM between fine root detritus and leaf litter. Yavitt et al. (2015) measured recovery of 13C labeled sugar maple leaf litter in MAOM for similar soils as IN A1 (Mardin series) and reported recovery in 0–5 cm mineral soil after one year of decay at 2.6 ± 0.4% (no recovery was detected below 5 cm). This value is analogous to the value of 2.3% recovery in MAOM observed here for fine root detritus in the IN A1 soil type. In both cases label recovery declined during the second year (Fig. 2; Yavitt et al. 2015) again indicating either or both decay or desorption of a portion of the new MAOM. After two years, recovery of leaf litter 13C was quite uniformly distributed among the four components of MAOM (free micro = 0.51%; free < 53 µm = 0.66%; occluded micro = 0.66%; and occluded < 53 µm = 0.42%), again similar to the pattern for fine root litter (free micro = 0.75%; free < 53 µm = 0.39%; occluded micro = 0.52%; and occluded < 53 µm = 0.24%). Thus, the close proximity of fine root detritus to mineral surfaces and aggregates did not appear to have strong effects on short-term incorporation into MAOM in comparison with leaf litter. Leaching of soluble organic C from decaying leaf litter was the principal mechanism delivering 13C label to mineral soil (Fahey et al. 2011), and the similarity of our results suggest the same mechanism is probably important for fine root detritus, at least in the initial stages of decay. Notably, however, recovery in mineral soil continued to increase in year 3 for leaf litter (Yavitt et al. 2015) whereas it levelled off for root litter before increasing later in year 6 and 7 (Fig. 2).
Over the seven-year course of the soil core incubations in the field site there were no significant changes in the weight proportions of the various aggregate fractions, and significant differences in aggregate abundances between the four soils were retained. Perhaps this is not surprising because root ingrowth into the cores was facilitated by access holes in the walls of the cores and abundant root growth was observed.
We emphasize the operational nature of our aggregate separation procedure. In particular, our definition of the size cutoff for microaggregates (53–250 µm) is arbitrary, and no doubt some proportion of the < 53 µm fraction actually consists of smaller microaggregates (Totsche et al. 2018). Thus, our procedure underestimates the actual percent recovery of fine root litter 13C in microaggregate fractions. Also, the coarse particulate component of the macroaggregates consisted almost entirely of fine sand as indicated by its low carbon concentration (average = 0.5%). This coarse particulate fraction constituted a small proportion of total recovery of 13C except for the first-year collection when it comprised 3.25% of 13C recovery. We surmise that this fraction initially included some particulate C that migrated in suspension and was deposited as surface coatings on fine sand particles.
In conclusion, in contrast with aboveground plant litter, fine root detritus is generated in immediate proximity to soil minerals, thereby maximizing the opportunity for stabilization in MAOM. Nevertheless, our observations suggest that only a small proportion of fine root litter is stabilized; after seven years only about 2% of the 13C added in fine root litter was recovered in MAOM. Most of the MAOM recovery was in microaggregates, especially those contained within macroaggregate structures. An additional 5% was recovered in low-density particulate organic matter, some of which could eventually be catabolized or converted to the MAOM fractions, as we observed in the last year of incubation. Therefore, although fine root turnover constitutes a large C input to forest soils, it does not appear be the principal contributor to the large, stabilized SOM stock. Our results support the conclusion of Sokol et al. (2019a) that rhizosphere carbon flux greatly exceeds both aboveground and belowground detritus as a source of temperate forest SOM.