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Carbon isotope dynamics during grass decomposition and soil organic matter formation.

INTRODUCTION

Stable C isotope ratios ([Delta]13C values) of soil organic matter (SOM) have been used to reconstruct past vegetation dynamics on geological (e.g., Kingston et al. 1994) and ecological time scales (e.g., Dzurec et al. 1985, McPherson et al. 1993) as well as to understand mechanisms of litter decomposition, SOM formation, and SOM turnover (e.g., Balesdent et al. 1987, Benner et al. 1987, Nadelhoffer and Fry 1988). Plants with the [C.sub.3] photosynthetic pathway are substantially more depleted in 13C ([Delta]13C commonly -26 to -28%) than [C.sub.4] plants ([Delta]13C commonly -12.0 to -13.0%) (Farquhar et al. 1989, Tieszen and Boutton 1989). Within these photosynthetic pathways, isotopic signatures of individual species or individual plants vary, particularly for [C.sub.3] plants, and are under both environmental and genetic control. However, these within-group differences are small relative to between-pathway differences, and when corrected for, do not strongly affect the ability to use [Delta]13C values to trace vegetation changes involving photosynthetic types (Tieszen and Archer 1990, McPherson et al. 1993). Recent studies suggest that [Delta]13C values of SOM reflect the isotopic values of associated plant litter inputs with high fidelity under conditions of vegetation stability after correcting for anthropogenic alterations of atmospheric C[O.sub.2] signals (Tieszen and Boutton 1989, Balesdent et al. 1993, Kingston et al. 1994).

However, this conclusion conflicts with current theoretical frameworks describing litter decomposition. The difficulty is that [Delta]13C signatures of plant litter should not remain constant during decomposition if the chemical fractions of plant tissue differ in [Delta]13C signature and decay at different rates (Melillo et al. 1982, Melillo et al. 1989, Berg et al. 1993). In particular, lignin, which decomposes slowly, is substantially depleted in [Delta]13C relative to bulk plant tissue (Benner et al. 1987). Many models of SOM development assume that the lignin fraction of plant litter becomes humus (the recalcitrant or decay-resistant SOM fraction) relatively unmodified by microbial decomposers (e.g., Parton et al. 1987). If lignin is disproportionately preserved during litter decomposition, then bulk tissue [Delta]13C values should decrease from that of the original litter and become more similar to that of the lignin fraction as overall mass loss occurs and the relative concentration of lignin increases. Accumulation of lignin should therefore result in SOM with a carbon isotope value more negative than bulk tissue. Benner et al. (1987) explained the presence of 13C depleted organic matter in Spartina marsh sediments on the basis of selective preservation of isotopically light refractory carbon.

A contrary hypothesis is that a slight enrichment (12%) of 13C occurs during the litter decomposition phase of the plant litter to SOM continuum (Melillo et al. 1989). This enrichment appears to occur at advanced stages of SOM decomposition and at greater depths in the soil horizon, perhaps with the conversion of humic to fulvic acids (Nadelhoffer and Fry 1988, Balesdent et al. 1993). It may also reflect the disproportionate loss of 13C-depleted chemical fractions of plant tissues during decomposition. Alternatively, the rapid turnover of SOM at the soil surface may result in the equilibration of this C with anthropogenically depleted 13C in atmospheric C[O.sub.2] (Keeling et al. 1979), giving the appearance of 13C enrichment in lower soil horizons (Von Fischer and Tieszen 1995).

In this study, we analyzed changes in the isotopic composition ([Delta]13C) of aboveground and belowground plant tissues of [C.sub.3] and [C.sub.4] grass species and associated changes in lignin concentration and isotopic composition during decomposition. We also monitored the short-term accretion of SOM and the contribution of decaying litter to the isotopic signal of SOM in monocultures of these same [C.sub.3] and [C.sub.4] species growing on soils of known initial organic matter content and isotopic composition.

METHODS

Study species and site

The five perennial grass species we studied are all abundant species in successional grasslands and native prairies at the research site, Cedar Creek Natural History Area (CCNHA) in east-central Minnesota (mean annual temperature = 5.5 [degrees] C, mean annual precipitation = 726 mm). The two [C.sub.4] species, Schizachyrium scoparium (Michx.) Nash-Gould (formerly Andropogon scoparius) and Andropogon gerardi Vitm. are native bunchgrasses that commonly dominate tallgrass prairie, savannah, and late-successional grasslands. Of the three [C.sub.3] species, Agrostis scabra Willd. is a native bunchgrass common in early successional grasslands, while Agropyron repens (L.) Beauv. and Poa pratensis L. are Eurasian rhizomatous grasses common in early and midsuccessional grasslands. We will refer to all species by genera. These species have been studied extensively at CCNHA (Pastor et al. 1987, Tilman 1988, Wedin and Tilman 1990, Tilman and Wedin 1991, Wedin and Pastor 1993).

