Alterations of ecosystem nitrogen status following agricultural land abandonment in the Karst Critical Zone Observatory (KCZO), Southwest China

Study area

The study area is located in the Chenqi catchment (26°15.779′–26°16.710′N, 105°46.053 ′–105°46.839′E), with an area of 1.54 km². The Chenqi catchment is one of the key research areas of the KCZO in Guizhou province, Southwest China. The climate in the catchment is mainly sub-tropical monsoonal, with a mean annual temperature (MAT) of 15.1 °C and mean annual precipitation (MAP) of 1,315 mm (Zhao et al., 2010). Seasonal evapotranspiration (mean: 260 mm in spring, 330 mm in summer, 185 mm in autumn, and 115 mm in winter) is much lower than seasonal precipitation (Gao et al., 2016). This catchment has a typical karst hoodoo depression landform, in which a valley is surrounded by three hills and a seasonal stream flows from east to west (Liu, Han & Zhang, 2020). The altitudes of these hills reach up to 1,524 m (above sea level) at their maximum, while the valley is only 1,310 m, with an average altitude of 1,350 m in this catchment (Yue et al., 2020). The soils on the hilltops and hillslopes are calcareous, mainly developed from the limestones of the upper and middle part of the Guanling Formation of the middle Triassic (Zhao et al., 2010), and classified as calcic Inceptisols in the soil taxonomy of the United States Department of Agriculture (USDA) (Soil Survey Staff, 2014). Quaternary deposits are mainly distributed on the valley floor, mainly deposited from the materials of surrounding hillslopes by soil erosion (Green, Dungait et al., 2019).

To restore the karst ecological environment, many sloping croplands with a low crop yield have been abandoned and naturally recovered under the GGP program since the 1990s (Wang et al., 2017). In the KCZO, a series of croplands were abandoned in different years, and gradually evolved into grassland, shrubland, and secondary forest lands in that order (Liu, Han & Zhang, 2020). The croplands (CL) on the valley floor remained in conventional cultivation or are in the fallow period, the shrub-grass lands (SG) at the foot of the hills have been transformed from terraced croplands for 3–8 years, and the secondary forest lands (SF) on the hillsides were converted from terraced croplands 50 years ago (Liu, Han & Zhang, 2020). Thus, the zone of croplands, shrub-grass lands, and secondary forest lands show a vertical distribution in the catchment. In the croplands, main crops such as maize (Zea mays), potato (Solanum tuberosum), oilseed rape (Brassica napus), and peanut (Arachis hypogaea) are planted in rotation from spring to autumn, while the fields lie fallow in winter (Qu & Han, 2022). N-P-K fertilizer and urea provide about 300 kg/ha N, 85 kg/ha P, and 6 kg/ha K per year for crop production, and farm manures are applied irregularly and non-quantitatively (Li et al., 2018). In the shrub-grass lands, the main plant species are herbaceous plants including Imperata cylindrical, Setaria viridis, and Miscanthus sinensis, and low shrubs or trees including Berchemia sinica, Ilex macrocarpa, and Pyracantha fortuneana. In the secondary forest lands, the main plant species are tall evergreen trees including Litsea pungens, Padus racemosa, Pinus tabuliformis, Cinnamomum camphora, Camellia japonica, and Cyclocarya paliurus. Above-ground biomass in the croplands, shrub-grass lands, and secondary forest lands are 2,020, 5,140, and 8,160 kg/ha, respectively; their below-ground biomass is 1,330, 2,740, and 1,570 kg/ha, respectively (Wang et al., 2021). The parent rock, soil type, and climate of different sites were similar in the small catchment. The average slope of the hills in the Chenqi catchment is more than 40° (Liu, Han & Zhang, 2020), which determines that the farmlands are designed as terraced croplands in order to avoid soil erosion. The topographic features of the different sites were similar, although there is a vertical distribution. In this study, the croplands, shrub-grass lands, and secondary forest lands are regarded as the different stages following secondary succession on abandoned croplands by the space-for-time substitution method (Blois et al., 2013). The photographs of the three stages following secondary succession are shown in Fig. 1.

