13C and 15N stabilization dynamics in soil organic matter fractions during needle and fine root decomposition

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Organic Geochemistry Organic Geochemistry 39 (2008) 465–477 www.elsevier.com/locate/orggeochem

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C and 15N stabilization dynamics in soil organic matter fractions during needle and fine root decomposition Jeffrey A. Bird a,b,*, Markus Kleber b, Margaret S. Torn b a

Queens College, CUNY, School of Earth and Environment Sciences, E-220 New Science Building, 65-30 Kissena Boulevard, Flushing, NY 11367, United States b Earth Sciences Division, Lawrence Berkeley National Laboratory, One Cyclotron Rd., Berkeley, CA 94270, United States Received 24 July 2007; received in revised form 15 November 2007; accepted 9 December 2007 Available online 15 December 2007

Abstract Little is known about how the chemical composition of plant litter affects the amount and nature of C and N stabilized as soil organic matter (SOM). We examined the fate of dual labeled 13C and 15N Ponderosa pine fine roots (<2 mm) and needles decomposing for 2 yr in situ in a temperate conifer forest soil in the Sierra Nevada, CA, USA. We compared the distribution and stabilization of litter derived C and N in four SOM pools using a density fractionation procedure followed by an alkaline extraction of the dense fraction into fulvic, humic, and humin fractions. The C turnover times (estimated with natural abundance 14C) of these SOM fractions were distinct and ranged from 5 yr (light fraction) to 260 yr (insoluble humin). Input of C as roots resulted in 28% more total C retained in soil when compared to inputs as needles. Twice as much root 13C was present in the particulate soil (>2 mm) than for needles, while bulk soil (<2 mm) 13C and 15N recoveries were similar between litters. SOM fractions provided greater sensitivity than bulk soil and showed significant differences between litters in both the amount and chemical composition of 13C and 15N compounds recovered within SOM fractions. More needle 13C was retained in humic and humin fractions than was 13C from roots. The chemical composition of stabilized organic molecules differed fundamentally between needle and root sources within the dense fraction SOM pools, especially during the first year. Root inputs were stabilized predominately as N-rich biomolecules in the humic and humin fractions, while needles contributed C-rich biomolecules to these dense fraction SOM pools. The large pulse of C-rich compounds from the more labile needles recovered in the humic and humin fractions did not persist after 1.5 yr, suggesting that low C:N ratio compounds derived from decomposing litters may stabilize more strongly and persist within the dense fraction SOM pools. The fundamental differences in C and N pathways during decomposition and stabilization of below ground (root) and above ground (needle) litters suggest that shifts in plant C allocation may influence the long term stability of plant-derived C in soil. Ó 2007 Elsevier Ltd. All rights reserved.

Abbreviations: SOM, soil organic matter; MRT, mean residence time; HS, humic substances; FS, fulvic substances; LF, light fraction; DF, dense fraction; P-GC–IRMS, pyrolysis-gas chromatography–isotope ratio mass spectrometry. * Corresponding author. Tel.: +1 718 997 3332; fax: +1 718 997 3299. E-mail address: [email protected] (J.A. Bird).

1. Introduction Soil C dynamics are critical to global C budgets because soil organic C represents two-thirds of the terrestrial C pool and is the primary energy source driving several of the biogeochemical processes that

0146-6380/$ - see front matter Ó 2007 Elsevier Ltd. All rights reserved. doi:10.1016/j.orggeochem.2007.12.003

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determine net primary productivity (Schlesinger, 1997). The chemical composition of plant inputs is considered important for predicting litter decomposition rates in forest soils (Aber et al., 1990; Aerts, 1997; Preston et al., 2000; Silver and Miya, 2001). Considerably less is known about how litter quality affects the amount and chemical nature of C stored long term as soil organic matter (SOM) (Ko¨gelKnabner, 2002). Carbon compounds are protected from biodegradation by two general types of processes in unsaturated and unfrozen soils: (i) spatial isolation from decomposer communities and their enzymes (Ekschmitt et al., 2005), and (ii) the formation of intimate associations with mineral surfaces and metal ions (Sollins et al., 1996; Torn et al., 1997; Kleber et al., 2007). However, the mechanisms underlying the processes that protect SOM remain unclear and the capacity for storage of C in soil is difficult to predict (v. Lu¨tzow et al., 2006). A better understanding of how the chemical characteristics of plant inputs affect the potential of soils to sequester organic C is needed to improve our present ecosystem scale C models by linking shifts in plant species or plant allocation patterns to the dynamics of soil C pools. Much of what we know about the importance of litter quality for C cycling in forest soils has been gained from studies determining the rates of mass loss from recognizable plant litters (Berg, 2000), which provide only limited knowledge about the amount of C retained in the soil after decay, the chemical and physical state of the retained C, or its resistance to further mineralization. High proportions of aromatic C, lipids, waxes, and cutin, and low Ca and N in litters generally retard their decomposition (Silver and Miya, 2001; Ko¨gelKnabner, 2002). Most of these litter quality indicators, however, were derived from studies of above ground foliar litter, with few studies examining the composition and fate of below ground plant C inputs including fine roots (Rasse et al., 2005; Bird and Torn, 2006). A recent review of the mechanisms important to C stabilization in soils contends that litter quality only provides information about the early stages of decomposition, and that organo-mineral interactions and spatial occlusion are more important at the later stages of C transformation and stabilization (v. Lu¨tzow et al., 2006). Few studies, however, have tested this hypothesis directly. A long term study that utilized stable isotope tracers suggests that vastly different starting materials, with dramatically different initial C loss rates (e.g., wheat