These five species were planted in monocultures and multispecies plots in an experimental garden in 1986. The garden was constructed by removing the top 0.7 m of soil from an abandoned field and then constructing 10 large soil mixtures (each 3 x 12 m) by mixing different amounts of black soil with the subsurface sand left at the garden site. CCNHA lies on a glacial outwash sandplain, and the subsurface sand averaged 93% sand, 3% clay, 4% silt, and 0.09% total C. The black soil, a Duelm sandy loam brought in from outside CCNHA averaged 72% sand, 4% clay, 24% silt, and 1.79% total C. The soil fractions were mixed thoroughly to a depth of 23 cm with a rototiller and ranged from 100% sand to 100% black soil. Details of seeding procedures and other methods are given in Wedin and Tilman 1990 and Tilman and Wedin 1991.

Litter bag decomposition study

Plant tissues for the litter bag decomposition study were collected from monocultures (each 1.5 x 2.4 m) growing on 100% black soil. Aboveground and below-ground tissues for Schizachyrium, Agropyron, and Poa were collected in late October 1989 in portions of the plots from which all previous years' litter had been removed. For Schizachyrium, aboveground tissue contained only recently senesced litter clipped at ground level, while for Agropyron and Poa aboveground tissues also contained 25 and 35%, respectively, green tissue. Belowground tissue samples were collected by washing soil blocks over 1-mm mesh sieves, and contained predominantly live, intact root systems. For Agropyron and Poa, the belowground samples contained both roots and rhizomes (40 and 33% rhizome by mass for Agropyron and Poa, respectively). For Schizachyrium, crowns (enlarged stem bases at the soil surface) were not included with belowground samples. Because Agrostis senesces aboveground at mid growing season and grows new leaves in the fall, recently senesced litter and belowground tissues were collected for Agrostis in early August 1990 with the same methods described above. Andropogon gerardi was not used for the litter decomposition study.

Approximately 9 g of aboveground litter were placed in each 12 x 30 cm litter bag. Aboveground litter bags were constructed with 100% polyester cloth bottoms and fiberglass screen tops (1.7-mm mesh). Below-ground litter bags were commercially available 10 x 20 cm 100% polyester rumen bags (Ankom, Fairport, New York, USA) and contained [approximately equal to]4 g of plant tissues. Aboveground litter was dried at 35 [degrees] C for 2-3 d and belowground litter at 35 [degrees] C for 5-7 d prior to weighing and placement in litter bags. All above- and below-ground litter bags were placed randomly in a common location in a CCNHA field dominated by Schizachyrium and Poa. Five 15 cm deep soil cores were collected at the site of the litter bag study, composited, and analyzed for stable C isotope composition. Above-ground bags were placed on the soil surface, while belowground bags were inserted vertically into cuts made in the soil with a flat spade. Bags for Schizachyrium, Agropyron, and Poa were placed in the field in mid-November 1989, while bags for Agrostis were placed in the field in mid-August 1990. Three replicate bags were collected at time zero for characterization of initial litter chemistry. Four replicate bags were sampled on each subsequent collection date for each of the eight tissue types over the first 2 yr of the study. There were eight sampling times for all species except Agrostis, which had six sampling times.

Initial masses for all bags were adjusted to oven-dry mass. Bags collected from the field were dried at 45 [degrees] C, and remaining tissues were weighed and ground with a cyclone mill (0.25-mm mesh screen). The three time zero replicates for each tissue type were analyzed separately for chemistry and [[Delta].sup.13]C to allow ANOVA analysis (see Table 1). For subsequent collection dates, only a composite sample for each tissue type from each sampling date was analyzed for tissue chemistry. Ash content was determined by heating at 450 [degrees] C for 4 h. Decomposition rate and tissue chemistry are calculated and presented on an ash-free basis. Tissue percent C and percent N was measured with a LECO CHN-800 analyzer. C chemistry was determined by sequential fractionation and extraction with a series of solvents (modified from Effland 1977, see McClaugherty et al. 1985, Geng et al. 1993). Following extraction of non-polar solubles with dichloromethane and polar solubles by [H.sub.2]O, the remaining sample was hydrolyzed in 72% sulfuric acid for 1 h, followed by autoclaving at 120 [degrees] C for 1 h in 1 mol/L sulfuric acid. Mass loss during this step is due to solubilization of complex carbohydrates, mainly cellulose and hemicellulose. The insoluble residue is assumed to be lignin. This fraction often contains compounds that are not true lignin in a biochemical sense, particularly suberin in root samples. Rather, our use of the term lignin is consistent with general useage by forage, wood, and ecosystem scientists (Van Soest 1982, Aber and Melillo 1991). Calculation of percent C for lignin samples on an ash-free basis was based on mass loss during acid digestion, the ash content of bulk litter, and the assumption that the ash fraction was not lost during sequential extraction and digestion. N dynamics of decomposing litter and the effect of litter quality parameters on rates of mass loss are discussed in D. A. Wedin and J. Pastor (unpublished manuscript).

