Soil organic carbon turnover following forest restoration in south China: Evidence from stable carbon isotopes

Forest Ecology and Management 462 (2020) 117988

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Soil organic carbon turnover following forest restoration in south China: Evidence from stable carbon isotopes

T

⁎

Xin Xionga,b,d, Huiling Zhanga,b,d, Qi Denga,b,c, , Dafeng Huie, Guowei Chua,b, Ze Menga,b, ⁎ Guoyi Zhouf, Deqiang Zhanga,b, a

Key Laboratory of Vegetation Restoration and Management of Degraded Ecosystem, South China Botanical Garden, Chinese Academy of Sciences, Guangzhou 510650, China b Center for Plant Ecology, Core Botanical Gardens, Chinese Academy of Sciences, Guangzhou 510650, China c Southern Marine Science and Engineering Guangdong Laboratory (Guangzhou), Guangzhou 510301, China d College of Resources and Environment, University of Chinese Academy of Sciences, Beijing 10049, China e Department of Biological Sciences, Tennessee State University, Nashville, TN 37209, USA f School of Applied Meteorology, Nanjing University of Information Science & Technology, Nanjing 210044, China

A R T I C LE I N FO

A B S T R A C T

Keywords: Reforestation Soil carbon stock Soil carbon turnover Stable carbon isotopes

As over half of the world’s tropical forests are reforested or afforested, understanding the resilience of carbon (C) pool in these forests is critical for global C balance. While most previous studies on the reforested lands have focused on C stock recovery, soil C turnover has largely been overlooked. We evaluated soil C turnover rate by calculating the isotopic enrichment factors of α (defined as the slope of the regression between the δ13C difference and ln-transferred C concentrations in mineral soil samples relative to the surface litter) and β (defined as the slope of the regression between δ13C and log-transferred C concentrations) along 0–30 cm soil profiles in a 400-year-old monsoon evergreen broad-leaved forest (MEBF), a 51-year-old mixed-native plantation (NP1), a 31-year-old mixed-native plantation (NP2), a 31-year-old Acacia mangium plantation (AP), a 31-year-old mixedconifer plantation (CP), and a 31-year-old secondary forest with natural restoration (SF). Results showed that soil C stocks did not differ among the six forests. The estimated α values ranged from 1.0023 to 1.0086 and increased in the order of MEBF = NP1 < NP2 = AP = CP < SF. The estimated β values ranged from −19.70 to −5.22 but showed an opposite trend to α values. Additionally, changes of the α and β values among these forests were mainly regulated by soil water content and bulk density. Our findings demonstrate that forest restoration could enhance soil C stock equivalent to the undisturbed old-growth forests within a few decades, but the rate of soil C turnover in these restored forests were still higher.

1. Introduction Tropical forests play a critical role in global carbon (C) balance, as they contain approximately 20% of the global soil C stock and account for roughly 35% of terrestrial net primary productivity (Jobbágy and Jackson, 2000). However, these forests are undergoing widespread loss as a result of agricultural expansion and soil erosion, resulting in lower productivity, decreases in biodiversity, and increased C release from soil (Benchimol et al., 2017; Seymour and Harris, 2019). To combat the ongoing losses, many projects have been implemented in different countries to increase the tropical forest area and the biomass of vegetation (Don et al., 2011; Deng et al., 2014; Chen et al., 2019). These projects were also expected to enable these forests to improve soil

properties, and for soil C pool to recover to an equivalent level of the undisturbed old-growth forests, which includes how much of C can be stored in the soils (defined as soil C stock) and how long the stored C can be sequestered (defined as soil C turnover rate). Studies regarding the influence of afforestation on soil C stocks have been conducted extensively in recent years (Paul et al., 2002; Don et al., 2011; Veloso et al., 2018). However, how forest restoration affects soil C turnover is still not well understood, and this may have important implications for long-term soil C sequestration in these restored forests (Richter et al., 1999). Evaluating the rate of soil C turnover is challenging. The traditional method is to collect field soils for laboratory incubation, but it is difficult to accurately simulate field conditions (Feng et al., 2016). The 13C

⁎ Corresponding authors at: Key Laboratory of Vegetation Restoration and Management of Degraded Ecosystem, South China Botanical Garden, Chinese Academy of Sciences, Guangzhou 510650, China. E-mail addresses: [email protected] (Q. Deng), [email protected] (D. Zhang).

https://doi.org/10.1016/j.foreco.2020.117988 Received 18 November 2019; Received in revised form 13 January 2020; Accepted 12 February 2020 0378-1127/ © 2020 Elsevier B.V. All rights reserved.

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2.2. Soil and litter sampling

natural tracer is an elegant approach in situ, which potentially provides a powerful tool for understanding the temporal dynamics of soil C (Bernoux et al., 1998; Balesdent et al., 2018). Broadly, the approach involves measuring the enrichments of 13C natural abundance (δ13C) with increasing depths in soils, concomitant with decreasing C concentrations (Balesdent et al., 1993; Wang et al., 2018). There are several hypotheses that have been proposed to explain the enrichment of δ13C with increasing depth, which can be grouped into two categories: (1) mixing of C from different sources with different isotopic compositions; and (2) kinetic fractionation of C isotopes during soil C decomposition (Wynn et al., 2006; Acton et al., 2013). Soil C dynamics are controlled by the complex interplay of climatic, edaphic, and biotic factors, which also influence the depth trends of soil δ13C (Nadelhoffer and Fry, 1988). As each δ13C depth profile holds the unique story of its development, examining its trends through time could reveal some important changes not evident by elemental concentration analysis alone (Diochon and Kellman, 2008). The vertical enrichment of soil δ13C has been proven to be related to soil C turnover rate, with linear regression slope between the logarithm of the C concentration and δ13C value in soil profile serving to approximate the C decomposition rate (Garten, 2006; Acton et al., 2013; Gautam et al., 2017). This method has been widely used to evaluate global soil C turnover (Wang et al., 2018), but very few studies have systematically examined C stable isotope dynamics in the soils following forest restoration in pure C3 ecosystems to understand processes driving soil C cycling. In this study, we examined the depth trends in soil C concentrations and δ13C values in an undisturbed old-growth forest (> 400 years) and five restored forests representing different restoration ages, different restoration strategies, and different tree species planted in south China. The main objectives of this study were to: (1) establish the validity of using δ13C and C concentration depth profiles to infer temporal changes in below-ground C cycling processes following forest restoration in pure C3 ecosystems; (2) confirm whether these restored forests can provide equivalent C stock and turnover rate in the soil to the undisturbed oldgrowth forests; and (3) investigate the effects of different reforestation ages and strategies on soil C storage and turnover.