straw and glucose), results in similar quantities of soil C remaining after 10 yr (Voroney et al., 1989). In temperate ecosystems, recent insights on the mechanisms driving C stabilization processes in soil have relied on fractionating soil into SOM pools that differ in compound composition and C mean residence time (MRT) (e.g., Rethemeyer et al., 2005; Sollins et al., 2006). In part, these operationally defined pools provide greater resolution to detect changes than can be observed in bulk soil (Bird et al., 2002; Rasmussen et al., 2005). The spike in atmospheric 14CO2 generated from nuclear weapons testing in the 1950s and 1960s, makes radiocarbon (14C) a robust tool for characterizing C turnover times of SOM fractions, particularly those that cycle on timescales of decades and longer (Trumbore, 1993; Trumbore and Zheng, 1996). Moreover, stable isotopic tracers (13C, 15N) in conjunction with a SOM fractionation approach can be used to directly measure the stabilization rates and pathways of specific substrates among defined SOM pools on shorter time scales (Bird et al., 2003; Loya et al., 2004; Swanston et al., 2005). The SOM fractionation scheme applied to soils in this study were chosen to account for the chemical transformations that occur when organic C and N are transferred from plant tissues into SOM. Surface soils typically contain considerable amounts of particulate vegetation fragments that retain visible plant structures and are better characterized as decomposing litter than as highly processed, amorphous SOM, which is why we isolated this ‘‘light” fraction from the bulk soil (<2 mm) and determined its specific properties independently. The degradation of organic materials in soils is an oxidative process that increases the number of ionizable oxygen containing functional groups associated with decomposing organic compounds. An alkaline extraction of soil dissociates protons attached to ionizable functional groups and thus renders the decomposing organic material charged and soluble in the polar solvent water (Hayes, 2006). Alkaline extraction can therefore be viewed as a means to specifically isolate organic compounds that are in the process of oxidative decomposition. It leaves behind a fraction (humin) that either (i) has no ionizable functional groups and is thus unable to dissolve in water or (ii) has its functional groups involved in bonds to mineral surfaces. Given that soil biota rely on aqueous biochemical systems for their metabolism, materials that are insoluble in water are likely difficult to utilize and thus turn over

J.A. Bird et al. / Organic Geochemistry 39 (2008) 465–477

slowly. Historically, the organic compounds soluble in alkali have been divided in those that re-precipitate upon acidification (humic substances, HS) and those that remain soluble even after the majority of their ionizable functional groups have been neutralized by acidification to a pH of 1.5 (fulvic substances, FS). There is no clear chemical boundary between these two, although the greater responsiveness to the polar solvent combined with an often smaller molecular size should theoretically provide for a faster turnover of carbon incorporated in FS as compared to HS (Stevenson, 1994). Our overall objective was to determine the effects of litter quality and plant allocation on the distribution and stabilization of litter derived C and N in four distinct SOM fractions. To achieve this, we observed the decomposition of needles and fine roots of Ponderosa pine (Pinus ponderosa) during the first 2 yr of decomposition in situ, using stable isotope tracers. The dual label approach (13C and 15 N labeled litters) was chosen to track separately the retention and transformation of C and N, and to describe the general molecular characteristics (i.e., C to N ratio, acid or base solubility) of the litter derived compounds stabilized in different SOM fractions. 2. Materials and methods 2.1. Field site and soil The study site is located in the Blodgett Experimental Forest, on the western slope of the Sierra Nevada in El Dorado County, CA, USA (120°380 3000 W; 38°5300 0000 N) at 1315 m a.s.l. The soil is classified as a sandy, mixed, mesic Ultic Haploxeralf and is derived from granite (Soil Survey Staff, 1999). The vegetation was a 90-yr = old conifer forest dominated by Ponderosa pine. A well developed organic O horizon was present at the site (8.2 cm depth) and contained 338 g C kg1 and 8.4 g N kg1. The climate is Mediterranean with warm, dry summers and cool, wet winters. Annual precipitation averaged 1774 mm (1962–2001) and was concentrated between November and April (88%), with much falling as snow. Daily mean soil moisture and soil temperature during 2001–2003 were reported for this site by Bird and Torn (2006). Background soil chemical and physical characteristics were analyzed on A horizon (0–10 cm depth) soil samples from control mesocosms excavated in November 2001. Particle size distribution

467

was 675 g sand kg1, 235 g silt kg1 and 90 g clay kg1. Cation exchange capacity was 9.0 cmol kg1 measured by barium acetate saturation and calcium replacement (Janitzski, 1986). Soil pH was 5.9 and electrical conductivity was 0.25 dS m1. Soil C content was 57 ± 6 g kg1 and N was 1.9 ± 0.1 g kg1. Bulk density averaged 0.82 ± 0.03 g cm3. 2.2.