Soil organic matter study

Seventy-five monocultures were used to test the effects of individual grass species on soil organic matter: five replicates for each of the five species at three different soil organic matter levels. All monoculture plots were 0.75 x 0.75 m and were separated by 25 cm deep sheet metal. The continuous soil gradient created in the garden was divided into three discreet levels for this analysis: low organic matter soils contained mostly sand (mean total C = 0.24%, range 0.0-0.5%), medium soils contained mixed sand and black soil (mean total C = 0.76%, range 0.5-1.25%), and high organic matter soils were predominantly black soil (mean total C = 1.65%, range 1.25-2.1%). In May 1986, prior to seeding, four 20 cm deep x 2.5 cm diameter cores were taken from each plot, pooled, dried at 50 [degrees] C, and passed through a 2-mm mesh sieve. Soils from the plots were resampled after four growing seasons (November 1989) with the same methods except that 16 cm deep cores were taken. Because the year 1 (1986) samples were taken immediately after soils were mixed, the greater depth sampled did not bias the year 1 vs. year 4 comparisons. Subsamples of all soil samples were dried at 60 [degrees] C, ground with a coffee mill, and analyzed for total soil C and N with a Carlo-Erba NA 1500 CHN analyzer. N results from these analyses are presented in Wedin and Pastor (1993).

To estimate the percentage of net primary production (NPP, as C) that was found in the total soil C pool after 4 yr, we used data from the monocultures on above- and belowground production for Schizachyrium and Andropogon (Wedin 1990, Tilman and Wedin 1991). Wedin (1990) concluded that peak standing crop live biomass for these two species (data in Tilman and Wedin 1991) represented 75% of estimated aboveground primary production. Belowground production in these species was estimated to be [approximately equal to]65% of NPP (Wedin 1990). These production estimates were based on 3-yr-old monocultures and were multiplied by 4 to estimate total NPP after four growing seasons. C concentration of biomass was assumed to be 45.6% (see Table 1). NPP estimates at the three soil organic matter levels (as grams of carbon per square metre per year) are 101, 172, and 356 for Schizachyrium and 86, 186, and 412 for Andropogon, respectively. To estimate soil C on an areal basis, bulk densities were measured in a subset of plots and estimated for the remainder based on a regression of bulk density vs. total soil C ([R.sup.2] = 0.98; see Wedin and Tilman 1990). A soil depth of 16 cm was used in calculations of total soil C per square metre.

Stable carbon isotope analyses

Bulk tissue samples and lignin fractions for the eight tissue types (above- and belowground tissues for each of four species) and ground soil samples from year 1 and year 4 of the study were analyzed by combustion in a Carlo-Erba CHN analyzer that separated C[O.sub.2] and [N.sub.2] chromatographically prior to cryogenic purification in a triple trap on a VG Micromass SIRA 10 isotope ratio mass spectrometer. The [[Delta].sup.13]C value is calculated as

[[Delta].sup.13]C = (13C/12[C.sub.sample] - 13C/12[C.sub.reference]) / (13C/12[C.sub.reference]) x 1000. (1)

[TABULAR DATA FOR TABLE 1 OMITTED]

Values are referenced to the Peedee formation belemnite carbonate standard (PDB) of the National Bureau of Standards.

Knowing the initial soil [[Delta].sup.13]C ([[Delta].sub.i]) and the final soil [[Delta].sup.13]C ([[Delta].sub.f]) after a period of growth by vegetation with a different [[Delta].sup.13]C signature ([[Delta].sub.n]), a simple mixing model can be used to estimate the proportion (X) of C in the soil coming from the new vegetation:

[[Delta].sub.f] = (1 - X)[[Delta].sub.i] + X[[Delta].sub.n]. (2)

In the analysis of [[Delta].sup.13]C dynamics during litter decomposition, Eq. 2 was also used to estimate the proportion of new C in decomposed litter or the lignin fraction of decomposed litter coming from microbial populations with a different [[Delta].sup.13]C signature. In this case, [[Delta].sub.i] and [[Delta].sub.f] are the initial and final litter [[Delta].sup.13]C, respectively, [[Delta].sub.n], is the soil organic matter [[Delta].sup.13]C, and X is the proportion of total litter bag C originating in soil organic matter.

RESULTS AND DISCUSSION

Lignin dynamics during decomposition

The four grass species differed significantly in belowground (range 9.5-22.5%) and aboveground (range 10.7-17.4%) lignin concentrations of undecomposed tissues (Table 1). During decomposition, the lignin concentration of all tissue types increased on an ash-free dry mass basis, converging on a value of 20-25% lignin in the late stages of decomposition ([ILLUSTRATION FOR FIGURE 1A OMITTED]). Decomposition studies with diverse plant materials (e.g., conifer litter, deciduous leaves, graminoid leaves) have observed that 20-30% of the initial litter mass remains at the end of long-term litter bag studies (Melillo et al. 1982, Melillo et al. 1989). This remaining litter fraction is thought to be the precursor of soil humus. For our discussion, we will assume that 70% mass loss is the end point of litter decomposition. This point was reached in our study within 2 yr by tissues with high litter quality (e.g., Agrostis roots, lignin: N ratio = 5.7), but not by tissues of low litter quality (e.g., Schizachyrium roots, lignin : N ratio = 46.1; see Pastor et al. 1987, Wedin and Tilman 1990).

