Soil and organic carbon redistribution in a recently burned Mediterranean hillslope affected by water erosion processes
Graphical abstract
Introduction
Organic matter (OM) in soil is the largest and most dynamic reservoir of carbon (C) on Earth and, thus, a key factor in global carbon cycling (Cerli et al., 2012, Kirkels et al., 2014). Organic C is stabilized in soils through key mechanisms as physical isolation (occlusion), chemical interaction of OM with reactive soil minerals, and preservation of recalcitrant compounds (Berhe, 2012, von Lützow et al., 2006, Wang et al., 2014). These mechanisms relate to soil aggregation by providing physical and chemical protection of soil organic carbon (SOC) against decomposition, so that OM becomes non-accessible for soil microorganisms and fauna (Doetterl et al., 2016, Six et al., 2002), and by aggregate formation through SOC sorption to pedogenic metal oxides, clay minerals or by co-precipitation with polyvalent cations (Mikutta et al., 2006, Schmidt et al., 2011). It has been found that aggregate formation appears to be closely linked with SOC storage and stability (Hu and Kuhn, 2016, Wiesmeier et al., 2019). On the other hand, soil erosion can promote breakdown of aggregates at the eroding landform positions leading to exposure of previously protected SOC (Doetterl et al., 2012, Lal, 2003), which potentially increases its mineralization rate (Berhe, 2012, Zhang et al., 2006) and releases soluble compounds (Badía et al., 2014, Caon et al., 2014). However, transported and deposited SOC can be protected from decomposition if efficiently buried in slow turn-over environments, leading to large C sinks in colluvial and alluvial sediments depending on the rate of burial, the time since burial, the nature and amount of mobilized C, and the post-depositional conditions (Doetterl et al., 2016, Kirkels et al., 2014).
There are a number of disturbances affecting soil aggregation and SOC as changes in land use (Martínez-Mena et al., 2012, Wiesmeier et al., 2019) or forest fires (Shakesby, 2011, Shakesby et al., 2015), among others. Forest fires are considered one of the main causes of soil degradation in the European Mediterranean region (Caon et al., 2014, Garcia-Ruiz et al., 2013, Shakesby, 2011), affecting their physical, chemical and biological soil properties (Badía et al., 2014, Bento-Gonçalves et al., 2012, Campo et al., 2008), as well as an increase in water runoff and sediment loss (Campo et al., 2006, Cawson et al., 2012). Particularly important are the effects of forest fires in the first few centimetres of topsoil, in relation to changes in SOC quantity and quality (González-Pérez et al., 2004), aggregate stability and erosion processes (Mataix-Solera et al., 2011, Shakesby and Doerr, 2006).
Soil erosion disturbs topsoils and, preferentially, removes SOC from upslope places, resulting in the redistribution and burial of SOC in depositional environments (Martínez-Mena et al., 2012, Wang et al., 2014). There is a number of factors that can influence the net effect of transport and deposition on eroded SOC (and consequently on its fate) as the rate and nature of soil erosion, the amount and nature of the eroded C, soil texture, soil aggregation, the transport distance, and terrain attributes such as slope gradient and surface roughness (Doetterl et al., 2016). Slow but long-range transport may lead to a higher degree of decomposition of mobilized SOC, whereas fast but short-range transport might lead to the burial of mobilized SOC, with a lower degree of decomposition, at the depositional site (Berhe and Kleber, 2013, Kuhn et al., 2009, Quinton et al., 2010). At this zone, the rate of decomposition of eroded SOC can be reduced by a combination of processes. These are biochemical (recalcitrance of organic constituents mainly those associated to pyrogenic carbon), physical (protection with burial, aggregation, and changing water, air, and temperature conditions) or chemical (mineral-OM associations). As stated by Berhe et al. (2007), in this scenario, eroded SOC remaining near the surface of foothills could contribute to enhanced decomposition, higher mineralisation rates (Quinton et al., 2010, van Hemelryck et al., 2010, van Hemelryck et al., 2011), whereas the decomposition rate of buried C stocks is likely to be reduced. In general, initial sediment C-enriched will be buried by the sediments transported during subsequent rains (depending on their magnitude and on the frequency of episodic heavy rainfall events or floods) (Nadeu et al., 2012).
Redistribution of SOC can be affected by forest fires, which can increase or reduce SOC stocks depending on several factors (González-Pérez et al., 2004, Shakesby et al., 2015). The role of soil aggregation and stabilization in SOC dynamics during erosion and deposition has attracted scientific attention in recent decades (Nadeu et al., 2011, De Nijs and Cammeraat, 2020). However, the effects of such processes on SOC stabilization have not been studied in Mediterranean forest soils, where bare-soil areas act as sources of runoff water, sediment, seeds and nutrients (including C) that move downslope and are captured by, and concentrated in, vegetation patches (Urgeghe and Bautista, 2015). Even less is known about SOC stabilization when disturbances, such as forest fires, lead to a decrease in vegetation cover and litter, therefore increasing the connectivity of runoff-source areas and the transport capacity of the flow, reducing the hillslope storage potential for water and sediments (Cammeraat, 2004, Mayor et al., 2011), and exposing the already fire-affected SOC.
