Soil erosion is seriously threatening ecosystem functioning in many parts of
the world. In this context, it is assumed that tree species richness and
functional diversity of tree communities can play a critical role in
improving ecosystem services such as erosion control. An experiment with
170 micro-scale run-off plots was conducted to investigate the influence of
tree species and tree species richness as well as functional traits on
interrill erosion in a young forest ecosystem. An interrill erosion rate of
47.5 Mg ha
Soil erosion is considered as one of the most severe environmental challenges
globally (Morgan, 2005). It is also a serious challenge in the PR China,
especially in the southern tropical and subtropical zone. Although important
improvements in erosion control have been achieved in this area in the last
decades (Zhao et al., 2013), the annual soil loss rates range between 0.28
and 113 Mg ha
Soil erosion can negatively influence biodiversity (Pimentel and Kounang, 1998), but it is assumed that this relationship also acts vice versa (Körner and Spehn, 2002; Geißler et al., 2012b; Brevik et al., 2015). It has been shown that a change in biodiversity can have remarkable effects on ecosystem functions and stability (e.g. Hooper et al., 2005; Scherer-Lorenzen, 2005). In many cases, increasing biodiversity enhanced ecosystem productivity and stability (Loreau, 2001; Jacob et al., 2010). In particular, tree species richness (the diversity of tree species) as well as functional diversity (the diversity of functional traits as morpho-physiophenological attributes of a given species; cf. Violle et al., 2007) of tree communities can play a critical role in improving ecosystem services such as water filtration or climate regulation (Quijas et al., 2012; Chisholm et al., 2013; Scherer-Lorenzen, 2014). As forests are generally considered beneficial for erosion control, afforestation is a common measure in soil protection (Romero-Diaz et al., 2010; Jiao et al., 2012). This also applies to the south-eastern part of China, which is known to be a hotspot of biodiversity and especially tree species richness (Barthlott et al., 2005; Bruelheide et al., 2011). Guo et al. (2015) showed that forests in this area experienced the lowest soil loss rates of all land use types. Considering that studies on soil erosion under forest have mostly focused on deforestation (Blanco-Canqui and Lal, 2008) and that counteracting measures such as afforestation often result in monoculture stands (Puettmann et al., 2009), it appears that the role of tree species richness for soil erosion has been largely disregarded. Zhou et al. (2002) and Tsujimura et al. (2006) demonstrated that tree monocultures have only limited mitigation potential for soil losses, but further research is scarce. Nevertheless, there is growing evidence that a higher species richness can reduce soil erosion (Körner and Spehn, 2002). Bautista et al. (2007) pointed out that an increase in functional diversity within a perennial vegetation cover decreased soil losses in a semiarid Mediterranean landscape. Pohl et al. (2009) showed that an increase in the diversity of root types led to higher soil stability on an alpine grassy hillslope, and most recently Berendse et al. (2015) found that a loss of grass species diversity reduced erosion resistance on a dike slope.
Conceivable mechanisms underlying positive species richness effects on soil erosion are that vegetation cover with a high number of species includes a high number of plant functional groups which complement one another. Thus, they are more effective in controlling erosion processes than vegetative cover with few species (Pohl et al., 2012). For example, high tree species richness can result in an increased stratification of canopy layers (Lang et al., 2010) and a higher total canopy cover (Lang et al., 2012). In addition, a highly diverse structure within the leaf litter layer on the forest floor seems to improve its protective effect (Martin et al., 2010). Further research on the influence of tree species richness on erosion control appears to be necessary, but the complex system of interacting functional groups within the vegetation cover is also of great interest.