The C concentration of lignin fractions from undecomposed tissues also differed significantly among species and tissue types (Table 1). The polyphenolic structure of lignin results in higher C concentrations for lignin and lignin-like materials than for complex carbohydrates, which comprise the bulk of plant tissues (Van Soest 1982, Benner et al. 1987). Whereas the C concentration of bulk undecomposed tissues averaged 46%, aboveground lignin fractions averaged 83% C and belowground 76% C. Note that the standard errors for our lignin percent C estimates are relatively high. Variance in the estimates of ash content for bulk litter was magnified through the series of sequential extractions and digestions.

As the relative lignin content of tissues increased during decomposition, the C concentration of the bulk litter increased significantly, reaching 50% C at 70% mass loss (average across eight tissue types). The C concentration of the lignin fraction did not change significantly during decomposition, however. This difference in C concentrations between lignin fractions and bulk litter must be accounted for when considering the C isotope dynamics of decomposing litter (Benner et al. 1987). On an ash-free dry mass basis, the lignin concentration of undecomposed tissue was 15% (average across eight tissue types), but on a C basis the lignin concentration was 27%. At the point of 70% mass loss, the decomposing tissues contained 35-40% lignin on a C basis ([ILLUSTRATION FOR FIGURE 1B OMITTED]).

Isotopic signatures of plant tissues

Tissue [[Delta].sup.13]C of Schizachyrium, the single [C.sub.4] grass analyzed, averaged - 12.0% across above- and belowground bulk tissues, while [[Delta].sup.13]C of the three [C.sub.3] species averaged -26.4% (Table 1). For three out of four species, belowground bulk tissues were significantly enriched in 13C compared to aboveground tissues (average difference of +0.5% across all species). Within the [C.sub.3] species, there were significant differences in [[Delta].sup.13]C, with a range of 1.7% for aboveground tissues among the three species.

Lignin fractions were consistently depleted in 13C compared to bulk tissues, with an average difference of -3.6% between lignin and bulk samples (Table 1). This consistent depletion of 13C in lignin fractions compared to bulk plant tissues has been previously documented (Benner et al. 1987) and might account for differences in bulk tissue [[Delta].sup.13]C among plant species that differ in lignin concentration. This was not the case in our study, however. Differences among species in estimated [[Delta].sup.13]C for the nonlignin tissue fraction were as large as those for bulk tissues.

Aboveground lignin fractions, averaging across all species, were -1.1% more depleted than belowground lignin fractions. The lack of a significant species by tissue interaction indicates that this difference was consistent across all species (Table 1). This pattern has not been previously reported, and may arise from the different biochemical origins of what we are calling above- and belowground lignin. Because lignin research has been dominated by forage and wood scientists, less attention has been given to the structure, biosynthesis, and degradation pathways of recalcitrant C fractions for roots than for aboveground tissues. This is unfortunate, considering that C inputs to SOM in many ecosystems, particularly grasslands, are dominated by belowground primary production.

Carbon isotope dynamics during plant tissue and lignin decomposition

For four of the eight bulk tissue types, there was a significant shift in [[Delta].sup.13]C values during decomposition (data in Fig. 2, regression analyses in Table 2). At 70% mass loss, the estimated shift ranged from 0.4 to 1.5% and, most importantly, occurred in opposite directions for [C.sub.3] and [C.sub.4] species. Bulk [[Delta].sup.13]C values for Schizachyrium, the [C.sub.4] species, decreased 1.5 and 1.0% for above- and belowground tissues, respectively, while, averaging across the three [C.sub.3] species, [[Delta].sup.13]C values increased 0.6% for both above- and belowground tissues.

For five of the eight tissue types, there was also a significant shift in [[Delta].sup.13]C values for lignin fractions during decomposition ([ILLUSTRATION FOR FIGURE 2 OMITTED], Table 2). Lignin [[Delta].sup.13]C for the [C.sub.4] species decreased while values for the [C.sub.3] species increased. Lignin [[Delta].sup.13]C shifts were larger for [C.sub.3] species than for the [C.sub.4] species. Averaging across the [C.sub.3] species, [[Delta].sup.13]C values increased 1.5% for aboveground lignin and 1.1% for belowground lignin (estimates for 70% mass loss, Table 2). For the [C.sub.4] species, [[Delta].sup.13]C decreased 1.3% for above- and 0.6% for belowground lignin.

To test if these [[Delta].sup.13]C shifts could have been caused by contamination of the litter bags by soil or dust during the 2-yr study, we asked if increases in the relative ash content of the bags were accounted for by mass loss of the nonash fraction, or if the ash content per bag had increased in absolute terms. As expected, no increases in absolute ash content were observed for the root litter bags, which were sealed polyester cloth. There were small, but consistent, increases in the absolute ash content of aboveground litter bags, which [TABULAR DATA FOR TABLE 2 OMITTED] had mesh tops, indicating some soil contamination. Assuming that this soil had a C content comparable to soil from the field in which the study took place, we estimate that from 0.1 to 0.5% of the total C in the litter bags at 70% mass loss was brought in with contaminating soil. This small amount of C cannot account for the changes in litter [[Delta].sup.13]C observed during decomposition.

