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Forest, Range & Wildland Soils

Soil Organic Matter Changes in a Spruce Ecosystem 25 Years after Disturbance

Lehrstuhl für Bodenkunde, Dep. für Ökologie, Wissenschaftszentrum Weihenstephan für Ernährung, Landnutzung und Umwelt, Technische Univ. München, D-85350 Freising-Weihenstephan, Germany

^* Corresponding author (spielvogel{at}wzw.tum.de)

	ABSTRACT

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS DISCUSSION CONCLUSIONS REFERENCES

 
Long-term effects of forest dieback on soil organic matter (SOM) pools in an unmanaged high-elevation Norway spruce [Picea abies (L.) H. Karst.] forest were investigated by comparing three dieback sites with three adjacent sites with healthy spruce. The aim was to assess the long-term biogeochemical impact of a major disturbance in unmanaged forest ecosystems due to potential modification of their soil organic carbon (OC) pool. The spruces at the dieback sites had been killed by a bark beetle (Ips typographus) infestation 25 yr ago. Total OC stocks in the topsoil of all sites were determined; different OC pools were quantified by density/particle-size fractionation and ¹³C cross polarization magic angle spinning nuclear magnetic resonance (CPMAS NMR) spectroscopy. Additionally, the aboveground and belowground plant-derived OC input was quantified and the OC species were analyzed. The forest floor at the dieback sites showed significantly lower OC stocks, storing <70% of the respective amount at the sites with healthy forest. In the forest floor and the light fraction of the Ah horizons at the dieback sites, the C/N ratio and the concentration of O-alkyl C in the OM were significantly lower, whereas the OM of the other fractions and of the bulk SOM in Ah horizons showed no changes. The observed SOM changes were caused primarily by an accelerated SOM turnover and only secondarily caused by changed litter input. Particularly the OC pool with a low degree of humification was reduced in its amount and increased in its decomposition status while the more stable C fractions showed no significant changes.

Abbreviations: CPMAS, cross polarization magic angle spinning • CWD, coarse woody debris • FWD, fine woody debris • NMR, nuclear magnetic resonance • OC, organic carbon • SOM, soil organic matter

	INTRODUCTION

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS DISCUSSION CONCLUSIONS REFERENCES

 
PREDICTED FUTURE CLIMATE CHANGE will probably be associated with an increased risk of severe disturbance in forest ecosystems, including stand dieback events, on large forested areas in the Northern Hemisphere (Dale et al., 2001; Classen et al., 2005). The loss of the tree canopy and the associated change in ground vegetation cover (e.g., increased growth of grasses and herbaceous plants) after forest dieback is thought to result in a changed OC input rate and input composition. Additionally, microclimate conditions are modified, affecting humus decomposition (Morecroft et al., 1998; Classen et al., 2005). Both effects are supposed to change the SOM stock and its composition. Different SOM compounds exhibit differences in their recalcitrance due to their specific properties regarding structural chemistry, solubility, aggregation capacity and behavior, and interaction with mineral surfaces. Moreover, the cation exchange capacity, the water storage capacity, and the sorption properties of forest soils for nutrients and pollutants are also strongly affected by their SOM pool. Thus changes in SOM stock size and composition are of major relevance for forest productivity, and for the cycling of water, nutrients, and pollutants in the respective ecosystem. On a larger scale, changes in SOM result in modified C turnover and C storage within forested landscapes. Therefore, detailed knowledge about responses of the organic matter pool of forest soils to natural forest disturbances and vegetation changes is of interest not only for soil scientists, but also for forest ecologists, ecosystem modelers, and hydrologists.

The influence of tree canopy loss on soil OC pools has been the issue of several publications in the last 30 yr. Covington (1981) described differences in forest floor OC stocks under northern hardwood forests that had been clearcut at different dates over the past century. He reported a decline of the forest floor organic matter stock by 50% within 20 yr following clearcutting. He interpreted this C loss to be a result of increased humus decomposition and changes in woody litter input after tree canopy removal and during establishment of the succeeding stand. The results of Covington (1981) have been confirmed by other researchers (Mattson and Smith, 1993; Brais et al., 1995). However, some recent studies (Federer, 1984; Mroz et al., 1985; Johnson et al., 1995) report only small forest floor OC losses after clearcut, and others (Huntington and Ryan, 1990; Johnson and Todd, 1998) did not find any differences between the OC stocks of forest floors before and after loss of the tree canopy. Some authors even reported OC gains in the forest floor after loss of the tree canopy (Johnson et al., 1985; Mattson and Swank, 1989; Hendrickson et al., 1989). Consequently, the key processes that had been hypothesized by Covington (1981) to modify the forest floor OC pool after canopy loss, namely changes in humus decomposition rate and in root litter input, have been challenged in the last years (Yanai et al., 2003). Other explanations for observed OC losses of the forest floor after canopy loss have been proposed (Yanai et al., 2003), including increased erosion, bioturbation, and leaching of dissolved organic matter, but also the transfer of forest floor material into the mineral topsoil. Forest operations often disturb soils, transferring mineral soil material to the soil surface and burying forest floor organic matter into the mineral soil. This mixing results in a reduction of the OC stock of the forest floor as typically defined and measured. Thus, in a recent review, Yanai et al. (2003) state that a loss of the tree canopy does not induce an acceleration of SOM decomposition as hypothesized by Covington (1981). However, a weakness of all previous studies is the confounding effect of logging treatments on the forest floor OC stock. This effect is supposed to be significant, but difficult to measure. The examination of canopy loss effects in an unmanaged forest ecosystem provides a unique chance to investigate the importance of changes in litter input and SOM decomposition rates in explaining SOM losses after disturbance unaffected by harvest disturbance.