In October 2015, at each site except the MEBF, six replicated plots (10 × 10 m) were sampled, and for the MEBF, eight replicated plots (10 × 10 m) were sampled. In each plot, ten soil cores were randomly collected at three depths (0–10 cm, 10–20 cm, and 20–30 cm) and then pooled by depth. Each soil sample was sieved through a 2-mm mesh while plant materials were removed, and then divided into two parts: one part was stored at 4 °C and used for determination of soil moisture and soil microbial biomass carbon (MBC); and the other part was airdried and used for chemical analyses. Simultaneously, litter samples were randomly collected in each plot. The collected litters were cleaned with a soft brush in the laboratory before being oven-dried at 65 °C. 2.3. Laboratory analyses Soil water content (SWC) was measured by oven-drying 15 g of fresh soil sample at 105 °C for 24 h. Soil MBC was determined by subjecting fresh soil samples to the chloroform fumigation-extraction method (Vance et al., 1987). Soil readily oxidized organic carbon (ROC) was measured by 333 mM KMnO4 oxidation-colorimetry (Blair et al., 1995). The subsamples of litter and soil were analyzed for their C and nitrogen (N) concentrations and C isotope compositions by an isotope ratio mass spectrometer (IsoPrime 100, IsoPrime, Manchester, UK) connected to an elemental analyzer (Vario isotope cube, Elementar, Hanau, Germany). C isotope ratios (R, 13C/12C) of subsamples were expressed in δ notation (in ‰ units):

R δ = ⎛ sam − 1⎞ × 1000 ⎝ Rstd ⎠ ⎜

⎟

13

(1) 12

13

12

where Rsam is the C/ C ratio of the sample and Rstd is the C/ C ratio of the Vienna Pee Dee Belemnite (VPDB) standard (Coplen et al., 2006). Analysis of internal laboratory standards ensured that the precision of the measurements was ± 0.1‰ for δ13C. 2.4. Soil bulk density and organic carbon stock

2. Materials and methods

Soil bulk density was determined in parallel with field sampling. Undisturbed soil cores were taken from five randomly selected locations at three depths (0–10 cm, 10–20 cm, and 20–30 cm) within each plot using a stainless steel corer (5.65 cm in diameter, 4 cm in depth, and 100 cm3 in volume). All soil cores were oven-dried at 105 °C to a constant weight. Soil bulk density was calculated as the ratio of total dry weight to total soil volume. Soil organic carbon (SOC) stock is the product of SOC concentration, layer thickness, and bulk density (Post et al., 2001). The SOC stock can be calculated as follows:

2.1. Study site The study site is located in the southwest of Guangdong Province, China. The climate of this region is strongly influenced by subtropical monsoons, with mean annual precipitation of 1400 to 1956 mm, of which nearly 80% falls during the wet season (April through September). The mean annual temperature ranges from 21.7 to 23 °C. Soils are classified as Ultisol and Oxisol according to the USDA soil classification system (Buol et al., 2003). The zonal climax vegetation is monsoon evergreen broad-leaved forest (MEBF), but most of the forest has been clear-cut in the past two centuries, and some of the area has even degenerated into ‘bare land’ without living trees or shrubs because of soil erosion. As a result, the government launched a series of afforestation campaigns starting in the early 1960s, and encouraged local residents to plant different tree species on the degraded lands or enable the deforested lands to naturally restore (Deng et al., 2014; Wang et al., 2017b). At present, the area of the secondary forest has exceeded half of the total forest area in this region (SFA, 2013). In this study, we examined the depth trends in soil C concentrations and δ13C values in an undisturbed old-growth forest (> 400 years) and five restored forests: a 51-year-old mixed-native plantation (NP1); a 31-year-old mixed-native plantation (NP2); a 31-year-old Acacia mangium plantation (AP); a 31year-old mixed-conifer plantation (CP); and a 31-year-old secondary forest with natural restoration (SF). Further information of the study site is presented in Table 1.

SOCstock = SOCconcentration × H × D

(2)

where H is soil layer thickness and D is bulk density. 2.5. Calculations of isotopic enrichment factors Isotope fractionations occur during biochemical reactions (e.g., SOC decomposition) or during other processes that are affected by mass. The kinetic isotope fractionation factor, denoted by α, is defined as the ratio of C isotope ratios in reactant and product (Farquhar et al., 1989):

α = Rr / Rp

(3) 13

12

where Rr and Rp are the C/ C ratios of reactant and product, respectively. During the process of SOC decomposition, α is defined as the ratio of C isotope ratios in SOC and respired CO2 (Diochon and Kellman, 2008):

α = RSOC / RCO2 2

(4)

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Table 1 Description of the study site. Forest

Age (Year)

Dominant species

Land-use history

MEBF

> 400

Undisturbed old-growth forest more than 400 years

NP1

51

NP2 AP CP SF

31 31 31 31

Aporusa yunnanensis, Schima superba, Blastus cochinchinensis, Castanopsis chinensis, and Cryptocarya chinensis Carallia brachiata, Acacia confusa, Syzygium rehderianum, Cinnamomum camphora, and Ficus chlorocarpa Schima superba, Castanopsis hystrix, Schima wallichii, and Michelia macclurei Acacia mangium Pinus massoniana, Pinus elliottii, and Cunninghamia lanceolata Melicope pteleifolia, Trema tomentosa, Ilex asprella, and Dicranopteris dichotoma

The forest was planted with native species in 1964 The forest was planted with native species in 1984 The forest was planted with Acacia mangium in 1984 The forest was planted with coniferous species in 1984 Natural succession has occurred in the forest since 1984

MEBF, monsoon evergreen broad-leaved forest; NP1, 51-year-old mixed-native plantation; NP2, 31-year-old mixed-native plantation; AP, Acacia mangium plantation; CP, mixed-conifer plantation; SF, secondary forest with natural restoration.

where RSOC is the 13C/12C ratio of SOC and RCO2 is that of the respired CO2. Defined in this way, α can be thought of as the ratio of the rate constants for 12C and 13C containing substrates, k12 and k13, respectively (Farquhar et al., 1989). Thus

α = k12/ k13

value and the log-transformed C concentration in well-drained undisturbed soils (Powers and Schlesinger, 2002; Acton et al., 2013):

δ 13C = βlog (C ) + b2

The slope of the regression (β) is a proxy of SOC turnover, and b2 is a constant. More negative β (steeper slope) was associated with higher turnover rate (Wang et al., 2018). The advantage of using the β value is its purely descriptive nature and the absence of underlying assumptions that confine use of the α value. The disadvantage is that the β value is not easily linked to the kinetic isotope fractionation that occurs during SOC decomposition (Garten, 2006). Both α and β are utilized in this study because there is no broad consensus on which method yields the most appropriate parameter for describing vertical changes in soil δ13C.

(5)

For most reactions, the light isotope molecules react faster than heavy isotope molecules. Thus, α is positively correlated with the rate constant of SOC decomposition. When the system is in steady state, we can determine the fractionation factor α using the linear relationship between C isotope composition and natural logarithm-transformed C concentration. In this study, α was calculated from the slope of the following equation modified from Poage and Feng (2004):

1 C Δδ 13C = 1000 ⎛ − 1⎞ ln ⎛ d ⎞ + b1 ⎝α ⎠ ⎝ C0 ⎠ ⎜

(7)

2.6. Data analyses All statistical analyses were conducted with the SPSS statistical software (SPSS Statistics 20, IBM Corp.). Data were transformed to meet the assumptions of normality and homogeneity of variances when necessary. Two-way ANOVA was performed to detect the effects of forests and depths on C concentrations and δ13C values. One-way ANOVA was conducted to examine the differences in C concentrations and δ13C of litter layer, C:N ratios of litter and soil, soil bulk density, SWC, ROC, MBC, SOC stock, and α and β values among different forests. Tukey’s multiple comparison test was further conducted to separate differences among means. Linear regression analysis was used to examine the relationships between α value and other soil properties. Statistical significance was determined at p < 0.05 unless otherwise stated.