13

C and

15

N plant litter

Two year old ponderosa pine saplings were grown and labeled with 13CO2 and 15 NO 3 under controlled greenhouse conditions to produce a season’s growth of uniformly labeled needles and fine roots (Bird and Torn, 2006). Needle clusters and fine roots (<2 mm) produced during the 13C labeling season were harvested by clipping. Litter material was dried at 25 °C and homogenized prior to addition to the field microcosms. Labeled roots were <2 mm in diameter and at least 7.6 cm long. The 13 C enrichment of needles was 3.8 atom% and that of roots was 3.3 atom%. The 15N enrichment of needles was 5.5 atom% and that of roots was 8.1 atom%. Pyrolysis-gas chromatography–isotope ratio mass spectrometry (P-GC-IRMS) fragments of litters had similar 13C and 15N enrichments, indicating uniform enrichment of C and N compounds (data not shown). The chemical composition differed between roots and needles according to a forest products determination of proximate C fractions (Bird and Torn, 2006) and P–GC–IRMS (data not shown). There was general agreement between the two methods. Proximate analysis showed a lower C:N ratio for needles (39) compared with roots (49). The acid resistant fraction:N ratio of needles (17) was also lower than for roots (29). Needles had 17% more water soluble C and a lower proportion of acid hydrolysable C and acid resistant C compounds than did fine roots (Bird and Torn, 2006). 2.3.

13

C and

15

N litter field study

The overall experiment was a 2  2 factorial design with four field replicates and was described in detail in Bird and Torn (2006). Factor 1 compared the C and N dynamics of added roots versus needles; factor 2 compared the effects of soil depth (O horizon versus A horizon). In this paper, we report the comparison between litter types added to the A (mineral) soil horizon (0–10 cm below O

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J.A. Bird et al. / Organic Geochemistry 39 (2008) 465–477

horizon). We do not focus on the soil depth factor here because soil depth did not affect the recovery of 13C or 15N during the first 1.5 yr in situ (Bird and Torn, 2006). Microcosms (10.2 cm diameter, 23 cm long) were inserted 21 cm into the soil in 2001 and allowed to equilibrate for 120 d prior to the application of 13 C and 15N labeled litter. Three microcosms per treatment per block were installed for harvest 152 d (5 months), 294 d (10 months) and 568 d (1.5 yr) after the application of the labeled litter substrates. Microcosms had two 5-cm diameter windows fit with 450 lm mesh to allow fungal hyphae and roots to penetrate the core. Microcosms were placed >1 m from large trees and >0.5 m from other microcosms. In November 2001, 13C and 15N labeled roots or needles were applied to the A horizon of microcosms at a depth of 2–4 cm below the O/A interface. The amount of labeled litter added to soil was 147 g m2 (dry matter) for needles and 135 g m2 for roots. Litter was placed in the microcosms and lightly mixed with soil in its placement depth. Additional microcosms served as controls and were either treated like those that received litter (i.e., minimally disturbed soil without litter addition) or left undisturbed. Soil microcosms were excavated intact from the field site on four dates: November 16, 2001, April 14, 2002, September 4, 2002 and June 6, 2003. Fungal hyphae and roots that had penetrated into the microcosms were severed at the microcosm exterior edge and retained in the soil sample. Soil was sieved using a 2 mm sieve to separate particulate soil material (>2 mm) and bulk soil (<2 mm). The resulting 2 size fractions were homogenized and analyzed separately. 2.4. Soil organic matter fractionation We fractionated soil samples into four operationally defined soil organic matter (SOM) fractions: (1) light fraction, LF; (2) humics substances, HS; (3) fulvic substances, FS; and (4) non-alkali extractable humics, humin. Soil samples from the A horizon (0– 10 cm depth) for all four microcosm excavation dates were fractionated by a sequential physical and chemical procedure described in Bird et al. (2002) and here briefly. The physical fractionation was performed on subsamples of air dried soil (25 °C). The LF was separated using NaI solution (specific gravity of 1.75 g cm3) and centrifugation. Light fraction was retained for analyses and rinsed

with deionized water on ashed GF/A filters. The remaining dense fraction (DF) was rinsed with 0.1 M HCl to remove salts and carbonates. Humic substances and FS were then extracted with eight aliquots of 0.4 M NaOH, under N2, and separated by centrifugation. All extractions were done using a 10:1 vol./soil dry wt ratio. Fulvic solution subsamples were dialyzed to remove Na and small molecular compounds in deionized water using 500 MW cellulose tubing (Spectrum Industries, Inc., Rancho Dominguez, CA). The remaining insoluble SOM fraction, humin, is bound in stable aggregates with metal polyvalent cations and clay minerals (Bird et al., 2002). The SOM fractions were analyzed after lyophilizing and homogenizing. Soil and SOM solutions were maintained at 4 °C during and after fractionation. 2.5. Isotope analyses Natural abundance 13C and 15N values of soil and SOM fractions were measured on samples taken in April 2002 and analyzed on a Europa Scientific Hydra 20/20 IRMS (PDZ Europa, Cheshire, UK). For enriched samples, isotopic enrichment of 13 C and 15N was determined on a Europa Scientific INTEGRA IRMS (PDZ Europa, Cheshire, UK). The relative stabilization efficiency of litter derived C to litter derived N added to microcosms was calculated for soil and SOM fractions and is referred to as the litter derived C:N recovery ratio. To measure 14C content, organic soil C was reduced to graphite in three steps: complete combustion to CO2 in the presence of CuO, cryogenic purification of CO2, and sealed tube zinc reduction to graphite (Vogel, 1992). Radiocarbon content was determined by accelerator mass spectrometry at Lawrence Livermore National Laboratory, with precision better than ±5.7‰ of D14C (Stuiver and Polach, 1977). For each SOM fraction, we estimated the mean C residence time with a one-pool stockflow model (Trumbore, 1993; Gaudinski et al., 2000; Torn et al., 2002) in which soil 14C in year i is calculated as soil 14C in the previous year plus the difference between the 14C added in plant inputs and that lost via decomposition and radioactive decay. At steady state carbon stock, the model can be simplified to the following equation:    1 Dt ð1Þ Ri ¼ Ri1 1  Dt k þ þ ðRatm;i-lag Þ s s

