HOME HELP FEEDBACK SUBSCRIPTIONS ARCHIVE SEARCH TABLE OF CONTENTS		QUICK SEARCH:   [advanced] Author: Keyword(s): Year:  Vol:  Page:

This Article Abstract Figures Only Full Text (PDF) Free Alert me when this article is cited Alert me if a correction is posted Services Similar articles in this journal Alert me to new issues of the journal Download to citation manager Citing Articles Citing Articles via Google Scholar Google Scholar Articles by Oelmann, Y. Articles by Weisser, W. W. Search for Related Content PubMed Articles by Oelmann, Y. Articles by Weisser, W. W. Agricola Articles by Oelmann, Y. Articles by Weisser, W. W. Related Collections Nitrogen Nutrient Cycling Nutrient Management Soil Chemistry

SOIL BIOLOGY & BIOCHEMISTRY

Soil and Plant Nitrogen Pools as Related to Plant Diversity in an Experimental Grassland

^a Inst. of Ecology, Dep. of Soil Science, Berlin Univ. of Technology, Salzufer 11-12, D-10587 Berlin, Germany
^b Geographic Inst., Professorship of Soil Geography/Soil Science, Johannes Gutenberg Univ., Johann-Joachim-Becherweg 21, D-55128 Mainz, Germany
^c Inst. of Chemistry and Dynamics of the Geosphere, ICGIII Phytosphere Inst., Jülich Research Centre, D-52425 Jülich, Germany
^d Inst. of Plant Sciences, ETH Zentrum LFW C56, Universitätsstrasse 2, CH-8092 Zurich, Switzerland
^e Max Planck Inst. for Biogeochemistry, P.O. Box 100164, D-07701 Jena, Germany
^f Inst. of Ecology, Friedrich Schiller Univ. of Jena, Dornburger Straße 159, D-07743 Jena, Germany

^* Corresponding author (yvonne.oelmann{at}uni-mainz.de).

	ABSTRACT

TOP NOTES ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS DISCUSSION CONCLUSIONS REFERENCES

 
Increasing plant species richness decreases soil NO₃^– concentrations in experimental plant mixtures, but the role of particular plant functional groups has remained unclear. Most analyses have focused on particular times of the year or were restricted to NO₃^–. We tested whether plant species richness or particular plant functional groups affect the size of plant-available N pools in soil (KCl-extractable NO₃^–, dissolved inorganic N and organic N [DON] and total dissolved N [TDN] in soil solution) and N concentrations and pools in aboveground biomass. Furthermore, we assessed seasonal variations in the effects of plant species richness and plant functional groups. The experimental grassland site had 86 plots with different combinations of numbers of species (1, 2, 4, 8, 16, and 60) and numbers of functional groups (1, 2, 3, and 4, being grasses, small nonlegume herbs, tall nonlegume herbs, and legumes). In the second year after establishment, increasing species richness reduced soil NO₃^– concentrations (ANOVA, 11% of sum of squares [SS]). The presence of legumes correlated positively with soil NO₃^– concentrations (17% of SS). The presence of grasses significantly decreased soil NO₃^– concentrations (11% of SS). Seasonality had no influence on the relationships between NO₃^– concentrations and species richness. Volume-weighted mean DON and TDN concentrations in soil solution correlated negatively with species richness. Nitrogen pools in plant mixture biomass correlated positively with species diversity (14% of SS), indicating that total N uptake increased with increasing diversity. We conclude that both diversity (either in species or functional groups) and functional composition of grassland mixtures are significant controls of soil and plant N pools. Plant communities with more diverse mixtures are liable to use limiting resources such as N more effectively.

Abbreviations: DON, dissolved organic nitrogen • SS, sum of squares • TDN, total dissolved nitrogen • vwm, volume-weighted mean

	INTRODUCTION

TOP NOTES ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS DISCUSSION CONCLUSIONS REFERENCES

 
Global loss of biodiversity has initiated a debate on its potential impacts on ecosystem functioning and stability (Schulze and Mooney, 1993; Chapin et al., 2000; Schwartz et al., 2000; Loreau et al., 2001). A number of manipulative field experiments have been conducted to investigate the effect of a loss of biodiversity on ecosystem processes such as N cycling (Schmid et al., 2002, Spehn et al., 2005). Because plants are major players in the cycling of N, most studies in which plant diversity was manipulated have addressed the effects of a loss in plant species richness on N pools (Tilman et al., 1996; Hooper and Vitousek, 1997; Loreau et al., 2001; Niklaus et al., 2001b; Scherer-Lorenzen et al., 2003). The underlying mechanistic hypothesis in these experiments is that plants can use available nutrients in a complementary way such that strong competition in diverse systems promotes niche differentiation in space and time, resulting in a more effective community resource use compared with less diverse systems (Hooper et al., 2005). In these experiments, the relationship between the number of plant species or functional plant groups and plant-available NO₃^– in soil was found to be either missing (Symstad et al., 1998) or negative (Tilman et al., 1996, 1997; Hooper and Vitousek, 1997; Hooper and Vitousek, 1998; Symstad et al., 1998; Niklaus et al., 2001a; Scherer-Lorenzen et al., 2003). Despite close correlations, however, it is not clear whether the decrease in plant-available N with increasing diversity was attributable to disproportionate effects of particular species. In manipulated experiments with random selection (with replacement) of species from a species pool, the probability of including a dominant or key species at high diversity levels, known as the "sampling effect" (Huston, 1997), has to be handled cautiously. As an example, legumes often have a strong impact on productivity and soil N and could therefore be such key species (Mulder et al., 2002; Spehn et al., 2002). Further investigations have to address not only the effect of legumes (Craine et al., 2002; Scherer-Lorenzen et al., 2003) but also the effects of other functional plant groups on the soil N pool. Grasses might have a negative effect on plant-available NO₃^– due to their extensive rooting system (Craine et al., 2002). Although most of the previous experimental designs included grasses as a functional plant group (Niklaus et al., 2001b; Mulder et al., 2002; Spehn et al., 2002; Scherer-Lorenzen et al., 2003), the effect of grasses was rarely tested (except for Hooper and Vitousek, 1998).