Sampling

Soil sampling was carried out in the Chenqi catchment in June 2016. At present, the proportional area of croplands, shrub-grass lands, and secondary forest lands within the Chenqi catchment is approximately 50%, 20%, and 30%, respectively. In total, 18 sampling sites from different land-use types were randomly selected related to the proportional area of each within the catchment, of which 8 sites were in the croplands, five sites were in the shrub-grass lands, and five sites were in the secondary forest lands. The location and land-use history of all sampling sites are shown in Table 1. The distance between two different sites under the same land-use type was 50 m to 100 m. A soil pit (0.5 m × 0.5 m × 0.5 m) was dug at each sample site, and three duplicate soil samples were chosen from the three sides of the pit. Soil samples were collected from the top to the bottom at 0–10, 10–20, and 20–30 cm depth. In total, 162 soil samples were collected. The duplicate samples in each pit at the same depth were mixed into one composite sample.

The dominant vegetation species under the three land use types were identified in the field. Leaf samples from at least five plants with the same species were selected under the same land-use type. The mature leaves of the dominant vegetation species were collected at the high, middle, and low tree heights. The leaves from the same species were mixed to be one sample. In total, 37 leaf samples were collected, including nine in croplands, 13 in shrub-grass lands, and 15 in secondary forest lands.

Sample analysis

After washing off the dust on the leaf surface with pure water 3 times, the leaf samples were dried at −40 °C in a freezer dryer, then ground into powder by an attritor. Soil samples were air-dried (25 °C) after removing obvious gravel and fresh coarse roots. A part of the samples was crushed by hand to make all particles pass through a 10 mesh-steel sifter (2 mm), which was stored as the sample of bulk soil (<2 mm). The remaining soil samples were not crushed and were used for soil aggregate separation by using the improved wet-sieving method (Six et al., 2002). Macro-aggregates (250–2000 µm), micro-aggregates (53–250 µm), and silt + clay sized fractions (<53 µm) were collected after passing through 250 µm and 53 µm sifters, respectively. The moist aggregate samples were dried in an oven at 55 °C until constant weight and then weighed to calculate their mass percent.

The samples of bulk soils and different-sized aggregates were ground in an agate mortar until all fine particles pass through a 200 mesh-nylon sifter. The powder samples (<75 µm) were soaked in a 2 mol/L KCl solution for 24 h to remove NO₃⁻, NH₄⁺, and dissolved organic N (Meng, Ding & Cai, 2005). The treated samples were washed repeatedly with pure water until all Cl⁻ was removed, then dried and ground into powder for analyses of N content and δ¹⁵N composition.

The foliar N contents and SON contents of bulk soils and aggregates were measured by a multi-elemental analyzer (Vario TOC cube; Elementar, Hessen, Germany) in the Surficial Environment and Hydrological Geochemistry Laboratory, China University of Geosciences (Beijing, China). Standard soil substances (OAS B2152) were repeatedly measured to monitor reproducibility. The relative standard deviations were less than 3%. Actual SON content could be calibrated because of sample mass reduction after removing dissolved inorganic and organic N, i.e., the measured SON content is multiplied by the ratio of the sample mass after and before treatment (Liu, Han & Li, 2021).

The N stable isotope ratio (¹⁵N/¹⁴N) of leaf, SON in bulk soils and aggregates were determined by an isotope mass spectrometer (MAT-253, Thermo, USA) in the Central Laboratory for Physical and Chemical Analysis, Institute of Geographic Sciences and Natural Resources Research, Chinese Academy of Sciences. The measurements are expressed in standard δ notation (‰) to indicate the differences between the ¹⁵N/¹⁴N ratio of the samples and accepted standard materials (atmospheric N₂), where (Han et al., 2020):

(1)${\displaystyle {\delta }^{15}{N}_{\mathrm{sample}}\left(‰\right)=\left(\left(R_{\mathrm{sample}}-R_{\mathrm{air}}\right)/R_{\mathrm{air}}\right)\times 1000,R{=}^{15}N{/}^{14}N.}$

Reference material (GBW04494, δ¹⁵N_Air: 0.24‰ ± 0.13‰) was monitored and repeatedly measured to evaluate the precision of the measurements (<0.1‰).