J.A. Bird et al. / Organic Geochemistry 39 (2008) 465–477

Where Ri = 14C content of SOM in year i (fraction modern); Dt = annual time step; k = 14C radioactive decay constant (0.000121 yr1); s = turnover time of SOM (yr). Ratm = 14C of northern hemisphere atmospheric CO2 in year i-lag (fraction modern). Lag = years between C being photosynthesized and entering the litter or soil (Torn et al., 2005); ponderosa pine needles typically live three years before entering the top of O horizon. We assumed lag = 5 yr for all SOM fractions in the A horizon. The spike of atmospheric 14C from nuclear weapons testing is a good tracer of decadal turnover times, but when this approach does not provide a unique estimate of turnover time, NPP-stock relationships may be used to select the best fit option (Torn et al., 2005). In this study, for the LF fraction, turnover times of 5 and 61 yr were consistent with the 14C data, but the 5 yr turnover time was a better fit with site NPP and stock data.

469

Table 1 Recovery of added 13C and 15N from bulk soil (<2 mm), particulate material (>2 mm) and total soil 568 d after application of the litters to the microcosms Litter type

Soil fraction Bulk (<2 mm)

Particulate (>2 mm)

Total soil

13

C recovered (% of applied after 568 d) Needles 18.1 (2.7) 23.9 (4.9) Fine roots 18.6 (2.5) 51.1 (2.4)

42.0 (2.2) 69.7 (1.7)

15 N recovered (% of applied after 568 d) Needles 40.0 (5.0) 41.2 (7.0) Fine roots 43.3 (2.7) 46.4 (2.0)

81.3 (2.9) 89.7 (3.3)

Means and standard errors shown, N = 4.

similar between litters on all sampling dates (Table 1). The recovery of 15N was unaffected by litter type in the total soil, particulate (>2 mm) and bulk soil (<2 mm) during 1.5 yr (Table 1). 3.2. Soil organic matter C and N fractions

2.6. Statistical analyses The main effect of litter type was tested using a general linear model (GLM). All data are expressed as least squares means with standard errors of indicated treatments. Fisher (F) statistics and P values are indicated in text and tables for GLM procedures. A significance level of P 6 0.05 was set a priori as the a-level, and P > 0.10 are noted as non-significant (NS). Post-hoc Tukey pairwise comparisons were performed for comparisons among C:N recovery ratios and the native C:N ratios of each SOM fraction. Studentized t-tests were performed among specific dates for each SOM fraction. Since proportional data are often not normally distributed and isotope recovery data often have non-homogeneous variance, recovery data were tested using Cochran’s test for homogeneity of variance, and analyses were performed after log transformation when needed. 3. Results 3.1.

13

C and

15

N dynamics in soil

Fine roots decomposed much more slowly than did needles and soils retained 28% more root 13C when compared to inputs than needles after 1.5 yr (Table 1). This was because root 13C retention in the particulate soil fraction (>2 mm) was nearly double that of needles after 1.5 yr. In contrast, 13C recovery in the bulk soil (<2 mm soil fraction) was

The four SOM fractions extracted from bulk soil (<2 mm) had distinct C and N properties (Table 2). The highest C and N contents were in the LF, with 48% of bulk soil C and 34% of bulk soil N. The C in the LF had the shortest MRT (5 yr). Within the DF, C pool sizes declined from the largest to smallest as follows: humin > HS > FS. The MRT among DF SOM fractions followed the reverse order of C pool size, increasing from FS to HS to humin C (Table 2). The C:N ratios of DF SOM fractions ranged from 16 to 26; each was significantly lower than that of the LF (34). 3.3.

13

C partitioning in SOM fractions

The largest pool of litter 13C isolated from the bulk soil (<2 mm) was the LF for all sampling dates, regardless of litter type. The percent recovery of applied 13C in the LF increased from 4.9% ± 0.9 of applied after 5 months to 14.4% ± 1.4 after 1.5 yr (averaged across litter type; Fig. 1). The proportion of bulk soil 13C recovered in the LF was not affected by litter type on any sampling date. Within the DF, 13C recovery was similar in magnitude in the HS and humin fractions; and both fractions had 2 times more 13C than did the FS fraction (Fig. 1). The 13C content of each of the three DF SOM pools was highest after 1.5 yr compared with 5 and 10 months after application (P 6 0.05). In contrast to the lack of a litter effect on the recovery of 13C in the bulk soil (<2 mm), 13C

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J.A. Bird et al. / Organic Geochemistry 39 (2008) 465–477

Table 2 Carbon and N characteristics of bulk soil (<2 mm) and organic matter fractions Soil fraction

Soil C

Soil N

13

C:N ratio

Proportion of soil (%)

Soil A < 2 mm Light fraction Fulvic Humic Humin

48.2 8.6 14.5 17.2

14

C

C

Mean residence time (yr)

Natural abundance

26.2 35.3 20.6 15.7 26.3

34.3 10.7 25.4 15.3

‰

D

25.36 25.92 23.90 25.54 25.01

101.6 133.9 122.3 58.7 12.2

85 5 68 140 260

Soil sampled in April 2002 from control microcosm, A horizon (0–10 cm depth). Means for C and N stable isotopes are natural abundance values, N = 4. Radiocarbon 14C values are from combined field replicates samples.