Vegetation cover is generally considered a key factor for the occurrence and extent of soil erosion (Thornes, 1990; Hupp et al., 1995; Morgan, 2005). A leaf litter layer on the forest floor, for example, protects the soil from direct raindrop impact and modifies the water flow and storage capacities at the soil surface (Kim et al., 2014). Moreover, forests can provide a multistorey canopy layer which largely influences rain throughfall patterns and leads to the capture of raindrops as well as the storage of water within the tree crown (Puigdefábregas, 2005). Nevertheless, large drops can be formed at leaf apexes of tall trees (Geißler et al., 2012a) and thus may increase the kinetic energy of throughfall in older forest stands by a factor of up to 2 to 3 compared to open fields (Nanko et al., 2008, 2015). This leads to considerable soil loss if the forest floor is unprotected, which may be the case if protective layers diminish, e.g. under shady conditions (Onda et al., 2010) or fast decomposition (Razafindrabe et al., 2010). While the effects of soil surface cover on soil erosion are well studied (Thornes, 1990; Blanco-Canqui and Lal, 2008), much less is known about the influence of species-specific functional traits of the tree layer such as crown or stem characteristics (Lavorel and Garnier, 2002; Guerrero-Campo et al., 2008). Moreover, most research on the latter aspects was performed in old, full-grown forests (e.g. Zhou et al., 2002; Nanko et al., 2008; Geißler et al., 2012a), whereas forests at an early successional stage are rarely mentioned. In these young forests, tree heights are lower than at later stages, but structural and spatial complexity is high and species-specific growth rates differ considerably (Swanson et al., 2011). It is assumed that these species-specific differences in structure and growth will influence soil erosion rates.
This research focused on the influence of tree species, tree species richness
and species-specific functional traits on interrill erosion in young forests,
when a leaf litter cover is not present. Testing for these effects on soil
erosion requires a common garden situation, in which confounding factors such
as different tree ages and sizes, inclination or soil conditions can be
monitored in detail. These requirements were met in the forest-biodiversity–ecosystem-functioning experiment in subtropical China
(BEF China; cf. Bruelheide et al., 2014). Within this experiment, 170
micro-scale run-off plots were established in a randomly dispersed and
replicated design. Thereby, the following hypotheses were postulated:
Increasing tree species richness decreases interrill erosion rates. Tree species differ in their impact on interrill erosion rates. The effects of different tree species on interrill erosion rates can be
explained by species-specific functional traits.
The study was conducted in Xingangshan, Jiangxi Province, PR China
(29
Mean characteristics of the 34 selected very intensively studied plots (VIPs) in 2013 in the BEF China experiment, Xingangshan, Jiangxi Province, PR China.
The experimental area has been used as a commercial forest plantation
(
Twenty-six selected tree species used in the experiment according to the
Flora of China web page (
To determine sediment delivery (as initial interrill erosion) and surface
run-off volume, micro-scale run-off plots (ROPs, 0.4 m
Measurement set-up showing a run-off plot (ROP,
0.4 m
At each ROP, tree crown cover, leaf area index (LAI), soil surface cover,
slope and rainfall amount were measured. Crown cover and LAI were determined
using a fish-eye camera system (Nikon D100 with Nikon AF G DX 180
Weather conditions were recorded by an on-site climate station (ecoTech data logger with Vaisala weather transmitter and ecoTech tipping bucket balance) at 5 min intervals. In 2013, the total precipitation in the study area was 1205 mm and lower than the mean of the preceding 3 years (1635 mm). In May and June, 10 rainfall events were captured with ROP measurements in the study area. Events were determined by breaks in rainfall of at least 6 h. Four of these events (E1–E4) were strong enough to trigger soil erosion (out of 33 events over the entire year of 2013) following Wischmeier and Smith (1978), who used an event threshold of 12.7 mm. The total rainfall amount from May to June was 185 mm, of which 135 mm fell during erosive rainfall events. The mean and peak intensities as well as the total rainfall amount (except for E4) increased from May to June (Table 3), reflecting a growing monsoon influence from the beginning to the middle of summer.
Characteristics of rainfall events considered erosive (threshold 12.7 mm) in Xingangshan, Jiangxi Province, PR China in May and June 2013.