Statistical analysis

One-way ANOVA with the least significant difference (LSD) test was conducted to identify the significant differences in SON contents of aggregates, δ¹⁵N values of leaves and SON in bulk soils and aggregates, and the EF values among the three land-use types at the significance level of P < 0.05. Linear and non-linear regression analyses were used to determine the variations of foliar N contents and C/N ratios following secondary succession on abandoned croplands. Statistical analyses were performed using the SPSS 18.0 (SPSS Inc., Chicago, IL, USA) software program and the figures were drawn using SigmaPlot 12.5 (Systat Software GmbH, Erkrath, Germany) software program.

Foliar N contents and C/N ratios

Foliar N contents in the croplands were higher than those in the shrub-grass lands and much higher than those in the secondary forest lands (Fig. 2). Foliar C/N ratios in the croplands were lower than those in the shrub-grass lands and much lower than those in the secondary forest lands (Fig. 2). Foliar N contents decreased while C/N ratios increased following secondary succession on abandoned croplands.

SON contents of bulk soils and aggregates

The SON contents of bulk soils and macro-aggregates at the 0–20 cm depth in the croplands were slightly lower than those in the shrub-grass lands, and significantly lower than those in the secondary forest lands (Fig. 3). While the SON contents of macro-aggregates and silt + clay sized fractions at the 0–20 cm depth were not significantly different among the three land-use types (Fig. 3). Moreover, in the soils at the 0–20 cm depth, the SON contents of bulk soils, macro-aggregates, micro-aggregates, and silt + clay sized fractions were not significantly different among the three land-use types (Fig. 3). SON contents of bulk soils and macro-aggregates in the surface soils (0–20 cm depth) significantly increased following secondary succession on abandoned croplands.

Foliar δ¹⁵N values and soil to plant ¹⁵N EF values

Foliar δ¹⁵N values in the croplands (mean 3.37‰) were significantly higher than those in the shrub-grass lands (mean −2.43‰), and much higher than those in the secondary forest lands (mean −7.06‰) (Fig. 4). Similarly, the soil to plant ¹⁵N EF values in the croplands (mean −3.19‰) were significantly higher than those in the shrub-grass lands (mean −8.48‰), and much higher than those in the secondary forest lands (mean –10.40‰) (Fig. 4). Foliar δ¹⁵N values and the EF values significantly decreased following secondary succession on abandoned croplands.

δ¹⁵N values of SON in bulk soils and aggregates

The δ¹⁵N values of SON in bulk soils at the 0 −10 cm depth under the croplands (mean 6.56‰) were significantly higher than those under the shrub-grass lands (mean 4.99‰) and secondary forest lands (mean 5.16‰), while the δ¹⁵N values in the soil of the 10–30 cm depth were not significantly different among the three land-use types (Fig. 5). The SON of top soils (0–10 cm depth) gradually enriched ¹⁴N following secondary succession on abandoned croplands. Additionally, the δ¹⁵N values of SON in different sized aggregate at all soil depths and under all land-use types varied following the order: micro-aggregates <macro-aggregates <silt + clay sized fractions (Fig. 5). Although these differences of δ¹⁵N values between different sized aggregates were not always significant statistically, the same feature of aggregate δ¹⁵N-discrepancies in all soil samples.