Fig. 3A shows the hypothesized dynamics of [[Delta].sup.13]C for both bulk and lignin fractions during decomposition assuming lignin is preserved intact (Benner et al. 1987). In this model, lignin [[Delta].sup.13]C values remain unchanged, while bulk litter [[Delta].sup.13]C values are predicted to decrease for both [C.sub.3] and [C.sub.4] species as nonlignin fractions less depleted in [[Delta].sup.13]C decay faster. The predicted shifts in bulk litter [[Delta].sup.13]C in Fig. 3A assume that the shifts occur linearly from time zero (fresh litter) through 70% mass loss. Estimates of bulk litter [[Delta].sup.13]C at 70% mass loss under the lignin preservation hypothesis were based on the lignin concentration of each tissue type at 70% mass loss (calculated with regressions for each tissue type fit to the data in Fig. 1) and the [[Delta].sup.13]C values of undecomposed lignin and nonlignin fractions (Table 1).

Our results do not support the lignin preservation hypothesis. For all [C.sub.3] species, [[Delta].sup.13]C values shifted in the opposite direction from that predicted, and lignin values showed relatively greater changes than the bulk tissue values ([ILLUSTRATION FOR FIGURE 3B OMITTED]). These results differ sharply from Benner et al.'s (1987) study of Spartina decomposition in salt marshes. Because most, if not all, lignin-digesting organisms are aerobic (Van Soest 1982), lignin degradation is minimal in anaerobic environments. Thus, disproportionate preservation of lignin could cause the observed depletion of bulk tissue [[Delta].sup.13]C observed by Benner et al. (1987). Under aerobic conditions, the increase in lignin concentration during litter decomposition may be largely caused by the addition of recalcitrant lignin-like compounds by microbial decomposers (Melillo et al. 1989, Berg et al. 1993).

Nor are our results reconcilable with the hypothesis that a small but consistent enrichment of 13C occurs during decomposition (Dzurec et al. 1985, Nadelhoffer and Fry 1988, Melillo et al. 1989). This hypothesis accounts for the [C.sub.3] but not the [C.sub.4] results. Enrichment may occur on much longer time scales once plant material has entered the highly resistant soil humus pool, especially at greater depths in the soil horizon (Balesdent et al. 1993). However, we do not see consistent evidence of it (i.e., across both [C.sub.3] and [C.sub.4] species) in the litter decomposition phase of the plant litter to soil organic matter continuum.

To explain these observed shifts in [[Delta].sup.13]C values during decomposition we suggest that C from both the original lignin and nonlignin fractions is being transformed by the microbial decomposers, and that, as this process occurs, C originating in soil microbial biomass is incorporated into the remaining, transformed litter. This C mixing hypothesis suggests that shifts in isotopic signature at advanced stages of litter decomposition are not due to differential loss of constituent C fractions or to fractionation during microbial transformation, but rather to mixing of external C with original litter C, presumably via fungal hyphae or microbial populations.

Given the initial and final [[Delta].sup.13]C values for litter and the [[Delta].sup.13]C value for the soil in which the litter bag study took place (-22.4%), we can estimate the magnitude of C mixing with the preexisting SOM pool during litter decomposition (see Eq. 2). Averaging across above- and belowground tissues, this model estimates that at 70% mass loss, 12% of the C in the bulk tissue and 17% of the C in the lignin fraction originated in the SOM pool for Schizachyrium, the [C.sub.4] species. For the [C.sub.3] species, the average estimate is that the bulk tissue contains 19% new C and the lignin fraction 17% new C. We stress that these estimates make numerous assumptions. We present them not as quantitative estimates of C mixing between SOM and plant litter during decomposition but as qualitative support consistent with the C mixing hypothesis. This model accounts for: (1) the opposite directions of [[Delta].sup.13]C shifts for [C.sub.3] and [C.sub.4] litter, (2) the relatively greater [[Delta].sup.13]C shift for lignin compared to bulk tissues for the [C.sub.3] species, and (3) the relatively smaller [[Delta].sup.13]C shift for lignin compared to bulk tissues for the [C.sub.4] species ([ILLUSTRATION FOR FIGURE 3B OMITTED]).
TABLE 3. Two-way ANOVAs testing the effect of grass species and
soil organic matter level on total soil carbon and soil
[[Delta].sup.13]C values after four growing seasons. See Fig. 2 for
means and standard errors for (A) and (D). One soil sample from year
1 and two from year 2 had insufficient C for [[Delta].sup.13]C
analyses and were treated as missing values.