Moreover, most previous research (Johnson et al., 1985, 1995; Mroz et al., 1985; Hendrickson et al., 1989; Mattson and Swank, 1989; Huntington and Ryan, 1990) has been performed to determine short-term and medium-term (0.5 to 8 yr) effects of canopy loss on forest floor OC stocks. Only a few studies (Covington, 1981; Mattson and Smith, 1993; Brais et al., 1995; Johnson and Todd, 1998) provide information about longer term changes (>10 yr) of SOM pools in forest floors after tree loss. For a better understanding of the soil C dynamics after forest dieback, it is furthermore necessary to know how different C pools and different C species are affected by a reduced or removed canopy and a changed ground vegetation cover. Using solid state ¹³C CPMAS NMR spectroscopy, different OC species in SOM can be quantified (Kögel-Knabner et al., 1988). The decomposition status of forest floor organic matter after disturbance can also be assessed using its alkyl-C/O-alkyl-C ratio as a measure (Baldock et al., 1997).

Based on the ambiguous results of previous clearcut studies, the objective of our study was to test the following hypotheses:

1. Organic C stocks in the forest floor and the topsoil at three study sites, where the stands have been killed by bark beetle infestation 25 yr ago, do not differ from the respective pools in the soil of three adjacent sites that have remained stocked with healthy Norway spruce forest.
   
2. In addition to the total OC stock, the chemical composition of the SOM in the forest floor and the topsoil is also the same in the dieback sites and the sites under healthy spruce forest.
   

If either of the two hypotheses should be refuted, differences between sites with and without forest dieback would be interpreted as an effect of forest dieback.

	MATERIALS AND METHODS

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS DISCUSSION CONCLUSIONS REFERENCES

 
Study Sites and Soils
The six study sites are located in the National Park Bayerischer Wald, Germany (54°42' N, 13°14' E; 54°26' N, 13°25' E). The climate at the sites is cool (mean annual temperature: 5.7°C) and humid, with the annual precipitation ranging from 1150 to 1300 mm (Elling et al., 1975). The elevation of the study sites ranges between 700 and 1150 m above sea level (a.s.l.), and the slope inclination between 0 and 25°. The natural forest association in higher elevations of the National Park Bayerischer Wald for forest below 1150 m a.s.l. is a Calamagrostio villosae-Piceetum barbilophozietosum, which is dominated by Norway spruce with single European beeches (Fagus sylvatica) and sporadic rowan trees (Sorbus aucuparia) (Elling et al., 1975). In 1970, the National Park Bayerischer Wald was established, and mechanical site treatment, insect control, and artificial stand regeneration were prohibited. Beginning in 1977, large areas of the National Park have been repeatedly infested with bark beetles resulting in widespread forest dieback. A considerable portion of the former forest has lost its canopy, and the vegetation at the sites is now dominated by grasses (Calamagrostis villosa, Avenella flexuosa, Luzula sylvatica) and herbaceous plants (e.g., Cirsium sp., Oxalis acetosella).

Soils in the study area comprise Typic Dystrudepts (Dystric Cambisols), Andic Dystrudepts (Dystric Cambisols with low bulk density), and Entic Haplorthods (Entic Podzols) (Soil Survey Staff, 2003; IUSS Working Group Reference Base, 1998). These soil types are typical for mountainous regions with temperate humid climate and cover large areas of the National Park Bayerischer Wald. Both Inceptisols were formed from quaternary deposits, the Spodosol from granite. A description of basic soil properties is given in Table 1. The soil horizons were classified according to the FAO Guidelines for Soil Description (FAO/UNESCO, 1990).

View larger version (52K): [in this window] [in a new window]   Fig. 1. Location and spatial distribution of the studied sites and sampled soil profiles in the National Park Bayerischer Wald (Germany) in July 2002. Map based on data obtained from the database of the High-Tech-Offensive der Bayerischen Staatsregierung (Regionalkonzept der Regierung von Niederbayern), Projects 33–1 and 33–6.