Recently, organic N forms have been observed to be important in soil solution N fluxes and plant N nutrition, and several researchers have stressed the importance of DON in grasslands (Miller and Bowman, 2002; Bardgett et al., 2003; Streeter et al., 2003; Jones et al., 2004; Weigelt et al., 2005). Dissolved organic N originates from leaching and mineralization of organic materials including plant litter and soil fauna. Recalcitrant compounds of DON can be leached to the groundwater, accounting for 60 to 90% of the total N export in forested catchments (Hagedorn et al., 2000; Hedin et al., 1995). Free amino acids can amount to 60% of water-extractable organic N compounds in a grassland system, which may be rapidly immobilized by direct plant or microorganism uptake (Bardgett et al., 2003). The amount and availability of DON may have consequences for plant competition and complementarity. Therefore, all N species in soil solution have to be considered to assess the effect of biodiversity on the complete plant-available N pool in soil. If plant species specialize in uptake of different forms of N, the potential for direct DON uptake by plants may facilitate niche differentiation (Jones and Kielland, 2002; McKane et al., 2002; Miller and Bowman, 2002). Therefore, highly diverse plant mixtures may take up N from soil more efficiently than less diverse mixtures. If more efficient N use contributes to increasing community biomass production (Mulder et al., 2002; Spehn et al., 2002), plant community biomass N may increase with increasing plant diversity even if N concentrations in plant tissue remain constant.

Plant-available NH₄⁺ and NO₃^– (mineral N) concentrations in soil and soil solution depend on the relation between mineralization (ammonification and nitrification), uptake by plants and soil organisms, denitrification, volatilization, and leaching (Corre et al., 2002; Schimel and Bennett, 2004). These processes are known to be highly variable in space and time and may differ among years, particularly during the growing season, because the activities of plants and soil organisms are governed by varying water availability and temperature (Jamieson et al., 1999; Hook and Burke, 2000). Therefore, the relationship between plant diversity and mineral N concentrations in soil and soil solution or aboveground N concentrations and N pools might vary among different sampling dates during the growing season.

The objectives of this study were to determine whether plant species richness, functional group richness, or the presence of particular functional plant groups controls (i) the size of mineral N pools in soil, (ii) the concentrations of different N species (organic or total N) in soil solution, and (iii) N concentrations and pools in aboveground plant material. Particular attention was paid to the influence of seasonality on the relationship between biodiversity and mineral N pools in soil and soil solution.

We hypothesized that increasing plant species richness decreases plant-available N in soil independent of the positive effect of legumes and the negative effect of grasses on plant-available N in soil. Our second hypothesis was that plant species richness negatively affects DON concentrations in soil solution, resulting in a negative effect on TDN concentrations in soil solution. Third, we hypothesized that aboveground N pools in plant mixtures correlate positively with plant species richness.

	MATERIALS AND METHODS

TOP NOTES ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS DISCUSSION CONCLUSIONS REFERENCES

 
This study is part of a research group working in a grassland experiment (known as "The Jena Experiment," www.the-jena-experiment.de [verified 22 Jan. 2007]) that addresses the role of biodiversity in element cycling and trophic interactions in a grassland ecosystem (Roscher et al., 2004).

Study Site
The field site is located close to the German city of Jena (50°55' N, 11°35' E; 130 m above sea level). Mean annual air temperature is 9.3°C, and mean annual precipitation amounts to 587 mm (Kluge and Müller-Westermeier, 2000). The soil is a Eutric Fluvisol developed from up to 2-m-thick fluvial sediments that are almost free of stones. Due to fluvial dynamics, the texture ranges from sandy loam near the river to silty clay with increasing distance from the river. This systematic variation in soil texture is considered in the statistical design of the experiment by arranging the plots in four blocks parallel to the river on texturally homogeneous subareas and including the block effect in the statistical analyses. Organic C concentrations range from 13 to 33 g kg^–1, organic C/N ratios from 8 to 15, and pH (H₂O) from 7.1 to 8.4. The soil contains some carbonates (15 g kg^–1 CO₃^2––C). The site was used as an arable field for the last 40 yr before the experiment. Converted from grassland in the early 1960s, the site was regularly plowed and fertilized during the last decades for crop production.

The experimental design is described in Roscher et al. (2004). Briefly, the main experiment comprises 86 plots (20 by 20 m, established from seed in May 2002) with different levels of species richness (0, 1, 2, 4, 8, 16, and 60), and different numbers (0, 1, 2, 3, and 4) of functional groups (grasses, small nonlegume herbs, tall nonlegume herbs, and legumes) chosen by the random replacement method from a species pool of 60 species from the Molinio-Arrhenatheretea meadows, Arrhenatherion community (Ellenberg, 1996). Each level of species richness is replicated four times per block, resulting in 12 plots per richness level except for the 60-species plots (n = 3). Scarcely or unestablished species were resown in November 2002. The following criteria were adopted to sow species a second time: (i) abundance <50 seedlings m^–2, or (ii) cover of species <5% in monoculture in September 2002 (Roscher et al., 2004). Knowledge about delayed germination of some species or essential low temperatures to induce seedling development (e.g., Apiaceae, Sanguisorba officinalis L., and Primula veris L.) was considered when deciding about resowing. Heracleum sphondylium L., Anthriscus sylvestris (L.) Hoffm., and Cardamine pratensis L. were sown with 100% of initial seed quantity (without adjustment following laboratory germination tests), because these species failed to germinate completely (in case of the Apiaceae) or established very poorly. Carum carvi L., Pastinaca sativa L., Primula veris L., Bromus erectus Huds., and Luzula campestris (L.) DC. were resown with half of the initial density because only a low number of individuals were established. Additionally Trifolium campestre Schreb. and T. dubium Sibth. were sown with 50% density because of the obligate short life cycle (annual) of these species (Roscher et al., 2004). In 2003, the realized species diversities in the main experiment were (mean ± SE): 1.0 ± 0.0 (n = 16) in the monocultures, 2.0 ± 0.0 (n = 16) in two-species mixtures, 3.8 ± 0.1 (n = 16) in the four-species mixtures, 7.5 ± 0.2 (n = 16) in the eight-species mixtures, 15.0 ± 0.2 (n = 14) in the 16-species mixtures, and 48.3 ± 0.9 (n = 4) in the 60-species mixtures. These numbers are based on an area of 18 m² out of 400 m² and give a rough estimate of the realized species richness. The management of all plots was adapted to extensive meadows used for hay production and mown twice a year in June and September. Plots were not fertilized during the experimental period. To maintain the sown species diversity level, plots were weeded regularly. To avoid disturbance during weeding, people perched on buckets turned upside down.