       Effect                             F        df       P


A) Total soil carbon in year 4


   Soil level                           222.4     2, 60    0.0001
   Species                                0.68    4, 60    0.6090
   Soil x species                         0.36    8, 60    0.9440


B) Soil [[Delta].sup.13]C in year 1


   Soil level                             1.505   2, 59    0.2303
   Species                                0.688   4, 59    0.6034
   Soil x species                         0.676   8, 59    0.7107


C) Soil [[Delta].sup.13]C in year 4


   Soil level                             5.677   2, 59    0.0056
   Species                               46.19    4, 59    0.0001
   Soil x species                         1.025   8, 59    0.4279


D) Soil [[Delta].sup.13]C in year 4 - soil [[Delta].sup.13]C in
year 1


   Soil level                             0.819   2, 58    0.4459
   Species                               21.574   4, 58    0.0001
   Soil x species                         1.936   8, 58    0.0717


This C mixing model also suggests that mass loss during litter decomposition (i.e., loss of original litter C) is occurring faster than indicated by the litter bag studies. If, on average, 15% of the C in our litter bags at the point of 70% mass loss is new carbon originating in SOM, the actual degree of mass loss is closer to 75%.

Shifts in [[Delta].sup.13]C of SOM under monocultures of [C.sub.3] and [C.sub.4] species

Initial (May 1986) [[Delta].sup.13]C values for soil organic matter did not differ significantly among plots that were later planted with the five species or among the three soil organic matter levels (n = 75, mean [+ or -] 1 SE = -23.92 [+ or -] 0.106%, range = -21.93 to -26.92%, see Table 3B for ANOVA). After four growing seasons, there was no significant effect of plant species on total soil C (Table 3A, Fig. 4A, see also Wedin and Pastor 1993). However, there was a highly significant species effect on soil [[Delta].sup.13]C by year 4 (P [less than] 0.0001, Table 3C). Averaging across the three soil organic matter levels, [[Delta].sup.13]C values increased 2.16% for Schizachyrium and 1.61% for Andropogon, the two [C.sub.4] species, while for the three [C.sub.3] species [[Delta].sup.13]C values shifted -0.45% for Agropyron, -0.002% for Poa, and +0.18% for Agrostis ([ILLUSTRATION FOR FIGURE 4B OMITTED]). The small shift in soil [[Delta].sup.13]C under the [C.sub.3] species is consistent with the small difference in [[Delta].sup.13]C between the initial soils (mean = -23.9%) and their bulk tissues (mean = -26.4%).

The large increase in soil [[Delta].sup.13]C under the two [C.sub.4] species reflects the large differences between their bulk tissue [[Delta].sup.13]C values and the initial soil values. Assuming that the isotopic signatures of bulk plant tissues do not change during decomposition and that the signature of the preexisting SOM has not changed in the 4 yr, we can estimate the percent of total soil C that had been contributed by the [C.sub.4] grasses during the 4-yr study using Eq. 2. This estimate is a linear transformation of the difference between year 4 and year 1 [[Delta].sup.13]C values ([ILLUSTRATION FOR FIGURE 4B OMITTED]). Averaging across the three soil organic matter levels, 17.8% of the soil C in Schizachyrium plots and 12.7% of the C in Andropogon plots appears to be new C ([ILLUSTRATION FOR FIGURE 4B OMITTED]). The estimated percentage of new [C.sub.4] C in SOM was higher in low soil organic matter plots (22.6%) and decreased with increasing soil organic matter content for Schizachyrium. This pattern was less clear in Andropogon plots.

This displacement of existing SOM C with new C is relatively rapid compared to other published [[Delta].sup.13]C SOM studies. Balesdent et al. (1987) estimated that 22% of total soil C turned over after 13 yr when corn (Zea mays, [C.sub.4]) was planted in soil with a predominantly [C.sub.3] signature. In tropical savannahs, Martin et al. (1990) estimated that 52-70% of soil C turned over in 16 yr and was replaced by new [C.sub.3] C when woody vegetation invaded a [C.sub.4] grassland site. The high turnover rate for soil C in our [C.sub.4] grass monocultures is probably a function of both the low total C status of these sandy soils and the high belowground C allocation of these tall-grass prairie dominant species (Wedin and Tilman 1990).

The difference between the two [C.sub.4] species in estimated C contribution to SOM appears to reflect a real difference in rates of C turnover rather than an inter-specific differences in isotopic signatures. Bulk tissue [[Delta].sup.13]C values of [C.sub.4] species generally differ less those of [C.sub.3] species (Tieszen and Boutton 1989). The mean value for aboveground tissue [[Delta].sup.13]C of Andropogon in our plots was - 11.7%, slightly less than the value of - 12.0% used for [C.sub.4] C in the calculations of Fig. 4B. Recalculation of the percentage of new [C.sub.4] C in SOM in Andropogon plots using - 11.7% rather than - 12.0% gave a mean value for the soil gradient of 12.4% new C compared to the value of 12.7% presented above.