From the Ah horizons of all soil profiles, disturbed and undisturbed samples were taken for chemical analyses and to calculate OC stocks. All samples were taken from soil pits, which were randomly distributed over the respective variants. Undisturbed samples were obtained using five stainless copper cores per horizon with an inner diameter of 8.8 cm and a height of 6.0 cm. Forest floor samples were taken at five randomly selected places within a 1-m distance around the soil profiles, using metal frames (20 cm x 20 cm). All samples were air dried to constant weight, and the coarse fraction > 2 mm was removed by dry sieving. Roots and coarse woody debris (CWD) were also removed. A subsample of the sieved soil material was ground in a ball mill to a particle size of <100 µm.

Density Fractionation
One Ah horizon sample per site was subjected to a density fractionation. Thirty grams of sieved fine earth were agitated in 500 mL of deionized water (density at 20°C: 0.998 g cm^–3) for 24 h. Then the agitation was stopped, and after another 24 h the material floating on the deionized water was obtained as the light fraction and freeze-dried. Deionized water was chosen for the density fractionation procedure after a preliminary study, in which we compared that fractionation procedure with the fractionation conducted with Na polytungstate according to Kaiser et al. (2002) with three different densities (1.4, 1.6, and 2.0 g cm^–3). In the preliminary study, we investigated the samples that had been fractionated with Na polytungstate with a scanning electron microscope (JEOL JSM 5900 LV, JEOL Corp., Tokyo, Japan). We observed that these samples contained a considerable amount of tungsten even after repeated washings with deionized water until the electrical conductivity was <50 µS cm^–1, making a reliable mass balance impossible. Since the C/N ratios and the solid state ¹³C NMR spectra of the different density fractions obtained by separation with Na polytungstate (1.4, 1.6, 2.0 g cm^–3) and with deionized water (0.998 g cm^–3) exhibited no differences, we decided to use deionized water for the density fractionation.

Particle-size Fractionation
After density fractionation, the remaining sediment was freeze-dried and subjected to a particle-size fractionation. In the first step the sediment was treated by ultrasound with a soil/water ratio of 1:5 to disperse macroaggregates (>200 µm) according to Amelung and Zech (1999). A Branson sonifier 250 with 200 W power output to the converter and equipped with a titanium probe (12.5-mm diam.) was used. The sonifier was operated at a frequency of 20 kHz and at 80% ultrasonic intensity for 3 min. According to Schmidt et al. (1999), the ultrasonic dispersion energy was calculated as energy output between 60 and 100 J mL^–1. The coarse sand and medium sand fractions were isolated by subsequent wet sieving (200 µm mesh size). The aqueous suspension containing the particles < 200 µm (soil/water ratio: 1:10) was subjected to a second ultrasonication process for 18 min, applying between 440 and 500 J mL^–1, which is sufficient for complete dispersion of microaggregates <200 µm (Rumpel et al., 2004). During this second ultrasonication step, the temperature was kept at 20–30°C using a water cooling jacket. After complete dispersion, the samples were separated into the fractions 2 to 200 µm (silt, fine sand), and <2 µm (clay). The separation was performed combining a wet sieving (20-µm mesh size) and a sedimentation procedure (Atterberg cylinder).

Determination of Soil Physical Variables
Bulk Density
The bulk density of the Ah horizons was determined by the core method (Blake and Hartge, 1986) with five replicates per horizon (diameter 8.8 cm, height 6 cm). The soil cores were dried at 105°C and subsequently weighed. The bulk density of all samples was calculated from their respective masses and the volume of the cores (364.9 cm³). The bulk density of the forest floors was determined by drying samples that had been taken with metal frames (20 x 20 cm, 5 replicates per profile) at 40°C to avoid loss of OC. The dried samples were weighed, and their bulk density was calculated from their respective masses, the area of the metal frame, and the measured forest floor thickness.

Coarse Fraction
The coarse fraction (>2 mm) of the soil samples taken with the cores was removed by dry sieving after determination of bulk density. The coarse material was weighed and its volume was calculated by dividing its mass by 2.65 g cm^–3 (density of quartz). The total stone content of the soils was determined for each horizon by excavating all stones from the profiles. The stones were cleaned, dried at 105°C, and subsequently weighed. The volume of the stones was determined by the displacement of water.

Assessment of Plant-Derived Organic Carbon Input
The input of plant necromass to the soils at the different study sites, including CWD and fine woody debris (FWD), root litter from trees, grasses, and shrubs, bark, twigs, tree foliage as well as leaves and other shoot compartments of annual grasses and shrubs, was either measured or estimated, depending on the litter source. Exact data about the organic matter input at the study sites over the entire 25-yr period after the bark beetle infestation in 1977 are not available. Therefore, the organic matter input at the studied sites, originating from different litter sources, was estimated from data acquired in 2001 (root litter: 2004). The change of the ecosystems at the dieback sites from a mature Norway spruce forest to a grass-dominated open area with interspersed defoliated trunks as remainders of the former trees occurred within a few months after the bark beetle attack. Therefore, our assessment of the relative contribution of different litter sources at the dieback sites as determined in 2001 and 2004 should provide fairly representative data for the 25-yr period after bark beetle infestation. Data for aboveground CWD and FWD as well as for ground vegetation biomass and litterfall in 2001 were obtained from the database of the National Park Bayerischer Wald (Nationalparkverwaltung Bayerischer Wald, 2001), data used to calculate the annual OC input from root necromass were gathered in 2004.