Improved soil N availability following secondary succession

Soil N availability reflects the bioavailable N supply capacity of the soil, which is mainly affected by SON contents and the reaction rate of N mineralization (Tang et al., 2017). SON is the main substrate derived from plant N and biological N fixation (Vitousek et al., 2013). In the croplands, extraneous N fertilizer is absorbed by crops to enhance its N content, thus foliar N contents were higher than those in the shrub-grass lands and secondary forest lands (Fig. 2). Although foliar N contents decreased after agricultural land abandonment, increased plant biomass enhance N input by litters and root exudates (Wang et al., 2021). Besides N inputs from plants, biological N fixation also make a considerable contribution to soil N inputs (Vitousek et al., 2013). Li et al. (2018) found the absolute abundances of N functional genes for N₂ fixation gradually increased following secondary succession in the study area, indicating enhanced soil N stock by biological N fixation. Thus, increases in N input from plant and biological N fixation promote SON accumulation in the surface soils after agricultural land abandonment (Fig. 3).

Soil NH₄⁺-N contents at the 0–30 cm depth significantly increased following secondary succession (Table S2), which suggests an accelerated N mineralization rate. The observed soil NH₄⁺-N content is not the gross yield of N mineralization, which is the remainder after deducting absorption by plant and soil microbes, as well as NH₄⁺ leaching loss (despite it being negligible) (Galloway et al., 2008; Robinson, 2001). However, absorbed NH₄⁺ by plant and soil microbes also increase due to increases in plant biomass and microbial biomass (Guo et al., 2021; Wang et al., 2021; Zhang et al., 2000). Thus, it is likely that the soil N mineralization rate is rapid after agricultural land abandonment. This is also supported by the changes in δ¹⁵N composition of SON and foliar N following secondary succession. Mean foliar δ¹⁵N values in the croplands, shrub-grass lands, and secondary forest lands were 3.37‰, −2.43‰, and −7.06‰, respectively (Fig. 4), while the δ¹⁵N values of SON in top soils (0–10 cm depth) under the three land-use types was 6.56‰, 4.99‰, and 5.16‰, respectively (Fig. 5). The δ¹⁵N composition of SON in top soils is influenced by the natural ¹⁵N-abundance of the source plants (Boeckx et al., 2005), and is also affected by δ¹⁵N fractionation during N mineralization (Baggs et al., 2003; Corre et al., 2007). Soil N mineralization leads to ¹⁵N enrichment in residual SON and produces ¹⁵N-depleted NH₄⁺ (Baggs et al., 2003; Corre et al., 2007). SON in top soils enriches more ¹⁵N compared to litter (or foliar) N following secondary succession, indicating an intensive N mineralization process. In summary, increases in N input from plants, biological N fixation, and soil N mineralization rate determine soil N availability during secondary succession.

Reduced soil NO₃⁻ loss following secondary succession

Foliar δ¹⁵N values and soil to plant ¹⁵N EF values significantly decreased following secondary succession (Fig. 4), indicating reduced soil N losses after agricultural land abandonment (Pardo et al., 2007). Wang et al. (2021) found denitrification was unlikely to be a major contribution to soil NO₃⁻ loss, based on the weak correlation between δ¹⁵N and δ¹⁸O of NO₃⁻ in this study area. Thus, soil NO₃⁻ leaching is the main cause of losses in this catchment. Furthermore, the (NO₃⁻-N)/(NH₄⁺-N) ratio gradually decreased following secondary succession in the study area (Table S1), suggesting a reduced risk of soil NO₃⁻ leaching (Wang et al., 2021). Soil NO₃⁻ loss potential decreased following secondary succession, mainly attributed to increased SON stabilization and soil inorganic N uptake capacity.