Although the [C.sub.3] species were more depleted in [[Delta].sup.13]C than the SOM, soil [[Delta].sup.13]C values increased an average of 0.45%o in [C.sub.3] monocultures on the richest soils ([ILLUSTRATION FOR FIGURE 4B OMITTED]). This suggests that [approximately equal to]3% of the total soil C in these plots is new C with a [C.sub.4] origin, which is not possible since only [C.sub.3] organic matter inputs occurred in these plots. Instead, this inconsistency indicates that the [[Delta].sup.13]C value of preexisting SOM did not remain constant over the 4-yr study. While there was a 2% gain in total soil C in low soil organic matter plots during the study, high soil organic matter plots lost, on average, 10% of their soil C. The medium soil organic matter plots lost, on average, 7% of total soil C. Increased SOM decomposition in these plots was probably caused by increased soil aeration and/or temperature, and disruption of the black soil's aggregate structure following repeated rototillering (Schimel 1986). We suggest that the anomalous positive [[Delta].sup.13]C shift in high soil organic matter [C.sub.3] plots was caused by differential decomposition and loss of C from various fractions of SOM (e.g., fulvic vs. humic acids). If labile and recalcitrant SOM fractions are not at isotopic equilibrium, a bias will occur in estimates of organic matter inputs to total soil C when SOM losses occur during the course of a study such as this (Martin et al. 1990, Bonde et al. 1992). This bias occurs in our estimates of [C.sub.4] C contributions to SOM in high soil organic matter plots (estimates of the percentage of new [C.sub.4] C are [approximately equal to]3% too high, Fig. 3B), but the estimates for the low soil organic matter plots appear to be unaffected.

Combining the estimates of percent new [C.sub.4] C in SOM with measured net primary production (NPP) of the two [C.sub.4] species allows estimation of the percentage of NPP found in SOM after 4 yr (see Methods: Soil organic matter study). On low soil organic matter plots, 33% of Schizachyrium NPP and 25% of Andropogon NPP became SOM after 4 yr. On high soil organic matter plots, the estimates, after correction for the positive fractionation bias discussed above, are 29% for Schizachyrium and 18% for Andropogon. These estimates are consistent with the conclusion from various litter bag studies that 20-30% of the original mass of decomposing litter remains as recalcitrant humus precursors after 2-3 yr of decomposition (Melillo et al. 1989). Because these [C.sub.4] grasses have large and heavily mycorrhizal root masses (Johnson et al. 1992), inputs of fixed C with characteristic [[Delta].sup.13]C signatures to SOM via exudates and rhizosphere symbionts may also be contributing to the observed shift in SOM signatures. The magnitude of these additional inputs is unknown.

CONCLUSION

We present a model of C mixing between litter and SOM during decomposition to account for the dynamics of [[Delta].sup.13]C in our litter bag study. No other hypothesis presented to date explains why [[Delta].sup.13]C values for bulk litter and, in particular, lignin fractions would shift in opposite directions for [C.sub.3] and [C.sub.4] species during decomposition. This model suggests that decomposers incorporate new C from SOM into both the lignin and nonlignin litter fractions and that this new C represents from 12 to 19% of the total litter C at 70% mass loss. If this model is correct, our results also indicate that the fidelity with which [[Delta].sup.13]C signatures are transferred from plant tissues to SOM is high under aerobic conditions.

Assuming that plant tissue [[Delta].sup.13]C signatures are faithfully transferred to SOM, we estimated that 22% of the soil C in Schizachyrium ([C.sub.4]) monocultures growing on sandy soils had turned over and been replaced by new C within 4 yr. To account for a turnover of this magnitude, 33% of the estimated NPP summed over 4 yr would need to remain in the SOM pool. Averaging across the two [C.sub.4] species and the three soil organic matter levels, 30% of the estimated NPP from the first 4 yr remained in SOM. In contrast to the apparently static behavior of bulk soil C in these plots ([ILLUSTRATION FOR FIGURE 3A OMITTED]), this isotopic analysis suggests that SOM is quite dynamic under [C.sub.4] prairie grasses, such as Schizachyrium and Andropogon. Testing the validity of this conclusion for other plants on other soil types (e.g., Martin et al. 1990) requires further research.

ACKNOWLEDGMENTS

This research was supported by the U.S. National Science Foundation (BSR-8811884), the Natural Sciences and Engineering Research Council of Canada, and the U.S. Department of Energy (NIGEC grant to L. L. Tieszen). We thank Mike Chapman, David Stemper, and Mark Larum for assisting in analyses.

LITERATURE CITED

Aber, J. D., and J. M. Melillo. 1991. Terrestrial ecosystems. Saunders College Publishing, Philadelphia, Pennsylvania, USA.

Balesdent, J., C. Girardin, and A. Mariotti. 1993. Site-related [[Delta].sup.13]C of tree leaves and soil organic matter in a temperate forest. Ecology 74:1713-1721.

Balesdent, J., A. Mariotti, and B. Guillet. 1987. Natural 13C abundance as a tracer for studies of soil organic matter dynamics. Soil Biology and Biochemistry 19:25-30.

Benner, R., M. L. Fogel, E. K. Sprague, and R. E. Hodson. 1987. Depletion of 13C in lignin and its implications for stable carbon isotope studies. Nature 329:708-710.

Berg, B., C. McClaugherty, and M.-B. Johansson. 1993. Litter mass-loss rates in late stages of decomposition at some climatically and nutritionally different pine sites. Long-term decomposition in a Scots pine forest. VIII. Canadian Journal of Botany 71:680-692.

Bonde, T. A., B. T. Christensen, and C. C. Cerri. 1992. Dynamics of soil organic matter as reflected by natural 13C abundance in particle size fractions of forested and cultivated oxisols. Soil Biology and Biochemistry 24:275-277.