Coarse Woody Debris
Coarse woody debris (woody debris >10 cm diam. and >1 m length) was estimated according to Harmon and Sexton (1996): The volume of lying CWD was determined by measuring the lengths of all boles and their diameters at both ends on three 0.1-ha plots. For standing dead wood, the diameter at breast height (DBH) was measured, and the volume was estimated from DBH, using relationships developed for living trees (Garman et al., 1995). The length of broken boles was estimated by experienced workers and their volume was calculated from their DBH and the estimated height. To convert these metric data into mass values, the mean density of about 20 boles and logs per 0.1-ha plot was determined and correlated to a site-specific decay class system (Harmon and Sexton, 1996). The density of the other trunks and boles was estimated by assigning them to one of the five decay classes, based on an assessment of (i) the presence of needles, twigs, and branches, (ii) bark cover, (iii) friability or crushability of wood, (iv) color of wood, and (v) movability of the branch stubs. According to Nationalparkverwaltung Bayerischer Wald (2001), 57% of the CWD at the Andic Dystrudept site, 49% of the CWD at the Entic Haplorthod site, and 41% of the CWD at the Typic Dystrudept site was classified in decay Class 2; the respective percentages for decay Class 3 were 29, 34, and 38%. For decay Class 2, the mean wood density was 0.383 g cm^–3, for decay Class 3 it was 0.338 g cm^–3. The remaining CWD in decay Classes 4 and 5 (wood density 0.247 g cm^–3 and 0.113 g cm^–3, respectively) is a result of previous dieback of individual stems and not due to the bark beetle infestation 25 yr ago, thus it is corresponding with old CWD at the healthy spruce forest sites. The density losses during wood degradation are comparable with the values reported by Harmon and Sexton (1996) for different tree species in the Olympics, the Cascades, and the Rocky Mountains under temperate climate.

Annual Litterfall
Annual litterfall rates including needle litter and FWD (<20 cm length and <1 cm diam.) from trees were estimated with five randomly distributed 1.6-m² collection traps per site that were collected weekly. Needle litter was manually separated from fine woody debris. The annual aboveground litter input from grass and herbaceous plants was considered to be equal to the total aboveground plant biomass, which was determined for each site at five randomly distributed plots (area 2 m²). In all cases the litter was dried to constant weight.

Litter Input from Fine Roots
Dead fine roots were obtained from five randomly selected forest floor samples using 20 cm x 20 cm metal frames. In a first step, roots with a diameter between 5 and 10 mm and 10 to 20 mm were removed manually. Then the forest floor samples were screened through a 2-mm sieve and roots > 2 mm were separated. The roots were differentiated into dead and live components based on their color and physical integrity, and all roots in the diameter classes > 2 mm, 5 to 10 mm, and 10 to 20 mm were combined into a single sample. The measured amount of dead fine roots was used as an estimate for the annual fine root litter input into the soil of the different sites. Dead fine roots with a diameter < 2 mm could not be separated quantitatively due to their close association with the soil material and were thus excluded from the estimation.

Sampling of Plant Material for NMR Analysis
Shoots of live grasses (Calamagrostis villosa, Avenella flexuosa) and of thistles (Cirsium spec.), which constitute a major herbaceous plant species at the dieback sites, were manually sampled at the time of leaf senescence in September 2004. Additionally, senescent spruce needles were cut from branches of trees randomly distributed around the soil profiles. Spruce bark was sampled by peeling from healthy spruce trees. Live fine roots were obtained from the metal frames used for forest floor sampling. Adhering soil material was separated from the roots by carefully washing the root samples. Dead roots were manually separated from roots considered as physiologically active. All plant material was dried for 48 h at 40°C.

Determination of Standard Chemical Variables
pH Measurements
The pH of the mineral soil samples was measured after 30 min equilibration in the supernatant of a 2.5/1 (w/w) 0.01 M CaCl₂ solution/soil suspension with a glass electrode. For pH measurements of the forest floor, a 10/1 (w/w) 0.01 M CaCl₂ solution/litter suspension was used.

Determination of Total Carbon and Nitrogen Concentration
The concentration of total C and N in the soil and plant samples was determined in two replicates with an Elementar Vario EL analyzer by dry combustion at 950°C. Since all soil samples were strongly acidic (Table 1) and free of carbonate, the measured total C concentration was equivalent to the OC concentration.