Sampling
Five soil cores (diameter 0.01 m) were taken at a depth of 0 to 0.15 m of the mineral soil from each of the experimental plots and pooled in September 2002, March, June, and October 2003, March and October 2004, and April 2005. Four months before the first collection period, suction plates (UMS, Munich, Germany, sintered glass, diameter 0.12 m, pore size 1–1.6 µm) were installed at a depth of 0.3 m and coupled with a permanent vacuum system to collect soil solution (at Blocks 1–3, excluding Block 4 at the largest distance from the river). The vacuum system was regulated with the help of manual measures of soil matric potential. In the second year after establishment, we collected cumulatively extracted soil solution fortnightly (n = 700 samples) from March 2003 to May 2004. During summer and autumn 2003 we could not collect any solution due to dry conditions. In September 2002, May and August 2003, and May and August 2004, aboveground plant biomass was harvested on all plots within a frame (0.2 by 0.5 m, height 0.03 m) at four randomly located sites per plot. Data on belowground N storage of the respective mixtures in our study are not yet available.

Extractions and Chemical Analyses
As an estimate of plant-available N, soil inorganic N (NO₃^– and NH₄⁺) concentrations were determined by extraction of soil samples with 1 M KCl solution. In the laboratory, 5 g of fresh soil was shaken with 50 mL of 1 M KCl on an orbital shaker for 0.5 h within 2 d of sampling. In the meantime, samples were stored at 8°C. After allowing particles to settle, the resulting solution was filtered through ash-free paper filters (no. 595, Schleicher & Schuell, Dassel, Germany, pore size 4–7 µm).

Plant biomass of mixtures from all plots was harvested and separated into species for all harvests (except August 2004 when samples were sorted into legume and nonlegume species). After oven drying (70°C, 48 h) to constant weight, plant material from each of the four quadrats collected within each plot was weighed, giving a dry weight for all species in all plots. To determine N in aboveground plant mixtures, living plant material from all samples per plot was pooled together per harvest campaign and ground with a Cyclotec 1093 Sample Mill (Foss Tecator, Hoganas, Sweden) using a 0.5-mm screen for chemical analysis. Twenty milligrams of the ground plant material was analyzed for plant N concentration with an elemental analyzer CE 1110 (Carlo Erba Instruments, Milan, Italy). Plant mixture N pools were then calculated using biomass of mixtures and N concentrations.

Nitrate and NH₄⁺ concentrations were measured in soil solution and soil extracts with a Continuous Flow Analyzer (CFA, Skalar, Breda, the Netherlands). Nitrate was analyzed photometrically after reduction to NO₂^– and reaction with sulfanilamide and naphtylethylenediamine-dihydrochloride to an azo-dye. Our NO₃^– concentrations contained an unknown contribution of NO₂^– that is expected to be small (Moutonnet and Fardeau, 1997). Simultaneously with the NO₃^– analysis, NH₄⁺ was determined photometrically as 5-aminosalicylate after a modified Berthelot reaction. The detection limits of NO₃^– and NH₄⁺ were 0.02 and 0.03 mg N L^–1, respectively. For the solid soil phase, NO₃^– concentrations were expressed in milligrams per kilogram oven-dry (105°C) soil after determining the water content of the samples gravimetrically. Total dissolved N in soil solution was analyzed by oxidation with K₂S₂O₈ followed by reduction to NO₂^– as described above for NO₃^–. Dissolved organic N concentrations in soil solution were calculated as the difference between TDN and the sum of mineral N (NO₃^– + NH₄⁺). In 5% of the samples, TDN concentrations were equal to or smaller than inorganic N concentrations. In these cases, DON concentrations were assumed to be zero. The mean difference between inorganic N and TDN concentrations in these cases was 4.8% after one outlier was removed. Thus, the difference between inorganic N and TDN concentrations was within the error of measurement of all N species (around 5%). To assess the precision of our measurements, we ran two replicate analyses of a selected number of samples (22 for NO₃^– and 15 for TDN) and we calculated the variances of the two replicates. Thereafter, all variances were averaged and the square root was drawn to calculate a mean standard deviation. Expressed as percentage of the averaged means of the two replicates, the mean standard deviations were 5.2 and 5.9% for NO₃^– and TDN analyses, respectively.

Calculations and Statistical Analyses
To eliminate concentration or dilution effects, we calculated volume-weighted mean (vwm) TDN, inorganic N, and DON concentrations for each plot for spring 2003 (March–May 2003), winter 2003/2004 (December 2003–February 2004), and spring 2004 (March–May 2004). Nitrogen concentrations were multiplied by the collected volume of each sampling date. The sum of these N masses divided by the sum of the collected volumes during the respective seasons is the volume-weighted mean. Alternatively, we used arithmetic means that did not change the main results (not shown).