⎟

(6)

where Δδ13C is the δ13C difference between mineral soil samples and the surface litter, Cd and C0 are the C concentrations of mineral soil samples and the surface litter, respectively, and b1 is a constant. Some studies have adapted the equation to describe the changes in δ13C values of organic matter with increasing soil depth (Garten, 2006, Wang et al., 2015; Gautam et al., 2017). This approach has the advantage of relating the process of SOC decomposition to kinetic isotope fractionation and theory regarding the isotopic relationship of a substrate (i.e., SOC) to its product (respired CO2). It should be pointed out that the application of the equation is based on the following two assumptions: (1) C isotopes are homogeneously distributed within the substrate and all components of SOC decompose at the same rate with depth in the profile (Wynn et al., 2006); (2) the isotopic enrichment of down-profile SOC is generally controlled by the kinetic isotope fractionation (Poage and Feng, 2004). Another empirical linear relation was observed between the δ13C

3. Results 3.1. C concentration, SOC stock and δ13C Both the C concentration and δ13C in the litter layer showed

Table 2 Carbon concentrations, bulk densities and soil organic carbon (SOC) stocks in the litter layer and upper 30-cm of soils in different forests. Forest

Carbon concentration

Bulk density

SOC stock

(%)

(g cm−3)

(Mg C ha−1)

Litter layer MEBF NP1 NP2 AP CP SF

41.69 45.49 46.94 51.00 48.12 47.42

± ± ± ± ± ±

0–10 cm c

1.66 0.44b 0.21b 0.17a 0.46b 0.36b

2.84 3.21 3.05 3.04 2.83 2.57

± ± ± ± ± ±

10–20 cm ab

0.09 0.09a 0.21ab 0.15ab 0.12ab 0.20b

1.59 1.27 1.40 1.34 1.29 1.23

± ± ± ± ± ±

20–30 cm a

0.06 0.07a 0.09a 0.14a 0.11a 0.09a

1.06 0.78 0.92 0.83 0.77 0.95

± ± ± ± ± ±

0–30 cm a*

0.05 0.09b* 0.05ab* 0.08ab* 0.07b* 0.06ab*

1.19 1.20 1.34 1.23 1.35 1.42

± ± ± ± ± ±

0–30 cm c

0.03 0.07c 0.03ab 0.03bc 0.03ab 0.03a

65.10 61.87 68.14 61.71 63.15 64.61

± ± ± ± ± ±

2.18a 1.99a 3.69a 3.19a 3.03a 4.24a

MEBF, monsoon evergreen broad-leaved forest; NP1, 51-year-old mixed-native plantation; NP2, 31-year-old mixed-native plantation; AP, Acacia mangium plantation; CP, mixed-conifer plantation; SF, secondary forest with natural restoration. Values are mean ± standard error (n = 6; MEBF, n = 8). Different lowercase letters indicate significant differences among different forests (p < 0.05). * indicates significant differences among soil depths (p < 0.01). 3

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Table 3 Carbon isotope compositions (δ13C) in the litter layer and upper 30-cm of soils in different forests. Forest

Litter layer

MEBF NP1 NP2 AP CP SF

−29.43 −31.43 −30.44 −31.79 −28.70 −29.98

Soil depth (cm) 0–10

± ± ± ± ± ±

−27.31 −28.65 −25.61 −27.45 −23.81 −24.85

b

0.29 0.11d 0.13c 0.12d 0.20a 0.33bc

10–20 ± ± ± ± ± ±

−25.93 −26.24 −20.47 −22.75 −17.68 −18.30

d

0.17 0.30e 0.44c 0.12d 0.27b 0.48c

20–30 ± ± ± ± ± ±

−25.04 −24.58 −18.71 −20.56 −17.21 −16.40

d

0.34 0.66d 0.74b 0.92c 0.25a 0.37a

± ± ± ± ± ±

0.35e* 0.65e* 0.47b* 0.32c* 0.60a* 0.28a*

MEBF, monsoon evergreen broad-leaved forest; NP1, 51-year-old mixed-native plantation; NP2, 31-year-old mixed-native plantation; AP, Acacia mangium plantation; CP, mixed-conifer plantation; SF, secondary forest with natural restoration. Values are mean ± standard error (n = 6; MEBF, n = 8). Different lowercase letters indicate significant differences among different forests (p < 0.05). * indicates significant differences among soil depths (p < 0.01)

differences among forests (p < 0.05), and increased in the order of MEBF = NP1 < NP2 = AP = CP < SF (Table 5). Linear regression of δ13C with the log-transformed C concentration fit the data well for all the forest sites, as judged by R2 and p values (Table 5). The estimated regression slope (β value) ranged from −19.70 to −5.22 and showed significant differences among forests (p < 0.05), which decreasing in the order of MEBF = NP1 > NP2 = AP = CP > SF (Table 5). In addition, a strongly negative correlation was observed between α and β (R2 = 0.99, p < 0.01; Fig. 1).

significant differences among forests (p < 0.05; Tables 2, 3). The C concentration of litter layer from different forests ranged from 41.69 to 51.00%, and the δ13C value ranged from −31.79‰ to −28.70‰ (Tables 2, 3). For the upper-20 cm of soils, C concentration in each of the restored forests (NP1, NP2, AP, CP, and SF) was similar to that in the MEBF (p > 0.05; Table 2). The C concentration of the 20–30 cm soil layer was highest in the MEBF (Table 2). The C concentration significantly decreased with depth in all the mineral soil profiles (p < 0.01; Table 2). Soil bulk density in the MEBF was similar to that in the NP1 (p > 0.05), but was significantly lower than that in other forests (p < 0.05; Table 2). There were no significant differences in SOC stock (0–30 cm) among different forests (p > 0.05; Table 2). Soil δ13C significantly increased with depth in all the mineral soil profiles (p < 0.01), and the vertical δ13C enrichment (Δδ13C) from different forests ranged from 2.27‰ to 8.45‰ (Table 3).

3.4. The relationships between α and soil properties Results of regression analyses showed that the α value of soil profile was positively correlated with soil bulk density (R2 = 0.81, p < 0.05; Fig. 2a), but negatively with SWC (R2 = 0.87, p < 0.01; Fig. 2b). However, the relationships between the α value and litter C:N ratio, soil C:N ratio, ROC and MBC were not significant (p > 0.05; Fig. 2c-f).