Soil Organic Matter Speciation
Hydrofluoric Acid Pretreatment
In soils with high Fe/C ratios, Fe can severely confound the NMR-spectroscopic assessment of different C species. Because Fe is paramagnetic and leads to shortening of the relaxation times of carbon atoms, certain C structures are underestimated by NMR spectroscopy in iron-rich samples (Skjemstad et al., 1997). Therefore, demineralization of soil samples by hydrofluoric acid (HF) before ¹³C CPMAS NMR spectroscopic analyses is applied routinely to remove paramagnetic substances and to enrich OC (Parfitt et al., 1999; Kiem et al., 2000). Also in this study, mineral soil horizon samples were treated with 10% (v/v) HF before NMR spectroscopy (Schmidt et al., 1997). Approximately 10 g of ground sample were shaken with 50 mL HF for 2 h. After centrifugation at 4°C, the supernatant was removed. The procedure was repeated five times. The remaining sediment was washed five times with 50 mL of deionized water and freeze-dried. For all samples, the HF treatment resulted in OC losses of <11% of the original content. Based on the low C losses and results of earlier studies showing that, within the resolution of solid-state ¹³C NMR, a treatment with HF does not lead to major changes in the structural composition of SOM in topsoil horizons (Preston et al., 1989; Skjemstad et al., 1994; Schmidt et al., 1997), we conclude that the HF pretreatment of our samples has not resulted in a confounding alteration of its OC composition.

Solid State Carbon-13 NMR Spectroscopy
Solid state ¹³C NMR spectra were acquired for all soil and plant samples using a BRUKER DSX 200 NMR spectrometer (resonance frequency: 50.323 MHz, Bruker Corp. Rheinstetten, Germany), applying the CPMAS technique (Schaefer and Stejskal, 1976). The sample was rotated at the magic angle (54.7°) with a spinning speed of 6.8 kHz to avoid line broadening due to orientation-dependent interactions during an NMR experiment (Wilson, 1987). The spectra were obtained with a pulse delay of 500 ms and using a ramped ¹H-pulse during the contact time of 1 ms to circumvent Hartmann-Hahn mismatches (Peersen et al., 1993). Depending on the sensitivity of the sample, the number of accumulated scans was between 10000 and 250000. To improve the signal/noise ratio, a line broadening between 10 and 15 Hz was used before Fourier transformation. The chemical shift is given relative to tetramethylsilan (=0 ppm) and was calibrated with glycine (176.03 ppm).

For quantification of different C species, the ¹³C NMR spectra were divided into four chemical shift regions that are assigned to specific C groups (Wilson, 1987; Knicker and Lüdemann, 1995): 0–45 ppm alkyl C (lipids, cutin, amino acids), 45–110 ppm O/N-alkyl C (carbohydrates, cellulose, methoxyl C, C-N of amino acids, hemicellulose), 110–160 ppm aryl C (lignin, tannin, aromatic compounds, olefines), 160–220 ppm carboxyl/amide and carbonyl C (carboxylic acids, amide, aldehyde, and ketone groups). The respective areas were quantified by integration. All NMR spectra were manually phased and baseline corrected. The variation of integration data of signals due to the treatment of a well-resolved FID (Fourier transformation, phasing, and baseline correction) is <5% (Knicker et al., 2000). The concentrations of different OC groups in the bulk soil material, normalized to dry weight (dw) were calculated by multiplying the proportion of the relative signal intensity in the respective chemical shift region of the ¹³C NMR spectra by the total OC content of the sample, fore example, for aryl C: aryl C (g kg^–1 sample) = [OC (g kg^–1 sample) · aryl C (%)]/100 (%). The ratio of alkyl C to O-alkyl C was calculated as a measure of the degree of SOM decomposition according to Baldock et al. (1997). This approach is based on the circumstance that chemical changes associated with an increasing degree of decomposition of forest floor OM are characterized by an increase in alkyl-C and a decrease in O-alkyl C contents, and thus an increased alkyl C/O-alkyl C ratio (Kögel-Knabner et al., 1988; Baldock et al., 1997).

Calculation of Organic Carbon stocks
For each profile, OC stocks (kg OC m^–2) in the forest floors and mineral soil horizons were calculated by multiplying the respective concentration (g OC kg^–1 forest floor) by its respective dry bulk density (g soil cm^–3; determined from the metal frames), and by the respective thickness of the different horizons. All soil C stocks were corrected for the coarse fraction of the respective horizon.

Statistics
The assumptions of normality and homogeneity of variances were tested for the measured parameters using Fractile Diagrams (Nielsen and Wendroth, 2003) and applying the Kolmogoroff-Smirnoff test. The significance of differences between treatments was tested with the U-test according to Wilcoxon, Mann, and Whitney (Sachs, 1999), since the parameters OC and N were not normally distributed and showed non-homogeneity of variance. All statistics were computed using SPSS Version 12.1.