The experiment is based on a factorial design including the different combinations of species number and number of functional groups. Nevertheless, the design is not completely balanced because the two factors depend on each other (Roscher et al., 2004). As the differences in soil N concentrations between bare ground and vegetated plots are mainly attributable to the presence or absence of vegetation regardless of its composition, we did not include the bare-ground plots in our statistical analysis aimed at elucidating a potential effect of vegetation diversity and composition. Before using an ANOVA procedure, the variance homogeneity of the data sets was tested with a Levene test (SPSS 11.5, SPSS Inc., Chicago, IL). All data sets showed variance homogeneity (P > 0.05), except plant N data in September 2002 and May 2004, which were log-transformed before analysis. To test for effects on the response variables with time, a repeated measures ANOVA was performed with block, species richness, functional group richness, presence or absence of legumes, presence or absence of grasses, presence or absence of tall nonlegume herbs, and presence or absence of small nonlegume herbs as between-subject factors and time as the within-subject factor (general linear model [GLM] in SPSS 11.5, Type I sum of squares). We distinguished between a log-linear (log-transformed species richness) and a nonlinear (no transformation of species richness) effect of species richness. The order in which factors were fitted is motivated to detect richness (species and functional groups) effects before the effects of respective functional groups (again fitted in the order of expected importance). The Greenhouse–Geisser correction was used to adjust significance values if the assumption of sphericity was not met. An univariate ANOVA was performed on data from specific time points, using the same variables as factors (GLM in SPSS 11.5, Type I sum of squares). The log-transformed species richness and in some cases the biomass of the legumes and the biomass of the grasses were fitted as covariables. In Type I linear models, changes in the order in which factors are fitted possibly influence the results. Therefore, results of different models were compared, to test if the order of the factors mattered.

	RESULTS

TOP NOTES ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS DISCUSSION CONCLUSIONS REFERENCES

 
Mineral Nitrogen in Soil Solid Phase
In the upper soil layer (0–0.15 m), NO₃^––N concentrations in the KCl extracts ranged from below the detection limit to 42 mg kg ^–1 soil for all sampling dates. Ammonium concentrations were below the detection limit in 70% of all extracts. Maximum NH₄⁺ concentrations were 4.5 mg kg ^–1. The KCl-extractable NO₃^– concentrations were significantly higher on plots with legumes than on plots without legumes at all sampling dates except for spring 2003 (Fig. 1 and 2 ). In addition to the presence of legumes (Table 1), the biomass of the legumes explained part of the variance in soil NO₃^– concentrations in an ANOVA of the specific sampling dates (7–15% of SS, P < 0.001). The presence of grasses decreased concentrations of NO₃^– and explained 11% of the variance in soil NO₃^–, whereas the biomass of grasses had no additional explanatory power (P > 0.05).

View larger version (16K): [in this window] [in a new window]   Fig. 1. Mean concentrations of NO3– in KCl extracts of the different sampling dates for plots (A) with and without legumes and plots (B) with and without grasses. Data are presented as mean ± standard error.

View larger version (21K): [in this window] [in a new window]   Fig. 2. Relationship between species richness differentiated according to the presence of (A) legumes or (C) grasses, or functional group richness differentiated according to the presence (B) of legumes or (D) grasses, and concentrations of NO3– in KCl extracts in autumn 2003. Data are presented as mean ± standard error. Note that there is no 60-species mixture without legumes.

The KCl-extractable soil NO₃^– concentrations differed significantly among sampling dates (Table 1). The effect of the presence of legumes on KCl-extractable soil NO₃^– concentrations varied with time without showing a consistent trend (Table 1). The effects of the presence of grasses or species richness, however, did not vary with time (repeated measures ANOVA, within-subject effects, P > 0.05). The number of functional groups consistently did not have a significant effect on soil NO₃^– concentrations (repeated measures ANOVA, within-subject effects, P > 0.05).

Organic and Total Nitrogen in Soil Solution
Maximum vwm concentrations of DON reached 6.0 mg L^–1. The presence of legumes and grasses affected vwm DON concentrations in soil solution (Fig. 3 , Table 2). The presence of legumes increased vwm DON concentrations. In contrast, vwm DON concentrations were significantly lower on plots with grasses than on plots without grasses. The log-linear increase in species richness was negatively correlated with vwm DON concentrations and contributed 12% to the SS. The interaction between the presence of grasses and species richness was not significant. The effect of functional group richness was not significant, even if fitted before plant species richness (P > 0.05).

View larger version (19K): [in this window] [in a new window]   Fig. 3. Relationship between species richness differentiated according to the presence of (A) legumes or (C) grasses, or functional group richness differentiated according to the presence of (B) legumes or (D) grasses, and volume-weighted mean (vwm) concentrations of dissolved organic N (DON) in spring 2003. Data are presented as mean ± standard error. Note that there is no 60-species mixture without legumes.

View larger version (13K): [in this window] [in a new window]   Fig. 4. Relationship between species richness differentiated according to the respective sampling periods (A–C) and volume-weighted mean (vwm) concentrations of total dissolved N (TDN) and dissolved organic N (DON).

The presence of legumes and grasses had a significant effect on vwm TDN concentrations (Fig. 4, Table 3), with higher vwm TDN concentrations under legume-containing mixtures than plots without legumes. Conversely, the presence of grasses significantly decreased TDN concentrations. Species richness was negatively, linearly correlated with vwm TDN concentrations in soil solution (Table 3). The effect of functional group richness was not significant, even if fitted before plant species richness (P > 0.05). Independent of the order in which interactions were fitted, we observed no significant interaction between species richness and the presence of legumes nor grasses.

View larger version (24K): [in this window] [in a new window]   Fig. 5. Relationship between species richness differentiated according to the presence of (A) legumes or (C) grasses, or functional group richness differentiated according to the presence of (B) legumes or (D) grasses, and volume-weighted mean (vwm) concentrations of total dissolved N (TDN) in spring 2003. Data are presented as mean ± standard error. Note that there is no 60-species mixture without legumes.

View larger version (21K): [in this window] [in a new window]   Fig. 6. Relationship between species richness and mean N pools in plant mixture (aboveground) in May 2003 differentiated according to the presence of (A) legumes or (C) grasses, and functional group richness differentiated according to the presence of (B) legumes or (D) grasses. Data are presented as mean ± standard error. Note that there is no 60-species mixture without legumes.

	DISCUSSION

TOP NOTES ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS DISCUSSION CONCLUSIONS REFERENCES

 
Mineral Nitrogen in Soil Solid Phase
Soil NO₃^– concentrations (0–42 mg NO₃^––N kg^–1 dry soil) were higher than or similar to comparable experiments (Scherer-Lorenzen, 1999: 0–10 mg NO₃^––N kg^–1 dry soil; Tilman et al., 1996: means of the particular species levels 0.2–0.4 mg NO₃^––N kg^–1 dry soil; Hooper and Vitousek, 1998: means of the particular species levels 0–5 mg NO₃^––N kg^–1 dry soil). In contrast to Hooper and Vitousek (1998) and Scherer-Lorenzen (1999), we found almost no water-soluble or KCl-extractable NH₄⁺ in soil, indicating that either plants almost completely took up NH₄⁺ or NH₄⁺ was rapidly nitrified.