3.2. C:N ratios of litter and soil, SWC, ROC and MBC The C:N ratios of litter and soil as well as SWC, ROC and MBC showed significant differences among forests (p < 0.05; Table 4). The C:N ratio of litter was lowest in the SF; there were no significant differences among other forests (p > 0.05; Table 4). The soil had the lowest C:N ratio in the MEBF; there were no significant differences among other forests (p > 0.05; Table 4). The pattern of ROC was consistent with the C:N ratio of soil (Table 4). SWC in the MEBF was similar to that in the NP1 (p > 0.05), but significantly higher than that in other forests (p < 0.05; Table 4). Soil MBC was significantly different among forests and it was highest in the AP (Table 4).

4. Discussion 4.1. Impacts of reforestation on SOC stock This study demonstrated that soil C stocks in the reforested lands, regardless of tree species planted, could be restored to the equivalent level of the undisturbed old-growth forest within a few decades. As the forest recovers, the consecutive input of fresh organic C (above- and below-ground) resulted in increases in SOC pools, although soil bulk density decreased (Table 2). Reforestation ages may play the key role in the increase of soil C stock, particularly for the sites with no intermediate land use after initial clearing (Marin-Spiotta and Sharma, 2013). Brown and Lugo (1990) found that there was a general pattern of increasing soil C stocks with increasing age of secondary forests, the 50-year-old secondary forests having approximately the same soil C stocks as the primary forests in subtropics. In this study, the length of time was shorter in accordance with the result reported by Veloso et al.

3.3. α and β values The linear regressions of Δδ13C against ln(Cd/Co) for all the forest sites were significant (p < 0.05) but showed disparate δ13C increasing trends with depth among different forests. The regression fits were pretty good for all the soil profiles (R2 > 0.85; Table 5). The estimated α value ranged from 1.0023 to 1.0086 and showed significant Table 4 Soil properties of the upper 30-cm in different forests. Forest

Litter C:N

MEBF NP1 NP2 AP CP SF

28.56 31.50 31.18 29.09 34.39 19.52

± ± ± ± ± ±

Soil C:N 2.42a 1.04a 1.09a 0.59a 1.32a 0.36b

12.05 13.67 14.01 14.13 13.60 14.30

± ± ± ± ± ±

ROC (g kg−1)

SWC (%) 0.11b 0.18a 0.24a 0.15a 0.17a 0.52a

25.10 23.32 21.08 21.15 18.16 13.98

± ± ± ± ± ±

0.58a 0.99ab 0.10b 0.37b 0.89c 1.32d

4.83 7.48 6.45 6.80 6.05 6.52

± ± ± ± ± ±

0.25b 0.50a 0.45a 0.57a 0.33a 0.59a

MBC (mg kg−1) 236.94 281.25 309.14 333.76 244.39 219.84

± ± ± ± ± ±

22.74bc 19.62abc 18.00ab 26.45a 16.17bc 11.91c

C:N, the ratio of carbon to nitrogen; SWC, soil water content; ROC, soil readily oxidized organic carbon; MBC, soil microbial biomass carbon. MEBF, monsoon evergreen broad-leaved forest; NP1, 51-year-old mixed-native plantation; NP2, 31-year-old mixed-native plantation; AP, Acacia mangium plantation; CP, mixedconifer plantation; SF, secondary forest with natural restoration. Values are mean ± standard error (n = 6; MEBF, n = 8). Different lowercase letters indicate significant differences among different forests (p < 0.05). 4

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Table 5 Parameters of regression equations between carbon isotope composition (δ13C) and logarithm-transformed carbon (C) concentration in different forests. Forest

Δδ13C vs. ln (Cd/C0) R2

Slope MEBF NP1 NP2 AP CP SF

−2.27 −2.99 −6.01 −5.70 −5.60 −8.56

δ13C vs. log (C)

± ± ± ± ± ±

0.18 0.51 0.44 0.84 0.53 0.46

> 0.96 > 0.97 > 0.92 > 0.97 > 0.85 > 0.96

p < 0.05 < 0.05 < 0.05 < 0.05 < 0.05 < 0.05

α

Slope (β)

1.0023 1.0030 1.0060 1.0057 1.0056 1.0086

± ± ± ± ± ±

c

0.0002 0.0005c 0.0004b 0.0009b 0.0005b 0.0005a

−5.22 ± 0.41 −6.89 ± 1.18a −13.83 ± 1.02b −13.13 ± 1.93b −12.89 ± 1.21b −19.70 ± 1.05c a

R2

p

> 0.96 > 0.97 > 0.92 > 0.97 > 0.85 > 0.96

< 0.05 < 0.05 < 0.05 < 0.05 < 0.05 < 0.05

Cd and C0 are the carbon concentration of mineral soil samples and the surface litter, respectively. MEBF, monsoon evergreen broad-leaved forest; NP1, 51-year-old mixed-native plantation; NP2, 31-year-old mixed-native plantation; AP, Acacia mangium plantation; CP, mixed-conifer plantation; SF, secondary forest with natural restoration. Values are mean ± standard error (n = 6; MEBF, n = 8). Different lowercase letters indicate significant differences among different forests (p < 0.05).

Brunn et al. (2017) found that the soil δ13C gradients with depth can form within 30 years following reforestation, which is in accordance with the results of this study. In addition, the depth trend of soil δ13C may be highly site-dependent. The mean enrichment of soil δ13C in the MEBF was 2.3‰, while the enrichment ranged from 4.1‰ to 8.5‰ in the restored forests (Table 3). The vertical enrichment of soil δ13C observed in Costa Rican rain forests has been reported as 5.9‰ (Powers and Schlesinger, 2002), in Amazonian forests as 5.0‰ (Desjardins et al., 1994), in a temperate forest as 3.4‰ (Brunn et al., 2014), and in a boreal forest as 1.2‰ (Flanagan et al., 1996). In the semi-arid grassland of northern China, the mean enrichment of soil δ13C has been reported as 3.7‰ (Wang et al., 2017a). On a global scale, the average variation of soil δ13C with depth was reported as 3.5‰ according to a meta-analysis of 176 soil profiles worldwide (Wang et al., 2018). Each soil profile holds the unique story of its development, which is controlled by the specific climatic, biotic and edaphic conditions. The difference in the vertical enrichment of soil δ13C may be attributed to the differences in the quality and quantity of plant C inputs and the consequences of C cycling processes in soil depth profiles. As summarized in the introduction, two main processes have been proposed to explain the δ13C enrichment with increasing soil depth: mixing of different C sources and soil C decomposition (Wynn et al., 2006; Acton et al., 2013). The Suess effect describes the isotopic depletion of atmospheric CO2 as a result of the combustion of 13C-depleted fossil fuels and biomass burning (Friedli et al., 1986), which could have caused the increasing trends of δ13C in the soil profiles (more 13C-enriched older SOC in the deep and more 13C-depleted recent SOC at the surface) (Diochon and Kellman, 2008). However, the average δ13C value of atmospheric CO2 decreased only about 1.1‰ during the period from 1964 (−7.3‰) to 2015 (−8.4‰) (Keeling et al., 2017). The decrease was much smaller than the vertical enrichments of δ13C in this study (Table 3), showing that there are other mechanisms besides the Suess effect. It is worth noting that while the gradual decrease of δ13C in atmospheric CO2 causes a corresponding decrease of δ13C in plants, the complex effects induced by climate change (e.g., warming or regional drought) might offset this subtle difference through constraining the carbon isotope discrimination in photosynthesis (Cornwell et al., 2018; Szejner et al., 2018). Torn et al. (2002) showed that the C isotopic signatures of a modern soil profile (collected in 1997) are similar to its 100-year-old archive soil (collected sometime between 1895 and 1903) from the same site, indicating that the change of soil δ13C with depth is not due to the isotopic depletion of atmospheric CO2. A few studies have proposed that roots are generally more 13C-enriched than leaves or branches from the same plant, and so the soil δ13C value at the surface may be lower than that in the deeper soil (Powers and Schlesinger, 2002). However, the mean difference documented between above- and below-ground biomass was no more than 1.5‰ (Wedin et al., 1995; Wynn et al., 2006). What is more, we did not find significant differences in δ13C values between roots and above-ground litter in the MEBF through ten years of observation (Fig.