Understanding SOC nature and reactivity upon changes, as those produced by forest fires, has become crucial to model and define the role of soil as source or sink of C, and to sustainably manage ecosystem services related to the soil resource (Faria et al., 2014). This knowledge starts by the identification of organic fractions with distinct chemical and biological functions and turnover times, characteristics strongly related to the form and/or the type of interactions with minerals (e.g. Rasmussen et al., 2005). Therefore, isolating fractions of OM occurring either inside or outside of aggregates or being part of organic–mineral associations, all of them different in terms of biochemical properties and functional relevance, has become a major research topic during the last two decades (Berhe, 2012, Cerli et al., 2012, Doetterl et al., 2012, Grünewald et al., 2006, Wang et al., 2014).
Physical fractionation by density has been proven useful to separate SOM and to identify meaningful soil fractions, which can be related to different stability and stabilization processes (von Lutzow et al., 2007, Wagai et al., 2009, Wagai et al., 2015, Nadal-Romero et al., 2016, Yeasmin et al., 2017). In contrast to chemical extractions, density fractionation allows for isolation of unmodified C fractions, and is theoretically related to the spatial arrangement and interactions of organic compounds and minerals (Cerli et al., 2012). The method separates light and heavy fractions (Christensen, 1992), taking advantage of the difference in density between minerals and organic material, and often by additional physical dispersion (e.g. sonication), aiming an aggregate disruption and subsequent release of the OM occluded therein (Golchin et al., 1994).
The free light fraction (FLF), which floats in a solution of given density without additional dispersion, comprises undecomposed, easily accessible OM, i.e. large organic fragments that underwent little physical and/or chemical transformation. The occluded light fraction (OLF) comprises much finer organic material with similar composition as the FLF but slightly more altered, stabilized by aggregation, and protected within aggregates, i.e., OM floating in the solution after aggregate disruption. The remaining OM fraction in sediments represents the heavy fraction (HF), in which C is strongly bound to minerals and cannot be completely separated from them, i.e. organic–mineral associations (e.g. Cerli et al., 2012, Kaiser and Guggenberger, 2007). Both floating (light) fractions are supposed to comprise mainly plant-derived debris (leaves, branches, and roots) plus some animal residues, charcoal, seeds, pollen, and microorganisms (Golchin et al., 1994, Wagai et al., 2009). The main differences between the two light fractions should be their size and location within the soil matrix. Together with size fractionation, this technique would be helpful to study how erosion, transport and deposition can induce transitions in SOC from one fraction to another (i.e. active to passive or vice versa) for example by aggregate disruption or deep burial, which may change C mineralisation rates (Wang et al., 2014).
The present study intends to obtain a better understanding of SOC accumulation and stabilization in a post-fire Spanish Mediterranean hillslope under soil erosion and deposition processes. The main hypothesis is that fire affects soil characteristics related to SOC stabilization and, together with erosion processes, can modify the SOC distribution within aggregates and in the burned hillslope. Accordingly, the main objectives are: (a) to determine differences caused by fire in soil aggregation, SOC content and stock, at hillslope scale; (b) to evaluate the influence of different variables as hillslope position, environment (under canopy and bare soils) and depth on the changes of soil aggregation and SOC distribution; (c) to use density and size fractionations in burned and unburned soils, as well as in sediments, to assess the effects of fire and erosion on SOC distribution within aggregates; and, (d) to estimate the role of fire and erosion in the possible changes of SOC accumulation and stabilization (i.e. SOC stock) in a Mediterranean hillslope, in order to shed some light in the discussion about the role of soil erosion as source or sink of C.
Section snippets
Study site and sampling
This work was carried out in the municipality of Azuébar, Natural Park of Sierra de Espadán, in the Province of Castellón, Spain (Fig. 1). Coupled hillslopes (burned: BU, and control: CO, ca. 0.25 ha each one) belonging to the coastal foothills of the Iberian Mountain System were selected (BU: 39°50′45.11″N, 0°22′20.52″W; CO: 39°51′08.7″N, 0°22′17.6″W). Both slopes are located on forested concave hillsides, with ENE aspect, 25 − 28° of slope and an altitude around 370 m a.s.L. (more information
Soil characteristics
Table 1 summarizes the soil properties determined in this study. In relation to the WDT, significant differences were only found between BU and CO, and between UC and BS (M−W, p < 0.05). In CO soils, generally, >200 drops were needed to break aggregates, and no differences could be observed for position, environment or depth (Table 1, Table 2). In BU soils, only UC (200 drops) and BS (120 drops) were significantly different (M−W, p < 0.05, Fig. S5). Results from the wet-sieving test also showed
Soil characteristics
Soils were in general very stable, but opposite results were observed (Table 1), which can be attributed to the slower wetting rate of the WDT as compared to the wet sieving. Imeson and Vis (1984) stated that WDT test is very suitable for soils of low aggregate stability, and the method would not be appropriate for the stable aggregates of this study, and other Mediterranean soils burned at high and moderate severity fires (Campo et al., 2008).
The response to AS to forest fires is complex since
Conclusions
Wildfire impact on Azuebar’s hillslope caused changes in several of the soil properties, which confirms the research hypothesis. Burned and control hillslopes showed differences in the stability of soil aggregates but trends depended on the size analysed and the method used. Factors as hillslope position (related also to hillslope steepness, length, and curvature), soil depth and environment (under canopy soil vs bare soil) influenced significantly the movement, and stock of OC in the
Declaration of Competing Interest
The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper.
Acknowledgements
This work has been supported by the VALi + d postdoctoral contract (APOSTD/2014/010) of the Generalitat Valenciana. J. Campo also wants to acknowledge C. Celi, J. Westerveld and J. Schoorl for the help with laboratory work; and A. Revynthi, B. Peñarroya, M.I. Montoya and P. Yousefi for their great support.
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