Our finding of significantly increased inorganic N concentrations in soils under legume-containing plant mixtures (Fig. 1) is expected and in line with observations of Scherer-Lorenzen et al. (2003). Hooper and Vitousek (1998) also found higher inorganic N concentrations in plots where N₂-fixers were seeded than in plots with other functional plant groups. The effect of legumes on soil N is attributable to lower soil N demand because of N₂ fixation and higher N input via litter and root exudates. The significant decrease in soil mineral N concentrations when grasses were present is probably related to their high root production. The mean proportion of grasses sown to the species mixtures correlated positively with the root length density in ingrowth cores left for 5 mo at our field site (data not shown; P < 0.001, September 2003–June 2004, H. Beßler and C. Engels, Humboldt Univ. of Berlin, personal communication). For legumes the reverse was true (data not shown; P < 0.001, September 2003–June 2004). Hooper (1998) found the highest root biomass in plots where perennial bunchgrasses were seeded. Additionally, the presence of bunchgrasses increased the root biomass of mixtures containing annuals.

Increasing plant species richness resulted in a significant decrease in soil mineral N concentrations (Table 1). As indicated by the significant interaction terms presence of legumes or grasses x species richness, this correlation is much more pronounced in legume-containing mixtures and in mixtures without grasses.

One important aspect of our study was the analysis of N pools at different times during 2 yr and the explicit inclusion of sampling date in the analysis. Nitrogen availability in soil differed among spring, summer, and autumn sampling dates (Fig. 1, Table 1) but the significant relationship between soil mineral N and species richness or the effect of grasses did not change. Niklaus et al (2001a) also found that there was a consistent effect of diversity on NO₃^– concentrations with time, although KCl-extractable NO₃^– concentrations varied significantly throughout different years and different seasons. In contrast, the effect of legumes changed with time without a consistent trend (Fig. 1), which might be attributable to differences in the contribution of legume biomass to total biomass among the sampling dates. Seasonality did not influence the relationship between plant-available N pools in soil and species richness or the effect of grasses. Niklaus et al. (2001a) also found that there was a consistent effect of diversity on NO₃^– concentrations with time, although KCl-extractable NO₃^– concentrations varied significantly throughout different years and different seasons. Legumes increased and grasses decreased NO₃^– concentrations in soil, probably due to their particular plant traits (fixation of atmospheric N₂ vs. extensive rooting system) irrespective of different seasons and years. Plant species richness explained NO₃^– concentrations in soil to a similar extent as each of the functional plant groups (grasses and legumes) and this relationship was consistent across sampling dates.

As our experiment was located in an alluvial floodplain with systematically varying soil texture (increasingly finer perpendicular to the river), we also had to account for possible effects of this variation on the relationship between plant diversity and N concentrations. We did this by assessing a possible block effect because the four blocks of our experiment reflected the variation in soil texture. There was no significant block effect. We conclude that soil mineral N concentrations are related to species richness and the presence of legumes and grasses over the investigated range of soil properties.

The N cycle is complex and we were not able to address all associated processes. All factors fitted in our ANOVA model explained up to 40% of the variance (if legume biomass was included). Thus, there are other important factors affecting NO₃^– concentrations in the studied soils including, e.g., leaching, denitrification, and microbial immobilization.

Organic and Total Nitrogen in Soil Solution
Similar to NO₃^– concentrations, DON concentrations in soil solution decreased with increasing species richness (Fig. 3b, Table 2, up to 18% explained variance). As it is reasonable to assume that part of DON is plant available, we suggest the following two explanations for our finding. First, complementary N acquisition strategies, probably including uptake and incorporation of organic N by plants and microbes (Bardgett et al., 2003), might have resulted in a more complete N uptake from soil. Streeter et al. (2000) reported direct uptake of amino acids by grasses and stressed the importance of grasses for uptake of organic N. In our study, the presence of grasses consistently decreased organic N in soil solution, while the presence of legumes increased organic N (Table 2). Nevertheless, the interaction term presence of grasses or legumes x species richness was not significant, indicating that the relationship between vwm DON concentrations and species richness holds true for mixtures with and without grasses or legumes. Alternatively, microbial degradation of organic N compounds and their transformation to inorganic N forms in soil solution might increase in highly diverse mixtures compared with less diverse mixtures. Little is known about the relationship between plant diversity and microbial activity (Wardle and Nicholson, 1996; Maire et al., 1999; Spehn et al., 2000; Gastine et al., 2003). Volume-weighted mean DON concentrations in soil solution showed a peak during winter 2003/2004. In contrast, Hagedorn et al. (2001) found high leaching of DON during summer and autumn due to stimulated microbial activity in a forested ecosystem. The summer of 2003 was exceptionally dry (Müller-Westermeier and Riecke, 2004), therefore reducing microbial activity across all experimental plots. The suction cups in 0.3 m got rewetted, i.e., again started collecting water in November to December 2003. Thus, DON produced in autumn 2003 reached the suction cups at 0.3-m soil depth in winter. In addition, we suspect that leaching (from litter, dead plant material, and decaying microorganisms) of organic N compounds that were not mineralized due to low winter temperatures caused the increased DON concentrations during winter. The increase in the contribution of vwm DON to vwm TDN concentrations is probably attributable to the decreasing effect of former fertilization (before the establishment of the experiment).

As higher DON contributions to TDN (>60%) only occurred in samples with low TDN concentrations (<1mg N L^–1), the overall effect of species richness on TDN was dominated by the effect on NO₃^– alone. Thus, the effects of grasses, legumes and species richness on TDN were similar to those on soil mineral N (Fig. 5) including the stronger species richness effect in legume-containing mixtures. Legumes and grasses had a similarly strong effect on TDN concentrations (together 22% explained variance) as on KCl-extractable NO₃^– concentrations, whereas species richness accounted for only 9% of SS (Table 3). Thus, species richness plays a more important role for NO₃^– than for TDN concentrations in soil solution.