Fig. 1. Relationship between β value and α value.

(2018), which stated that reforestation can restore the initial soil C stock relative to a subtropical natural forest after 32 years. In a previous study of changes in soil C stocks with land-use change (Detwiler, 1986), soil C stocks in forest fallows after abandonment of agricultural land were modeled to increase steadily to full recovery of tropical primary forest levels in the first 35 years. With a meta-analysis, Martin et al. (2013) found similar soil C stocks in secondary and undisturbed tropical forests, which showed little relationship with time. However, an earlier meta-analysis by Don et al. (2011) suggested that the increased C stocks in afforested soils are still lower than those in undisturbed forests, probably because their analysis included far more human-disturbed plantations than undisturbed forests. No significant differences in soil C stock were observed among restored forests with different tree species after 31 years of reforestation (Table 2), indicating that forest composition made a minor contribution to soil C restoration. The result is supported by other studies (e.g., Hoogmoed et al., 2014; Kolbe et al., 2016). Generally, N-fixing species may sequester more soil C stocks than the other species owing to the extra N input (Kasel et al., 2011), but that the response may be site- and species-specific (Hoogmoed et al., 2014). In this study area, atmospheric N deposition and soil N availability are high, and the development of forest ecosystems is not limited by N status (Lu et al., 2014). 4.2. Changes in δ13C with depth in the mineral soils Our results showed the enrichments of soil δ13C with depth at all the forest sites, concomitant with decreasing C concentrations (Tables 2, 3). It is consistent with the earlier observations in well-drained forest soils (Nadelhoffer and Fry, 1988; Powers and Schlesinger, 2002; Acton et al., 2013). As forest recovers, soil δ13C becomes depth-dependent because of organic matter inputs from vegetation (Billings and Richter, 2006). 5

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Fig. 2. Relationships between the α value and soil properties. a, relationship between the α value and soil bulk density; b, relationship between the α value and soil water content (SWC); c, relationship between the α value and litter C:N ratio; d, relationship between the α value and soil C:N ratio; e, relationship between the α value and soil readily oxidized organic carbon (ROC); f, relationship between the α value and soil microbial biomass carbon (MBC). The error bars indicate standard errors of mean.

S1). In summary, the existing evidences imply that the vertical δ13C enrichment along soil profiles that can be attributed to the mixing of different C sources should be small. The other explanation is that the enrichment of δ13C with soil depth resulted from the kinetic isotope fractionation during SOC decomposition (Balesdent et al., 1993; Powers and Schlesinger, 2002; Brunn et al., 2014). Heavy isotope atoms with extra neutrons make bonds that are harder to break, and move slowly relative to light isotope atoms (Farquhar et al., 1989). Soil microbes have been shown to discriminate against 13C during the degradation of both simple and complex C compounds, resulting in the gradual 13C-enrichment in the residual SOC (Fernandez et al., 2003; Menichetti et al., 2015). As depth increases, the decay degree of SOC increases, and subsequently, the SOC in deep horizons becomes more enriched in 13C. If isotopic fractionation is the dominate mechanism for δ13C enrichment along soil profile, the increasing trend of soil δ13C could be used as a robust indicator for soil C turnover dynamics (Poage and Feng, 2004; Acton et al., 2013).

4.3. Impacts of reforestation on SOC turnover Both α and β have been proposed as proxies for SOC turnover rate at different spatial scales, as higher α or lower β implies higher turnover rates (Wang et al., 2015, 2018). Relative to the increasingly wide utilization of the β value, the application of the α value is confined by the assumption that the 13C enrichment of down-profile SOC is generally controlled by the kinetic isotope fractionation (Poage and Feng, 2004). In this study, there was a perfect linear relationship of β with α (R2 = 0.99; Fig. 1), providing evidence that the assumption stated in the previous sentence is true. Therefore, the two values can be used interchangeably to describe the turnover rate of soil C in the study region. The mean α value estimated in the MEBF was 1.0023 (Table 5), which is in accordance with other observations in tropical forests (Laskar et al., 2016). The mean α value in the NP1 was similar to that in the MEBF (Table 5), indicating that the soil C turnover rate had 6

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measurements; DH revised the draft; GC and ZM collected the samples; GZ discussed the draft with XX. All authors contributed critically to the drafts and gave final approval for publication.

approached the level of undisturbed old-growth forests after 51 years of reforestation on the degraded lands. However, the turnover of soil C in the restored forests of 31 years was significantly faster than that in the undisturbed old-growth forest, regardless of tree species planted (Table 5). Our results favor the traditional opinion that soil C sequestration is limited by rapid decomposition in the early stage of forest restoration (Richter et al., 1999). The results also demonstrated that tree species has little impact on soil C turnover rate (Table 5). The α value observed in the SF was significantly higher than that in the plantations of the same age (Table 5), suggesting that artificial restoration management can effectively slow down soil C turnover rate relative to natural secondary succession. Reforestation mainly regulates the turnover of soil C by altering soil physical structure and water content. Our results showed that soil C turnover rate was positively correlated with soil bulk density, but negatively correlated with SWC (Fig. 2). This suggests that the decreasing bulk density and increasing SWC inhibited the decomposition of soil C following reforestation. As forest recovers, the input of biomass C increases, accompanied by the development and agitation of roots, which promotes the formation of soil aggregate structure, which in turn facilitates the preservation of organic C (Bronick and Lal, 2005; Paul, 2016). High soil porosity and moisture accelerate the vertical transport of SOC to deep soil layers (Zhou et al., 2019). Moreover, high SWC may help to maintain anaerobic condition and inhibit microbial decomposition of SOC (Schuur et al., 2001). Results in this study showed that soil C:N ratio and ROC had no direct effects on soil C turnover rate, providing further support for the viewpoint that SOC turnover is governed by accessibility rather than recalcitrance (Dungait et al., 2012). Some studies have indicated that the decomposition of SOC was closely associated with the soil C:N ratio, and suggested soil chemical characteristics dominated the process of soil C turnover (Powers and Schlesinger, 2002; Xu et al., 2016). We did not detect direct evidence for the impact of litter C:N ratio on soil C turnover rate, although it profoundly affects the fate of fresh organic C and SOC accumulation (Huang et al., 2011; Zhou et al., 2019). Soil microbial biomass did not show a regular trend with the forest development (Table 4). The growth and activity of soil microbes might be regulated by the complex interplay of multiple factors, such as exogenous C input, soil N availability and soil moisture (Skopp et al., 1990; Fontaine and Barot, 2005; Meyer et al., 2018), which caused a weak correlation between soil C turnover rate and microbial biomass in this study.