Linear mixed effects models with restricted maximum likelihood were used with
R 3.0.2 (R Core Team, 2013) and “lmerTest” (Kuznetsova et al., 2014) to
investigate the influences on sediment delivery. Models were fitted with
crown cover, leaf area index, tree height, stem diameter, crown width, slope,
surface cover, SOM, amount of precipitation and tree species richness as
fixed effects. As random effects, precipitation event (E1–E4) nested in
plot, tree composition (species pool), site (A or B) and ROP nested in plot
were used. Nesting was introduced to avoid pseudoreplication considering the
degrees of freedom in our hypotheses tests. Tree and crown characteristics
were fitted one after the other because they were highly correlated.
Contrasts between diversity levels (div0 to div1–div24, div1 to div8–div24)
were introduced to quantify the effects of bare plots vs. tree plots and tree
monocultures vs. mixtures, respectively. The effect of individual tree
species (div1) was tested separately against the mean sediment delivery using
crown cover, slope, surface cover, SOM and amount of precipitation as fixed
factors and site and ROP nested in plot as random factors (
The results were based on 334 ROP measurements out of a total of 378
measurements. Invalid measurements were caused by technical constraints such
as plugged tubes or toppled rainfall gauges. Sediment delivery over all VIPs
and rainfall events ranged from 14 to 920 g m
Results of the basic linear mixed effect model for sediment delivery
(
Mean sediment delivery in g m
Sediment delivery and growth characteristics (means) of tree species with significant differences in delivery at the experimental site in Xingangshan, Jiangxi Province, PR China.
Tree species richness did not affect sediment delivery or run-off volume
(Table 4 and Fig. 2). Sediment delivery and run-off volume did not differ
between bare plots (div0) and plots with trees (div1–div24) nor between
monocultures (div1) and species mixtures (div8, div16, div24). The standard
deviations of sediment delivery (g m
Sediment delivery and run-off volume at five diversity levels based
on four rainfall events in May and June 2013 in Xingangshan, Jiangxi
Province, PR China (n.s.: not significant;
Individual tree species in monocultures showed significant differences in
sediment delivery (Fig. 3) ranging from 90 g m
Sediment delivery under 20 tree species in monocultures based on four rainfall events in May and June 2013 in Xingangshan, Jiangxi Province, PR China. Dashed line indicates mean sediment delivery of all 20 species. Horizontal lines within box plot represent medians, and diamonds represent mean values found for a respective species.
The mean sediment delivery is 199 g m
Effects of species-specific functional traits and site characteristics on sediment delivery. Analyses were based on four rainfall events in May and June 2013 in Xingangshan, Jiangxi Province, PR China. Black lines represent linear trends.
Crown cover was highly correlated with LAI, tree height, stem diameter and
crown width (
Growth characteristics of the 20 tree species in monocultures analysed and associated plot characteristics in Xingangshan, Jiangxi Province, PR China (mean, standard deviation (SD), maximum (max) and minimum (min)).
Crown cover: proportion of soil surface area covered by crowns of
live trees (%); leaf area index: one-sided green leaf area per unit soil
surface area (dimensionless); tree height: distance from stem base to apical
meristem (m); stem diameter: cross-section dimension of the tree stem at
5 cm above ground (m); crown width: length of longest spread from edge to
edge across the crown (m); soil surface cover: proportion of soil surface
area covered by stones, biocrusts and litter (%); soil organic matter:
fraction of organic carbon containing substances in the soil (%); slope:
inclination (
Growth characteristics were highly variable between tree species, which was reflected by high standard deviations of the respective variables. In contrast, site characteristics of these plots showed a low variability (Table 7).