D EC FE B AP R JU N AU G O C T D EC JA N M AR JU N AU G O C T

D EC FE B AP R JU N AU G O C T D EC JA N M AR JU N AU G O C T

Month (2001 - 2003) 2.5

20.0

Fulvic C 2.0

15.0

1.5 10.0 1.0 5.0

0.5

0.0

0.0

Humic C

720

660

600

540

480

420

360

300

240

180

120

60

720

660

600

540

0.0

480

0.0

420

0.5

360

0.5

300

1.0

240

1.0

180

1.5

120

1.5

60

2.0

0

Humin C

2.0

0

13

C recovery (% of applied 13C)

Light fraction C

Days in situ Fig. 1. The recovery of 13C applied to soil microcosms from needles (triangles) or roots (square) among soil organic matter fractions. Errors shown are standard errors (N = 4).

recovery differed significantly between litter types in SOM fractions (Fig. 1; Table 3). After 5 months in situ, more needle 13C was recovered in the humin fraction than root 13C. The same trend between litter types was apparent in the FS and HS fractions at 5 months, however they were not significant at P 6 0.05 (Table 3). The HS fraction was the only SOM fraction affected by litter type after 10 months, and contained more 13C from needles than from roots.

3.4.

15

N partitioning in SOM fractions

The distribution of litter derived 15N recovered in SOM fractions was similar to that of 13C, with LF representing the largest pool of 15N during the initial 1.5 yr. After 1.5 yr in situ, an average of 58.1% ± 3.8 of the 15N recovered in the bulk soil was isolated as LF; and there was no difference between litter types (Fig. 2; Table 4). The percent of applied 15N recovered in the LF increased from

J.A. Bird et al. / Organic Geochemistry 39 (2008) 465–477

quantities of 15N recovered in FS and humin fractions. Like 13C, the largest increase in the amount of 15N isolated in the DF SOM pools occurred after the initiation of the second wet season, with the recovery of 15N at 1.5 yr exceeding that recovered at 5 or 10 months (Fig. 2). The amount of litter derived 15N partitioned into the different SOM fractions was generally similar for the two litter substrates (Table 4). The largest differences occurred after 10 months, when more root derived 15N was recovered in the humin pool than from needles.

Table 3 Summary of statistical parameters of the effect of litter type (needles versus fine roots) on the recovery of 13C among SOM fractions during the first 1.5 yr in situ Sample date (days after application)

13

C/15N

Soil fraction

Statistic

152

294

568

Light fraction

F P

0.1 NS

1.0 NS

1.5 NS

Fulvic

F P

8.1 0.066

0.8 NS

2.4 NS

Humic

F P

9.9 0.051

54 0.005

9.4 0.055

Humin

F P

10.1 0.050

0.2 NS

0.3 NS

471

3.5. Litter derived C:N ratios in SOM fractions The chemical composition of the biomolecules stabilized in SOM fractions was described by their 13 15 C: N stoichiometry, which we describe as the litter derived C:N recovery ratio. As a reference, we compared the overall native C:N ratio for SOM fractions to the litter derived C:N recovery ratios from each litter type (Fig. 3; Table 5). There were clear differences in the ratio of litter derived C to

Means and standard deviations shown in Fig. 1, N = 4.

9.1% ± 1.9 at 5 months to 23.6% ± 2.1 at 1.5 yr (averaged across litter type; Fig. 2). Within the DF, the HS fraction had the largest amount of 15 N present during the initial 1.5 yr, with smaller

30

10

N) N recovery (% of applied

15

Fulvic N

Light fraction N

25 20

8 6

15

4 10 2

5 0

0

720

660

600

540

480

420

360

300

240

180

120

0

720

660

600

0

540

0

480

2

420

2

360

4

300

4

240

6

180

6

120

8

60

8

60

Humin N

Humic N

0

15

D EC FE B AP R JU N AU G O C T D EC JA N M AR JU N AU G O C T

D EC FE B AP R JU N AU G O C T D EC JA N M AR JU N AU G O C T

Month (2001 - 2003)

Days in situ Fig. 2. The recovery of 15N applied to soil microcosms from needles (triangles) or roots (square) among soil organic matter fractions. Errors shown are standard errors (N = 4).

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J.A. Bird et al. / Organic Geochemistry 39 (2008) 465–477

son in the second year of the study, root litter derived C:N recovery ratios in the LF increased to 35, similar to that of the native LF fraction. In contrast, the needle litter derived C:N recovery ratio in the LF was similar to that of the native LF C:N ratio after 5 months, but declined to 20 after 10 months (Fig. 3). Litter derived C:N recovery ratios were consistently lower for roots than for needles in each of the DF pools (Fig. 3), meaning that relatively more N than C from roots was stabilized in the dense SOM fractions. In the humin fraction, the needle derived C:N recovery ratios was 50:1 after 5 months, and declined to 28 at 1.5 yr. In contrast, the recovery ratios for roots were very low initially and increased to converge with the value for needles and that of the native humin fraction after 1.5 yr (Fig. 3). In the FS and HS fractions, both litter substrates had similar recovery ratios after 1.5 yr. In both the FS and HS fractions, the litter derived C:N recovery ratios were 7–15 units lower than those of the native SOM fraction.