Similar to mineral N in the soil solid phase, vwm DON and TDN concentrations in soil solution were controlled by the presence of legumes and grasses and decreased with increasing plant species richness. These relationships were strongest at the beginning of the study period.

Plant Mixture Nitrogen Concentrations and Pools
Because plant and soil samples were not taken on exactly the same date, relations between plant and soil N pools must be discussed cautiously. In our experiment, the biomass of mixtures was positively correlated with plant species richness (Roscher et al., 2005). Thus, if N concentrations in plant tissues remain constant, this should lead to the negative effect of plant species richness on plant-available NO₃^– concentrations in soil.

Legumes increased both aboveground N pools in the biomass of mixtures and soil N concentrations (Fig. 1 and 5). The additional N source of legumes has been shown not only to increase the N concentrations in non-N₂–fixing plants growing with legumes (Mulder et al., 2002; Spehn et al., 2002; Lee et al., 2003) but also to increase soil NO₃^– concentrations on plots with legumes (Scherer-Lorenzen et al., 2003).

In contrast, the presence of grasses negatively affected soil N concentrations (Fig. 1). Thus, grass-containing mixtures are supposed to have taken up more N than mixtures without grasses, assuming similar N leaching; however, aboveground N pools and N concentrations in biomass of mixtures even decreased in the presence of grasses (Fig. 6, Tables 4 and 5). Mineralization rates of grass litter and internal N cycling in grasses also influence plant N uptake. Because these data were not measured, we cannot directly relate soil to plant N. Restricted to our measurements, we hypothesize that the observed difference in soil N between plots with and without grasses might be better reflected in belowground plant N pools. In our experiment, the proportion of grasses sown in the mixtures significantly increased root biomass (data not shown; P < 0.001, H. Beßler and C. Engels, Humboldt Univ. of Berlin, personal communication). Thus, a substantial proportion of N might be stored in the roots of grasses, explaining the lower N pool size in grass-containing mixtures in spite of the negative correlation between species richness of grass-containing plots and soil N forms.

For all harvests, N concentrations and N pools in aboveground plant mixture biomass were positively correlated with increasing species diversity (Table 5). Functional group diversity had an effect on N pools in aboveground biomass, but not on N concentrations (Tables 4 and 5). The significant interaction effect of grass presence and functional group richness on N pools shows that complementary resource use is probably playing a role. The logic behind this is that a sampling effect (whereby increasing the number of functional groups increases the likelihood of including a dominant one) would produce no interaction between grass (or other functional group) presence and functional group richness. In contrast, the observed interaction suggests that a combination of grasses and other functional groups means better N uptake of mixtures than grass or other functional groups alone. Legumes increased N pools in aboveground biomass and mineral N concentrations in soil through fixation of atmospheric N₂ in symbiosis with bacteria. Grasses decreased aboveground N pools and mineral N concentrations in soil, probably due to their extensive rooting system. Decreasing N availability in soil with increasing plant species richness might partly be explained by increased uptake of N in more diverse mixtures.

	CONCLUSIONS

TOP NOTES ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS DISCUSSION CONCLUSIONS REFERENCES

 

1. The presence of grasses decreased and the presence of legumes strongly increased soil solution N concentrations. Increasing species richness resulted in significantly decreased plant-available N in soil after functional group richness and the effects of particular plant functional groups have been accounted for. Although differing among seasons, these effects were consistent with time.
   
2. Similar to mineral N concentrations in the soil solid phase, organic and total N concentrations in soil solution were controlled by the presence of legumes or grasses and species richness. The effects were strongest, however, in the beginning of the study period, which was probably related to the strong effect of legumes or grasses and species richness on N pools in aboveground biomass.
   
3. Legumes increased and grasses decreased N pools in aboveground biomass because of fixation of atmospheric N₂ (legumes) and probably because of the increased N pools belowground (grasses). Increasing plant diversity of mixtures resulted in increasing uptake of N in aboveground biomass.
   

While legumes are key species that generally increase soil solution N, grasses tend to use limiting resources such as N more effectively, especially in plant communities with moderate to high richness. If both above- and belowground plant N pools are considered, we suggest that complementary N acquisition gains in importance as the species richness increases in the community.

	ACKNOWLEDGMENTS

 
Thanks to Gerd Gleixner and Sibylle Steinbeiss, Max Planck Inst. for Biogeochemistry, Jena, for their substantial input to this work. We thank the many people who helped with the management of the experiment, especially the gardeners S. Eismann, S. Junghans, B. Lenk, H. Scheffler, and U. Wehmeier, and many student helpers, especially M. Bärwolff, F. Beer, C. Möller, P. Theuring, F. Walsh, and K. Würfel, assisting in the plant sample preparation for N analysis. Thanks also to all the helpers during the weeding campaigns. W. Wilcke acknowledges the Heisenberg grant of the DFG (Wi 1601/3-2). The Jena Experiment is funded by the Deutsche Forschungsgemeinschaft (DFG, FOR 456), with additional support from the Friedrich Schiller Univ. of Jena and the Max Planck Society. Furthermore, we thank Bernhard Schmid and Michael Scherer-Lorenzen for their helpful constructive comments on this work.

	NOTES

TOP NOTES ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS DISCUSSION CONCLUSIONS REFERENCES

 
All rights reserved. No part of this periodical may be reproduced or transmitted in any form or by any means, electronic or mechanical, including photocopying, recording, or any information storage and retrieval system, without permission in writing from the publisher. Permission for printing and for reprinting the material contained herein has been obtained by the publisher.

Received for publication June 1, 2006.