Declaration of Competing Interest We declare that we have no conflict of interest. Acknowledgements The research was financially supported by the National Natural Science Foundation of China (41773088, 41573077, 31870461), the “Hundred Talent Program” of South China Botanical Garden at the Chinese Academy of Sciences (Y761031001), the Key Research Program of the Chinese Academy of Sciences (QYZDJ-SSW-DQC003), and by Key Special Project for Introduced Talents Team of Southern Marine Science and Engineering Guangdong Laboratory (Guangzhou) (GML2019ZD0408). Data availability statement The data supporting the results will be available on request. Appendix A. Supplementary material Supplementary data to this article can be found online at https:// doi.org/10.1016/j.foreco.2020.117988. References Acton, P., Fox, J., Campbell, E., Rowe, H., Wilkinson, M., 2013. Carbon isotopes for estimating soil decomposition and physical mixing in well-drained forest soils. J. Geophys. Res-Biogeo. 118, 1532–1545. https://doi.org/10.1002/2013jg002400. Balesdent, J., Basile-Doelsch, I., Chadoeuf, J., Cornu, S., Derrien, D., Fekiacova, Z., Hatte, C., 2018. Atmosphere-soil carbon transfer as a function of soil depth. Nature 559, 599–602. https://doi.org/10.1038/s41586-018-0328-3. Balesdent, J., Girardin, C., Mariotti, A., 1993. Site-related δ13C of tree leaves and soil organic matter in a temperate forest. Ecology 74, 1713–1721. https://doi.org/10. 2307/1939930. Benchimol, M., Talora, D.C., Mariano-Neto, E., Oliveira, T.L.S., Leal, A., Mielke, M.S., Faria, D., 2017. Losing our palms: The influence of landscape-scale deforestation on Arecaceae diversity in the Atlantic forest. For. Ecol. Manage. 384, 314–322. https:// doi.org/10.1016/j.foreco.2016.11.014. Bernoux, M., Cerri, C.C., Neill, C., de Moraes, J.F.L., 1998. The use of stable carbon isotopes for estimating soil organic matter turnover rates. Geoderma 82, 43–58. https://doi.org/10.1016/s0016-7061(97)00096-7. Billings, S.A., Richter, D.D., 2006. Changes in stable isotopic signatures of soil nitrogen and carbon during 40 years of forest development. Oecologia 148, 325–333. https:// doi.org/10.1007/s00442-006-0366-7. Blair, G.J., Lefroy, R.D.B., Lise, L., 1995. Soil carbon fractions based on their degree of oxidation and the development of a carbon management index for agricultural systems. Aust. J. Agric. Res. 46, 1459–1466. https://doi.org/10.1071/ar9951459. Bronick, C.J., Lal, R., 2005. Soil structure and management: a review. Geoderma 124, 3–22. https://doi.org/10.1016/j.geoderma.2004.03.005. Brown, S., Lugo, A.E., 1990. Effects of forest clearing and succession on the carbon and nitrogen content of soils in Puerto Rico and US Virgin Islands. Plant Soil 124, 53–64. https://doi.org/10.1007/bf00010931. Brunn, M., Brodbeck, S., Oelmann, Y., 2017. Three decades following afforestation are sufficient to yield C-13 depth profiles. J. Plant Nutr. Soil Sci. 180, 643–647. https:// doi.org/10.1002/jpln.201700015. Brunn, M., Spielvogel, S., Sauer, T., Oelmann, Y., 2014. Temperature and precipitation effects on delta C-13 depth profiles in SOM under temperate beech forests. Geoderma 235, 146–153. https://doi.org/10.1016/j.geoderma.2014.07.007. Buol, S.W., Southard, R.J., Graham, R.C., McDaniel, P.A., 2003. Soil Genesis and Classification, fifth ed. Iowa State Press, Ames, pp. 339–347. Chen, C., Park, T., Wang, X., Piao, S., Xu, B., Chaturvedi, R.K., Fuchs, R., Brovkin, V., Ciais, P., Fensholt, R., Tommervik, H., Bala, G., Zhu, Z., Nemani, R.R., Myneni, R.B., 2019. China and India lead in greening of the world through land-use management. Nat. Sustain. 2, 122–129. https://doi.org/10.1038/s41893-019-0220-7. Coplen, T.B., Brand, W.A., Gehre, M., Groning, M., Meijer, H.A.J., Toman, B., Verkouteren, R.M., 2006. New guidelines for delta C-13 measurements. Anal. Chem. 78, 2439–2441. https://doi.org/10.1021/ac052027c. Cornwell, W.K., Wright, I.J., Turner, J., Maire, V., Barbour, M.M., Cernusak, L.A., Dawson, T., Ellsworth, D., Farquhar, G.D., Griffiths, H., Keitel, C., Knohl, A., Reich, P.B., Williams, D.G., Bhaskar, R., Cornelissen, J.H.C., Richards, A., Schmidt, S., Valladares, F., Korner, C., Schulze, E.D., Buchmann, N., Santiago, L.S., 2018. Climate

5. Conclusion This study demonstrated that the vertical δ13C gradients along 0–30 cm soil profiles can form within about 30 years following reforestation on degraded lands; these are mainly attributed to the kinetic isotope fractionation during the decomposition of soil organic matter. Either the α or β value can be used to describe the soil C turnover rate in the tropical forests of China. The restored forests can provide equivalent C stock in the soil to the undisturbed old-growth forests within a few decades, while the rate of soil C turnover rates were still higher in these restored forests except for the 51-year-old mixed-native plantation only. Additionally, the α values in the natural restoration forests were significantly higher than those in the artificial restoration forests over 31 years, while there was no difference among the three artificial restoration forests with different tree species planted. This may suggest that reforestation ages rather than tree species constrained soil C dynamics following reforestation. Relative to natural secondary succession, artificial restoration management may be more beneficial to slow down soil C turnover during ecological restoration. 6. Authors’ contributions QD and DZ conceived and designed the study; XX analyzed the data, drew the figures and wrote the primary draft; HZ did the 7