Table 4 Summary of statistical parameters of the effect of litter type (needles versus fine roots) on the recovery of 15N among SOM fractions during the first 1.5 yr in situ Sample date (days after application)

13

C/15N

Soil fraction

Statistic

152

294

568

Light fraction

F P

8.4 0.063

0.1 NS

0.1 NS

Fulvic

F P

3.9 NS

8.2 0.064

0.5 NS

Humic

F P

1.0 NS

4.3 NS

0.5 NS

Humin

F P

3.6 NS

32.7 0.011

0.5 NS

Means and standard deviations shown in Fig. 2, N = 4.

N isolated in each SOM fraction. During the first year, the root litter derived C:N recovery ratio in the LF was only 20 as compared with the native LF C:N ratio (36). After the onset of the rainy sea-

Light fraction C:N

40

30

30

20

20

10

10

0

Fulvic C:N

0

Humin C:N

Humic C:N

720

660

600

540

480

420

360

300

240

0

720

660

600

0

540

0

480

10 420

10 360

20

300

20

240

30

180

30

120

40

60

40

180

50

120

50

D EC FE B AP R JU N AU G O C T D EC JA N M AR JU N AU G O C T

50

40

0

Litter derived C:N recovery ratio

50

60

60

60

D EC FE B AP R JU N AU G O C T D EC JA N M AR JU N AU G O C T

Month (2001 - 2003)

Days in situ Fig. 3. The litter derived C:N recovery ratios from needles (triangles) or roots (square) and native C:N ratios of soil organic matter fractions (grey diamonds). Errors shown are standard errors (N = 4).

J.A. Bird et al. / Organic Geochemistry 39 (2008) 465–477 Table 5 Summary of statistical parameters of the effect of litter type on the litter derived C:N recovery ratios from needles versus fine roots during the first 1.5 yr in situ Sample date (days after application)

13

C/15N

Soil fraction

Statistic

152

294

568

Light fraction

F P

8.2 0.064

0.9 NS

33.0 0.01

Fulvic

F P

12.0 0.040

1.2 NS

0.5 NS

Humic

F P

69.5 0.004

38.0 0.009

15.4 0.029

Humin

F P

57.8 0.005

5.1 NS

0.3 NS

Means and standard deviations shown in Fig. 3, N = 4.

4. Discussion The SOM fractions had distinct stabilities as estimated by natural abundance 14C: a labile LF pool, a more stable DF containing two dynamic pools (HS and FS) and the most recalcitrant pool, humin. The C turnover times of the four SOM fractions ranged from 5 to 260 yr, which were similar to the 14C derived MRT for comparable SOM fractions from an agricultural surface soil (Wang and Chang, 2001). Using a similar SOM extraction method, Rethemeyer et al. (2005) reported slightly higher but similar 14C values for humic and humin fractions from agricultural and grassland surface soils to those measured in this forest surface soil. The SOM fractions differed in C and N contents and C:N ratios. As has been reported for other temperate forest soils (Spycher et al., 1983; Gaudinski et al., 2000; Gru¨newald et al., 2006), the LF in this study contained most of the soil C and N, and had the highest C:N ratio, the lowest mineral content, and the fastest C turnover (shortest MRT) of all SOM fractions. The LF is a mixture of partially decomposed plant and microbial residues, char, and other low density materials (Sollins et al., 1996). The LF isolated in our study may be described as the ‘‘free LF” because it is thought to reside in inter-aggregate space or in weak aggregates dispersed by the dense Na solution with agitation (Golchin et al., 1994; Sollins et al., 2006). Due to its fast turnover and close connection to litter inputs, the LF is the most responsive SOM pool to management (Golchin et al., 1995; Bird et al., 2003).

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The majority of litter 13C was recovered in the labile LF pool, even after more than half of the needle C had been lost from soil at 1.5 yr. This finding suggests that much of the retained litter C in the soil was not physically protected within aggregates and may not persist for longer than the MRT of 5 yr would predict. The LF can be conceptualized as two pools, that existing outside aggregates or associated with weak aggregates (free LF) and that found within stable aggregates (occluded LF). The occluded LF has high C:N ratios like litter and free LF, but much older 14C values than free LF (Golchin et al., 1994; Swanston et al., 2005), indicating that it is stabilized within aggregates. In this study, we did not isolate separate ‘‘free” and ‘‘occluded” light fraction material. However, the two LF pools were separated in a companion study on soil sampled within 100 m of our site (Rasmussen et al., 2005). In that study, 38% of total C was in the free LF and 27% in the occluded LF. Our LF content of 48% of total soil C was smaller than the total LF (occluded and free LF) isolated by Rasmussen et al. (2005) but higher than they obtained for free LF alone. This is likely due to two reasons. First, since we used a higher specific gravity and sampled a shallower soil depth, our samples yielded a larger free LF than reported by Rasmussen et al. (2005). Second, because we did not physically disrupt the aggregates by sonication, it is likely that much of the occluded LF was recovered in the humin fraction. Evidence of occluded LF in our humin was observed by the slightly higher than expected C:N ratio of the humin. The presence of occluded LF in the humin fraction may have given it a longer residence time than would have been observed without an occluded LF contribution. In these soils the occluded LF has a longer MRT than does the mineral fraction (Rasmussen et al., 2005). The DF, with its high mineral content, contained SOM bound by interactions at mineral surfaces (Golchin et al., 1994; Sollins et al., 1996). The FS and HS pools extracted from the DF using alkali had low C:N ratios, and are considered to be relatively polar in nature (Stevenson, 1994; Sutton and Sposito, 2005). The insoluble humin is thought to be stabilized most strongly by mineral associations (Stevenson, 1994) and contained a lower C:N ratio than the LF in this study. The scope of our study considered only the litter quality effect of root versus shoot tissue and did not include the potential fate of root exudates. Stabilization of root exudates may contribute significantly to