	RESULTS

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS DISCUSSION CONCLUSIONS REFERENCES

 
Forest Floor
All soils considered, the thickness of the forest floor was significantly smaller under bark beetle-infested stands compared with the soils at the forest floor with healthy spruce forest (p 0.05). This was mainly due to a significantly (p 0.01) smaller thickness of the Oa layer in all soil types (Table 2), whereas the the Oi and Oe layers showed no significant differences between the dieback sites and the sites with healthy spruce forest. The difference in Oa layer thickness between dieback site and respective healthy spruce forest site was most pronounced for the Typic Dystrudepts. Here, the Oa layer at the dieback site was <0.5 cm thick for three of the five studied profiles. In contrast to the sites with healthy spruce forest, at the dieback sites the boundaries between Oe and Oa layers as well as between the forest floor and the mineral topsoil were diffuse and the layers hardly separable as a result of an intensive penetration of the humic topsoil by grass roots. The pH value of the forest floor at the dieback sites was slightly but not significantly higher if all soil types were considered. However, the difference was significant (p 0.05) for the Inceptisol sites (Table 1). For the Spodosols, the differences between morphological forest floor properties at the dieback site and the site with healthy spruce forest were less pronounced, and the pH values did not differ at all (Tables 1 and 2).

View larger version (35K): [in this window] [in a new window]   Fig. 2. Total C and N concentration and OC stock of the forest floor and the Ah horizons of the studied soils, sampled in the National Park Bayerischer Wald (Germany) in July 2002 (arithmetic mean value ± standard deviation of values acquired from 5 profiles per site). Asterisks indicate significant differences (p 0.01) tested with U-test according to Wilcoxon, Mann, and Whitney.

View larger version (21K): [in this window] [in a new window]   Fig. 3. Comparison of CPMAS 13C NMR spectra of the forest floor (a, b) and the Ah horizon (c, d) of a Typic Dystrudept under healthy Norway spruce forest and under an infested stand in the National Park Bayerischer Wald (Germany), sampled in July 2002.

View larger version (38K): [in this window] [in a new window]   Fig. 4. Stocks of different carbon species (a) in the forest floor and (b) in the Ah horizons of the studied soils in the National Park Bayerischer Wald (Germany), sampled in July 2002 (arithmetic mean value ± standard deviation of values acquired from 5 profiles per site). Asterisks indicate significant differences (*: p 0.05; **: p 0.01) tested with U-test according to Wil–coxon, Mann and Whitney (Sachs, 1999).

A Horizons
Bulk Soil Samples
The total OC stocks in the A horizons of the soils under healthy spruce forest and bark beetle-infested stands were not significantly different (Fig. 2). On average, the A horizons contained 4.2 kg OC m^–2 at the dieback sites and 3.7 kg OC m^–2 at the sites with healthy spruce forest. With 5.7 kg OC m^–2 at the site with healthy spruce forest and 7.7 kg m^–2 at the dieback site, the Andic Dystrudepts contained the largest OC stocks. The Typic Dystrudepts contained 3.7 (healthy stand) and 2.9 (infested stand) kg OC m^–2, the Entic Haplorthods only 1.8 (healthy stand) and 1.9 (infested stand) kg OC m^–2 (Fig. 2). Also the C and N concentration in the A horizons of the soils under healthy spruce forest and bark beetle-infested stands exhibited no significant differences. For the Inceptisols, due to the large bulk density and thickness of the A horizons, the OC stock in that horizon is larger than in the forest floor. In contrast, for the Spodosols, which are characterized by a thick forest floor and a shallow A horizon (Table 1), the OC stock in the forest floor exceeds that of the Ah horizons.

The ¹³C CPMAS NMR spectra of the A horizons of the Typic Dystrudepts under healthy spruce forest (Fig. 3c) and at the bark beetle-infested site (Fig. 3d) are almost identical. From the spectra it can be concluded that in contrast to the forest floor there is no difference between the chemical composition of the C pool in the A horizons at the sites with healthy spruce forest and the bark beetle-infested sites. For the Typic Dystrudept (Fig. 3), but also for the other soil types (spectra not shown), the relative contribution of alkyl C to the total OC stock is increased in the Ah horizons compared with the forest floor (Fig. 4b). On the other hand, the relative contribution of O-alkyl C is decreased. According to Baldock et al. (1997), this finding demonstrates the increased decomposition status of the OC stock in the Ah horizons relative to the forest floor. Twenty-five years after forest dieback, no significant differences in the chemical composition of the OC in the Ah horizons between soils under healthy spruce forest and bark beetle-infested stands could be found for any of the soil types (Fig. 4b).

Light Fraction and Different Particle-Size Fractions of the A Horizons
The light fraction of the Ah horizons was of particular interest for our study, since this fraction consists of relatively young, highly decomposable organic material (Christensen, 1992), and thus might be particularly influenced by a strongly reduced or totally removed canopy and a changed ground vegetation cover. The contribution of OC stored in the light fraction to the total OC stock of the A horizon is about 10% for the dieback and the healthy spruce forest sites (Table 3). Similar to the forest floor organic matter, the C/N ratio of the light fraction under bark beetle-infested stands is significantly lower in all soil types (p 0.01) compared with the respective healthy stands, whereas the C/N ratios of the bulk soil in the Ah horizons show no significant differences between healthy spruce forest and bark beetle-infested stands.