The soil loss rate determined by micro-scale ROPs
(47.5 Mg ha
Tree species richness did not affect sediment delivery or run-off volume, and thus the first hypothesis has to be rejected. Nevertheless, a trend of decreasing sediment delivery and run-off volume from diversity level 0 to 8 was visible. However, both parameters were nearly the same at diversity level 1 and 24 and standard deviations were high. In contrast to tree growth patterns in monocultures, which were highly variable, mixed stands indicated a more balanced development (cf. Kelty, 2006). All species mixtures in this experiment ensured a high level of tree height and ground coverage after 4 to 5 years of tree growth, whereas in monocultures the canopy cover was lower and highly tree-species-specific. Thus, several monoculture plots were excluded before measurements because some species could not provide enough ground coverage. At the same time, sediment delivery in 8- and 16-species mixtures was lower than in monocultures. Nevertheless, contrasts in the model could not show any statistical difference between monocultures and mixtures or bare and covered plots.
The absence of a species richness effect on interrill erosion is likely attributable to the early successional stage of the forest experiment with low tree ages. Full canopy cover with high stratification and overlap has not yet been developed at the study site and the trees were far from reaching terminal height (Goebes et al., 2015b; Li et al., 2014). It is assumed that these vegetation characteristics will change with increasing tree age and tree species richness may become evident in adult stands. Young trees are functionally more equivalent to one another than older trees (Barnes and Spurr, 1998), and specific crown traits may emerge more distinctly in later successional stages. Geißler et al. (2013) found that the erosion potential was higher in medium and old, full-grown forests than in young forests. This effect is caused by raindrop transformation processes during the canopy passage, resulting in higher throughfall kinetic energy under forest than on fallow land (Geißler et al., 2010) and has only been proved for advanced successional forest stages (Nanko et al., 2008; Geißler et al., 2013). As the experiment progresses and tree height increases, increasing throughfall kinetic energy is expected, which in turn increases the general soil erosion potential if an understorey is missing.
Trees in monocultures differed in their impact on interrill erosion and thus hypothesis 2 can be confirmed. In a study on common European tree species, Augusto et al. (2002) showed that the tree species composition of forests has an impact on chemical, physical and biological soil properties. Several studies revealed that individual plants are important for erosion control in arid and semi-arid Mediterranean landscapes (e.g. Bochet et al., 2006; cf. Durán Zuazo and Rodríguez Pleguezuelo, 2008) and Xu et al. (2008) showed that different plant morphologies may control soil loss and improved soil properties in a dry river valley in China.
In this study, four tree species (
Tree species differed widely in canopy characteristics and sediment delivery was significantly related to crown cover, LAI and tree height. Therefore, the species-specific effects of interrill erosion can be partially attributed to species-specific functional traits, which confirms hypothesis 3. The falling velocities of throughfall drops are highly variable under different tree species due to the species-specific growth pattern and crown characteristics (Goebes et al., 2015a). Frasson and Krajewski (2011) showed that the mechanisms of interception are manifold even within a single canopy, and varying canopy levels create different drop size distributions.
Increasing crown cover and LAI were mitigating interrill erosion in this early ecosystem stage. The magnitude of canopy cover determines the proportion of raindrops intercepted (Blanco-Canqui and Lal, 2008), and it has been shown that drop size distributions differ between different canopy species (Nanko et al., 2006). High crown cover and leaf area increase the interception of rain drops and the storage capacity of water in the canopy (Aston, 1979; Geißler et al., 2012a), which can lead to higher stemflow and thus decreasing throughfall (Herwitz, 1987). Nevertheless, Herwitz (1987) also showed that canopy drainage can lead to larger throughfall drops and thus to increasing throughfall kinetic energy depending on the leaf species (Hall and Calder, 1993; Geißler et al., 2012a; Goebes et al., 2015a). In any case, LAI showed a weaker significance than crown cover, probably because many trees had not yet developed a multi-layered canopy structure.