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Menichetti, L., Houot, S., van Oort, F., Katterer, T., Christensen, B.T., Chenu, C., Barre, P., Vasilyeva, N.A., Ekblad, A., 2015. Increase in soil stable carbon isotope ratio relates to loss of organic carbon: results from five long-term bare fallow experiments. Oecologia 177, 811–821. https://doi.org/10.1007/s00442-014-3114-4. Meyer, N., Welp, G., Rodionov, A., Borchard, N., Martius, C., Amelung, W., 2018. Nitrogen and phosphorus supply controls soil organic carbon mineralization in tropical topsoil and subsoil. Soil. Biol. Biochem. 119, 152–161. https://doi.org/10. 1016/j.soilbio.2018.01.024. Nadelhoffer, K.F., Fry, B., 1988. Controls on natural N-15 and C-13 abundances in forest soil organic-matter. Soil Sci. Soc. Am. J. 52, 1633–1640. Paul, E.A., 2016. The nature and dynamics of soil organic matter: Plant inputs, microbial transformations, and organic matter stabilization. Soil. Biol. Biochem. 98, 109–126. https://doi.org/10.1016/j.soilbio.2016.04.001. Paul, K.I., Polglase, P.J., Nyakuengama, J.G., Khanna, P.K., 2002. Change in soil carbon following afforestation. For. Ecol. Manage. 168, 241–257. https://doi.org/10.1016/ s0378-1127(01)00740-x. Poage, M.A., Feng, X.H., 2004. A theoretical analysis of steady state delta C-13 profiles of soil organic matter. Globa. Biogeochem. Cy. 18, GB2016. https://doi.org/10.1029/ 2003gb002195. Post, W.M., Izaurralde, R.C., Mann, L.K., Bliss, N., 2001. Monitoring and verifying changes of organic carbon in soil. Clim. Change 51, 73–99. https://doi.org/10.1023/ a:1017514802028. Powers, J.S., Schlesinger, W.H., 2002. Geographic and vertical patterns of stable carbon isotopes in tropical rain forest soils of Costa Rica. Geoderma 109, 141–160. https:// doi.org/10.1016/s0016-7061(02)00148-9. Richter, D.D., Markewitz, D., Trumbore, S.E., Wells, C.G., 1999. Rapid accumulation and turnover of soil carbon in a re-establishing forest. Nature 400, 56–58. https://doi. org/10.1038/21867. Schuur, E.A.G., Chadwick, O.A., Matson, P.A., 2001. Carbon cycling and soil carbon storage in mesic to wet Hawaiian montane forests. Ecology 82, 3182–3196. https:// doi.org/10.1890/0012-9658(2001) 082[3182:Ccascs]2.0.Co;2. Seymour, F., Harris, N.L., 2019. Reducing tropical deforestation. Science 365, 756–757. https://doi.org/10.1126/science.aax8546. SFA, 2013. Reports of Chinese Forestry Resource: The 8th National Forest Resource Inventory. Chinese Forestry Publishing House, Beijing, China. Skopp, J., Jawson, M.D., Doran, J.W., 1990. Steady-state aerobic microbial activity as a function of soil-water content. Soil Sci. Soc. Am. J. 54, 1619–1625. https://doi.org/ 10.2136/sssaj1990.03615995005400060018x. Szejner, P., Wright, W.E., Belmecheri, S., Meko, D., Leavitt, S.W., Ehleringer, J.R., Monson, R.K., 2018. Disentangling seasonal and interannual legacies from inferred patterns of forest water and carbon cycling using tree-ring stable isotopes. Global Change Biol. 24, 5332–5347. https://doi.org/10.1111/gcb.14395. Torn, M.S., Lapenis, A.G., Timofeev, A., Fischer, M.L., Babikov, B.V., Harden, J.W., 2002. Organic carbon and carbon isotopes in modern and 100-year-old-soil archives of the Russian steppe. Global Change Biol. 8, 941–953. https://doi.org/10.1046/j.13652486.2002.00477.x. Vance, E.D., Brookes, P.C., Jenkinson, D.S., 1987. An extraction method for measuring soil microbial biomass-C. Soil. Biol. Biochem. 19, 703–707. https://doi.org/10.1016/ 0038-0717(87)90052-6. Veloso, M.G., Dieckow, J., Zanatta, J.A., Bayer, C., Higa, R.C.V., Brevilieri, R.C., Comerford, N.B., Stoppe, A.M., 2018. Reforestation with loblolly pine can restore the initial soil carbon stock relative to a subtropical natural forest after 30 years. Eur. J. For. Res. 137, 593–604. https://doi.org/10.1007/s10342-018-1127-y. Wang, C., Houlton, B.Z., Liu, D.W., Hou, J.F., Cheng, W.X., Bai, E., 2018. Stable isotopic constraints on global soil organic carbon turnover. Biogeosciences 15, 987–995. https://doi.org/10.5194/bg-15-987-2018. Wang, C., Wei, H.W., Liu, D.W., Luo, W.T., Hou, J.F., Cheng, W.X., Han, X.G., Bai, E., 2017a. Depth profiles of soil carbon isotopes along a semi-arid grassland transect in northern China. Plant Soil 417, 43–52. https://doi.org/10.1007/s11104-017-3233-x. Wang, F.M., Ding, Y.Z., Sayer, E.J., Li, Q.L., Zou, B., Mo, Q.F., Li, Y.W., Lu, X.L., Tang, J.W., Zhu, W.X., Li, Z.A., 2017b. Tropical forest restoration: Fast resilience of plant biomass contrasts with slow recovery of stable soil C stocks. Funct. Ecol. 31, 2344–2355. https://doi.org/10.1111/1365-2435.12925. Wang, G.A., Jia, Y.F., Li, W., 2015. Effects of environmental and biotic factors on carbon isotopic fractionation during decomposition of soil organic matter. Sci. Rep. 5, 11043. https://doi.org/10.1038/srep11043. Wedin, D.A., Tieszen, L.L., Dewey, B., Pastor, J., 1995. Carbon-isotope dynamics during grass decomposition and soil organic-matter formation. Ecology 76, 1383–1392. https://doi.org/10.2307/1938142. Wynn, J.G., Harden, J.W., Fries, T.L., 2006. Stable carbon isotope depth profiles and soil organic carbon dynamics in the lower Mississippi Basin. Geoderma 131, 89–109. https://doi.org/10.1016/j.geoderma.2005.03.005. Xu, X., Shi, Z., Li, D.J., Rey, A., Ruan, H.H., Craine, J.M., Liang, J.Y., Zhou, J.Z., Luo, Y.Q., 2016. Soil properties control decomposition of soil organic carbon: Results from dataassimilation analysis. Geoderma 262, 235–242. https://doi.org/10.1016/j.geoderma. 2015.08.038. Zhou, G., Xu, S., Ciais, P., Manzoni, S., Fang, J., Yu, G., Tang, X., Zhou, P., Wang, W., Yan, J., Wang, G., Ma, K., Li, S., Du, S., Han, S., Ma, Y., Zhang, D., Liu, J., Liu, S., Chu, G., Zhang, Q., Li, Y., Huang, W., Ren, H., Lu, X., Chen, X., 2019. Climate and litter C/N ratio constrain soil organic carbon accumulation. Natl. Sci. Rev. 6, 746–757. https:// doi.org/10.1093/nsr/nwz045.