	DISCUSSION

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS DISCUSSION CONCLUSIONS REFERENCES

 
Our results show that 25 yr after forest dieback due to bark beetle infestation, the forest floor of the dieback sites is characterized by (i) a significantly smaller thickness of the Oa layer, accompanied by (ii) significantly smaller C/N ratios, and (iii) significantly smaller forest floor OC stocks (Table 2 and Fig. 2). Thus, Hypothesis 1, claiming that OC stocks in the forest floor of the dieback sites do not differ from the respective stocks in the forest floors of those sites which have remained stocked with Norway spruce, must be refuted.

To detect whether lower OC litter input rates are responsible for the smaller OC stocks in the forest floor of the dieback sites, we compared the annual input of FWD, needle litter, grass litter, litter from herbaceous plants, and fine root litter to the forest floor of both the dieback sites and the sites with healthy spruce forest, based on data acquired in 2001 and 2004. Additionally, we estimated the input of CWD within the last 25 yr.

At the sites with healthy spruce forest, the CWD input is associated with normal tree mortality and not a result of a single disturbance event. As a consequence, the number of standing and downed dead trees and logs is considerably lower in healthy stands compared with the dieback sites with values as low as 45 dead trees ha^–1 at the Entic Haplorthod site and 76 dead trees per ha at the Andic Dystrudept site. Furthermore, due to the slow sequential dieback of individual trees the different dieback times of the single logs and boles, the range of CWD decay classes at the healthy spruce forest sites comprises all five decay classes.

The estimated amount of CWD at the dieback sites ranges from 105 Mg dry wood mass ha^–1 (about 720 dead trees) at the Entic Haplorthod site to 286 Mg dry wood mass ha^–1 (about 1250 lying and standing dead trees) at the Andic Dystrudept site. The larger amount of CWD at the two Dystrudept sites compared with the Haplorthod site is due to higher stem numbers ha^–1 before dieback. At all three sites about 80% of the CWD show no significant signs of decay even 25 yr after the dieback event. In fact, 86% of the CWD at the Andic Dystrudept site, 83% of the CWD at the Entic Haplorthod site, and 79% of the CWD at the Typic Dystrudept site was classified in decay Class 2 or 3. Based on these data, the mean mass loss of CWD for the period between 1977 and 2001 is about 20% of the original coarse woody biomass, corresponding to an annual OC input of 100 g OC m^–2.

The annual aboveground OC input from FWD, needles, grasses, and herbaceous plants was significantly lower at the dieback sites compared with the aboveground input at the healthy spruce forest sites (Table 5). In contrast, the annual belowground OC input into the topsoil at the dieback sites significantly exceeded the belowground input at the healthy spruce forest sites due to an intense penetration of the dieback sites with grass roots and effective production of grass root litter. Considering the larger root litter input and the estimated additional input of about 100 g OC m^–2 yr^–1 from CWD at the dieback sites, we conclude, that the total annual OC input to the soils of the dieback sites is slightly (about 10%) larger compared with the respective sites with healthy spruce forest.

Improved microclimate conditions for microbial litter decomposition, as reported for forested areas after canopy reduction due to insect infestation by Classen et al. (2005), seems to have occurred at the dieback sites in the National Park Bayerischer Wald. This is indicated by the observed differences in humus type, which is biologically inactive mor under healthy spruce stands, but a more active moder at the dieback sites. The differences between our results and the results of earlier studies on clear-cut stands (Mattson and Swank, 1989; Hendrickson et al., 1989; Huntington and Ryan, 1990; Johnson and Todd, 1998) may be due to the circumstance that in unmanaged forests like the National Park Bayerischer Wald, the recovery of the forest floor OC pool may take more time because of a retarded establishment of the next stand generation. At the dieback sites, the establishment of a new forest stand is strongly retarded by the lack of fructiferous old spruces, the small amount of well-decayed CWD with direct soil contact acting as seed-bed, and heavy competition by ground vegetation.

The observed smaller OC stock in the forest floor of the dieback sites compared with the sites with healthy spruce forest is predominantly due to a lower forest floor thickness, and not a decrease in OC concentration (Table 2 and Fig. 2). This indicates that any enhanced bioturbation at the study sites has not resulted in a substantially increased mixing of forest floor and mineral soil material and an associated reduction of the forest floor SOM pool. The lack of a decrease in the OC concentration of the forest floor corresponds with the fact that the OC concentration and the OC stock in the Ah horizons are not significantly increased at the dieback sites compared with the sites with healthy spruce stands.

Beyond the observed changes in the forest floor OC stock, changes in the content of OC species was also apparent in bark beetle-infested stands. For example, the O-alkyl C stock in the forest floor of the dieback sites is up to 40% smaller compared with the forest floor of the healthy spruce forest sites, whereas the stocks of alkyl C and carboxyl C are identical in the forest floors under infested and healthy stands. The changed site conditions after forest dieback apparently induced a preferential loss of easily decomposable O-alkyl C species, such as polysaccharides. As a consequence, the alkyl C/O-alkyl C ratio of the forest floor at the dieback sites, a sensitive indicator for the decomposition status of SOM (Baldock et al., 1997), is significantly smaller compared with the forest floor under healthy Norway spruce stands, indicating an increased decomposition status.