	REFERENCES

TOP NOTES ABSTRACT INTRODUCTION MATERIALS AND METHODS RESULTS DISCUSSION CONCLUSIONS REFERENCES

 

• Bardgett, R.D., T.C. Streeter, and R. Bol. 2003. Soil microbes compete effectively with plants for organic-nitrogen inputs to temperate grasslands. Ecology 84:1277–1287.[Web of Science]
• Chapin, F.S., E.S. Zavaleta, V.T. Eviner, R.L. Naylor, P.M. Vitousek, H.L. Reynolds, D.U. Hooper, S. Lavorel, O.E. Sala, S.E. Hobbie, M.C. Mack, and S. Diaz. 2000. Consequences of changing biodiversity. Nature 405:234–242.[CrossRef][Medline]
• Corre, M.D., R.R. Schnabel, and W.L. Stout. 2002. Spatial and seasonal variation of gross nitrogen transformations and microbial biomass in a northeastern US grassland. Soil Biol. Biochem. 34:445–457.[CrossRef]
• Craine, J.M., D. Tilman, D. Wedin, P. Reich, M. Tjoelker, and J. Knops. 2002. Functional traits, productivity and effects on nitrogen cycling of 33 grassland species. Funct. Ecol. 16:563–574.[CrossRef]
• Ellenberg, H. 1996. Die Vegetation Mitteleuropas mit den Alpen in ökologischer, dynamischer und historischer Sicht. 5th ed. Ulmer, Stuttgart, Germany.
• Gastine, A., M. Scherer-Lorenzen, and P.W. Leadley. 2003. No consistent effects of plant diversity on root biomass, soil biota and soil abiotic conditions in temperate grassland communities. Appl. Soil Ecol. 24:101–111.[CrossRef]
• Hagedorn, F., J.B. Bucher, and P. Schleppi. 2001. Contrasting dynamics of dissolved inorganic and organic nitrogen in soil and surface waters of forested catchments with Gleysols. Geoderma 100:173–192.[CrossRef][Web of Science]
• Hagedorn, F., P. Schleppi, P. Waldner, and H. Flühler. 2000. Export of dissolved organic carbon and nitrogen from Gleysol dominated catchments—the significance of water flow paths. Biogeochemistry 50:137–161.
• Hedin, L.O., J.J. Armesto, and A.H. Johnson. 1995. Patterns of nutrient loss from unpolluted, old-growth temperate forests—evaluation of biogeochemical theory. Ecology 76:493–509.[CrossRef][Web of Science]
• Hook, P.B., and I.C. Burke. 2000. Biogeochemistry in a shortgrass landscape: Control by topography, soil texture, and microclimate. Ecology 81:2686–2703.[CrossRef][Web of Science]
• Hooper, D.U. 1998. The role of complementarity and competition in ecosystem responses to variation in plant diversity. Ecology 79:704–719.[CrossRef][Web of Science]
• Hooper, D.U., F.S. Chapin, J.J. Ewel, A. Hector, P. Inchausti, S. Lavorel et al. 2005. Effects of biodiversity on ecosystem functioning: A consensus of current knowledge. Ecol. Monogr. 75:3–35.[CrossRef][Web of Science]
• Hooper, D.U., and P.M. Vitousek. 1997. The effects of plant composition and diversity on ecosystem processes. Science 277:1302–1305.[Abstract/Free Full Text]
• Hooper, D.U., and P.M. Vitousek. 1998. Effects of plant composition and diversity on nutrient cycling. Ecol. Monogr. 68:121–149.[CrossRef][Web of Science]
• Huston, M.A. 1997. Hidden treatments in ecological experiments: Re-evaluating the ecosystem function of biodiversity. Oecologia 110:449–460.[CrossRef][Web of Science]
• Jamieson, N., R. Monaghan, and D. Barraclough. 1999. Seasonal trends of gross N mineralization in a natural calcareous grassland. Global Change Biol. 5:423–431.[CrossRef]
• Jones, D.L., and K. Kielland. 2002. Soil amino acid turnover dominates the nitrogen flux in permafrost-dominated taiga forest soils. Soil Biol. Biochem. 34:209–219.[CrossRef]
• Jones, D.L., D. Shannon, D.V. Murphy, and J. Farrar. 2004. Role of dissolved organic nitrogen (DON) in soil N cycling in grassland soils. Soil Biol. Biochem. 36:749–756.[CrossRef]
• Kluge, G., and G. Müller-Westermeier. 2000. Das Klima ausgewählter Orte der Bundesrepublik Deutschland: Jena. Berichte des Deutschen Wetterdienstes 213, Offenbach, Germany
• Lee, T.D., P.B. Reich, and M.G. Tjoelker. 2003. Legume presence increases photosynthesis and N concentrations of co-occurring non-fixers but does not modulate their responsiveness to carbon dioxide enrichment. Oecologia 137:22–31.[CrossRef][Web of Science][Medline]
• Loreau, M., S. Naeem, P. Inchausti, J. Bengtsson, J.P. Grime, A. Hector, D.U. Hooper, M.A. Huston, D. Raffaelli, B. Schmid, D. Tilman, and D.A. Wardle. 2001. Biodiversity and ecosystem functioning: Current knowledge and future challenges. Science 294:804–808.[Abstract/Free Full Text]
• Maire, N., D. Borcard, E. Laczko, and W. Matthey. 1999. Organic matter cycling in grassland soils of the Swiss Jura mountains: Biodiversity and strategies of the living communities. Soil Biol. Biochem. 31:1281–1293.[CrossRef]
• McKane, R.B., L.C. Johnson, G.R. Shaver, K.J. Nadelhoffer, E.B. Rastetter, B. Fry, A.E. Giblin, K. Kielland, B.L. Kwiatkowski, J.A. Laundre, and G. Murray. 2002. Resource-based niches provide a basis for plant species diversity and dominance in arctic tundra. Nature 415:68–71.[CrossRef]
• Miller, A.E., and W.D. Bowman. 2002. Variation in nitrogen-15 natural abundance and nitrogen uptake traits among co-occurring alpine species: Do species partition by nitrogen form? Oecologia 130:609–616.[CrossRef][Web of Science]
• Moutonnet, P., and J.C. Fardeau. 1997. Inorganic nitrogen in soil solution collected with tensionic samplers. Soil Sci. Soc. Am. J. 61:822–825.[Abstract/Free Full Text]