and soils together regulate photosynthetic carbon isotope discrimination within C-3 plants worldwide. Global Ecol. Biogeogr. 27, 1056–1067. https://doi.org/10.1111/ geb.12764. Deng, L., Liu, G.B., Shangguan, Z.P., 2014. Land-use conversion and changing soil carbon stocks in China's 'Grain-for-Green' Program: a synthesis. Global Change Biol. 20, 3544–3556. https://doi.org/10.1111/gcb.12508. Desjardins, T., Andreux, F., Volkoff, B., Cerri, C.C., 1994. Organic carbon and 13C contents in soils and soil size-fractions, and their changes due to deforestation and pasture installation in eastern Amazonia. Geoderma 61, 103–118. https://doi.org/10.1016/ 0016-7061(94)90013-2. Detwiler, R.P., 1986. Land use change and the global carbon cycle: the role of tropical soils. Biogeochemistry 2, 67–93. https://doi.org/10.1007/bf02186966. Diochon, A., Kellman, L., 2008. Natural abundance measurements of (13) C indicate increased deep soil carbon mineralization after forest disturbance. Geophys. Res. Lett. 35, L14402. https://doi.org/10.1029/2008gl034795. Don, A., Schumacher, J., Freibauer, A., 2011. Impact of tropical land-use change on soil organic carbon stocks - a meta-analysis. Global Change Biol. 17, 1658–1670. https:// doi.org/10.1111/j.1365-2486.2010.02336.x. Dungait, J.A.J., Hopkins, D.W., Gregory, A.S., Whitmore, A.P., 2012. Soil organic matter turnover is governed by accessibility not recalcitrance. Global Change Biol. 18, 1781–1796. https://doi.org/10.1111/j.1365-2486.2012.02665.x. Farquhar, G.D., Ehleringer, J.R., Hubick, K.T., 1989. Carbon isotope discrimination and photosynthesis. Annu. Rev. Plant Physiol. Plant Mol. Biol. 40, 503–537. https://doi. org/10.1146/annurev. pp. 40.060189.002443. Feng, W.T., Shi, Z., Jiang, J., Xia, J.Y., Liang, J.Y., Zhou, J.Z., Luo, Y.Q., 2016. Methodological uncertainty in estimating carbon turnover times of soil fractions. Soil. Biol. Biochem. 100, 118–124. https://doi.org/10.1016/j.soilbio.2016.06.003. Fernandez, I., Mahieu, N., Cadisch, G., 2003. Carbon isotopic fractionation during decomposition of plant materials of different quality. Globa. Biogeochem. Cy. 17, 1075. https://doi.org/10.1029/gb001834. Flanagan, L.B., Brooks, J.R., Varney, G.T., Berry, S.C., Ehleringer, J.R., 1996. Carbon isotope discrimination during photosynthesis and the isotope ratio of respired CO2 in boreal forest ecosystems. Globa. Biogeochem. Cy. 10, 629–640. https://doi.org/10. 1029/96gb02345. Fontaine, S., Barot, S., 2005. Size and functional diversity of microbe populations control plant persistence and long-term soil carbon accumulation. Ecol. Lett. 8, 1075–1087. https://doi.org/10.1111/j.1461-0248.2005.00813.x. Friedli, H., Lotscher, H., Oeschger, H., Siegenthaler, U., Stauffer, B., 1986. Ice core record of the C-12/C-13 ratio of atmospheric CO2 in the past 2 centuries. Nature 324, 237–238. https://doi.org/10.1038/324237a0. Garten, C.T., 2006. Relationships among forest soil C isotopic composition, partitioning, and turnover times. Can. J. Forest Res. 36, 2157–2167. https://doi.org/10.1139/x06115. Gautam, M.K., Lee, K.S., Song, B.Y., Bong, Y.S., 2017. Site related delta C-13 of vegetation and soil organic carbon in a cool temperate region. Plant Soil 418, 293–306. https:// doi.org/10.1007/s11104-017-3284-z. Hoogmoed, M., Cunningham, S.C., Baker, P.J., Beringer, J., Cavagnaro, T.R., 2014. Is there more soil carbon under nitrogen-fixing trees than under non-nitrogen-fixing trees in mixed-species restoration plantings? Agr. Ecosyst. Environ. 188, 80–84. https://doi.org/10.1016/j.agee.2014.02.013. Huang, Y.H., Li, Y.L., Xiao, Y., Wenigmann, K.O., Zhou, G.Y., Zhang, D.Q., Wenigmann, M., Tang, X.L., Liu, J.X., 2011. Controls of litter quality on the carbon sink in soils through partitioning the products of decomposing litter in a forest succession series in South China. For. Ecol. Manage. 261, 1170–1177. https://doi.org/10.1016/j.foreco. 2010.12.030. Jobbágy, E.G., Jackson, R.B., 2000. The vertical distribution of soil organic carbon and its relation to climate and vegetation. Ecol. Appl. 10, 423–436. https://doi.org/10. 1890/1051-0761(2000) 010[0423:Tvdoso]2.0.Co;2. Kasel, S., Singh, S., Sanders, G.J., Bennett, L.T., 2011. Species-specific effects of native trees on soil organic carbon in biodiverse plantings across north-central Victoria, Australia. Geoderma 161, 95–106. https://doi.org/10.1016/j.geoderma.2010.12. 014. Keeling, R.F., Graven, H.D., Welp, L.R., Resplandy, L., Bi, J., Piper, S.C., Sun, Y., Bollenbacher, A., Meijer, H.A.J., 2017. Atmospheric evidence for a global secular increase in carbon isotopic discrimination of land photosynthesis. P. Natl. Acad. Sci. USA 114, 10361–10366. https://doi.org/10.1073/pnas.1619240114. Kolbe, S.E., Miller, A.I., Townsend-Small, A., Cameron, G.N., Culley, T.M., 2016. Impact of land-use history and forest trees on soil organic carbon and nitrogen stocks. Soil Sci. Soc. Am. J. 80, 1089–1097. https://doi.org/10.2136/sssaj2015.12.0426. Laskar, A.H., Yadava, M.G., Ramesh, R., 2016. Stable and radioactive carbon in forest soils of Chhattisgarh, Central India: Implications for tropical soil carbon dynamics and stable carbon isotope evolution. J. Asian Earth Sci. 123, 47–57. https://doi.org/ 10.1016/j.jseaes.2016.03.021. Lu, X.K., Mao, Q.G., Gilliam, F.S., Luo, Y.Q., Mo, J.M., 2014. Nitrogen deposition contributes to soil acidification in tropical ecosystems. Global Change Biol. 20, 3790–3801. https://doi.org/10.1111/gcb.12665. Marin-Spiotta, E., Sharma, S., 2013. Carbon storage in successional and plantation forest soils: a tropical analysis. Global Ecol. Biogeogr. 22, 105–117. https://doi.org/10. 1111/j.1466-8238.2012.00788.x. Martin, P.A., Newton, A.C., Bullock, J.M., 2013. Carbon pools recover more quickly than plant biodiversity in tropical secondary forests. P. Roy. Soc. B-Biol. Sci. 280, 20132236. https://doi.org/10.1098/rspb.2013.2236.

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