It has been shown that tree height is an import factor for sediment
detachment under forest (Geißler et al., 2013), mostly due to increasing
drop falling heights (Gunn and Kinzer, 1949). As trees had not yet reached
adult height (mean height
Stem diameter and crown width did not seem to influence erosion processes in early-stage forest ecosystems. Several other tree-related functional traits (Pérez-Harguindeguy et al., 2013) could be used to explain sediment delivery such as branching architecture, specific leaf area and root system morphology. Especially studies on leaf traits (Nanko et al., 2013) as well as belowground stratification (Gyssels et al., 2005; Stokes et al., 2009) showed the potential of these features to influence soil loss and highlighted the complexity of factors mitigating soil erosion in forest ecosystems.
Results showed that soil surface cover and SOM affect interrill erosion. Even though a leaf litter cover was not present in this experiment, the remaining soil surface cover by stones and biological soil crusts was the most important driver to reduce sediment delivery. This finding underlines the general importance of covered soil surfaces for erosion control (cf. Thornes, 1990; Morgan, 2005) and shows that the protective effect of leaf litter could not only be replaced by soil skeleton but also by topsoil microbial communities in young forest stands. The mitigating effect of leaf litter on soil losses has not been in the focus of this experimental approach, but it is presumed that the fall of leaves even in young forests reduces soil erosion considerably compared to bare land (Blanco-Canqui and Lal, 2008; Seitz et al., 2015). Furthermore, SOM reduced interrill erosion, which could be explained by its ability to bind primary particles into aggregates (Blanco-Canqui and Lal, 2008). If we assume that SOM increases with increasing species richness, as was recently demonstrated in a grassland study by Cong et al. (2014), an indirect effect of biodiversity on soil erosion could be supposed. Finally, slope angle did not affect interrill erosion due to the short plot length that limits run-off velocities (cf. Seitz et al., 2015).
An experiment with 170 micro-scale run-off plots was conducted to investigate the influence of tree species and tree species richness as well as species-specific functional traits on interrill soil erosion processes in a young forest ecosystem. The results led to the following conclusions.
Tree species richness did not affect sediment delivery and run-off volume, although mixed stands showed a more balanced and homogenous vegetation development than monocultures. This finding was ascribed to the young successional stage of the forest experiment. Future research should concentrate on how erosion rates change with increasing stand age. Therefore, long-term monitoring of soil erosion under closing tree canopies is necessary.
This study provided evidence that different tree species affect interrill erosion processes. Different tree morphologies have to be considered when regarding erosion in young forest ecosystems. The appropriate choice of tree species for afforestation as a measure against soil erosion becomes important already at an early successional stage.
Species-specific functional traits and site characteristics affected interrill erosion rates. High crown cover and leaf area index reduced soil erosion, whereas it was slightly increased by increasing tree height. Thus, low tree stands with high canopy cover were effectively counteracting soil loss in initial forest ecosystem. In further studies, a wider range of functional tree traits such as leaf habitus or belowground stratification should be taken into consideration. Moreover, investigations into the influence of biological soil crusts, topsoil microbial communities and their impact on organic-matter accumulation will open the way to new insights on soil erosion processes.
Thomas Scholten, Peter Kühn and Steffen Seitz designed the experiment and Steffen Seitz carried it out. Steffen Seitz, Philipp Goebes and Helge Bruelheide developed the model code and performed the statistics. Ying Li and Werner Härdtle provided data on tree growth and species-specific functional traits. Steffen Seitz prepared the manuscript with contributions from all co-authors.
This study was financed by the German Research Foundation (DFG FOR 891/2) in cooperation with the Chinese Academy of Science (CAS). We are grateful to the Sino-German Center for Science Promotion for organising summer schools and providing travel grants (GZ 1146). Thanks go to Chen Lin and Zhiqin Pei for organisation and translation in China, Milan Daus and Kathrin Käppeler for assistance during field work, Bertram Bläschke for the installation of the first ROPs, Shunhe Lian, Yangmeng Liu and Wuchai Liu for technical support in China and finally to our numerous, tireless Chinese field workers. Edited by: P. Fiener