Dzurec, R. S., T. W. Boutton, M. M. Caldwell, and B. N. Smith. 1985. Carbon isotope ratios of soil organic matter and their use in assessing community composition changes in Curlew Valley, Utah. Oecologia 66:17-24.

Effland, M. 1977. Modified procedure to determine acid-soluble lignin in wood and pulp. Tappi 60:143-144.

Farquhar, G. D., J. R. Ehleringer, and K. T. Hubick. 1989. Carbon isotope discrimination and photosynthesis. Annual Review of Plant Physiology 40:503-537.

Geng, X., J. Pastor, and B. Dewey. 1993. Decay and nitrogen dynamics of litter from disjunct, congeneric tree species in old-growth stands in northeastern China and Wisconsin. Canadian Journal of Botany 71:693-699.

Johnson, N. C., D. Tilman, and D. Wedin. 1992. Plant and soil controls on mycorrhizal fungal communities. Ecology 73:2034-2042.

Keeling, C. D., W. G. Mook, and P. P. Tans. 1979. Recent trends in the 13C/12C ratio of atmospheric carbon dioxide. Nature 277:121-123.

Kingston, J. D., B. D. Marino, and A. Hill. 1994. Isotopic evidence for neogene hominid paleoenvironments in the Kenya Rift Valley. Science 264:955-959.

Martin, A., A. Mariotti, J. Balesdent, P. Lavelle, and R. Vuattoux. 1990. Estimate of organic matter turnover rate in a savanna soil by 13C natural abundance measurements. Soil Biology and Biochemistry 22:517-523.

McClaugherty, C. A., J. Pastor, J. D. Aber, and J. M. Melillo. 1985. Forest litter decomposition in relation to soil nitrogen dynamics and litter quality. Ecology 66:266-275.

McPherson, G. R., T. W. Boutton, and A. J. Midwood. 1993. Stable carbon isotope analysis of soil organic matter illustrates vegetation change at the grassland/woodland boundary in southeastern Arizona, USA. Oecologia 93:95-101.

Melillo, J. M., J. D. Aber, A. E. Linkins, A. Ricca, B. Fry, and K. J. Nadelhoffer. 1989. Carbon and nitrogen dynamics along the decay continuum: plant litter to soil organic matter. Plant and Soil 115:189-198.

Melillo, J. M., J. D. Aber, and J. F. Muratore. 1982. Nitrogen and lignin control of hardwood leaf litter decomposition dynamics. Ecology 63:621-626.

Nadelhoffer, K. J., and B. Fry. 1988. Controls on nitrogen-15 and carbon-13 abundances in forest soil organic matter. Soil Science Society of America Journal 52:1653-1660.

Parton, W. J., D. S. Schimel, C. V. Cole, and D. S. Ojima. 1987. Analysis of factors controlling soil organic matter levels in Great Plains grasslands. Soil Science Society of America Journal 51:1173-1179.

Pastor, J., M. A. Stillwell, and D. Tilman. 1987. Little blue-stem litter dynamics in Minnesota old fields. Oecologia 72:327-330.

Schimel, D. S. 1986. Carbon and nitrogen turnover in adjacent grassland and cropland ecosystems. Biogeochemistry 2:345-357.

Tieszen, L. L., and S. Archer. 1990. Isotopic assessment of vegetation changes in grassland and woodland systems. Pages 293-321 in C. B. Osmond, L. F. Pitelka, and G. M. Hidy, editors. Plant biology of the basin and range. Ecological Studies 80. Springer-Verlag, Berlin, Germany.

Tieszen, L. L., and T. W. Boutton. 1989. Stable carbon isotopes in terrestrial ecological research. Pages 167-195 in P. W. Rundel, J. R. Ehleringer, and K. A. Nagy, editors. Stable isotopes in ecological research. Ecological Studies 68. Springer-Verlag, Berlin, Germany.

Tilman, D. 1988. Plant strategies and the dynamics and structure of plant communities. Princeton University Press, Princeton, New Jersey, USA.

Tilman, D., and D. Wedin. 1991. Plant traits and resource reduction for five grasses growing on a nitrogen gradient. Ecology 72:685-700.

Van Soest, P. J. 1982. Nutritional ecology of the ruminant. Cornell University Press, Ithaca, New York, USA.

Von Fischer, J. C., and L. L. Tieszen. 1995. Carbon isotope characterization of soil organic matter from four tropical forests in Luquillo, Puerto Rico. Biotropica, in press.

Wedin, D. A. 1990. Nitrogen cycling and competition among grass species. Dissertation. University of Minnesota, Minneapolis, Minnesota, USA.

Wedin, D. A., and J. Pastor. 1993. Nitrogen mineralization dynamics in grass monocultures. Oecologia 96:186-192.

Wedin, D. A., and D. Tilman. 1990. Species effects on nitrogen cycling: a test with perennial grasses. Oecologia 84:433-441.
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Author:Wedin, David A.; Tieszen, Larry L.; Dewey, Bradley; Pastor, John
Publication:Ecology
Date:Jul 1, 1995
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