• Mulder, C.P.H., A. Jumpponen, P. Högberg, and K. Huss-Danell. 2002. How plant diversity and legumes affect nitrogen dynamics in experimental grassland communities. Oecologia 133:412–421.[CrossRef][Web of Science]
• Müller-Westermeier, G., and W. Riecke. 2004. Die Witterung in Deutschland 2003. p. 71–78. In DWD Klimastatusbericht. Deutscher Wetterdienst, Offenbach, Germany.
• Niklaus, P.A., E. Kandeler, P.W. Leadley, B. Schmid, D. Tscherko, and C. Körner. 2001a. A link between plant diversity, elevated CO₂ and soil nitrate. Oecologia 127:540–548.[CrossRef][Web of Science]
• Niklaus, P.A., P.W. Leadley, B. Schmid, and C. Körner. 2001b. A long-term field study on biodiversity x elevated CO₂ interactions in grassland. Ecol. Monogr. 71:341–356.[CrossRef][Web of Science]
• Roscher, C., J. Schumacher, J. Baade, W. Wilcke, G. Gleixner, W.W. Weisser, B. Schmid, and E.-D. Schulze. 2004. The role of biodiversity for element cycling and trophic interactions: An experimental approach in a grassland community. Basic Appl. Ecol. 5:107–121.[CrossRef]
• Roscher, C., V.M. Temperton, M. Scherer-Lorenzen, M. Schmitz, J. Schumacher, B. Schmid, N. Buchmann, W.W. Weisser, and E.-D. Schulze. 2005. Overyielding in experimental grassland communities—irrespective of species pool or spatial scale. Ecol. Lett. 8:419–429.[Medline]
• Scherer-Lorenzen, M. 1999. Effects of plant diversity on ecosystem processes in experimental grassland communities. Bayreuther Forum Ökologie 75, BITÖK, Bayreuth, Germany.
• Scherer-Lorenzen, M., C. Palmborg, A. Prinz, and E.-D. Schulze. 2003. The role of plant diversity and composition for nitrate leaching in grasslands. Ecology 84:1539–1552.[Web of Science]
• Schimel, J.P., and J. Bennett. 2004. Nitrogen mineralization: Challenges of a changing paradigm. Ecology 85:591–602.[CrossRef][Web of Science]
• Schmid, B., J. Joshi, and F. Schläpfer. 2002. Empirical evidence for biodiversity–ecosystem functioning relationships. p. 120–150. In A. Kinzig et al. (ed.) Functional consequences of biodiversity: Experimental progress and theoretical extensions. Monogr. in Popul. Biol. 33. Princeton Univ. Press, Princeton, NJ.
• Schulze, E.-D., and H.A. Mooney. 1993. Biodiversity and ecosystem function. Ecol. Stud. 99. Springer, Berlin.
• Schwartz, M.W., C.A. Brigham, J.D. Hoeksema, K.G. Lyons, M.H. Mills, and P.J. van Mantgem. 2000. Linking biodiversity to ecosystem function: Implications for conservation ecology. Oecologia 122:297–305.[CrossRef][Web of Science]
• Spehn, E.M., A. Hector, J. Joshi, M. Scherer-Lorenzen, B. Schmid, E. Bazeley-White et al. 2005. Ecosystem effects of biodiversity manipulations in European grasslands. Ecol. Monogr. 75:37–63.[CrossRef][Web of Science]
• Spehn, E.M., J. Joshi, B. Schmid, J. Alphei, and C. Körner. 2000. Plant diversity effects on soil heterotrophic activity in experimental grassland ecosystems. Plant Soil 224:217–230.[CrossRef][Web of Science]
• Spehn, E.M., M. Scherer-Lorenzen, B. Schmid, A. Hector, M.C. Caldeira, P.G. Dimitrakopoulos et al. 2002. The role of legumes as a component of biodiversity in a cross-European study of grassland biomass nitrogen. Oikos 98:205–218.[CrossRef][Web of Science]
• Streeter, T.C., R. Bol, and R.D. Bardgett. 2000. Amino acids as a nitrogen source in temperate upland grasslands: The use of dual labelled (C-13, N-15) glycine to test for direct uptake by dominant grasses. Rapid Commun. Mass Spectrom. 14:1351–1355.[CrossRef][Web of Science][Medline]
• Streeter, T.C., R.F. King, and B. Raymond. 2003. Organic nitrogen in soil water from grassland under different land management strategies in the United Kingdom—a neglected N load to upland lakes? Agric. Ecosyst. Environ. 96:155–160.
• Symstad, A.J., D. Tilman, J. Willson, and J. Knops. 1998. Species loss and ecosystem functioning: Effects of species identity and community composition. Oikos 81:389–397.[CrossRef][Web of Science]
• Tilman, D., J. Knops, D. Wedin, P. Reich, M. Ritchie, and E. Siemann. 1997. The influence of functional diversity and composition on ecosystem processes. Science 277:1300–1302.[Abstract/Free Full Text]
• Tilman, D., D. Wedin, and J. Knops. 1996. Productivity and sustainability influenced by biodiversity in grassland ecosystems. Nature 379:718–720.[CrossRef]
• Wardle, D.A., and K.S. Nicholson. 1996. Synergistic effects of grassland plant species on soil microbial biomass and activity: Implications for ecosystem-level effects of enriched plant diversity. Funct. Ecol. 10:410–416.[CrossRef]
• Weigelt, A., R. Bol, and R.D. Bardgett. 2005. Preferential uptake of soil nitrogen forms by grassland plant species. Oecologia 142:627–635.[CrossRef][Web of Science][Medline]

This Article Abstract Figures Only Full Text (PDF) Free Alert me when this article is cited Alert me if a correction is posted Services Similar articles in this journal Alert me to new issues of the journal Download to citation manager Citing Articles Citing Articles via Google Scholar Google Scholar Articles by Oelmann, Y. Articles by Weisser, W. W. Search for Related Content PubMed Articles by Oelmann, Y. Articles by Weisser, W. W. Agricola Articles by Oelmann, Y. Articles by Weisser, W. W. Related Collections Nitrogen Nutrient Cycling Nutrient Management Soil Chemistry