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high retention of root-derived C in soil. In addition, roots added to soil in this study did not naturally penetrate aggregates and were not associated with soil minerals prior to decomposition. Rasse et al. (2005) concluded that about one fourth of the longer MRT of root C over shoot C was due to the higher chemical recalcitrance of roots compared to that of shoots, with physio-chemical protection playing a large role. Therefore, our results may underestimate the greater overall contribution of roots to SOM storage compared with shoots because root exudates and natural aggregate associations were not considered in our study. The largest influence of litter quality on C retention in the soil was a much greater amount of roots than needles present in the particulate soil fraction (A horizon > 2 mm). In contrast, there were no differences in the recovery of 13C or 15N as LF in the bulk soil (<2 mm) between litters. This result was unexpected because the main source of LF is considered the particulate soil fraction (>2 mm), thus we expected to see more root derived C and N in the LF compared with needles. Most of the root biomass remaining in the particulate fraction appeared relatively unaltered (visible inspection), and the C:N ratios for roots were consistently similar to the original tissue in the particulate soil (>2 mm). In contrast, the C:N ratio of needles in the particulate soil fraction (>2 mm) dropped by half from the original value after 1 yr, indicating there had been significant degradation (Bird and Torn, 2006). Movement out of the particulate soil fraction (>2 mm) into the LF appeared to be significantly retarded for roots compared with needles. Once roots had been physically degraded enough to enter the LF (i.e., become <2 mm), the C and N dynamics in the LF were similar for both litters. During the initial stages of decomposition (0– 5 months), soluble and easily decomposable compounds were liberated from the litters and produced the highest 13CO2 respiration rates observed during the 2 yr study (Bird and Torn, 2006). The 13CO2 fluxes were significantly larger from needles than roots only during the initial 5 months, likely due to the greater amounts of soluble C present in needles than roots (Bird and Torn, 2006). Litter quality difference between roots and shoots influenced both the amount and the chemical nature of the compounds in the more stable SOM within the DF. These differences were already present after only 5 months in situ. During this period, significant amounts of litter derived compounds were stabilized

in the FS, HS and humin pools. More needlederived 13C was recovered in the DF than from root-derived 13C. This trend is opposite to the results for total 13C recovery. It appears that a greater proportion of 13C is incorporated into recalcitrant SOM fractions from a faster decomposing litter, like needles, because more needle biomass was processed by microorganisms than for roots (Bird and Torn, 2006). However, these needlederived compounds were initially very C-rich (50:1 in humin), compared with more N-rich, litter derived compounds from roots. Wide differences in the types of compounds stabilized did not persist over time, however, suggesting that the C-rich compounds were mainly contributed to these fractions during the first 5 months. Alternatively, they may have been weakly stabilized and subsequently remineralized. We posit that the source of the N-poor material (i.e., high C:N) from needles in humic and humin fractions was composed of polar and mildly non-polar C compounds (e.g., amphiphiles and carbohydrates) leached from needles, which had 17% more water soluble C than did roots. These types of compounds have been shown to associate and stabilize on mineral surfaces; and may make up a significant portion of humic and humin C (Sutton and Sposito, 2005). Alternatively, it is possible that the N-poor needle material recovered in the humin could have been partially decomposed needle material that had become occluded in aggregates with high density and/or were not acid or base soluble. However, the high C:N recovery ratios did not persist, suggesting that the litter derived compounds recovered as humin were more likely from DOC and were fairly labile. While we cannot differentiate which stabilized compounds were of microbial versus plant origin, litter derived C:N ratios >25 from needles supports a significant direct contribution from the needle tissues. In contrast, the N-rich, root derived material present in the HS and humin fractions was more likely of microbial origin. Nitrogen enrichment observed in DF and humin has been hypothesized to arise in part because of the low C:N ratios (5– 15) of the N-containing secondary microbial products formed from the original plant tissues (Baisden et al., 2002). The significant and rapid accumulation of 13C and 15N in the most stable fraction, humin, suggests that humin contains a sub-pool that cycles more rapidly then the MRT would suggest. Dense fractions with long turnover times have been reported