Foliar N contents decreased, while C/N ratios increased following secondary succession (Fig. 2), which could cause similar variations in the N contents and C/N ratios of litter. Generally, the decomposition of litter with a low N concentration and high C/N ratio will be restrained at the initial stage of litter decomposition (Xia et al., 2021). Thus, the fresh SON derived from litter in the shrub-grass lands and secondary forest lands is more stable than that in the croplands. SON contents of bulk soils and macro-aggregates in the surface soils significantly increased following secondary succession, but SON contents of micro-aggregates and silt + clay sized fractions did not significantly vary (Fig. 3). Moreover, macro-aggregate proportions and the contribution of macro-aggregates to SON stock in the surface soils also significantly increased following secondary succession (Table S1). These results indicate that added SON after agricultural land abandonment is mainly stored within macro-aggregates. Physical protection of macro-aggregates by isolating oxygen, water, microbes, and enzymes also enhances SON stabilization (Six et al., 2002; Wright & Hons, 2005; Zhu, Deng & Shangguan, 2018). In addition, the δ¹⁵N values of SON in micro-aggregates were always lower than that in macro-aggregates (Fig. 5), which was closely linked to the different degrees of N mineralization. Due to ¹⁵N enrichment in residual SON during the N mineralization process (Baggs et al., 2003; Corre et al., 2007), it can be inferred that the SON in micro-aggregates is fresher than that in macro-aggregates, which is consistent with the findings based on δ¹³C values of SOC (Liu, Han & Zhang, 2020). According to the aggregate hierarchy model (Six, Elliott & Paustian, 2000), micro-aggregates are formed within macro-aggregates, resulting in more fresh-SON combined within micro-aggregates. The SON within micro-aggregates is more stable because of the multiple protections of micro-aggregates and macro-aggregates (Beare, Hendrix & Coleman, 1994). The stability of SON is improved after agricultural land abandonment due to the increased C/N ratio of litter and aggregate protection. During secondary succession, the total N mineralization rate is improved, as reflected by the increased NH₄⁺ contents (Table S2). The ratio of mineralized SON in total SON is likely reduced due to increased SON stability, which may restrict N mineralization and subsequent nitrification to produce NO₃⁻ in unit time and unit SON. Thus, improved SON stability is meaningful to reduce soil NO₃⁻ loss potential following secondary succession.

Soil NO₃⁻ production and consumption (mainly including uptake by plants and microbes, leaching loss, and residual NO₃⁻ in soils) at all stages of secondary succession must be balanced (Dai et al., 2020; Soper et al., 2018). Nitrification causes ¹⁵N enrichment in residual NH₄⁺ and produces ¹⁵N-depleted NO₃⁻ (Lim et al., 2015). The δ¹⁵N values of NH₄⁺ increased and the δ¹⁵N values of NO₃⁻ decreased following secondary succession (Table S2), indicating an intensive nitrification process and increased NO₃⁻ production after agricultural land abandonment (Wang et al., 2021). Soil NO₃⁻ leaching loss and residual soil NO₃⁻ (Table S2) gradually decreased following secondary succession. Thus, it is likely that soil NO₃⁻ uptake by plants and microbes is enhanced after agricultural land abandonment, which is mainly attributed to increases in plant biomass and microbial biomass. N is a bioelement; the growth and reproduction processes of plant and soil microbes need to assimilate bioavailable N (including NO₃⁻-N) from soils (Zhang et al., 2000). In the study area, plant biomass and soil microbial necromass increased following secondary succession (Guo et al., 2021; Wang et al., 2021), increasing demands for bioavailable N.

In the present study, the diagram exhibits the changes in plant and soil N content and δ¹⁵N values during secondary succession after agricultural land abandonment under GGP as shown in Fig. 6, which reflects improved soil N availability and reduced soil NO₃⁻ loss following secondary succession. Our study highlights the positive implications of the GGP program for karst ecosystem restoration. As the GGP program in China progresses, this conceptual model can provide basic scientific guidance for policymakers in the karst region, and even in the non-karst region. However, considering the differences in climate, soil type, and vegetation restoration approach in different regions, the key factor that affected soil N transformation processes and N loss may differ (Osborne et al., 2017). For future research, the conceptual model should be improved to make it applicable in other environments, particularly non-karst.