The observed differences in forest floor OC speciation for the dieback sites relative to the healthy spruce forest sites supports our conclusion drawn from the observed lower total OC stocks and higher C/N ratios in the dieback sites. In both cases the differences between the forest floors of the dieback sites and the healthy spruce forest sites are primarily a result of accelerated organic matter decomposition and only secondarily a result of changed litter input. In fact, changes in plant-derived OC input cannot be a cause of the observed decrease in the alkyl C/O-alkyl C ratio of the forest floor organic matter at the dieback sites, since the alkyl C/O-alkyl C ratio of the investigated plant-derived OC input at the dieback and healthy sites is about the same (Table 6).

View larger version (22K): [in this window] [in a new window]   Fig. 5. 13C CPMAS NMR spectra of (a) aboveground and (b) belowground plant material with significant contribution to the litter input for the studied soils in the National Park Bayerischer Wald (Germany); sampled at the healthy and the bark beetle-infested sites in September 2004.

Our study also shows that the major part of the SOM pool in the mineral topsoil remained unaltered 25 yr after dieback of the original stand due to bark beetle infestation. An exeption to this generally unaltered condition is a change in the OC of the light fraction of the Ah horizons. Similar to the bulk SOM in the forest floor (Tables 3 and 4), the C/N ratio and the O-alkyl C concentration in the light fraction of the Ah horizons are significantly smaller at the dieback sites compared with the sites with healthy spruce stands (Tables 3 and 4), although the total OC stock in the light fraction is identical for both sites. The similarity of the bulk SOM in the forest floor and the light fraction in the Ah horizon can be explained by the fact that, like the SOM in the forest floor, the light fraction of the Ah horizons also mainly consists of root necromass supplemented with litter fragments and thus of relatively young and easily decomposable OC compounds. The discrepancy between this observed change in the light fraction in the absence of a change in bulk SOM can be explained by the fact that the light fraction comprises only 5 to 12% of the total OC stock of the Ah horizons (Table 3), and the predominating soil particle-size fractions in the A horizons do not show different C/N ratios at the dieback sites and the sites with healthy spruce forest.

	CONCLUSIONS

TOP ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS DISCUSSION CONCLUSIONS REFERENCES

 
The forest floor of the studied sites that had been killed by bark beetle infestation 25 yr ago showed significantly smaller OC stocks compared with the forest floor at the sites with healthy spruce forest, refuting the hypothesis that the OC stocks in the forest floor at sites with canopy loss do not differ from the respective stocks in the forest floors of three adjacent sites which have remained stocked with healthy Norway spruce forest. In contrast, our results show that the forest floor C stock of unmanaged temperate forests may respond strongly to 25 yr of changed litter input and changed site conditions. The forest dieback also resulted in a change of the humus type from mor to moder. No differences between the total C stocks in the mineral topsoil of the dieback sites and the sites with healthy spruce forest were detected, demonstrating that the OC pool of the mineral topsoil is fairly stable against changes in litter input and canopy cover.

The smaller forest floor OC pool in the forest floor of the dieback sites was associated with a different chemical composition of the forest floor organic matter compared to the sites with healthy spruce forest, refuting our second hypothesis that the chemical composition of the SOM in the forest floor is the same for the dieback sites and the sites under healthy spruces. Significantly smaller C/N ratios, O-alkyl C stocks, and aryl C stocks (the latter only at the Haplorthod site) resulted in a larger relative contribution of alkyl C to the forest floor OC pool under the bark beetle infested stands, indicating an increased decomposition status of the forest floor organic matter at the infested sites. The C/N ratio and the O-alkyl C concentration of the light fraction (<0.998 g cm^–3) of the Ah horizons, comprising 5 to 12% of total OC in these horizons, was also smaller at the dieback sites, whereas the organic matter in bulk soil and in the particle-size fractions of the Ah horizons showed no significant differences in their chemical composition.

The hypothesis that differences in SOM pools between respective sites with and without forest dieback show the same patterns for all three investigated soil types could not be discarded, confirming that the long-term effects of disturbances of natural forest ecosystems on the organic matter pool in the forest floor and the light fraction of the Ah horizons are not substantially modified by the soil type.

In summary, the results of our study show that the forest dieback and the canopy loss at the infested sites resulted in major long-term (>20 yr) changes of the SOM stock and composition of the forest floor. Hence the predicted future climate changes and the associated increased risk of severe disturbances of temperate forest ecosystems, including dieback events on large areas, can be of considerable relevance for forest productivity, nutrient, water, and pollutant cycling, C turnover and C storage by substantially changing the organic matter pool of the humic topsoil.

	ACKNOWLEDGMENTS

 
Funding for this study was provided by the Regionalkonzept der Regierung von Niederbayern im Rahmen der High-Tech-Offensive der Bayerischen Staatsregierung. We thank Heike Knicker, Angelika Kölbl and Ingo Schöning for helpful discussions and Gabi Albert and Vivika Volar for help in the laboratory.

Received for publication January 21, 2005.

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