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to contain a portion of rapidly cycling C (Golchin et al., 1994; Trumbore and Zheng, 1996; Swanston et al., 2005). In an upland oak forest, the dense fraction (DF) remaining after removal of free and occluded LF contained a sizable amount of actively cycling C, although the majority of this DF was very stable (Swanston et al., 2005). The active component of the DF has been shown by the production of similar CO2 fluxes from LF and from forest soil DF incubated in the laboratory (Swanston et al., 2002) and during rice straw decomposition (Bird et al., 2003). The classical conceptual model of C and N stabilization processes at mineral surfaces has been revised recently to include a fundamentally different view of the structure of humic substances in soils (Sutton and Sposito, 2005). The view held for most of the last century was that humics and humin are composed of high molecular weight macromolecules, with covalent linkages, form 3-dimensional conformations and maintain a slow and limited exchange of compounds (Sutton and Sposito, 2005). Unlike the classical model, the ‘‘multilayer” model suggests that amphiphilic organic fragments attach to mineral surfaces in a discrete zonal sequence (Kleber et al., 2007). A major emphasis of this new model is the perception that organic fragments in direct contact with the mineral surface (‘‘contact zone”) are likely to achieve much longer residence times than those that are part of an outer ‘‘kinetic” zone that are held together by cation bridges and hydrogen bonding. The patterns of litter 13C and 15N stabilization observed in this study are more consistent with the dynamic ‘‘multilayer” theory than with the idea of adsorbed polymeric macromolecules. Moreover, the ‘‘multilayer” model helps to explain why the large pulse of C rich litter derived material from needles into the DF pools did not persist during year 2. The dynamic nature of the 13C and 15N movement into and through SOM pools with slower residence times is difficult to reconcile with the incorporation of litter derived materials into rigid and high molecular weight humic molecules. Rather, the rapid decline in the litter derived C:N ratios suggests the existence of weaker associations in outer layers of organic coatings on mineral surfaces. The relatively narrow litter derived C:N ratios observed in all DF pools after 1.5 yr supports the view that N-rich compounds may potentially become more strongly associated with mineral surfaces than are C-rich compounds. Sollins et al.

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(2006), using a sequential density fractions with C and N isotopes, observed a decrease in C:N ratios that were associated with an increase in MRT as mineral density increased. The likely source of these N-rich compounds is secondary metabolites produced by microorganisms (Gleixner et al., 2002; Ko¨gel-Knabner, 2002). Consequently, the long-term effects of litter quality on SOM may be more associated with the amount and macromolecular structure of the N-rich compounds formed from litter (via microbial activity) than from the initial plant tissue chemistry. We suggest that future studies use molecular approaches, such as P-GC-IRMS, to determine the main source of stabilized, N-rich compounds in SOM fractions with slow MRT. 5. Summary A novel use of dual isotopic tracers allowed us to quantify the decomposition and stabilization of root and needle C and N in different SOM fractions in a forest soil. We found that litter quality differences between needles and roots were apparent for both the initial decomposition of intact plant biomass and the amount and types of chemical compounds stabilized in stable SOM fractions. As expected, the higher litter quality of needles increased the rate of decomposition and overall C loss compared with roots, which had less labile constituents. A significant rate limiting step in root decomposition was the physical breakdown step from intact plant residues into the bulk soil (<2 mm). Root C tended to remain in the particulate soil fraction (>2 mm) much longer than did needle C. The root material that was broken down to <2 mm then behaved much like the needles in terms of total retention as SOM. More significantly, litter type influenced the main forms of C and N retained in soil. More needle C was respired, but also more needle C was transformed to FS, HS, and humin fractions. The needle material rapidly incorporated into DF pools had a wide C:N ratio, which may be indicative of plant tissue; whereas the narrow C:N ratio of root material in DF pools suggests that the majority of it was from microbial degradation products. Our 13 C results indicate that some components of humin cycles more rapidly than the MRT of the whole fraction would suggest, supporting the idea that humin and most dense fractions are composed of at least two functional SOM pools. The time course litter 13C and 15N incorporation into SOM

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fractions supported the new dynamic ‘‘multilayer” model of SOM stabilization. Carbon-rich compounds from needles built up quickly in the most stable SOM fractions but did not persist. In contrast, N-rich compounds from roots built up slowly but to a higher level during the first 1.5 yr in situ. This is inconsistent with the conventional view of individual, polymeric macromolecules adsorbed to mineral surfaces. It is consistent, however, with a multilayered architecture where N-containing amphiphilics are a significant part of the stable inner layers surrounded by C-rich compounds that cycle quickly. Acknowledgements This research was supported by the Climate Change Research Division, Office of Science, Biological and Environmental Research, US Department of Energy, under Contract No. DE-AC0376SF00098. This work was assisted by the cooperative efforts of the University of California, Berkeley, Center for Forestry, Blodgett Forest Research Station. We recognize the generous contributions to this work from research assistants D. Williard, A. Rowan, J. Westbrook, L. Andrews, J. Harrison, and collaborators W. Horwath, J. Gaudinski, T. Dawson and S. Mambelli. Special thanks to D. Maddocks for her support. Stable isotope analyses by D. Harris and M. Madhav. Associate Editor—I. Kogel-Knabner References Aber, J.D., Melillo, J.M., McClaugherty, C.A., 1990. Predicting long-term patterns of mass loss, nitrogen dynamics and soil organic matter formation from initial litter chemistry in temperate forest ecosystems. Canadian Journal of Botany 68, 2201–2208. Aerts, R., 1997. Climate, leaf litter chemistry and leaf litter decomposition in terrestrial ecosystems: a triangular relationship. Oikos 79, 439–449. Baisden, W.T., Amundson, R., Cook, A.C., Brenner, D.L., 2002. Turnover and storage of C and N in five density fractions from California annual grassland surface soils. Global Biogeochemical Cycles 16, 1–16. Berg, B., 2000. Litter decomposition and organic matter turnover in northern forest soils. Forest Ecology and Management 133, 13–22. Bird, J.A., Torn, M.S., 2006. Fine roots versus needles: a comparison of 13C and 15N dynamics in a ponderosa pine forest soil. Biogeochemistry 79, 361–382. Bird, J.A., van Kessel, C., Horwath, W.R., 2002. Nitrogen dynamics in humic fractions under alternative straw manage-

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